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An evaluation of leeches as in situ biomonitors of chlorinated phenolic compounds discharged from bleached… Prahacs, Steven Michael 1994

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A N EVALUATION OF LEECHES AS IN SITU BIOMONTTORS OF CHLORINATED PHENOLIC COMPOUNDS DISCHARGED FROM BLEACHED KRAFT PULP MILLS by STEVEN MICHAEL PRAHACS B.Sc, Concordia University, 1986 A THESIS SUBMITTED IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF MASTER' OF SCIENCE in THE FACULTY OF GRADUATE STUDIES RESOURCE MANAGEMENT AND ENVIRONMENTAL STUDIES PROGRAMME W e Accep t this thesis as conforming to the required standard The University of British Columbia November 1994 © Steven Michael Prahacs, 1994 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 ^ € S a u R O £ t>\ftsAGgMgAjT. . ~ - w The University of British Columbia Vancouver, Canada D a t e f 3 ^ Z DE-6 (2/88) 11 ABSTRACT Leeches were evaluated as b iomoni to r s o f ch lo r ina ted pheno l i c compounds d ischarged f rom bleached kraft pu lp m i l l s u s ing integrated labora tory and f i e l d invest igat ions . Semi-stat ic laboratory bioassays o f one week durat ion were carr ied out i n order to de te rmine h o w e n v i r o n m e n t a l factors such as con t aminan t concentration, water p H , water temperature and suspended sediments as w e l l as b iot ic factors such as leech weight and species affect b ioconcen t ra t ion . In order to evaluate leeches under v a r y i n g in situ cond i t i ons , three f i e l d m o n i t o r i n g t r ia l s dur ing summer (July) , f a l l (October) and winter (February) were conducted on the Fraser R ive r , downstream of three bleached kraft pulp mi l l s at Prince George B . C . There was a s t rong l i nea r co r r e l a t i on ( r 2 = 0.89 - 0.96) between water contaminant concent ra t ion (0.1 - 10 u g / L ) and b i o c o n c e n t r a t i o n o f c h l o r i n a t e d gua i aco l s , w i t h s low depura t ion rates > 28 days) . B i o c o n c e n t r a t i o n o f chlorinated guaiacols was inverse ly related to p H (5.1 - 9.0), but was on ly weak ly correlated wi th the concentrat ion o f the undissociated compound, ind ica t ing that the i o n i z e d compound also contr ibuted to the b ioava i l ab le f ract ion. B ioconcen t r a t i on increased between 4.4 and 11.8 ° C , but d id not show a significant change from 11.8 to 20.0 ° C , i n d i c a t i n g a b i -phas i c mode l for the temperature - b ioconcen t r a t i on re la t ionship . N o clear re la t ionship between suspended sediment concentrat ion (0 -0.15 g / L ) and bioconcentra t ion was observed, al though the presence o f 5% organic material i n suspended sediments reduced b ioava i l ab i l i t y o f the chlor inated phenol ics . Wate r s t i r r i ng i n the suspended sediment b ioassays inc reased b i o c o n c e n t r a t i o n relat ive to quiescent semi-stat ic condi t ions . There was strong inverse re la t ionship ( r 2 = 0.90 - 0.94) between leech weight and bioconcentra t ion wi th no clear trend between two different species o f leech (Nephelopsis obscura and Percymoorensis marmorata). Ill Field monitoring revealed chlorinated phenolics in pulp mill effluent (0.6 -17 ug/L). and water (0.002 - 0.073 pg/L) and suspended sediments (0.36 - 176 u g / k g ) during the study periods. Leeches were effective biomonitors of tri- and tetrachlorinated guaiacols under diverse seasonal conditions, with bioconcentration factors ranging from 465 - 6000 and were accurate indicators of the relative proportions of these chlorinated contaminants in both pulp mill effluent and the Fraser River, 40 km downstream of pulp mill outfalls. iv T A B L E OF C O N T E N T S P a g e Abstract 1 1 Table of Contents i v List of Figures • • v i i i List of Tables x * List of Abbreviations • x v Acknowledgments x y i i Dedication , xv i i i 1. General Introduction 1 2. Literature Review 5 2.1. Bleached Kraft Mill Technology 5 2.1.1. Introduction 5 2.1.2. Kraft Pulping, and Washing 5 2.1.3. Pulp Bleaching 8 2.1.4. Effluent Treatment 12 2.2. Environmental Fate of Chlorinated Phenolics 15 2.2.1. Introduction 15 2.2.2. Sources.: 16 V Page 2.2.3. Partitioning to the Water Phase 17 2.2.4. Partitioning to the Sediment Phase .' 18 2.2.5. Partitioning to Biota 20 2.2.5.1. The Effect of Octanol-Water Partition Coefficient 22 2.2.5.2. The Effect of pH 23 2.2.5.3. The Effect of Dissolved and Particulate Solids... 24 2.2.5.4. The Effect of Temperature 26 2.2.5.5. The Effect of Multi-Contaminant Exposure 27 2.2.6. Biotransformation 28 2.2.7. Fate of Chlorinated Phenolics in the Fraser River 31 2.3. Biomonitoring 35 2.4. Leech Biology 41 3. Methodology 43 3.1. Experimental Organisms 43 3.2. Laboratory Studies 44 3.3. Field Studies 47 3.3.1. Study Area 47 3.3.2. Study Periods 49 3.3.3. Sampling Procedures 50 3.4. Analytical procedures 52 3.4.1. Chemicals and Reagents 52 3.4.2. Extraction Procedures 53 3.4.2.1. General Rational 53 3.4.2.2. Effluent Sample Preparation 54 3.4.2.3. Water Sample Preparation 55 3.4.2.4. Sediment Sample Preparation... 56 v i Page 3.4.2.5. Leech Sample Preparation 57 3.4.2.6. Reference Standard Preparation 58 3.4.3. Instrumental analysis 58 3.4.4. Quality Control 59 4. Results and Discussion 65 4.1. Laboratory Studies 65 4.1.1. Bioconcentration as a Function of Contaminant Concentration 65 4.1.2. Bioconcentration as a Function of Leech Weight 76 4.1.3. Bioconcentration as a Function of Water pH 83 4.1.4. Bioconcentration as a Function of Water Temperature. 94 4.1.5. Bioconcentration as a Function of Suspended Sediment Load 102 4.2 Field Studies 113 4.2.1. Temporal Variations in Contaminant Concentrations in Effluent, Water, Suspended Sediments and Leeches 113 4.2.2. In Situ Bioconcentration as a Function of Contaminant Concentration 125 4.2.3. In Situ Bioconcentration as a Function of Leech Weight. 132 4.2.4. Interspecies Differences in Bioconcentration 136 4.2.5. Estimation of In Situ Water Contaminant Concentrations Using Laboratory Bioconcentration Relationships 139 5. Summary, Recommendations and Applications 142 6. References 151 v i i Page 7. Appendices 167 Appendix 1: Leech Identification 168 Appendix 2: Calculations of Reported Chlorinated Phenolics Concentrations 170 Appendix 3: Raw Data - Laboratory Bioassays 171 Appendix 4: Raw Data - Field Monitoring Trials 181 Appendix 5: Laboratory Predictions of In Situ Water Contaminant Concentrations - Sample Calculation 192 V l l l LIST OF FIGURES Page 1.1: Chemical structures and abbreviated names of compounds tested 3 3.1: Fraser River biomonitoring study area 48 3.2: Separation of chlorinated phenolic compounds on a 30 m DB-5 capillary column 60 3.3: Separation of chlorinated phenolic compounds on a 30 m DB-1701 capillary column 61 4.1: The effect of chlorinated phenolic exposure concentration on leech (N. obscura) bioconcentration of 3,4,5-TCG and 3,4,5-TCVer 68 4.2: The elimination of chlorinated phenolics from leeches (N. obscura) over a four week period 73 4.3: Bioconcentration of 3,4,5-TCG and 3,4,5-TCVer by leeches (N. obscura) of differing weights 79 4.4: The effect of water pH on the bioconcentration of 3,4,5-TCG and 3,4,5-TCVer by leeches (N. obscura) 85 4.5: Dissociation and leech (TV. obscura) bioconcentration of chlorinated guaiacols plotted against water pH 87 ix Page 4.6: Dissociation and leech (TV. obscura) bioconcentration of chlorinated phenols plotted against water pH 88 4.7: Relationship between leech. (TV. obscura) bioconcentration and log Kow of chlorinated phenolics at pH 5.1 92 4.8: Alternative interpretations of the effect of water temperature on the bioconcentration of 3,4,5-TCG and 3,4,5-TCVer by leeches (TV. obscura) 97 4.9: The effect of suspended sediment concentration on leech (TV. obscura) bioconcentration of 3,4,5-TCG and 3,4,5-TCVer 105 4.10: The effect of the presence of organic material in suspended sediments on the bioconcentration of chlorinated phenolics 110 4.11: Relationship between chloroguaiacol pKa and the ratio of leech (TV. obscura) bioconcentration in turbid bioassay water of 0% organic particulate content versus 5% organic particulate content Il l 4.12: Relative concentrations of chlorinated phenolics in leeches and water samples from Stoner, B.C. and effluents from two Prince George BKM outfalls, over three seasonal monitoring periods 124 4.13: Leech (TV. obscura) bioconcentration factor as a function of water concentration for 4,5,6-TCG detected in field studies for Jul., Oct., 1991 and Feb. 1992 and laboratory studies 126 X Page 4.14: Leech (N. obscura) bioconcentration as a function of water concentration for 4,5,6-TCG detected in field studies for Jul., Oct.,1991 and Feb. 1992 and laboratory studies 126 4.15: Leech (N. obscura) bioconcentration as a function of water concentration for 3,4,5-TCG detected in field studies for Jul., Oct., 1991 and Feb. 1992 and laboratory studies 127 4.16: Bioconcentration of 3,4,5-TCG (a) and TeCG (b) by leeches (N. obscura) of differing weights, exposed under both laboratory and field conditions : 134 x i LIST OF TABLES P a g e 2.1: Major classes and sources of chlorinated phenolic compounds found i n softwood B K M effluent 10 2.2: The effect o f percent chlorine d ioxide substitution and chemical charging order on the formation of chlorinated phenolics 13 2.3: Compar i son o f effluent quali ty parameters for various sequences used for bleaching softwood pulp 13 2.4: The effect o f b io logica l treatment on effluent quality parameters 14 2.5: Dissoc ia t ion constants (pKa) and percent dissociat ion o f chlorinated phenolics i n water of p H 7.8. 17 2.6: Bioconcent ra t ion o f chlorophenols by various aquatic organisms sampled from Canagagigue Creek, Ontario, Canada 39 3.1: Me thod detection l i m i t ranges for the analysis o f chlorinated phenolics i n effluent, water, suspended sediments and leeches 64 4 .1 : The bioconcentrat ion of chlorinated phenolics by leeches (N. obscura) after seven days exposure to concentrations ranging from 0 . 1 - 1 0 ug /L 67 Xl l Page 4.2: The average proportion of 4,5,6-TCG found in leeches after seven days of semi-static exposure at various concentrations 69 4.3: Bioconcentration factors of chlorinated phenolics, after seven days exposure to concentrations ranging from 0.1 - 10 ug/L 71 4.4: Bioconcentration of chlorinated phenolics by leeches (N. obscura) of different weights 77 4.5: The effect of water pH on bioconcentration of chlorinated phenolics by leeches (N. obscura) 84 4.6: The effect of water temperature on bioconcentration of chlorinated phenolics by leeches (N. obscura). 95 4.7: The effect of suspended solids concentration bioconcentration of chlorinated phenolics by leeches (N. obscura). 104 4.8: Field conditions for July 8-15, October 17-24, 1991 and February 19-26, 1992 at two monitoring stations on the Fraser River 115 4.9: Pulp mill process conditions for July 8-15, October 17-24, 1991 and February 19-26, 1992 at Prince George, B.C 115 4.10: Flow weighted average effluent concentrations of chlorinated phenolics discharged from Prince George bleached kraft mill outfalls. .. 116 Xl l l P a g e 4.11: Seven day mean water concentrations o f chlorinated phenolics detected i n the Fraser R ive r at Stoner, B . C 117 4.12: Suspended sediment concentrations o f chlorinated phenol ics and percent organic content in the Fraser R ive r 118 4.13: Predic ted sediment-water part i t ion coefficients (Ksw) for 3 ,4 ,5 -TCG and T e C G in the Fraser River at Stoner, B . C . i n July 1991 and February 1992 120 4.14: M e a n tissue concentrations (|ag/kg) o f chlorinated phenol ics i n leeches exposed in the Fraser R ive r at Prince George, B . C 122 4.15: Cross channel differences i n tissue concentrations o f chlor inated phenolics i n leeches (TV. obscura) exposed for seven day periods i n the Fraser River at Stoner, B . C 129 4.16: Tissue concentration o f chlorinated phenolics i n leeches (TV. obscura) of different weights, after seven days exposure i n the Fraser R ive r 133 4.17: Predicted decreases i n bioconcentrat ion o f chlor inated phenol ics by leeches (TV. obscura) between 0.5 g and 1.0 g under both laboratory and field conditions 135 4.18: Bioconcentrat ions o f chlorinated phenolics i n two leech species, Nephelopsis obscuraand Percy mo or en sis marmorata exposed i n the Fraser River at Stoner, B . C 138 xiv Page 4.19: Bioconcentrat ions o f chlorinated phenolics i n two leech species, Nephelopsis obscuraand Percymoorensis marmoratq after seven days exposure to a water concentration of 1.0 ug/L 138 4.20: Compar ison of measured seven day f ie ld water concentrations o f 3 , 4 , 5 - T C G and T e C G to those predicted from laboratory leech bioconcentrat ion relationships 140 5.1: Summary o f effects of environmental and biot ic factors on leech (N. obscura) bioconcentration of 3 ,4 ,5 -TCG and 3 ,4 ,5 -TCVer 147 X V LIST OF ABBREVIATIONS A O X a d s o r b a b l e o r g a n i c h a l i d e B C F b i o c o n c e n t r a t i o n f a c t o r B K M b l e a c h e d k r a f t m i l l B K M E b l e a c h e d k r a f t m i l l e f f l u e n t B C D b i o c h e m i c a l o x y g e n d e m a n d B S W b r o w n s t o c k w a s h i n g C c h l o r i n e b l e a c h i n g s t a g e D c h l o r i n e d i o x i d e b l e a c h i n g s t a g e D t C c h l o r i n e / c h l o r i n e d i o x i d e b l e a c h i n g s t a g e E a l k a l i n e e x t r a c t i o n b l e a c h i n g s t a g e E 0 o x y g e n r e i n f o r c e d a l k a l i n e e x t r a c t i o n E C O e x t r a c t a b l e o r g a n i c c h l o r i n e f o c f r a c t i o n o f o r g a n i c c a r b o n C C g a s c h r o m a t o g r a p h y G P C g e l p e r m e a t i o n c h r o m a t o g r a p h y H h y p o c h l o r i t e b l e a c h i n g s t a g e K D W o c t a n o l - w a t e r p a r t i t i o n c o e f f i c i e n t K s w s e d i m e n t - w a t e r p a r t i t i o n c o e f f i c i e n t M C C m o d i f i e d c o n t i n u o u s c o o k i n g x v i M D L m e t h o d d e t e c t i o n l i m i t N D n o n e d e t e c t e d P O X p u r g a b l e o r g a n i c h a l i d e S S s u s p e n d e d s o l i d s T O Q t o t a l o r g a n i c c h l o r i n e 1° p r i m a r y 2° s e c o n d a r y XVII ACKNOWLEDGEMENTS This research was made poss ible through the co l labora t ive efforts o f people from academic and government institutions as we l l as the pulp and paper industry. P . P a r k i n s o n , S. Harper and T. M a , E n v i r o n m e n t a l E n g i n e e r i n g Labora to ry , Dept . o f C i v i l Eng inee r ing , U n i v e r s i t y o f B r i t i s h C o l u m b i a , are grateful ly acknowledged for their analyt ical , technical support. Dr . L . L a v k u l i c h , Dept. o f S o i l Science, Un ive r s i ty of B r i t i s h C o l u m b i a , k i n d l y performed the X - r a y d i f f rac t ion analysis o f sediment samples. Dr . R . W . Davies , Dept. o f B i o l o g y , Univers i ty o f Ca lgary , provided expert leech species ident i f icat ion. W . Duncan and D . Sutherland, B r i t i s h C o l u m b i a M i n i s t r y o f E n v i r o n m e n t , L a n d s and Pa rks , P r i n c e George branch , con t r ibu ted exce l l en t l o g i s t i c a l , technica l and advisory aid for Fraser R i v e r f i e ld studies. G . Derksen , E n v i r o n m e n t Canada , E n v i r o n m e n t a l Pro tec t ion , Nor th V a n c o u v e r k i n d l y shared his knowledge and data concerning Fraser R i v e r po l lu t ion chemistry. R . Parne l l , Canfor L t d . , Pr ince George Pulp D i v i s i o n and J. N y l u n d , Nor thwood Pulp and Paper Company , P r i n c e George , a long w i t h m i l l t echn ica l staff were ins t rumenta l i n p r o v i d i n g effluent samples du r ing f i e l d studies per iods . S. Prahacs Sr . , S .P . B i o s p h e r e Internat ional Inc, generously lent his t ime and expertise for consul ta t ions i n the area o f pulp and paper po l lu t ion abatement technology. Thesis committee members, Dr . G . Be l lward , Dr . T. Northcote, Dr . H . Schreier and Dr . L . L a v k u l i c h provided useful insight and commentary on this research. Special thanks is extended to Dr . K . J. H a l l , Dept. o f C i v i l Engineer ing, Univers i ty o f B r i t i s h C o l u m b i a , for guidance i n a l l aspects of this research and science i n general as w e l l as endless patience i n awai t ing the comple t ion o f this thesis. Th i s research was funded by a Na t iona l Sciences and Engineer ing Research C o u n c i l o f Canada operating grant to Dr . K . J . H a l l . XV111 DEDICATION T h i s t h e s i s i s d e d i c a t e d w i t h ' a l l m y h e a r t a n d s o u l t o B r i g i t a G r a z y s , t h e b e s t f r i e n d a n y o n e c o u l d h a v e . 1 1 . INTRODUCTION On a broad scale, resource management can be defined as the development and execution of plans for the integrated, sustainable use of natural resources; a balance between human use and conservation. Comprehensive management should address and weigh social, economic and biophysical impacts of resource use, although the nature and relevance of each aspect is unique to a given situation. Management decision making requires the integration of a vast array of needs with often complex and uncertain information. In the case of water resources, both water quantity and quality issues must be examined. Areas of concern are diverse and may include: water demand in water-deficient regions, municipal sewage treatment infrastructure, protection of drinking water quality and maintenance of aquatic habitat. The latter two concerns are closely linked with management of toxic contaminants discharged by point and non-point anthropogenic pollution sources. Water pollutant management requires a comprehensive monitoring program to track the spatial and temporal distribution of contaminants. Monitoring can identify the pollutants of concern as well as determine sensitive areas and time periods when potential environmental impacts may be manifested. Aquatic monitoring data serve to aid in the assessment of the risks to both humans and affected ecosystems as well as the effects of management decisions concerning regulation of pollutant discharge, industrial process changes and site remediation efforts. There are always many uncertainties and unknowns to contend with in environmental decision making; therefore, it is desirable to have a comprehensive data base from which to effect informed decisions. With the high economic cost associated with comprehensive field monitoring, it is necessary to select monitoring techniques which yield the most pertinent information at the lowest cost. In the case of bleached kraft mill (BKM) sources, target areas for monitoring include: effluent discharges, receiving waters, sediments 2 and aquatic organisms. In most cases monitoring is carried out separately in one or more of these areas. Biomonitoring organisms can provide useful information in a number of areas, including seasonal bioavailability, as well as relative water and effluent concentrations (Phillips 1978). This thesis involves the development and assessment of an aquatic, biological monitoring tool for the purpose of routine monitoring of organochlorine contaminants specific to the discharge of bleached kraft pulp mills. Specifically, leeches are evaluated as in situ biomonitors of chlorinated phenolic compounds (Figure 1.1) through an integrated program of laboratory and field studies. This thesis will show the leech species, Nephelopsis obscura, to be a hardy and sensitive pollutant monitor, with the ability to provide reliable information concerning the presence, relative proportions and bioavailability of chlorinated phenolic compounds by assessment of tissue bioconcentrations. In order to provide adequate background information the thesis contains a literature review covering: 1) state of the art kraft mill process and pollution abatement technology, 2) environmental fate of chlorinated phenolic compounds, 3) biomonitoring theory, 4) leech biology. Many environmental and biotic factors influence the bioconcentration of organic contaminants, such as: contaminant concentration and temperature (Barron 1990), water pH (Saarikoski and Viluksela 1981; Hall and Jacob 1988), organic particulate concentration (Lee et al. 1993), uptake/depuration rate (Ellgehausen et al. 1980) and body size (Connell 1991). Since, in situ conditions show wide temporal and spatial variations, it is necessary to quantify the nature and magnitude of environmental factors on bioconcentration, in order to get accurate, long term monitoring information. The effect of the following environmental variables on bioconcentration of chlorinated phenolic compounds was investigated in the laboratory: contaminant concentration, water pH, water temperature and suspended sediment concentration. Ideally, routine biomonitoring should employ organisms of uniform age, size and condition from a laboratory controlled population. However, in reality, wild populations of differing life stage and origin must often be used. 3 Therefore, it is necessary to identify and quantify interspecies variations in bioconcentration. The effect of leech size on bioconcentration was investigated in the laboratory. Figure 1.1: Chemical structures of 1) Chlorinated phenols (CPs): DCP = dichlorophenol TCP = trichlorophenol TeCP = tetrachlorophenol PCP = pentachlorophenol 2) Chlorinated guaiacols (CGs): DCG = dichloroguaiacol TCG = trichloroguaiacol TeCG = tetrachloroguaiacol 3) Chlorinated catechols (CCs): DCC = dichlorocatechol TCC = trichlorocatechol TeCG = tetrachlorocatechol 4) Chlorinated vanillins (CVs): DCV = dichlorovanillin TCV = trichlorovanillin 5) Chlorinated veratroles (CVers): DCVer = dichloroveratrole TCVer = trichloroveratrole TeCVer = tetrachlorveratrole chlorinated phenolic compounds. OH 4 Leeches were evaluated as in situ b iomoni to r s i n the Fraser R i v e r sys tem, under a var ie ty o f seasonal f i e l d c o n d i t i o n s . T e m p o r a l va r ia t ions i n l e e c h b ioconcent ra t ion were assessed re la t ive to changing effluent, water and suspended sediment contaminant levels and vary ing environmental condi t ions . In addi t ion , the p r imary test species , Nephelopsis obscura was evaluated against a second leech species, Percymoorensis marmorata,under both f ie ld and laboratory condi t ions . In order to test the accuracy of laboratory data, f ie ld bioconcentrat ions, predicted from laboratory der ived relat ionships, were compared to measured f ie ld values. 5 2. L I T E R A T U R E R E V I E W 2.1. B L E A C H E D K R A F T M I L L T E C H N O L O G Y 2 .1 .1 . In t roduct ion A var ie ty o f w o o d p u l p i n g technologies are used w o r l d w i d e , i n c l u d i n g : m e c h a n i c a l , t h e r m o - m e c h a n i c a l , c h e m i - t h e r m o - m e c h a n i c a l , s e m i - c h e m i c a l , sulphate, sulfite and kraft processes. Throughout Canada and B r i t i s h C o l u m b i a , kraft pu lp ing is the dominant pulp produc ing process, w i th 25.2 m i l l i o n tonnes produced (86% of total pulp production) i n 1989 (Sinc la i r 1990)-, 94% o f wh ich was bleached (Celgar 1990). B K M s are responsible for a s ignif icant amount o f organic pollutant l o a d i n g into Canad i an waters , ( S i n c l a i r 1990). Classes o f pol lu tants d i scharged inc lude : suspended sol ids (SS) , b i o c h e m i c a l oxygen demand ( B O D ) , co lour , wood ext rac t ives such as res in and fatty acids and terpenes plus an assortment o f ch lor ina ted organic compounds . The bleached kraft process can be broken down into the f o l l o w i n g stages: wood pu lp ing , pulp washing and bleaching, wi th much o f the chemica l reagent be ing recovered and reprocessed for further use. The major processes i n v o l v e d are summarized be low, a long wi th descript ions o f the technology used i n chlor ina ted organic p o l l u t i o n abatement. 2.1.2. Kraf t Pu lp ing and Wash ing The goal o f pu lp ing is to remove the fibre b ind ing organic compounds, known as l i gn in s , f rom wood , to y i e l d ce l lu lose f ibres, the pr imary ingredient o f paper. L i g n i n is a large and complex aromatic po lymer , composed o f many substi tuted benzene r ings l i n k e d together to fo rm an extensive support structure for w o o d fibres. The pulp ing process breaks apart l i g n i n molecules into various size fractions 6 ranging from single r i n g phenol ic type compounds up to c o l l o i d a l macro-molecules (O 'Conner and Voss 1992). It is the various products of l i gn in breakdown that provide a substrate for the synthesis o f most o f chlor inated organics dur ing pulp b leach ing (Voss et a l . 1980). Therefore, the m a x i m a l removal o f l i g n i n f rom pulp is an important step i n the overa l l reduction o f A O X from effluent. The major focus o f pu lp ing technology is the development o f processes w h i c h a l low for the m a x i m u m degradation and remova l o f l i g n i n , wh i l e main ta in ing the in tegr i ty and strength o f the ce l lu lose fibres. Kraf t , also known as alkaline pulp ing , is a chemical form of pu lp ing , in wh ich l i g n i n is separated from wood fibres i n a caustic chemica l soup, k n o w n as white l iquor , under h igh temperature (160 - 180 °C) and pressure (690 - 1140 kPa) . Whi te l i quo r is composed o f a mixture var ious inorganic sulphur and sod ium con ta in ing compounds; p r inc ip le active ingredients are N a O H and N a 2 S ( M c C u b b i n 1983). After the c o o k i n g process, the spent c o o k i n g l i quor , w h i c h is termed b lack l i quo r , is drained and the wood fibres are desegregated i n a b low tank. The black l iquor , wh ich is an thick, odoriferous syrup, is recycled to recover the o r ig ina l cook ing chemicals . Kra f t pu lp ing can be carr ied out by either batch or continuous p u l p i n g processes. The pr inc ip le difference between the two forms o f pu lp ing being the way i n wh ich wood chips are exposed to the white l iquor . Ba tch pu lp ing involves the one time exposure o f wood to a single large dose o f alkaline l iquor for a period of t ime, whi le i n a continuous process the wood chips pass through a mul t i -zone c o o k i n g system, each zone wi th characterist ic exposure condi t ions . In the last decade, new pu lp ing technology has resulted i n 40 - 50% reductions i n residual l i g n i n i n pulp. Th is achievement is based on the pr inc ip le o f extended d e l i g n i f i c a t i o n or longer pu lp c o o k i n g . N o r m a l l y extended c o o k i n g results i n decreased pulp y ie lds and reduced pu lp strength, s temming f rom the degradat ive effects o f p ro longed exposure to b l ack l i q u o r ( K o c u r e k 1989). The m o d i f i e d cont inuous c o o k i n g ( M C C ) process c i rcumvents this p rob lem, through r emova l o f par t ia l ly spent l iquor and re-exposure to fresher cook ing l iquor i n a counter current system. Chemica l s such as anthraquinone and ^ O ? have also been introduced as 7 additives to increase the rate of l i g n i n removal ( M c C u b b i n 1984; A l l i s o n and Gra t z l 1987). Addi t ives are a cheaper alternative, but have a relat ively minor effect. F o l l o w i n g kraft d iges t ion, the pu lp , now ca l led brownstock, is very dark i n co lour and heavi ly impregnated wi th b lack l iquor wh ich must be washed out before further processing. F r o m an economic perspective, pulp washing pr ior to b leach ing is impor t an t , s ince b l e a c h i n g e f f i c i e n c y increases and c h e m i c a l c o n s u m p t i o n decreases wi th cleaner pulp ( M c C u b b i n 1983). F r o m an environmental perspective i t is also important to clean pulp as thoroughly as possible, since this removes much of the B O D , toxic wood extractives and l i gn in , the primary substrate for the formation o f chlor inated organic compounds. Ideally i t is desirable to remove a l l o f the solutes from the pulp pr ior to b leaching, but i n practice this is never achieved. T r i n h and Crotogino (1987) were able to show that solute removal from pulp is bi-phasic. There is an i n i t i a l rapid removal o f l i g n i n associated wi th black l iquor surrounding wood fibres, f o l l owed by a very s low leaching o f interfibre bound solutes las t ing on the order o f days. B rowns tock washing ( B S W ) general ly requires volumes o f water greater than the capaci ty o f the in-house recovery systems and much o f the organic mater ia l r insed from the pu lp is discharged to the environment . W a s h i n g effectiveness is dependent upon the avai labi l i ty o f water, the d i lu t ion factor of the rinse used and the number o f r ins ing stages employed. In addi t ion the characteristics o f the pulp are important; softwood pulps have a re la t ive ly high concentrat ion o f saponifiable wood extract ives , such as res in and fatty acids , w h i c h promote foam format ion i n the system, reducing B S W eff ic iency. T w o basic washing strategies are c o m m o n l y employed i n B S W , upon w h i c h several specif ic processes have been designed to match a par t icular economic and regulatory s i tuat ion. The older , more t radi t ional strategy, vacuum drum wash ing , i n v o l v e s the vacuum suct ion d r iven displacement o f b lack l i q u o r and r inse water f rom a pulp mat, m o v i n g over a sequential set o f perforated, rotating drums. T h i s strategy is re la t ive ly cheap and effective, but has certain drawbacks such as: h igh water requirement, air entrainment i n pulp , foam formation, increased inc idence o f 8 spills and greater overall effluent discharge (McCubbin 1983). The second strategy, which is employed in most modern mills is diffusion washing. This process relies on the positive pressure of water pumped through the pulp mat, to displace black liquor through a perforated wall. Diffusion washers are designed to operate as closed systems with minimal water usage and air entrainment. The advantages of diffusion washing include: low circulating water requirement, little or no foaming, decreased incidence of spills and effluent discharge and increased chemical recovery. One variant of diffusion washing, termed pressure diffusion washing, is of particular importance, because of its relatively low capital cost, ease of installation and ability to handle pulp directly from the digesters. Pressure diffusion washing has been the method of choice for many mills required to upgrade their systems in response to more stringent government effluent regulations. 2.2.3. Pulp Bleaching Currently Canadian BKMs cater to a worldwide market demand for high quality fully bleached pulp, which can function on high speed printing machines and stand up to the addition of various paper additives (B.C. Environment 1992). Pulp which is bleached with little or no chlorine is of inferior strength and at this time has limited market with environmentally conscious printing firms, mostly in western Europe. Future markets for non-chlorine bleached pulp are expected to rise, but not enough, within the next decade, to match the capacity of Canadian mills (B.C. Environment 1992). Effluent content of chlorinated organic compounds may be reported on an individual compound basis or as a part of integrative parameters such as adsorbable organic halide (AOX), purgable organic halide (POX) for volatile compounds, total organic chlorine (TOC1) and extractable organic chlorine (EOC1). TOC1 normally ranges 10 - 30% below AOX levels, while POX is usually 5% of AOX values (McCubbin et al. 1990). AOX is the most common parameter used in chlorinated organic regulation 9 in Canada, and must be measured periodically by Canadian BKMs. Theoretical loadings of AOX may also be calculated using the following formula (Celgar 1990): AOX = 0.11 (Cl 2 + 0.5 H + 0.2 C102) where: AOX is in kg of C l 2 per tonne pulp and C l 2 (molecular chlorine), H (hypochlorite) and C10 2 (chlorine dioxide) represent consumption of chlorine bleaching chemicals per tonne pulp. There is a danger associated with relying on a sum parameter such as AOX to gauge environmental risk. For example a modification to the oxygen delignification process, nitrogen dioxide pretreatment, resulted in a 10% reduction in effluent AOX (Brannland et al. 1989). However, toxicity tests carried out on several organisms including: microtox bacteria, daphnid species and zebra fish (Brachydanio rerio), revealed that the reduced AOX levels in effluent associated with N 0 2 pretreatment did not correspond to any decrease in lethal toxicity (Brannland et al. 1989). In fact, toxicity to microtox bacteria increased significantly, even though AOX concentrations were lower. A large variety of chlorinated organic compounds are synthesized in the bleaching process. Roughly 250 of these compounds have been identified (Suntio et al. 1988), accounting for only 10 - 20% of the organically bound chlorine (Earl and Reeve 1989). Of those compounds identified, most are of low molecular weight (< 1000 daltons), while the majority of the uncharacterized compounds are represented by chlorinated lignin fractions of varying molecular size, upwards of 1000 daltons (Kringstad and Lindstrom 1984). Each bleaching stage releases characteristic proportions of certain chlorinated organic compounds. In general most of the compounds of < 10,000 daltons are discharged in the first stage (CD) of bleaching, while the fraction of molecular weight > 10,000 daltons appears in the first extraction stage (E) effluent (Heimberger et al. 1988a). Some of the very low molecular weight classes of compounds identified from BKM effluents include: chlorinated resin and fatty acids, single ring chlorinated aromatic compounds (chlorobenzenes, chlorinated phenolics) and dioxins. Dioxins and furans are a special case since they 10 are not believed to be the products of lignin or wood extractive chlorination, but originate from non-chlorinated dibxin precursors, present as contaminants in defoamers, treated wood chips and ambient air (Voss et al. 1988). The major classes of chlorinated phenolics along with their bleach plant sources are presented in Table 2.1. The chlorinated phenolics represent only 2 - 4% of the total effluent AOX (Heimberger et al. 1988a) Table 2.1: Major classes and sources of chlorinated phenolic compounds found in softwood BKM effluent. Chlorinated Phenolic Class Number of Possible Isomers Number of Isomers Detected in BKME* Bleach Plant Sources** Percent of Total Chlorinated Phenolics Chlorophenol (CP) 2 9 1 4 D+C 7 Chlorocatechol (CC) 1 6 1 5 D+C 3 6 Chloroguaiacol (CG) 15 5 E 1 4 3 Chlorovanillin (CV) 7 7 E1 1 4 * From Suntio et al. 1988. " From Gergov et al. 1988. D+C = First chlorination stage of bleaching. E1 = First alkaline extraction stage of bleaching. Bleaching is the chemical treatment of pulp used to increase brightness to a desired level. Bleaching consists of two phases: the dissolution and removal of residual lignin and colour and the chemical oxidization (brightening) of pulp. Bleaching is carried out in multiple stages, involving the sequential addition of such bleaching chemicals as : molecular chlorine (C), chlorine dioxide (D), sodium hydroxide (E), hypochlorite (H), hydrogen peroxide (P) and oxygen (O). Canadian mills commonly follow a five stage sequence such as C D E D E D or DEEDED, where C D represents low chlorine dioxide substitution (5 - 20%) and represents high chlorine dioxide substitution (>20%) (Pryke 1989). The first two stages work to remove lignin and colour and generate the majority of chlorinated organic material. The final three stages serve to brighten pulp. E stages both represent alkaline 11 extraction stages, but with slightly different functions; the first E stage is used to extract acidic lignin fractions and the second E stage serves to increase the reactivity of pulp to the final D stage brightening (Axegard et al. 1984). Two bleaching technologies are in wide operation, the multiple tower-drum washer process and the displacement bleaching process. The former employs a separate bleaching tank for each stage, with pulp drum washers interspersed between the stages. Displacement bleaching takes place in a single closed tower, with multiple chambers for each reaction stage. Other than a lower fresh water requirement for displacement bleaching, the two methods show no difference in the total amount of BOD, colour or AOX discharged (McCubbin 1984). There are several bleaching process modifications currently employed by Canadian mills to reduce AOX discharge to the environment. Oxygen delignification or pre-bleaching, prior to first stage chlorine bleaching is a technology, which can result in 30 - 50% effluent AOX reductions (Heimberger et al. 1988b; Presley 1990). The process functions by removing up to 50% of the lignin from pulp before it reaches the bleach plant and has a chance to react with chlorine. Since the effluent from the oxygen delignification process is non-chlorinated, it can be recycled through the mill recovery system. One variation of the oxygen delignification process, called PRENOXR, involves the addition of nitrogen dioxide (N0 2 ) prior to oxygen treatment. A further 10% reduction in effluent AOX can be achieved using this modification, but pulp quality seems to be variable (Lindqvist et al. 1986). The dramatic reduction in AOX seen with oxygen delignification has a heavy capital cost, since the system requires a separate reaction tower with associated mixing and pumping equipment. Chlorine multiple is the amount of chlorine administered to pulp with a given percent lignin content. Increasing chlorine multiple results in higher AOX generation (Pryke 1989). A reduction in the loading of chlorine to pulp can be achieved though, as long as machines used for mixing bleaching chemicals into pulp are optimized. In particular, high intensity mixing technology has been introduced to increase mixing efficiency to the level where chlorine multiple may be reduced 12 by 25 - 40% (Earl and Reeve 1989). Increased chlorine dioxide substitution is one of the most common bleaching process modifications used to achieve lower effluent AOX levels, while maintaining pulp brightness and strength. The most efficient delignification is achieved if chlorine dioxide is charged to the pulp before molecular chlorine, however the greatest reduction in AOX effluent levels occurs when the order is reversed. Intermediate results are observed when chemical charging is simultaneous (Axegard 1986). While increasing chlorine dioxide substitution generally reduces effluent AOX, the effects on individual chlorinated compounds are more variable. Table 2.2, from data presented in Liebergott et al. (1990) shows that chlorinated phenolics concentrations may rise with initial increases in chlorine dioxide substitution and that the situation is further complicated by chemical charging order. From these data we can see that the CGs and CVs are the most prominent chlorinated phenolics in bleachery effluents, even at high chlorine dioxide substitution. Voss et al. (1980) characterized BKM effluents for chlorinated phenolics and found that the following CGs and CVs were present in the highest concentrations: 4,5-DCG, 3,4,5-TCG, 4,5,6-TCG, TeCG, 6-CV and 5,6-DCV. These particular compounds make good candidates for tracing the extent of pulp mill derived chlorinated phenolic contamination. Effluent colour decreases with increasing chlorine dioxide substitution (Pryke 1989), while BOD values are variable and resin acid levels are unaffected (Heimberger et al. 1988b). Table 2.3 summarizes the changes in effluent colour, BOD and TOC1 as a function of bleaching sequence. 2.1.4. Effluent Treatment Pulp mills employ a number of effluent treatment techniques to reduce TSS, BOD, resin and fatty acids, AOX and effluent toxicity. An effluent treatment program can include pretreatment (screening, neutralization) primary (1°) clarification to remove suspended solids, secondary (2°) biological treatment (aerated lagoon, 13 ZJ Q . CO g "5 c CD X : Q-" O <D oS c 'i— o o sz cn o T— H— : o CC CD o CO O E » ° x i S3 II °> to O J CO - ~ X <1) O "CJ CO o II ° 6 •a c CO J o "3 CD S c Xt £ ZJ x CO cj 0} II S O o II — D> x: c o z •~ O s s e s 0) Q-2" o c Q) CO "8 5 .s CM c\i a> JO o CJ) a o LE IO T J O o O ) o o o co O OX - -b o h-CD O C O o Chi 5 CO q co d o c\i Is- CO ^ CM m cn CO M (A o o J Z O O CO 2 CO CO D o 5? o o o o J Z J Z O O o c 0) o O o cvi d CO CJ CM CO CO CO CO O) CO CM CO CJ> o in O) CD co IO CO iq d CM CD CO CJ CO CO CO r~-CO CD 6 CO •<* CO CJ) -a- CM d CO CM CM •q- m CM -3- o o> CJ in cc T J SZ CO C O O 6 | co n o CO > a. ZJ Q. T3 O O 5 CO Q- O CO & cn 1 5- o cr nj .2 xi iz CD CD "5 c o — 8 CO Q . E o O co cvi XI co xi CO CO a> co *— CD cn v_ CD XI E CD I E o c CD « cr CD co CO Q o o „ O a. ZJ C L H—' -a co "&» ZJ XI o o o CM r~ CM T -co in I O in CD o c CD cr CD C O o» c lc o CO _® m T- o CO CO o o in o C O m | a LU Q o" LU Q O Q LU Q o LU o Q O LU Q LU a LU a O o Q LU Q o O Q O O 14 activated sludge, anaerobic treatment), 2° clarification for contaminated biosolids and tertiary (3°) colour removal. For the removal of BOD, toxic wood extractives and chlorinated organics, 2° biological treatment and 2° clarification are most effective. The majority of Canadian BKMs discharging to inland waters have some form of biological treatment, usually aerated lagoons alone or in combination with activated sludge facilities. Coastal mills have been historically remiss in providing 2° effluent treatment, but with more stringent effluent AOX regulations, 2° treatment will become a necessity for all BKMs. Secondary clarification has been used most commonly in the recycling of biosolids lost from activated sludge systems (NCASI 1989), but research in the area of biosolids contamination by Derksen (personal communication) has shown that biosolids entering the Fraser River from B K M aerated lagoons contain adsorbed dioxins1 . Since many hydrophobic chlorinated organic compounds tend to partition to organic solids, it is likely that other chlorinated compound are also adsorbed to discharged biosolids. Secondary clarification could be one solution to this particular problem. The effect of biological treatment on effluent quality parameters is summarized in Table 2.4. Table 2.4: The effect of biological treatment on effluent quality parameters. Data from Heimberger et al. 1988a. | Production (kg/adt pulp) Chlorinated Bleaching Sequence TOCI BOD Colour Phenolics (CD)(E0)DED 5.5 25 200 0.10 With Bio-Treatment 3.3 2.5 200 0.08 O(CD)(E0)DED 3.5 20 100 0.07 With Bio-Treatment 2.1 2.0 100 0.06 Under optimal conditions both aerated lagoon and activated sludge treatments can reduce BOD (85 - 95%) (NCASI 1989), resin and fatty acids (71 - 100%) (Celgar 1 George Derksen, Environment Canada, Environmental Protection Branch, 224 West Esplanade, North Vancouver, B C . V 7 M 3H7, 1992. 15 1990), A O X (48 - 65%) and chlorinated phenolics (15 - 90%) (Gergov et al 1988; Heimberger et a l . 1988a). Anae rob ic treatment is not general ly prac t iced , but i s currently being investigated for future applications to B K M effluent. One Anaerob ic system, E N S O F E N O X R process has been attempted on a m i l l scale wi th some success ( N C A S I 1989). E N S O F E N O X R is a dual anaerobic ( f l u i d i z e d bed reactor) aerobic (biofi l ter) process. Reduc t ion i n tox ic i ty and mutagenicity as w e l l as dechlor ina t ion o f chlor inated phenolics has been achieved us ing this system. 2.2. E N V I R O N M E N T A L F A T E O F C H L O R I N A T E D P H E N O L I C S 2 .2 .1 . In t roduct ion A n integral part o f assessing the r i sk of environmental impact of pollutants is the determination o f environmental fate. What are the inputs? H o w is i t transported, distributed and cleared from the b iophys ica l system? H o w persistent is i t? These are al l questions that must be addressed in order to determine the environmental fate o f a pol lutant . In real i ty on ly l i m i t e d in format ion is avai lable from w h i c h to answer these questions. Since there are many interacting variables affecting fate and g iven that b iophys i ca l systems are often h igh ly variable both temporal ly and spat ia l ly , the answers become h igh ly complex . W i t h i n the context o f the fate of chlorinated organic compounds i n the aquatic env i ronment , there are three major e n v i r o n m e n t a l compar tments amoung w h i c h par t i t ioning takes place, namely water, sediments and biota . These compartments are not p h y s i c a l l y e x c l u s i v e , since pore water is present w i t h i n sediment beds and sediments suspended are transported wi th in water, whi le biota may be present at a l l c o m p a r t m e n t in te r faces . F a c t o r s i n f l u e n c i n g the t ranspor t , d i s t r i b u t i o n and c learance o f ch lo r ina ted organic compounds f a l l in to four ca tegor ies : po l lu tan t chemis t ry , p h y s i c a l and chemica l character is t ics o f the r e c e i v i n g envi ronment and b i o l o g i c a l i n t e rac t ions . 16 2.2.2. Sources of Chlorinated Phenolics The inputs of chlorinated phenolics to the aquatic environment are mostly from anthropogenic sources, but some natural processes contribute chlorinated phenolics to the environment. Grimvall et al. (1991) found elevated concentrations of AOX (200 (j.g Cl/L) in Scandinavian lakes considered to be unpolluted. Their laboratory studies gave strong evidence for enzyme mediated chlorination of humic substances in natural soils and waters. The only chlorinated phenolic detected from these reactions has been 2,4,6-TCP. The O-methylation product of 2,4,6-TCP, 2,4,6-trichloroanisole, was also detected in the same systems and is believed to be the result of in situ bacterial transformation. Anthropogenic sources of chlorinated phenolics include: saw mill and wood processing facilities using chlorinated phenolic treated wood (Hall and Jacob 1988), chlorinated herbicide industries and facilities carrying out organic combustion processes such as municipal waste incinerators and fossil fuel burning plants (Paasivirta et al. 1985). Vehicles for the input of chlorinated phenolics include: industrial effluent discharged directly into receiving waters, leaching and surface runoff from terrestrial sources and atmospheric deposition on the surface of dust particles. In BKM effluent receiving waters, direct discharge is the most important vehicle. The major classes of chlorinated phenolics, CPs, CCs, CGs and CVs (Figure 1.1) are derived from different industrial" sources. The CPs and CCs are formed in organic combustion processes and pulp bleaching. The CPs, particularly tetra and penta substituted forms, are also associated with wood treated with these compounds as an antifungicide. Currently, government restrictions on wood treating agents is reducing input from this source in North America and Western Europe. Air pyrolysis of the ubiquitous PCBs has also been, shown to result in formation of mono and di-chlorinated phenols (Paasivirta et al. 1985). A relatively small but unquantified percentage of CGs seem to be formed in the combustion of organic fuels, such as peat and coal, but for the most part the CGs along with the CVs are restricted to BKM 17 sources (Paasivirta et al. 1985), making these classes of chlorinated phenolics good tracers of BKM chlorinated organic pollution. 2.2.3. Partitioning to the Water Phase The chlorinated phenolics differ from many other lipophilic organic contaminants in that they are acidic and therefore dissociate under alkaline pH, to the phenolate ion form. Therefore, water pH is one of the factors governing the behaviour of chlorophenols in receiving waters. The undissociated form is more lipophilic and tends to partition to organic phases, such as living organisms (Carey 1988). Table 2.5 gives the dissociation constants and percent dissociation in water of pH 7.8, which is close to that of Fraser River water, for selected chlorinated phenolics. Note that the acidity of chlorinated phenolics increases, while pKa decreases with increasing chlorine substitution. Table 2.5: Dissociation constants (pKa) and percent dissociation of chlorinated phenolics in water of pH 7.8. Data from Xie (1983). COMPOUND pKa PERCENT DISSOCIATION (pH 7.8) 50\0~ 98.4 99.6 99.7 5.9 38.7 71.6 98.4 2.4- DCP 7.8 2,4,6-TCP 6.0 2,3,4,6-TeCP 5.4 PCP 5.3 4.5- DCG 9.0 3.4.5- TCG 8.0 4.5.6- TCG 7.4 TeCG 6.0 18 Volatilization of chlorinated phenolics is influenced by the molecular weight, ambient temperature and mixing conditions. There is a greater tendency for lighter, less chlorine substituted compounds to volatilize with increasing temperature and water surface agitation. Volatilization favours the unbound, undissociated form of chlorinated phenolics (Jacob 1986). In the Fraser River, the heavy tri- and tetra-chlorinated phenolics are mostly in the dissociated form and would not be expected to volatilize at a significant rate, but the lighter mono- and di-chlorinated phenolics may be lost to a measurable degree through volatilization. Chlorinated phenolics undergo photolytic degradation, particularly the phenolate form, which absorbs light at different wavelengths than the undissociated form (Boule et al. 1982). Photolysis rate varies with geographical latitude, time of day, season and light admitting characteristics of the particular water body. In the Fraser River photolysis reactions would be at a minimum in winter, under ice cover and may not be of great significance in spring and summer, since the suspended sediment load in the Fraser River hinders light penetration. Halogenated aromatic compounds are generally resistant to hydrolysis and this type of reaction is not considered important in the fate of chlorinated phenolics (Jacob 1986). 2.2.4. Partitioning to the Sediment Phase The interaction of chlorinated organic compounds with the sediment compartment is one of the key processes influencing their environmental fate. Pollutant chemistry and environmental factors such as octanol water partition coefficient (Kow), and the fraction of organic carbon in sediments (foe) exert a strong influence on the sediment-water partition coefficient (Ksw) of hydrophobic organic compounds (Karickhoff et al. 1979; Jaffe 1991). Voice et al. (1983) also pointed out a third variable, the concentration of the solid phase in the system, as a factor influencing Ksw. Their laboratory research revealed that as suspended 19 sediment concentration increased, sediment partitioning of chlorinated aromatic compounds appeared to decrease. This phenomenon was explained by the presence of a third phase in the system, colloidal macromolecules and microparticles, too small to separate from the water phase. As suspended sediment concentration was increased, the amount of sediment binding microparticles introduced also increased, making it appear that there was an inverse relationship between sediment concentration and K s w . The contribution of the microparticle and colloidal phase in the environment is difficult to quantify at this time, but with the discharge of suspended biosolids and colloidal lignins from effluent treatment lagoons, this process requires further investigation. Partitioning of chlorinated organic compounds to the larger, settleable sediment fraction does not remain constant over the whole particle size range. Particles of the silt and clay fractions, < 63 um tend to have the greatest adsorptive capacity, due to the greater surface area per unit mass (Schellenberg et al. 1984) and in some cases a greater amount of associated organic carbon (Voice et al. 1983). Equations used to estimate Ksw, from Kow and foe have been developed (Karickhoff et al. 1979); Ks w = Koc * foe and Koc =Kow* 0.63 where Koc is the carbon normalized partition coefficient. These relationships are only valid where organic content is > 0.1% (foe > 0.001); for organic poor sediments, inorganic interactions may be of overriding importance (Schellenberg et al. 1984). The term sorption is generally used to describe the binding of compounds to the solid phase. Sorption to a two dimensional surface is usually referred to as adsorption, while the term absorption describes partitioning into the three dimensional matrix of a sorbent (Schwarzenbach 1985). Dissolved nonpolar, hydrophobic organic compounds tend to associate with solids of similar hydrophobic character (ie. organic 20 particulates) in the thermodynamic ally favourable process of hydrophobic bonding, a non-covalent form of bonding of moderate strength (von Oepen et al. 1991). The relatively simple partitioning model describing the Ksw for hydrophobic organic compounds is only partially applicable to the ionizable chlorinated phenolics. Sediment partitioning of chlorinated phenolics is also dependent on water pH and ionic strength (Westall 1985). Hydrophobic compounds, which possess ionizable functional groups, may interact with both organic and inorganic sorbents, by processes such as ligand exchange, ion bonding (von Oepen et al. 1991) and ion pairs, which can in turn, move to organic phases (Schwarzenbach 1985). Phenolate anion sorption can take place on surfaces with positively charged sites, such a clay surfaces (Westall 1985), but may only be significant, under neutral pH, for the more acidic highly chlorine substituted chlorinated phenolics (Schellenberg et al. 1984). However, the mono and di-substituted chlorinated phenolics, which tend to be more in the undissociated phase would expect to be attracted to inorganic binding sites to a much lesser degree. Further complexities arise between the various groups of chlorinated phenolics. For example, the CCs were shown by Remberger et al. (1993), to form readily reversible complexes with metal cations (Fe^+, Al^ + ) , while the CGs showed evidence of increased proportion of covalently bound compound. From the many studies investigating sediment partitioning of chlorinated phenolics, it is difficult to make generalizations as to the nature of the partitioning, since it seems to vary greatly among various studies and is probably highly specific to sediment composition, water composition and laboratory conditions. 2.2.5. Partitioning to Biota Partitioning of contaminants to biota takes place through the processes of bioconcentration and bioaccumulation. Bioconcentration can be defined as the absorption of a substance by direct contact with the ambient environment through 21 oral, percutaneous or respiratory routes (Jaffe 1991). Inclusion of uptake through dietary sources is termed bioaccumulation. From this definition of bioaccumulation, it follows that contaminant concentration may be magnified through higher trophic levels of the food chain; this phenomenon is termed biomagnification. The ratio of the quantity of a contaminant bioconcentrated to the original exposure concentration in the water is related through the term bioconcentration factor (BCF). Strictly speaking, the BCF equals the steady state equilibrium tissue concentration divided by the water concentration (Ellgehausen et al. 1980). Since hydrophobic compounds tend to passively partition into lipid phases, the mechanism of uptake is usually passive diffusion through a lipid membrane, across a concentration gradient. Uptake sites are most commonly associated with respiratory surfaces exposed directly to contaminated water, such as gills in fish and respiratory pores or gill like structures for invertebrates (Barron 1990). Skin absorption is also an important uptake route in some fish species. Saarikoski et al. (1986) found that skin uptake accounted for 25 - 40% of PCB bioconcentration in guppies (Poecilia reticulata). Bioconcentration is a function of the competing processes of uptake and elimination (Ellgehausen et al. 1980). In turn, these processes are controlled by the chemical nature of the substance and the interaction of external environmental factors with both the physical-chemical behaviour of a substance and species specific biology, specifically uptake, distribution and elimination kinetics (Barron 1990). Chemical properties found to influence bioconcentration include both molecular size (Barron 1990) and hydrophobicity (Neely et al. 1974; McKim and Schmieder 1991). External environmental variables documented to affect bioconcentration of organic contaminants in aquatic organisms include contaminant concentration (Ellgehausen et al. 1980; Hall and Jacob 1988), water temperature (Veith et al. 1979; Barron et al. 1987b; Sijm 1991), pH (Saarikoski et al. 1986; Hall and Jacob 1988) and suspended and dissolved solids (Opperhuizen and Stokkel 1988; Lee et al. 1993). Biophysical characteristics such as lipid content (Ernst et al. 1991) and 22 surface to volume ratio (Saarikoski et al. 1986) have been reported to affect bioconcentration. 2.2.5.1. The Effect of Octanol-Water Partition Coefficient The octanol-water partition coefficient (Kow) defines the relative tendency of a chemical to partition to the organic phase (Carey 1988). A positive linear relationship between bioconcentration and Kow has been reported for both neutral lipophilic organic compounds and the undissociated form of the acidic chlorinated phenolic compounds. Hawker and Connell (1986) reported a strong linear correlation between bioconcentration and log K o w of organochlorine pesticides and chlorinated biphenyls in fish, mollusc and daphnid species over the log K o w range 2 - 6. Saarikoski et al. (1986) reported similar relationships for chlorinated phenolics in the unionized form (i.e. pH > 1 unit below pKa), but found that the plot of bioconcentration vs. log Kow started to level off at log Kow values > 4. The upper limit of the relationship appears to be limited by molecular size; large super-lipophillic compounds bioconcentrate to less than predicted values due to steric hindrance (Barron 1990; Jaffe 1991). The apparent strong correlation between bioconcentration and Kow has led many investigators to attempt to use this relationship as a predictive tool (Mackay 1982; Isnard and Lambert 1988; McKim and Schmieder 1991). There are several reasons why such a practice should be regarded with caution. Jaffe (1991) pointed out that structural differences between octanol and bio-lipids lead to differing thermodynamics of partitioning. Furthermore, close inspection of large data sets (Mackay 1982; McKim and Schmieder 1991) reveals that the relationship is only of a general nature; compounds of similar Kow can differ in bioconcentration by > 1.5 orders of magnitude. Finally, predictions of bioconcentration based upon Kow fail to account for interspecies differences in toxicokinetics and response to variations in external environmental factors (Barron 1990). 23 2.2.5.2. The Effect of pH The chemical nature of neutral organic compounds (e.g. 3,4,5-TCVer) does not change with water pH. However, the acidic family of chlorinated phenolic compounds dissociate into the less lipophilic phenolate anion form as water pH approaches pKa. TeCG (pKa = 6.0) is more than 90% ionized at pH 7.0, while 4,5-DCG (pKa = 9.0) is < 1% dissociated at the same pH. The ionized species is water soluble and therefore, less bioavailable to aquatic organisms (Barron 1990). The Kow for PCP is reduced by almost two orders of magnitude at pH > 7 (Carey 1988). However, when pH is > 1 unit below pKa, pH has a negligible affect on bioconcentration (Saarikoski et al. 1986). The assumption that the ionized form of a chlorinated phenolic compound does not represent a significant pool of bioavailable compound has been demonstrated to be incorrect in several studies. Laboratory studies have revealed an inverse relationship between water pH and both chlorophenol toxicity and bioconcentration, however the magnitude of change in bioconcentration is less than what would be predicted from chemical dissociation curves. Saarikoski and Viluksela (1981) reported decreased toxicity of tri- and pentachlorophenols to guppies (Poecilia reticulata) as pH increased (pH = 5 - 8 ) , with a significant toxic effect still apparent at pHs corresponding to > 90% ionization. The toxicity of 4-chlorophenol, which remained virtually unionized at all water pHs showed a slight trend towards increasing toxicity at elevated pH. Investigating the effect of water pH on the bioconcentration of 17 phenolic compounds, Saarikoski et al. (1986) again found that bioconcentration decreased less at elevated pH than predicted from pH - pKa dissociation curves, indicating that the phenolate ion represents a significant pool of bioavailable compound. Support for this conclusion comes from the finding that fish bioconcentrate relatively high levels of PCP (pKa = 5.0) and TeCG (pKa = 6.0) in the Fraser River (pH = 7.5 - 8.5), even though they are present at > 92% in the ionized form (Carey 1988; Rogers et al. 1988; Servizi et al. 1988; Dwernychuk et al. 1991). Variation in external pH, not only alters chemical speciation of acidic compounds, but also affects bioaccumulation physiology. Increased water pH resulted in 24 decreased metabolic clearance of PCP by goldfish (Carassius auratus) (Stehly and Hayton 1990). A similar affect would explain the trend towards increased toxicity of 4-chlorophenol to guppies at increased water pH, observed by Saarikoski and Viluksela (1981) and the increase in leech (Nephelopsis obscura) bioconcentration of 2,4-DCP over the pH interval 5.0 - 7.5 (Hall and Jacob 1988). Decreased bioavailability of chlorinated phenolic compounds at high water pH may be offset to some degree by reduced elimination rates. Theoretically, such an effect could be monitored by including a neutral organic compound, such as 3,4,5-TCVer in the laboratory assessments. 2.2.5.3. The Effect of Dissolved and Particulate Solids Adsorptive partitioning of neutral and ionized organic compounds to both organic (Schellenberg et al. 1984; Schwarzenbach 1985) and inorganic (Remberger et al. 1993; Xing et al. 1993) particulates as well as colloidal and dissolved organic material (Kukkonen and Oikari 1991) has been documented. Suspended and dissolved material represent potentially significant competitors to bioconcentration, but the phenomenon has only received limited investigation. The degree of the effect is dependent upon particle size, the percentage of organic material, its chemical character (Schwarzenbach 1985). and water pH (Westall 1985). The variable character of suspended and dissolved organic material present in natural receiving waters makes the task of assessing competitive effects upon bioconcentration highly complex and difficult to generalize. Kukkonen and Oikari (1991) assessed the bioavailability of neutral (benzo(a)pyrene; naphthalene; tetrachlorobiphenyl) and acidic (dehydroabietic acid) organic compounds to Daphnia magna in the presence of surface waters from 20 different locations in Eastern Finland. Dissolved organic material could be divided into hydrophobic acids, hydrophobic neutrals and hydrophilic material, each with a characteristic affinity for organic contaminants. Partitioning of benzo(a)pyrene to dissolved organic material was correlated to the 25 hydrophobic acid content (humic material), however no significant correlation between the other test compounds and any measured characteristic of dissolved material was evident. Bioconcentration by Daphnia magnawas strongly correlated to hydrophobic acid content for benzo(a)pyrene and tetrachlorobiphenyl and showed weak correlations for naphthalene and dehydroabietic acid. Physical-chemical properties of a contaminant, such as Kow, can influence the magnitude of competitive inhibition of bioconcentration by particulate and dissolved material. Opperhuizen and Stokkel (1988) reported that highly lipophilic compounds such as chlorinated biphenyls have a greater tendency to partition to the particulate phase (Chromosorb), resulting in significant decreases in bioconcentration by guppies (Poecilia reticulata) Bioconcentration of more water soluble, moderately lipophilic chemicals such as di- and trichlorinated benzenes was not affected significantly by particulate material. This is consistent with the observation that PCP toxicity to Daphnia magna and zebrafish (Brachydanio rerio)is not affected by the presence of dissolved humic material (Lee et al. 1993). In studies aimed at predicting the fate of chlorinated phenolics in the Fraser River by Carey (1988), it was concluded that most of the compound would remain dissolved in the water column, with suspended particulates causing little inhibitive effect on bioconcentration. However, colloidal lignins and organic particulates from pulp mill effluent treatment lagoons may reduce the amount of chlorinated organics available to aquatic organisms. Derksen (personal communication) found evidence of this in studies examining the association of chlorinated dioxins with very fine biosolids centrifuged out of BKM effluent from mills in Prince George, B.C 2 . Evidence suggests that adequate characterization of the amount and nature of organic material and their effect on bioconcentration, in receiving waters is necessary for accurate prediction of partitioning to biota. 2 George Derksen, Environment Canada, 1992. 26 2.2.5.4. The Effect of Temperature Ambient temperature may influence bioconcentration in a number of ways. Since fish and aquatic invertebrates are poikiolothermic, temperature will affect the rate of physiological processes. Rates of contaminant uptake, metabolism and elimination usually increase with temperature increases (Barron 1990). Over the temperature range 4 - 22 °C, Hall and Jacob (1988) observed a strong positive correlation between leech (N. obscura) bioconcentration of chlorophenols and temperature. The time to reach steady state tissue concentrations was inversely related to water temperature. At 4 °C equilibrium bioconcentrations of 2,3,4,6-TeCP were reached after four days, while steady state was not reached after seven days at 22 °C. The effect of temperature seems to be dependent upon the species specific optimal tolerance range. Veith et al. (1979) tested the effect of temperature (T = 5 - 25 °C) on PCB bioconcentration by three fish species. Rainbow trout (Salmo gairdneri) showed no increase in bioconcentration between 5 and 10 °C, with a relatively sharp increase over the temperature range of 10 - 20 °C. Both the fathead minnow (Pimephales promelas) and the green sunfish (Lepomis cyanellus), showed the opposite trend. In some cases, very slight effects of temperature on bioconcentration have also been observed. For example, Sijm (1991) reported only small differences in bioconcentration of PCBs in guppies (Poecilia reticulata)at1 = 13 - 33°C. Species specific relationships between bioconcentration and temperature can be linked to the effect of temperature on physiological function and accumulation kinetics. Barron et al. (1987a) observed linear increases in both cardiac output and blood flow rate for rainbow trout (Salmo gairdneri) exposed to water temperatures of 6, 12 and 18 °C. Uptake of organic contaminants by many aquatic organisms takes place mostly across highly blood profused respiratory surfaces (e.g. fish gill) (Barron 1990). Therefore, physiological variables such as cardiac output, ventilatory volume and blood flow rate are potential uptake rate limiting factors. Greater blood flow to the gill can result in a greater uptake rate, while increased ventilatory volume increases apparent exposure at the uptake site. Elevated temperature also 27 generally leads to increased metabolic elimination rates of organic compounds (Barron 1990). Over the temperature range 6 - 18 °C Barron et al. (1987b) observed an increase in the total body clearance of di-2-ethylhexyl phthalate in rainbow trout. However, uptake rate and distribution to deep storage compartments remained dominant over the enhanced elimination rate. Similarly, Jiminez et al. (1987) observed 5.8 times increase in uptake rate and 3.6 times increase in elimination rate of benzo(a)pyrene in sunfish species, over the temperature range 13 - 23 °C. The lack of increase in bioconcentration of PCBs by rainbow trout over the temperature range 5 - 1 0 °C, observed by Veith et al. (1979) is consistent with the idea that increased uptake rate due to elevated cardiac output was balanced by an increased rate of elimination. A sharp increase in bioconcentration of PCBs above 10 °C can be explained by dominance of uptake and deep compartment storage over elimination. Temperature affects other physiological determinants, such as membrane permeability and lipid composition, which changes to maintain a constant fluidity as temperature changes (Barron 1990). Furthermore, the degree to which each physiological process will be affected will vary with size, age and species (Barron 1990). Temperature can affect the chemistry of organic compounds. Solubility of compounds is generally increased at higher temperature (Phillips 1978), resulting in more available compound, but this is not expected to be environmentally relevant, since chlorinated phenolics, discharged from BKMs, are not expected to reach high enough concentrations to be in the undissolved form. 2.2.5.5. The Effect of Multi-Contaminant Exposure Investigators have been trying to elucidate the potential effects of simultaneous exposure to many toxic contaminants on the bioconcentration of individual compounds with conflicting results. Exposure to chemical mixtures has shown synergistic, additive or independent effects, depending upon the compounds 28 and animal species tested. F rede r i ck (1975) found no effect on b ioconcent ra t ion when white suckers (Catostomus commersoni) were exposed to A r o c h l o r 1232 and d ie ld r in , both i n d i v i d u a l l y and i n combinat ion . A s imi la r result was observed when At l an t i c salmon fry (Salmo salar) were exposed to mixtures o f the pesticides a ldr in and D D T (Addison et a l . 1976). Contrary to these results i t was reported (Phi l l ips 1978) that h igher leve ls o f both D D T and d i e l d r i n were accumulated by ra inbow trout (Salmo gairdneri) when the compounds were added to water as a mixture. P h i l l i p s (1978) suggests that an organism be e l iminated from considerat ion as a b iomoni tor , s h o u l d any k n o w n in t e rac t ive effects be tween o r g a n o c h l o r i n e s be e s t ab l i shed . C o n s i d e r i n g the wide spectrum o f poss ib le contaminant in teract ions o c c u r r i n g i n any f i e l d s i tua t ion , e spec ia l ly w i th respect to B K M water p o l l u t i o n , this seems u n r e a l i s t i c , s i nce i t i s l i k e l y there w i l l a l w a y s be i n t e r a c t i o n s a f f e c t i n g b ioconcent ra t ion . A more rea l i s t ic approach may be to conduct m u l t i - c o m p o u n d exposures at levels which are l i ke ly to occur i n the f ield. 2 .2 .6 . B i o t r a n s f o r m a t i o n C h l o r i n a t e d phenol ics are metabol ized by aquatic organisms through enzyme mediated t ransformation react ions. M i c r o b i a l t ransformation p robab ly accounts for most o f the the removal o f chlor inated phenol ics from aquatic systems. The rate m i c r o b i a l d e g r a d a t i o n is h i g h l y s i te s p e c i f i c and dependen t u p o n many env i ronmen ta l va r iab les , such as temperature, l i g h t penet ra t ion , water f l ow and b i o a v a i l a b i l i t y . In one case, Carey et a l . (1984) reported rapid dech lor ina t ion o f chlorophenols by stream per iphyton ; ha l f l i ve s were on the order o f four to s ix h o u r s . O n e o f the m o s t e n v i r o n m e n t a l l y s i g n i f i c a n t c h l o r i n a t e d p h e n o l i c transformations i n v o l v e s m i c r o b i a l O-methy la t ion o f phenol ic compounds , w i t h the concomi tan t format ion o f tox ic anisoles ( C P der iva t ive) and veratroles ( C C , C G derivative). The result ing derivatives are more l i poph i l i c and toxic to fish than the 29 parent compounds (Neilson et al. 1984). These transformations are performed by a number of classes of microbes including gram positive and gram negative bacteria as well as several species of fungi (Neison et al. 1987; Harper et al. 1989). Allard et al. (1988) investigated the rate of transformation of 3,4,5-TCG to 3,4,5-TCVer in a number of bacterial strains. It was discovered that high concentrations (100 ppb) of 3,4,5-TCG resulted primarily in the formation of the veratrole derivative, but decreasing substrate concentrations led to a complex array of metabolites (chlorosyringols, chloromethoxybenzenes). However, metabolism seemed to shift back towards chloroveratrole production as cell densities decreased to low concentrations, as would be expected in the environment. Relatively rapid O-methylation of CCs, to CGs and veratrole derivatives has also been observed in natural sediments and soils (Brezny et al. 1992; Remberger et al. 1986). Incubation of cell cultures with BKM derived lignin has also yielded both tri- and tetra-chloroveratrole derivatives (Allard et al. 1988), however it is not certain whether products were derived from attack and cleavage of lignin precursors or from adsorbed CG and CC molecules. Removal of methyl groups (de-O-methylation) has been reported to occur under anaerobic conditions in sediments (Neilson et al. 1984; Allard et al. 1988), however the environmental significance has not yet been determined. Dechlorination reactions are carried out by both aerobic and anaerobic bacteria. Aerobic processes most commonly proceed by hydroxylation followed by reductive dehalogenation and aromatic ring cleavage. Investigations of this process were carried out by Haggblom et al. (1988), using the bacteria Rhodococcus chlorophenolicus . This species prefers to attack the chlorophenols, but will also degrade other related compounds possessing a methoxy group in ring position 2 or 6. Therefore it is not a good degrader of the CCs. Fastest reaction rates seem to be for the more highly chlorine substituted compounds, but is also dependent upon the position of the chlorine atoms (eg. rate of reaction for 3,4,6-TCG > 4,5,6-TCG). Anaerobic dechlorination of CCs has been observed in natural sediment mesocosm studies (Neilson et al. 1989). The CGs can also be metabolized by this route, provided they undergo de-O-methylation to yield the corresponding CC. Unlike 30 aerobic dehalogenation, which can completely remove all chlorine atoms from the parent molecule, anaerobic dechlorination appears to halt after a single dechlorination (Neilson et al. 1987). Higher organisms, such as fish tend to carry out enzyme mediated biotransformations of chlorinated phenolics in detoxifying tissues located mostly in the liver and kidneys, yielding more water soluble conjugates, which are eliminated through the urine or intestinal bile secretions. Parent chlorinated phenolics are also excreted, but in concentrations roughly ten times lower than the transformed products (Kennedy 1989). Conjugation, in fish species, most often involves the enzyme catalysed (UDP-glucuronyl transferase) transfer of a glucuronic acid molecule (glucose derivative) to a reactive hydroxyl site. Sulfate group conjugation is also found in most fish species, but to a lesser extent (Kennedy 1989). For example, CGs and CCs were metabolized to glucuronide and sulfate conjugates by zebra fish (Brachydanio rerid), with quantities of the former ranging from three to seven times greater (Neilson et al. 1989). The more lipophilic chloroveratroles were also metabolized by zebra fish, first by single or double de-O-methylation to the corresponding CG or CC, followed by conjugation and excretion (Allard et al. 1988). Half lives of chlorinated phenolics in fish vary with species, but generally appear to be < 3 days. Kennedy (1989) reported a half life of 65 h for PCP in rainbow trout (Salmo gairdneri), while a half life of 6.2 - 23 h was reported for PCP for the same fish species in another study (Metcalf et al. 1988). Renberg et al. (1980) reported the half life of 4,5,6-TCG and TeCG in the brackish water fish species (Alburnus alburnus)as being less than 48 h. These results are of importance, because they indicate that there is little likelihood of biomagnification of chlorinated phenolics at higher aquatic trophic levels. 31 2.2.7. Fate of Chlorinated Phenolics in the Fraser River In the Fraser River upstream of Hope B.C., the major source of chlorinated phenolics are five BKMs, discharging approximately 15,000 kg of AOX and 56 g of chlorinated phenolics per day into the river (Schreier et al. 1991). In the heavily industrialized Lower Fraser Valley there are a wide variety of point and non-point sources of chlorinated phenolics, most commonly the CPs. The relative contribution of upstream BKM sources of chlorinated phenolics to the Fraser River Estuary is of the greatest overall importance, but varies seasonally. Carey and Hart (1988) found that during spring and summer high river flow periods, BKM derived chlorinated phenolics were present in higher concentrations in the Estuary. However, during winter low flow periods concentrations of chlorophenols in the Fraser Estuary area, the north arm of the Fraser River in particular, showed sporadic high episodes due to runoff from riverside lumber treatment fascilities. Fraser River waters and sediments show measurable levels of CPs (2,4,6-TCP, 2,3,4,6-TeCP, PCP), CCs (3,4,5-TCC, TeCC) and CGs (3,4,5-TCG, TeCG) throughout the year (Carey 1988; Dwernychuk et al. 1991). Water concentrations of 3,4,5-TCG in the Lower Fraser River were found to range from 0.03 ug/L at high flow to 0.07 pg/L at low flow (Carey and Hart 1988). TeCG concentrations were roughly 50% of those recorded for 3,4,5-TCG, which is in agreement with the estimated relative proportions of these compounds in bleachery effluents (Voss et al. 1980). Recent monitoring of Fraser River sediments (Dwernychuk et al. 1991) from both the middle and lower reaches revealed concentrations of 3,4,5-TCG, TeCG, 3,4,5-TCC and TeCC ranging from 1 - 19 ug/kg, however very little TeCP was detected in middle Fraser River sediments and levels up to 6 pg/kg were measured in Lower Fraser River sediments. A variety of fish species have been shown to bioconcentrate tri- and tetra-chlorinated phenolics. Both bottom fish (largescale sucker; Catostomus macrocheilus) and open water predatory fish (northern squawfish; Ptycocheilus oregonensis, rainbow trout), sampled from Prince George south 150 km to Marguerite, B.C., showed CP, CC and CG contamination in liver and white muscle 32 (Dwernychuk et al. 1991; Schreier et al. 1991). Liver concentrations were usually highest, ranging from 5 - 4 0 pg/kg for tri- and tetra-chlorinated phenols and even higher (20 - 200 ug/kg) for tri- and tetra-chlorinated guaiacols. Muscle tissue concentrations were generally an order of magnitude lower. Rogers et al. (1988a) detected 3,4,5-TCG, 4,5,6-TCG and TeCG in juvenile chinook salmon (Oncorhynchus tshawytschd), over wintering in the Middle Fraser, and 3,4,5-TCG and TeCG in the same species, in the Lower Fraser at Agassiz, B.C. Over wintering stages of salmonid species may be especially at risk, since they are present during the winter, when effluent dilution is at a minimum and environmental conditions are at an extreme. In the spring, Pacific eulachons (Thaleichthys pacificus) enter the Fraser River estuary to spawn. These fish were investigated for sex related differences in chlorinated phenolic bioconcentration by Rogers et al. (1988b). Male eulachons bioconcentrated statistically significant greater amounts of 3,4,5-TCG and TeCG in their liver and reproductive organs. In addition, chlorinated phenolic tissue concentration was found to increase in proportion to the residence time in the estuary. The toxicological significance of these findings to the pacific eulachon were not determined. At the average pH of the Fraser River of 7.8 (Carey 1988), chlorinated phenolics are present mostly in the ionized, water soluble form, except for 3,4,5-TCG, which is only about 40% dissociated at pH 7.8. Also, as explained earlier, neither photolysis or hydrolysis are expected to be significant degradative processes in the Fraser River. The suspended sediment concentrations in the Fraser River are relatively high, but vary considerably between high and low flow periods, ranging from 30 - 130 mg/L in the Lower Fraser at Hope, B.C. (Hall et al. 1991). The foe, calculated as 40% of the loss on ignition, is estimated to range from 1.6% at high flow to 6.0% at low flow (Carey 1988). The lower percentage of organic particulates during high flow is due to dilution with inorganic sediments introduced during the spring freshet. Partitioning of various CPs to the sediment phase in the Fraser River was found to be moderate (Ksw = 100 - 1000) by Carey (1988). Considering the chlorinated phenolics tend to be in the ionized, water soluble form at the ambient pH and that the 33 actual sediment load in the Fraser is low (30 - 130 ppm) relative to the water volume, it is expected that most of the chlorinated phenolic load would be carried in the water phase; environmental fate would probably be governed by processes affecting the water phase more than the sediment phase (Carey 1988). For similar reasons discussed above, it does not seem likely that a very large portion of the total chlorinated phenolic loading from BKMs would partition into the biota, which does not, however, imply that exposure levels would not be biologically significant. The primary mode of uptake of chlorinated phenolics by Fraser River biota would be direct absorption from the water, of undissociated and ion paired forms, rather than accumulation through the food chain (Carey 1988). Many aquatic organisms clear these compounds rapidly from their bodies, thus there may never be a build up of chlorinated phenolics. Overall, persistence of chlorinated phenolics in water, sediment and biota should be short, considering the relatively short hydraulic residence times in the Fraser River, the flushing action of the spring and early summer high flows and the rapid metabolism of chlorophenols by most aquatic organisms (Carey 1988). However, since there are constant inputs of chlorinated phenolics from the five upstream BKMs, there is still concern about the impact of long-term chronic exposure to the biota. Direct evidence of ecological impact of chlorinated phenolic pollution, from BKM sources, in the Fraser River has not been found. Benthic invertebrate sampling is done routinely on the Fraser River, at various locations above and below BKM outfalls. Shifts in benthic population structure to more pollution tolerant invertebrate species has been documented downstream of the Prince George mills (Schreier et al. 1991). However, it is impossible to link any observed effects to any one class of compounds present in BKM effluent, since there are so many other known toxins as well as uncharacterized components in BKM effluent. Cytochrome P-450 enzyme induction is another impact monitoring parameter which may be used in the future, but also suffers the same uncertainties. Assuming that there are some impacts attributable to chlorinated phenolics or BKM effluents in general, the broad ecological relevance remains uncertain. 34 The uncertaint ies concern ing the fate and e c o l o g i c a l effects o f ch lo r ina ted phenol ics , together w i th their measurable presence i n r ece iv ing waters suggest that this class o f compounds w o u l d make good targets for routine in situ m o n i t o r i n g . Though chlor ina ted phenol ics make up on ly a smal l f ract ion o f the total organic halides discharged from B K M s , they have been shown to bioconcentrate i n a l l classes o f aquatic organisms, from microbes to f ish species, sampled hundreds o f ki lometres from k n o w n point sources (Schreier et a l . 1991). Ch lo r ina ted phenol ics are also acutely toxic i n h igh concentrations (200 - 1000 u g / L ) ( M c L e a y 1987) and manifest a wide variety o f sublethal metabol ic , phys io log i ca l disturbances i n f ish species, such as altered l i ve r metabol i sm of b i l i r u b i n and steroid hormones, g lycogen metabol i sm imbalances and decreased growth rate ( O i k a r i et a l . 1988). M u c h focus on the envi ronmenta l effects o f ch lor ina ted pheno l ics has also lead to deta i led c h e m i c a l c h a r a c t e r i z a t i o n and deve lopmen t o f sound a n a l y t i c a l t echn iques , m a k i n g the ch lor ina ted pheno l ics r e l a t ive ly easy and cheap to moni tor i n a wide var ie ty o f env i ronmen ta l med ia . F o r m o n i t o r i n g programs focus ing on B K M sources o f chlorinated phenol ics , the C G s and C V s make especial ly p romis ing target compounds, since they are present i n re la t ive ly h igh concentrations i n B K M effluents and seem to be h igh ly specif ic to the B K M industry. Since the chloroguaiacols are readi ly t ransformed to t o x i c , l i p o p h i l i c ch lo rove ra t ro l e s , by bo th mic robes and h igher vertebrates, it is also prudent to include these derivatives in moni tor ing. Based upon studies focusing on discharge and fate o f chlorinated phenolics from B K M sources, i t is suggested that the f o l l o w i n g compounds w o u l d make good choices for rout ine moni tor ing: chloroguaiacols ( 4 , 5 - D C G , 3 , 4 , 5 - T C G , 4 , 5 , 6 - T C G , T e C G ) , ch lo rovan i l l i n (5,6-D C V ) , chloroveratrole ( 3 , 4 , 5 - T C V e r ) . 35 2.3. Biomonitoring Aquatic environmental pollutant monitoring can be as defined as the tracking of the spatial/temporal distribution of contaminants. Monitoring can identify the pollutants of concern as well as determine sensitive areas and time periods where and when potential environmental impacts may be manifested. Aquatic monitoring data aid in the assessment of the associated risks to both humans and affected ecosystems as well as the effects of management decisions concerning regulation of pollutant discharge, industrial process changes and site remediation effort. There are always many uncertainties and knowledge gaps to contend with in environmental decision making, therefore it is desirable to have a comprehensive data base from which to draw conclusions. With the high economic cost associated with comprehensive field monitoring, it is necessary to select monitoring techniques which yield the most pertinent information at the lowest cost. However, the variability and complexity of temperate aquatic systems conspire against this goal. The primary tool employed in aquatic pollutant monitoring is water sampling, often conducted by periodic single sample (grab sampling) collection. Direct water monitoring is, however, of low sensitivity, since organic contaminants are usually present in very low (ppb - ppt) concentrations. In the case of chlorinated phenolics discharged from BKM sources, target compounds are detected with only variable success, often only within several kilometres of the point source (Oikari et al. 1985; Metcalf and Hayton 1989). Further complications occur due to temporal variations in pollutant concentrations, arising from seasonal water flow changes and fluctuations in pollutant discharges. For example water flow on the Lower Fraser River varies, on the average, from 800 m /^sec to 5700 m /^sec between spring high flow and winter low flow periods (Hall et al. 1991), resulting in possible seven fold differences in water contaminant concentrations. Further complications arise, since pollutant distribution between water and suspended sediments may change as suspended sediment loads shift seasonally in quantity and character. Temporal variations in 36 point source pollutant discharge can also occur, on a da i ly , weekly or monthly basis depending upon the specific cause. B K M process changes, such as chlor ine d iox ide subst i tut ion, is an example o f a process change w h i c h can be implemented on a week ly basis, resul t ing i n wide variat ions i n chlor inated phenol ic output (Liebergot t et a l . 1990). Acc iden t a l spi l l s o f pu lp ing and b leaching filtrates as w e l l as whole effluent are examples o f more sporadic short term variat ions i n pollutant discharge. In the re la ted lumber indus t ry , sporadic release o f ch lo ropheno l s , f rom treated lumber was found to be related to on-site wash down act ivi t ies , w h i c h often took place at late hours i n the night, when water sampling might not have been l i k e l y to take place ( H a l l and Jacobs 1988). The above sources o f temporal variat ions i n contaminant water concentrat ions point out the necessity to careful ly consider the appropriate t i m i n g and frequency o f mon i to r ing . Comprehens ive moni to r ing also needs to take into account spatial d is t r ibut ion of target pollutants. In the case o f dynamic r iver systems, such as the Fraser, we may expect that the re la t ive ly water so luble chlor ina ted phenol ics , for the most part, t ravel downst ream w i t h the m a i n f l o w , un t i l reach ing the Fraser R i v e r Es tuary . H o w e v e r , there are numerous back water habitats and s loughs, w h i c h serve as impor tant aquatic habitats as w e l l as reposi tor ies for contaminated sediments . Spat ia l dis t r ibut ion o f r iver borne contaminants may also change wi th season, due to chang ing r i ve r f l ow dynamics . F o r example , under the i ce the dynamics and deposi t ional tendencies o f sediments under d i f fer ing f low condi t ions have not been adequately studied i n the Fraser R i v e r . W i t h the great spa t ia l / tempora l var ia t ions and c o m p l e x i t i e s , d i f f i c u l t i e s i n d e t e c t i n g l o w c o n c e n t r a t i o n s a l o n g w i t h the h i g h cos t a s s o c i a t e d w i t h comprehensive moni tor ing of aquatic systems, there is a need for a l o w cost method of generating time integrated results, cover ing , on the average, the whole spectrum of var ia t ions i n water pol lutant concentrat ions. A l s o needed is some method o f l i n k i n g water moni to r ing data to b i o l o g i c a l r i sk and effect. Th i s is not d i rec t ly achieved through s imple water moni to r ing , since as descr ibed i n sect ion 2.2.5. the uptake and effects o f many organic pollutants are dependent on environmental 37 variables other than water concentration. There is strong evidence which suggests that, as a compliment to standard water monitoring, biological organisms could be used as reliable indicators of B K M derived chlorinated phenolic distribution. This method, properly termed as biomonitoring, can broadly include measurement of both toxic effects on organisms as well as measures of pollutant levels in receiving waters. This thesis focuses on the use of leeches for in situ identification and spatial -temporal tracking of chlorinated phenolic contaminants. Biomonitors, as pollutant indicators, function by taking up contaminants in their tissues, the major advantage being that they can concentrate tissue levels far in excess of water concentrations, making it possible to detect contaminants which were too low to detect in the ambient water. Biological organisms are also effective time integrators of pollutant levels, since they are constantly taking up small samples of the ambient environment. This phenomenon is of the greatest importance when pollutant discharge or environmental conditions vary on a relatively short time scale. Living organisms also provide a measure of the bioavailability of a given compound, which may not remain constant under changing environmental conditions. This is important for compounds such as the chlorinated phenolics, which may exist in the more bioavailable non-dissociated form or the less bioavailable ionized form; depending on water pH. Also the presence of contaminant binding entities, such as organic particulates may decrease the effective concentration of a compound by competitive partitioning. This implies that a single water quality criterion may not be appropriate for every aquatic system and that site specific biomonitoring could aid in development of adjustable water quality cr i ter ia . There are numerous species from which to select a potential biomonitor of chlorinated phenolic contamination. Phillips (1978), in a comprehensive review of the use of bioindicator organisms to quantitate organochlorine contamination, provided a useful set of guidelines to aid in the selection of an appropriate 38 biomonitor. The selection criteria suggested that the potential biomonitor should be: 1) representative of the region to be monitored, 2) hardy enough to be maintained and tested under laboratory conditions, 3) long lived enough to allow sampling of more than a single age class, if desired, 4) large enough to provide an adequate sized tissue sample for analysis, 5) demonstrate a high bioconcentrating capacity, without being killed under a wide range of exposure conditions, 6) show a direct correlation between bioconcentration and average pollutant concentration. Phillips also suggested that the biomonitor should be tolerant of brackish water, which adds to the versatility of the biomonitor, but is not strictly necessary under freshwater conditions. To this list it is also important to add the additional characteristic of a slow rate of contaminant transformation and elimination, since the more a compound is retained in a biomonitor, the greater the bioconcentration and the better the correlation between water concentration and tissue concentration. It would also be desirable to select an organism which can be propogated at low cost, under laboratory conditions, in order to provide a homogeneous and uncontaminated population. Various aquatic species have been shown to bioconcentrate chlorinated phenolics in the field. Comprehensive sampling of an industrially polluted creek in Ontario, revealed marked differences in the bioconcentrations of chlorophenols in aquatic organisms (Table 2.6) (Metcalf et al. 1984). This study was especially informative, since there were a wide variety of organisms sampled, under identical field conditions, serving to illustrate that bioconcentration of chlorinated phenolics varies considerably between species and that leeches may be particularly sensitive indicators of chlorophenol contamination. Further investigations by Metcalf et al. 39 (1988), comparing biomonitoring potential the of three leech species, Dina dubia, Erpobdella punctata and Helobdella stagnalis, revealed that all three species of leech showed a high bioconcentrating capacity (BCF = 600 - 16700), slow elimination rates ( > 25 days) and final tissue bioconcentrations which were in the same relative proportion to exposure concentrations. The exceptions were two of the trichlorophenois (2,3,6-TCP, 2,4,5-TCP), which seemed to be preferentially bioconcentrated by all three leech species. Interspecies variability was observed in both bioconcentration and elimination. Bioconcentration varied between species, without an apparent pattern for the different chlorinated phenolics, while elimination was most rapid for H. stagnalis and slowest for E. punctata Table 2.6: Bioconcentration of chlorophenols by various aquatic organisms sampled from Canagagigue Creek, Ontario, Canada. Data from Metcalf et al. 1984. Bioconcentration Factor Organism 2,4-DCP 2,4,6-TCP 2,3,4,6-TeCP Leech (Dina dubia) 173800 6 8 7 2 7 122000 Leech (Erpobdella punctata) 163733 100405 145750 Aquatic Worms (Oligochaeta) 39200 9 5 9 5 9250 Dragonfly Larvae (Anisoptera) - 7 8 4 3 4 7 5 0 Caddisfly Larvae (Pvcnopsvche SD.) 2533 2 0 5 4 12750 Clams (Ferrissia sp.) - 2 7 0 3 7 5 0 Snails (Phvsa so.) - 3 9 6 1666 Crayfish (Orconectes DroDinauus) 16 52 2 5 0 Bull Frog Tadpole (Ftana catesbeiana) - 52 -Rock Bass (Ambloplites rupestris) 3 7 9 36 4 5 0 Aquatic mussel species, have been favoured as biomonitors of both metal (Tessier et al. 1984) and organic pollutants (Kauss and Hamdy 1985), because they have a high bioconcentrating capacity for certain of these pollutants and have well developed laboratory maintenance and monitoring protocols. In field comparisons 40 between the mussel species (Elliptio complanata) and the leech species (Nephelopsis obscura), downstream of a BKM source in the Rainy River, Ontario, the leeches proved the better biomonitor of chlorophenols, based on their higher bioconcentrating capacity and ability to indicate the relative proportion of chlorophenol congeners (Metcalf and Hayton 1989). Further investigations, conducted by Hall and Jacob (1988) showed N. obscura and a second leech species, Percymoorensis marmoratato be effective in providing time integrated monitoring information on the sporadic discharge of chlorophenols, from lumber storage facilities, into the Lower Fraser River. These studies also indicated interspecies differences in bioconcentration as well as hinting at differences due to leech size. Laboratory testing of the above pair of leech species also revealed the importance of environmental factors in biomonitoring. Leech bioconcentration was found to be positively correlated to water temperature and inversely correlated to water pH. Overall, one can conclude from the above investigations that leeches make good biomonitoring organisms and sensitive indicators of chlorophenol contamination. Based upon past success in using leeches as biomonitors of chlorophenols in both laboratory and field conditions, it seemed likely that leeches could be effective biomonitors of the other related classes of compounds, such as the chloroguaiacols, chlorovanillins and chloroveratroles. While several species of leeches are good chlorinated phenolic biomonitoring candidates, one species in particular, N. obscura, is readily available in the current study region, encompassing the Fraser River Watershed, is easy to maintain in the laboratory and has proven to be an effective biomonitor under both laboratory and field conditions. P. marmorata, though not tested as extensively as N. obscura, also makes a good biomonitoring candidate, since it is also available in the study region and has been used successfully in Fraser River chlorophenol biomonitoring. 41 2.4. Leech Biology Leeches are generally hardy organisms, able to tolerate a wide variety of environmental conditions. N. obscura, of the family Erpobdellidae, is a common cold water species of leech, inhabiting waters from the mid latitudes of Canada south to the northern United States and Rocky Mountain states (Linton et al. 1983). An extensive survey of leech species in Colorado (Herrmann 1970) revealed that N. obscura is generally associated with quiescent lentic waters of productivities ranging from oligotrophic to eutrophic. N. obscura has been found in waters ranging from 0.5 °C to 24 °C and pH 6.0 - 9.8 (Herrmann 1970; Metcalf and Hayton 1989). N. obscura has been found in waters of total ash and organic solids contents of approximately 10 - 1000 mg/L and 10 - 300 mg/L respectively, and total alkalinities of 10 - 100 mg/L (Herrmann 1970). N. obscura is a predator and scavenger, hunting aquatic invertebrates, such as snails (Gastropoda) and insect larvae (Chironomid, Coleopteran groups) and feeding on any dead organisms it can locate (Anholt 1986). Life history studies conducted by Davies and Everett (1977) on the species N. obscura, in Alberta, revealed that this species lives from 12 to 19 months in the wild and at least two different generations are present in any one period. The first consists of those individuals which hatch in the spring and are sexually mature by the next spring; these individuals may die off early in the summer. There is also a second generation of leeches, which hatch in the summer; these individuals are not mature until the following summer or fall and sometimes carry over another year before maturing sexually. The critical factor triggering breeding by N. obscura seems to be body weight. Further, the body weight at which any given leech species reaches maturity seems to vary with the environmental characteristics of the particular habitat that population lives in (Davies and Everett 1977; Peterson 1983). Therefore, leech generations may carry over to a third season if they did not reach an appropriate body weight to trigger sexual maturity by the second season. For information concerning identification of N. obscura refer to Appendix 1. 42 P. marmorata of the family Hirudinidae, is one of the widest ranging of North American leech species, being found from Alaska through to the mid northern United States (Sawyer 1972). P. marmorata has been found associated with both lotic and lentic habitats, but shows less tolerance for oligotrophic waters than N. obscura. P. marmorataaad has a narrower optimal temperature (5 - 25 °C) and pH (7.2 - 8.2) range, but has wider tolerance for dissolved inorganic (20 - 2800 mg/L) and organic (50 - 1000 mg/L) solids content and total alkalinity (10 -150 mg/L) (Herrmann 1970). Like N. obscura, P. marmorata preys on many aquatic invertebrates, including N. obscura, as well as being an opportunistic scavenger (Sawyer 1989). Breeding takes place somewhat later than for N. obscura, being triggered more by water temperature, than leech weight. Egg laying takes place in mid to late summer and the young hatch in the same year and overwinter in the lake (Sawyer 1989). For information concerning identification of P. marmorata refer to Appendix 1. 43 3. METHODOLOGY 3.1. EXPERIMENTAL ORGANISMS Leeches were collected in August 1990 and June and September 1991, from the littoral zone of Black Lake, located about 150 km east of Princeton, B .C. This lake did not support a fish population, due to winter anoxic conditions, but had a rich invertebrate population. Leeches were captured using 10 crn^ wire mesh traps (5 mm mesh) baited with calves liver. Typically 5 - 1 0 traps were placed along the shoreline at depths ranging from 0.5 - 2.5 m and then retrieved after 15 - 90 min. Two species of leech appeared to be common in the catch from the lake. These were tentatively identified in the laboratory as either Nephelopsis obscura Verrill (TV. obscura) and Percymoorensis [Haemopsis] marmorata '(P.marmorata) Davies, on the basis of external morphology, texture, behaviour and location of male and female gonopores (Appendix 1). Leech species identification was later confirmed as correct by Dr. R. Davies (Department of Biology, University of Calgary). In general, specimens of P. marmorata ranged to a larger size (0.5 - 6.0 g) than did N. obscura (0.1 g - 2.0 g). N. obscura was the more numerous of the two leech species, outnumbering P. marmorata by approximately 10:1 in the leech traps. Leech populations in Black Lake consisted of a mixture of different age classes, ranging from juveniles up to spawned out mature adults. Leeches were maintained in the laboratory, in 38 L aquaria containing aerated native lake water (depth = 5 - 10 cm) and a sand bottom littoral zone habitat. Leech loading ranged from 20 to 50 leeches per aquarium. Scattered rocks and driftwood were provided for secure hiding places. Water temperatures ranged from 10 - 15 °C. The leech stock was fed calves liver, free of chlorinated phenolics, at seven day intervals. P. marmoratawas observed to prey upon N. obscura, therefore, the two species were kept in separate aquaria. There was no observable mortality in the laboratory stocks of P.marmorata however, N. obscura populations suffered 20 - 40% 44 mortality during the month after capture. Mortality occurred mostly in the larger, sexually mature adult leech population. Black Lake is a small pothole lake, which is in an area isolated from any direct sources of BKM pollution, however its proximity to a highway was of concern with respect to potential chlorinated phenolic contamination. Laboratory analysis revealed that none of the test compounds were present in leech tissues. 3.2. LABORATORY STUDIES Laboratory leech bioassays were conducted to measure pollutant depuration rate and assess the effects of pollutant concentration, water temperature, pH, turbidity, leech weight and interspecies differences on bioconcentration. The general protocol followed throughout the laboratory bioassays is described below. Conditions specific to each bioassay accompany the results for each experiment. In keeping with earlier leech monitoring studies by Hall & Jacob (1988), which had indicated that steady state tissue concentrations of chlorophenols were reached after about seven days, a seven day exposure period was selected for the present study. The semi-static bioassay protocol used by Hall & Jacob (1988), where water was replaced completely once every 24 h, was followed. This exposure protocol is a compromise between strictly static bioassays, which are characterized by constantly decreasing pollutant concentrations and complex flow through bioassay systems, which allow for relatively constant exposure concentration. Laboratory bioassays were spiked with a 100 pL mixture of the following test compounds, at an appropriate concentration, using a methanol carrier; 4,5-dichloroguaiacol (4,5-DCG), 3,4,5-trichloroguaiacol (3,4,5-TCG), 4,5,6-trichloroguaiacol (4,5,6-TCG) and tetrachloroguaiacol (TeCG), 5,6-dichlorovanillin (5,6-DCV) and 3,4,5-trichloroveratrole (3,4,5-TCVer). Control bioassays were spiked with 100 pL of methanol carrier. Simultaneous exposure to all compounds was adopted over single compound exposure, since field studies by Carey and Hart (1988) indicted that aquatic 45 organisms are exposed to a mixture of chlorinated phenolics in the Fraser River and indeed many other BKM receiving waters. Following spiking, all bioassay waters were stirred vigorously with a glass rod. Leech bioassays took place in one or two litre volumes of water, contained within one or two litre Mason Jars, which had been rinsed three times with hexane. Since leeches are somewhat amphibious, and will try to climb out of bioassay jars, plexiglass covers (3 mm thickness) with 9 air holes (1.5 mm i.d.) drilled through the surface of the plastic, were used to cover jars. Leech loading into bioassay chambers ranged from about four to six grams of leeches (N. obscura, n = 5; P marmorata, n = 2 - 3) per 1 L water. Bioassay water was made from reconstituted laboratory distilled water of medium - hard formulation. Bioassay water was made up using the following analytical grade chemicals, supplied by the BDH Chemical company; NaHCO^ (110 mg/L), CaS0 4 • 2 H 2 0 (60 mg/L), MgS0 4 • 7H 2 0 (120 mg/L) and KC1 (4 mg/L). Water pH of this formulation ranged from 7.7 - 7.9, but was adjusted to the desired pH for each bioassay by drop-wise addition of 2M HC1 and 5M KOH, which had been pre-extracted twice with hexane. Temperature control was maintained by conducting bioassays in a temperature control chamber. Vancouver City well water was provided by the Department of Fisheries and Oceans, for the flow through experiments. Turbidity bioassays followed a slightly different protocol from other laboratory bioassays. In order to maintain the sediments in each bioassay chamber in the suspended form, bioassay water was continuously stirred using a magnetic stirring apparatus. The leeches appeared to be irritated by the continuous stirring action, producing more mucous than in previous unstirred bioassays. Therefore, a control bioassay was run without any stirring action, in order to determine the effect of this disturbance on leech bioconcentration. Turbidity experiments were designed to mimic the composition and range of turbidities (18 - 147 mg/L) encountered in the Middle Fraser River throughout one seasonal cycle (Hall et al. 1991). Turbidity bioassays focused on the particle size fraction of < 63 um, which is thought to represent the most significant adsorptive surface for organic pollutants (Voice et al. 1983: Schellenberg et al. 1984). Field 46 sampling indicated that the organic fraction of Middle Fraser River accounted for about 4 - 7% percent of total suspended sediments. Therefore we selected a suspended sediment composition of 95% inorganic to 5% organic content for bioassays. To model the character of Fraser River suspended sediment load more closely , an X-ray diffraction analysis of the < 63 pm fraction was carried out by Dr. Les Lavkulich, of the University of British Columbia Soil Science Department. With X-ray diffraction it was possible to identify and estimate the relative amounts of the clay minerals present in sediment samples. The procedure is able to measure the spatial orientation of atoms in a three dimensional crystal lattice, which is unique to each clay mineral. The principle clay components were chlorite and illite, in approximately 60% : 40% ratio. For the purposes of our bioassay we selected a 1:1 ratio of chlorite to illite to represent the inorganic fraction of the suspended sediments. Both chlorite and illite were supplied, in gravel form, by the U.B.C. Soil Sciences Department. Clays were washed with distilled water and methanol solvent and then processed into a fine powder by pulverization in a ring grinder at the U.B.C. Geology Department. The selection of an appropriate source of organic material for spiking into the turbidity bioassays is not straight forward. The exact nature of the organic content of suspended sediments in the Fraser River downstream of Prince George is difficult to assess and probably varies with seasonal changes in environmental conditions. Both anthropogenic and natural sources account for suspended organic particles. Municipal and industrial discharges of biosolids and woody plant material, as well as plant material entering the system from surface runoff and stream bank erosion are examples of outside sources. Natural populations of aquatic bacteria, planktonic plants and animals are important internal sources. In order to simplify the bioassays, we selected a single organic component for laboratory bioassays. Dried, crushed alder tree leaves (Alnus rubra) (< 63 pm) were selected to represent the organic fraction of suspended sediments. The leaves were collected in green form from trees located on the University of British Columbia Endowment Lands, dried at 104° C and ground into a fine powder, using a mortar and pestle. 47 Suspended sediment s p i k i n g solut ions o f k n o w n concentra t ion were prepared as carriers to introduce sediments to leech bioassays. Appropr ia te amounts o f the inorgan ic c lay so lu t ion o f (0.05 g / L ) and the organic so lu t ion (0.01 g / L ) were introduced to bioassay jars pr ior to chlor inated phenolic sp ik ing . Bioassays wi th an inorgan ic /o rgan ic par t ic le rat io o f 9 5 % - 5% were conducted at total suspended sediment levels o f 0.025 g / L (4.5 - 6.0 nephlometric turbidity units - N T U ) , 0.08 g /L (18 - 22 N T U ) , 0.15 g /L (34 - 39 N T U ) . A bioassay of 100% inorganic clay content was run at 0.08 g /L (9 - 15 N T U ) . Contro l bioassays using clear bioassay water (0.5 - 1.0 N T U ) and untreated turbid water 0.08 g /L (19 - .45 N T U ) were also conducted. 3.3. F I E L D S T U D I E S 3.3.1 Study Area The Fraser R i v e r near P r ince George i n central plateau reg ion o f B r i t i s h C o l u m b i a , Canada, was selected for thi§ moni tor ing study, because o f its p rox imi ty to three B K M s discharging direct ly into the Fraser R i v e r (Figure 3.1). A l l three Pr ince G e o r g e B K M s p rac t i ce p r i m a r y c l a r i f i c a t i o n and secondary ef f luent t reatment (aerated l agoon) ; two o f the m i l l s , Can fo r and In tercont inenta l , share a s ing le effluent out fa l l . Under pressure from new p rov inc i a l government p o l i c y in i t i a t ives , a l l m i l l s were unde rgo ing major b l e a c h i n g process changes , i n the f o r m o f inc reas ing C I O 2 subst i tut ion, w h i c h general ly results i n reduced A O X format ion i n the first ch lor ina t ion stage of b leaching. N o B K M sources are present upstream o f the Pr ince George m i l l s . The moni tor ing area covers about a 52 k m stretch o f the Fraser R ive r , north from Shelley, B . C . down to Stoner Creek. W i t h i n this reach o f the Fraser there is the entry o f a major tributary, the Nechako R i v e r at Pr ince George and an area o f strong water m i x i n g at R e d R o c k Canyon, about 30 k m down stream of P r i n c e George . Contro l sampling took place at Shelley B . C . , about 6 k m upstream of the 48 F i g u r e 3 . 1 : F r a s e r R i v e r b i o m o n i t o r i n g s t u d y a r e a . 49 Northwood BKM, where a Water Survey of Canada sampling station is located. The site is located near an abandoned sawmill facility, which could be a potential source of tri-, tetra- and pentachlorophenol, however there should not be any of the pulp mill specific chlorinated guaiacols, vanillins or veratroles associated with this site. Upstream of the Shelley site are several other active sawmill yards which are potential contributors of chlorophenol contamination. For this reason two chlorophenol compounds, 2,4,6-TCP and 2,3,4,6-TeCP were monitored at the field stations. One sampling station was set up at Shelley, on the east side of the River, in a zone of moderate water flow. The downstream monitoring station was located at Stoner B.C., about 40 km downstream of Prince George. This site was chosen since it had been used for routine monitoring conducted by the British Columbia Ministry of the Environment. Since the site is about 10 km below a strong mixing zone (Red Rock Canyon), it was hoped that water from the Nechako and Fraser Rivers along with the BKM effluent would be completely mixed upon reaching the Stoner station. Sampling stations were set up on both sides of the River, about 1.5 km downstream of the entry point of Stoner Creek to the Fraser. This distance was beyond the reach of any back eddy generated by the confluence on the Fraser River and Stoner Creek. 3.3.2. Study Periods In order to evaluate the leeches under a wide variety of environmental conditions, monitoring trials were conducted during three different seasonal periods: summer (July 8-15, 1991), fall (October 17-24, 1991) and winter (February 19-26, 1992). Monitoring was carried out during each trial for seven day periods, based on earlier findings, indicating that leeches attained steady state tissue concentrations of chlorophenols after seven day exposures at 4 and 12 °C (Hall and Jacob 1988). Environmental parameters such as water temperature, pH and turbidity were monitored at the beginning and end of each monitoring trial. Water turbidity 50 samples were collected in 50 mL glass screw-cap tubes with a teflon liner and taken back to the laboratory for turbidity measurement using a HACH Model 2100A turbidometer (HACH Chemical Co., Ames Iowa, U.S.A.). Turbidity was measured in nephlometric turbidity units (NTU). 3.3.3. Sampling Procedures Effluent samples were collected (200 mL/day) from the Canfor/Intercontinental and Northwood mill outfalls, by mill employees, during five days of every week long monitoring trial and composited into one single five day sample from each of the two outfalls. One litre, solvent rinsed amber bottles were used for sample collection and storage. All samples were shipped back to the laboratory about eight days after the collection of the first sample, in ice packed coolers and stored in a dark cold room at 4 °C in the laboratory. Upon receipt, effluent samples were either extracted immediately or preserved with concentrated H2SO4 (2 mL/L of effluent) for later analysis. Water samples were collected using both grab sampling and automatic water sampling techniques. Automatic water sampling was carried out at the Shelley control site in July 1991 and at Stoner in both July and October 1991. One automatic water sampler broke down before the October trial so grab sampling was used at the Shelley station from this point onward. Below freezing air temperatures may impair sampler electronics and mechanics as well as causing blockage due to frozen water, so only grab samples were collected in February. Winter water samples were collected on only four of seven days at both stations, namely February 19, 21, 24 and 26. The automatic water sampler, an ISCO Model 2900 (ISCO Inc., Lincoln, Nebraska, U.S.A.) consisted of a 10 m teflon lined sampling tube, which was attached to a battery (12 volt NiCad) operated programmable peristaltic pump. The samples are deposited into 24, 500 mL plastic bottles. Previous research by Jacob (1986) showed no 51 significant difference in the quantities of chlorinated phenolics between water samples stored in the plastic bottles or standard amber glass sampling jars. The automatic water sampler was programmed to take a 400 mL sample from a depth of about 0.5 m, every 8 h. Therefore, over each seven day sampling run a total of 21 water samples where collected. To minimize cross contamination between samples, a purge cycle was programmed prior to and after each sampling. To preserve the samples concentrated H2SO4 (2 mL/L of water) was added to each bottle prior to sampling. In the laboratory water samples were composited for analysis into seven 1200 mL water samples, representing one sample every 24 h period. Grab samples were collected from surface waters in solvent rinsed 1 L amber glass bottles and preserved in concentrated H2SO4 (2 mL/L of water). All water samples were shipped back to the laboratory in ice packed coolers and stored in a dark cold room at 4 °C. Suspended sediments were sampled at only the downstream Stoner test site alone in July 1991 and at both Shelley control and Stoner test sites in October and February 1992. No suspended sediments were recovered from the October trial, due to the disappearance of the sediment traps. Suspended sediment traps, supplied by the B.C. ministry of Environment, Prince George branch, consisted of four plexiglass tubes, 30 cm long, with a 4 cm wide opening at one end, fixed in a plastic holder. The trap was suspended at about a i m depth, from a float with the long axis of the tubes perpendicular to the flow of the water and the openings facing the water surface. The traps were then anchored to the river bottom with two approx. 15 kg concrete blocks. Where possible the concrete blocks were also tied off to shore. It should be noted that the dynamics of this particular type of sediment trap change as the trap collects sediment, resulting in a slightly biased sample. As the sediment collection tubes become full, the force of the downward eddy, created at the mouth of the tube decreases, resulting in the selective sampling of progressively heavier sediment fractions as the tube fills. At the bottom of the trap there tends to be a more representative sample, while at the top there is a bias towards larger sediments. This phenomenon becomes significant during medium and high flow periods on the 52 Fraser, when sediment loads are high and sediment traps attain nearly full capacity after a seven day period. One solution would be to remove sediments from the trap at short intervals rather than at the end of the sampling run, however this was not possible, due to logistical difficulties. For Fraser River biomonitoring, leeches were housed in cylindrical stainless steel wire mesh (< 1.0 mm mesh size) cages, measuring 15 cm by 6 cm dia. Leeches (n = 10/cage) were suspended from anchored floats at depths from 0.5 - 1.0 m. Control leeches were placed near the east side of the Fraser at Shelley. Test leeches, at Stoner, were stationed near both east and west sides of the River in July and October, to test for across river differences in bioconcentration. In February leeches were only stationed at the east side of the River at Stoner. N. obscura was used for all monitoring trials and P. marmoratav/as used as a measure of interspecies variability at the Stoner site during the July trial. All leech samples were frozen and shipped on ice to the laboratory about 24 h after the end of the bioassay. In the laboratory leeches were wrapped in solvent rinsed aluminium foil and frozen at -15 °C, until analysis. 3.4. ANALYTICAL PROCEDURES 3.4.1. Chemicals and Reagents Hexane and methanol organic solvents used in preparation of gas chromatography (GC) standards, spiking standards and for extraction of chlorinated phenolics solvents were all of gas chromatographic pesticide grade and purchased from BDH Chemical Co. Reagents used in extraction and derivitization (K2CO3, acetic anhydride, KOH and HC1) were analytical grade and purchased from BDH Chemical Co. All reagent solutions were pre-extracted twice with hexane in order to remove any remaining organic contaminants. Compounds selected for laboratory and field assay included 4,5-53 dichloroguaiacol (4,5-DCG), 3,4,5-trichloroguaiacol (3,4,5-TCG), 4,5,6-trichloroguaiacol (4,5,6-TCG) and tetrachloroguaiacol (TeCG), 5,6-dichlorovanillin (5,6-DCV) and 3,4,5-trichloroveratrole (3,4,5-TCVer). In addition two chlorophenol compounds, 2,4,6-trichlorophenol (2,4,6-TCP) and 2,3,4,6-tetrachlorophenol (2,3,4,6-TeCP) were selected for analysis in field investigations. Analytical standards, for laboratory analyses of the chlorinated guaiacols, vanillin and veratrole were obtained in 99%+ pure form, from Helix Biotech in Richmond, B.C. Chlorophenols, surrogate standard (2,4,6-tribromophenol) and internal standard (2,6-dibromophenol) were ordered from the Aldrich Chemical Co. (Milwaukee, Wisconsin, U.S.A.). Stock solutions of chlorinated phenolics were prepared for use as GC reference standards and spiking solutions. Individual stock solutions of each compound, ranging from 1000 pg/mL to 3000 pg/mL, were prepared in methanol. The moderate polarity of methanol made it a good choice for preparation of the standards, since it allowed for easy dissolution of the slightly polar chlorinated phenolics and acted as a good intermediary solvent for the introduction of organic analytes to water based extractions and bioassays. From individual stock solutions, mixed spiking standards containing all the test compounds were prepared at concentrations of 10 ug/mL and 1.0 ug/mL. All stock solutions and spiking standards were stored in teflon-sealed amber screw cap vials at -15 °C. Standard solutions were never used for more than 6 months. 3.4.2. Extraction Procedures 3.4.2.1. General Rational Pulp mill effluent, water, sediments and leech extractions were all carried out following the same general method with certain modifications specific to each sample phase. Since chlorinated phenolics are in the completely ionized form at pH > 12, the samples were first extracted into a basic solution of hCjCOg. The neutral 54 veratrole remained unchanged i n this step. Due to their h igh polar i ty and l o w vapour pressure, under iv i t i zed chlor inated phenolics are not part icularly amenable to capi l la ry G C analysis (Morales et a l . 1992). G C chromatograms o f pure phenol ic compounds often reveal asymmetric peak shape and peak ta i l ing and poor separation. Samples were der iv i t ized by acetylat ion, us ing acetic anhydr ide . The a lka l ine med ium, used to extract the ch lor ina ted phenol ic compounds , ionizes the compounds to the more react ive phenolate form, w h i c h is susceptible to attack by h igh ly reactive acetyl groups. However , this procedure may not be sui table for a l l classes o f ch lo r ina ted pheno l i c s , s ince the ch lo r ina t ed catechols are especial ly susceptible to oxida t ion in a lkal ine solut ions, resul t ing i n the format ion o f the respective qu ino id compounds (Starck et a l . 1985, M or a l e s et a l . 1992). Neu t r a l compounds such as ch lor ina ted veratroles are unaffected by the ace ty la t ion reac t ion . M e t h y l a t i o n , u s ing d iazomethane gas is another c o m m o n d e r i v i t i z a t i o n r eac t i on used i n c h l o r i n a t e d p h e n o l i c ana lys i s ( K e n n e d y 1989) . H o w e v e r , m e t h y l a t i o n was not the method o f cho i ce i n our case, s ince the methyla t ion o f 3 , 4 , 5 - T C G produces 3 , 4 , 5 - T C V e r , mak ing it imposs ib le to d is t inguish between these two compounds. In addi t ion, diazomethane is known to be a h igh ly potent ca rc inogen . F o l l o w i n g ace ty la t ion , the ch lo r ina ted pheno l i c acetates were extracted w i t h the o rgan ic so lvent hexane and concentra ted to an appropr ia te vo lume before G C analysis w i th electron capture detect ion ( E C D ) . The e lect ron capture detector i s h i g h l y sens i t ive to compounds w i t h h i g h e lec t ron af f in i t ies ( M o r a l e s et. a l . 1992) and a l l ows for the lowes t poss ib le detect ion l i m i t s for ch lor ina ted phenol ic compounds , w h i c h is o f impor tance when dea l ing w i t h very l o w e n v i r o n m e n t a l concen t ra t ions . 3.4.2.2. Eff luent Sample Preparat ion Bleached kraft m i l l effluent ( B K M E ) week ly composite samples were extracted by a modif ied form of the procedure recommended by Starck et al . (1985). This 55 method was selected because of its general high recovery of chlorinated phenolics (> 80 %) and relatively low variability (< 30%). Effluents were mixed to resuspend any settled solids prior to extraction. A 50 mL volume was measured into a 100 mL graduated cylinder and decanted into a 250 mL separatory funnel. Residual solids were backwashed from the graduated cylinder with 20 - 30 mL of effluent from the separatory funnel. Samples were spiked with 2,4,6-TBP surrogate and mixed by hand. Effluent pH, measured using EM Scientific pH 0 - 14 pH indicator paper, generally ranged from pH 6 - 7 prior to extraction. The pH of the samples was adjusted to pH 8, using 5 M KOH solution. 5M K 2 C 0 3 (1 mL/50 mL effluent) was added to each sample to make a final K2CO3 concentration of 0.1 M. Effluents were acetylated by addition of acetic anhydride (1.0 mL/50 mL effluent) to the separatory funnel, followed by vigorous manual shaking and venting of gas for 60 sec. Samples were then allowed to stand for 15 min. Acetylated chlorinated phenolics were then extracted by the addition of 10 mL of hexane, followed by vigorous manual shaking for 3 min. Vigorous extraction generally led to emulsion of hexane, which was broken up by the addition of 10 drops of methanol, followed by centrifugation at about 3000 rpm. for 15 min., in conical glass centrifuge tubes covered with teflon caps. The hexane extracts were then transfered into 15 mL graduated test tubes and concentrated under a gentle stream of nitrogen gas, in a 30 -35 °C water bath. Extracts were stored in the dark at -15 °C, prior to GC analysis. 3.4.2.3. Water Sample; Preparation Water samples were extracted using a procedure similar to that used for effluent extraction, but at a larger scale. Samples were mixed vigorously to resuspend any sediments on the bottom of the sample bottles. Whole, unfiltered water samples were extracted, unless other wise indicated. Water samples (1 - 1.2 L) were spiked with surrogate adjusted to pH > 12 with 5M KOH solution and extracted with 100 mL of hexane to remove base neutral compounds. This basic extraction 56 served to clean up the samples, allowing for greater concentration of the final extract; this solvent fraction also contained the neutral compound 3,4,5-TCVer. Waters were then neutralized to pH 7 using 2M HC1, and extracted in 2.0 L separatory funnels, by the same method per effluent samples, with the exception that three sequential hexane (100 mL) extractions were carried out. No significant emulsion problems were encountered with either Fraser River water or laboratory bioassay water. To break the minor emulsions encountered, 1.0 mL of methanol was added drop-wise to the sample in a 500 mL Ehrlenmyer flask, while swirling the extract. Base neutral and acetylated acid hexane fractions were filtered separately through #4 Whatman Filter Paper, into 500 mL round bottom flasks. Hexane fractions were concentrated by roto-evaporation at 40 °C, transferred to 15 mL graduated test tubes, further concentrated under nitrogen gas and stored in the dark at -15° C, for later GC analysis. 3.4.2.1. Sediment Sample Preparation Frozen (-15 °C) sediment samples were thawed, homogenized by mixing with a solvent rinsed metal spatula and weighed into 500 mL teflon screw cap bottles. Sub-samples were dried at 104 °C for percent moisture determination. Sediments were rotary extracted (10 rpm) in 250 mL of 0.1M rv jCOg for 15 h. It was found that 3,4,5-TCVer was not recovered in this procedure, so no sediment results for this compound are reported. h ^ C O g extracts, containing ionized chlorinated phenolics, were centrifuged at about 3000 rpm. for 15 min., in conical glass centrifuge tubes covered with teflon caps, to separate suspended sediments from the aqueous fraction. K 2 C O g extracts were decanted into 2.0 L separatory funnels, extracted, concentrated and stored as per water samples. 57 3.4.2.4. Leech Sample Preparation Leeches used in the monitoring of chlorophenols by past researchers (Metcalf et al. 1984; Jacob and Hall 1988) have been extracted using a procedure involving the following steps; acid digestion of leech tissue in concentrated HC1, sequential hexane extraction, back-extraction into O.IM hC^COg, acetylation and hexane extraction of chlorophenol derivatives. However, early attempts to apply this method to the chlorinated guaiacols and chlorinated vanillins resulted in low analyte recovery (< 50%). It was necessary to come up with a modified method of chlorinated phenolic extraction. Two different extraction methods were developed, both eliminating the need for a back-extraction from hexane to K2CO3; 1) the acid digest method, 2) the polytron extraction method. 1) The acid digest method: frozen (-15 °C) leeches were thawed, blotted dry with a paper towel and weighed to the nearest 0.001 g. Leeches were placed into 40 mL amber screw-cap tubes with teflon lined caps, spiked with 2,4,6-TBP surrogate and digested for 2.5 hours in concentrated HC1 (1 mL/0.2 g leech). Digests were decanted into a 250 mL separatory funnel and neutralized to pH 7 with 5M KOH (approx. 7 mL). Samples where then made up to a volume of 50 mL with pH 7 laboratory distilled water. 1 mL of 5M K2CO3 per 50 mL of sample was added. Samples were derivitized by addition of 1.2 mL acetic anhydride, followed by manual shaking for 90 sec. Samples were allowed to stand for 15 min. after which they were extracted with 10 mL of hexane for 3 min., by vigorous manual shaking. Samples generally emulsified after hexane extraction and had to be centrifuged as per effluent samples. 2) The polytron method: frozen (-15 °C) leeches were thawed, blotted dry with a paper towel and weighed to the nearest. 0.001 g. Leeches were placed into 40 mL test tubes with 2.0 mL of 0.1M K2CO3 surrogate spiked with 2,4,6-TBP. Leeches were then pre-extracted for 30 seconds, using the grinding action of a Brinkmann Model PT 10/35 polytron, fitted with a Model PT 10-TS probe. The polytron is supposed to break up 58 tissue at the cellular level, releasing cell contents into the K2CO3 extraction medium. Another 18 mL of 0.1M K2CO3 was added to the test tube and the sample was extracted for another 90 sec. at medium , speed. The extract was decanted into a 250 mL separatory funnel. The test tube was then rinsed with 30 mL of 0.1M K 2 C 0 3 , which was added to the separatory funnel. The resulting 50 mL volume of 0.1M K2CO3 leech extract was acetylated and extracted as per the above acid digest method. 3.4.2.5. Reference Standard Preparation Fresh acetylated reference standards, ranging from 0.01 - 0.1 pg/ml were prepared from 10 pg/ml mixed stock solutions, for GC analysis of each series of samples from a given bioassay. Standards were prepared by spiking 0.1 - 1.0 pg from mixed stock solutions of chlorinated phenolics and surrogate into a separatory funnel containing 50 mL of 0.1M K2CO3. Standards were acetylated by adding 1.0 mL of hexane extracted acetic anhydride to the separatory funnel, followed by vigorous manual shaking and venting of gas for 60 sec. Samples were then allowed to stand for 15 min. Acetylated chlorinated phenolics were then extracted by the addition of 10 mL of hexane, followed by vigorous manual shaking for 3 min. Hexane extracts were collected into 15 mL screw-cap tubes and stored in the dark, at -15 °C. 3.4.3. Instrumental analysis Acetylated chlorinated phenolic and veratrole compounds were quantitatively analysed by capillary gas chromatography on a Hewlett Packard model 5880A GC, equipped with an electron capture detector (temperature 310 °C) and a model 7672A automatic sampler and a splitless injection port (temperature 250 °C). Helium carrier gas (60 - 70 cm/second flow rate) and nitrogen detector make-up gas (30 - 35 cm/second flow rate) were used. The following oven temperature program was successful in separating all test compounds: 75 °C (3 min.) to 120 °C at 15 °C/min.; 59 120 °C (0.1 min.) to 200 °C at 3 °C/min.; 200 °C (5 min.) to 220 °C at 5 °C/min.; 220 °C (0.1 min.) to 265 °C at 20 °C/min. for 4 min. A J & W Scientific 30 metre DB-5 capillary column (i.d. 0.32 mm), supplied by Chromatographic Specialties Company of Brockville, Ontario, provided satisfactory separation of all eight test compounds as well as a number of other chlorinated phenolic compounds (Figure 3.2). The apparent detection of trace amount of organic compounds using a single chromatographic column is not generally accepted as strong enough evidence of their presence, especially in unknown field samples. The current method of choice for confirmation of trace organic contaminants is mass spectrometry (MS), however we were unable to make use of this method, since the levels of contaminants we were dealing with were generally far below the limits of detection of the MS system available to us. Therefore, we chose to confirm our findings on a second GC capillary column of greater stationary phase polarity; a J & W Scientific 30 metre DB-1701 (i.d. 0.32 mm) also supplied by Chromatographic Specialties Company (Figure 3.3). Results were calculated according to the response factor of each compound to the 2,6-dibromophenol internal standard (Appendix 2). 3.4.4. Quality Control Most glassware used in this research was kept aside for the exclusive use of this project, in order to prevent undue contact with outside sources of contamination. Before glassware of unknown origin was used it was fired in a muffle furnace at 530 °C for 60 minutes to remove volatile organic compounds. Prior to extractions all glassware and equipment was first rinsed in distilled water, followed by one methanol rinse and three hexane rinsings. Glassware rinsings were periodically analysed for chlorinated phenolic contamination by GC-ECD. Chlorinated phenolics have been shown to be susceptible to photodegradation (Tratnyek and Holgne 1991), therefore all sample extracts were stored in the dark at -15 °C and GC analysis was carried out using amber coloured GC vials. 60 Figure 3.2: Separation of chlorinated phenolic compounds on a 30 m DB-5 capillary column. 1) 2,4,6-Trichlorophenol 2) 2,6-Dibromophenol (internal standard) 3) 4,5-Dichloroguaiacol 4) 2,3,5,6-Tetrachlorophenol 5) 2,3,4,6-Tetrachlorophenol 6) 3,4,5-Trichloroveratrole 7) 2,3,4,5-Tetrachlorophenol 8) 3,4,5-Trichloroguaiacol 9) 2,4,6-Tribromophenol(surrogate standard) 10) 4,5,6-Trichloroguaiacol 11) 5,6-Dichlorovanil l in 12) Pentachlorophenol 13) 3,4,5-Trichlorocatechol 14) Tetrachloroguaiacol 15) Tetrachlorocatechol 61 Figure 3.3: Separation of chlorinated phenolic compounds on a 30 m DB-1701 capillary column. 1) 2,4,6-Trichlorophenol 2) 2,6-Dibromophenol (internal standard) 3) 2,3,5,6-Tetrachlorophenol 4) 2,3,4,6-Tetrachlorophenol 5) 4,5-Dichloroguaiacol 6) 3,4,5-Trichloroveratrole 7) 2,3,4,5-Tetrachlorophenol 8) 2,4,6-Tribromophenol (surrogate standard) 9) 3,4,5-Trichloroguaiacol 10) 4,5,6-Trichloroguaiacol 11) Pentachlorophenol 12) 5,6-Dichlorovanil l in 13) Tetrachloroguaiacol 14) 3,4,5-Trichlorocatechol 15) Tetrachlorocatechol 62 Background levels of contamination and chlorinated phenolic recovery where monitored through extraction and analysis of both blank control and spiked sample matrices. The exception was bleached pulp mill effluent, where no blank effluent matrix was available, however recovery could be gauged by spiking of a subsample of effluent of known contamination level and subtraction of the known chlorinated phenolics levels in the sample from the total concentration after spiking. Quality control of water samples was monitored using laboratory distilled water; the spiking level used was 1.0 ug/L. For sediments hexane extracted beach sand from the Similkameen River at Bromley Rock provincial park, British Columbia was used. Matrix spikes of Fraser River sediments of known chlorinated phenolic concentration where also examined. Sediment spiking level was 0.1 ug/g. Quality control of leech extractions was carried out by analysis of untreated laboratory leeches and spiked leech tissue samples; spiking level was 1.0 ug/g. To further assess extraction performance, surrogate compounds were spiked into each sample prior to extraction. Monitoring of surrogate recovery in each sample provided a warning signal or internal check to assess the success of the derivitization and extraction. The surrogate of choice should be a compound which is chemically, closely related to the target compounds, not likely to be already present in any sample and does not interfere in the analysis. An accepted surrogate for chlorinated phenolic analysis is the compound 2,4,6-tribromophenol, a halogenated phenolic of intermediate substitution level (Morales et al. 1992). Surrogate spiking level for each sample depended upon the expected final volume of each sample, but generally ranged from 0.25 - 1.0 ug/L for water and effluent and 0.25 - 1.0 ug/g for sediments and leeches. To assess and correct for the injection and overall GC-ECD performance an internal standard was added to each sample prior to GC analysis. A good internal standard should have the same qualifications as outlined for a surrogate compound. On the basis of method development research by Xie (1983), 2,6-dibromophenol was selected as an internal standard. Since this compound is a polar phenolic, it was necessary to pre-derivitize a batch of 2,6-dibromophenol (1.0 ug/L). Variability 63 between batches of internal standard was not measured, but as long as the same batch was used for any single set of samples and standards there was no overall effect on the final result. However qualitative examination did not reveal any variation greater than approximately 20% between batches of internal standard. The method detection limit (MDL) was estimated for each sample matrix (Table 3.1). The range in MDL was dependent upon a number of factors including; sample weight or volume, concentration of extracts, background level of interference and sensitivity of the analyte to the electron' capture detector. It was necessary to calculate detection limits for each individual compound of interest and modify them according to the extraction and GC running conditions used. Effluent MDLs ranged from 0.1 ug/L for tri- and tetra-chlorinated compounds to 0.5 pg/L for di-chlorinated compounds. Water MDLs were about ten times lower than for effluents, since we analysed greater volumes of water and they where cleaner than the effluents. Sediment and leech MDLs ranged from 10 and 20 pg/kg for dichlorinated compounds, down to 0.2 and 2.0 pg/kg for tri- and tetra-chlorinated compounds in sediments (dry weight) and leeches (wet weight) respectively. Generally, the main factor limiting the MDL was the cleanliness of the organic solvent extracts. Effluent and sediment samples contained many organic interferences. Leech samples contained lipid compounds, which precluded analysis of highly concentrated samples. Column chromatography, using silica gel or florisil was not used to cleanup extracts, because of poor recovery of target compounds. Gel permeation chromatography (GPC) is the method of choice for lipid removal, however the large number of samples and the lack of access to GPC equipment precluded its use. However, at the water concentrations bioassayed in this study, interference from lipid compounds proved to be of minor importance in laboratory and field bioassays using the leech species Nephelopsis obscura. Greater lipid interference was encountered in the leech species Percymoorensis marmorata . More highly chlorine substituted molecules such as tri- and tetra-chlorinated phenolics are more responsive the ECD detector. Therefore, the less chlorine substituted dichlorinated compounds (4,5-DCG, 5,6-DCV) were limited by their ECD sensitivity. 64 Table 3.1: Method detection limit ranges for the analysis of chlorinated phenolics in effluent, water, suspended sediments and leeches. Detection limits based on sample dry weight for suspended sediments and sample wet weight for leeches. Suspended Leeches Leeches Compound Effluent Water Sediments (N. obscura} (P. marmorata) (ug/L) (u-g/kg) (ng/kg) (ng/kg) 4,5-DCG 0.5 0.02 - 0.05 8.0 - 12.0 3.0 - 5.0 15 - 30 3,4,5-TCG 0.1 0.002 - 0.005 0.3 - 1.0 1.0 - 3.0 15 - 30 4,5,6-TCG 0.1 0.002 - 0.005 0.2 - 1.0 1.0 - 3.0 5.0 - 15 TeCG 0.1 0.002 - 0.005 0.3 - 1.0 1.0 - 3.0 5.0 - 15 3,4,5-TCVer 0.2 0.002 - 0.005 NA 1.0 - 3.0 5.0 - 15 5,6-DCV 0.5 0.01 - 0.05 1.0 - 5.0 20 - 50 100 - 200 2,4,6-TCP * 0.2 0.005 - 0.01 0.5 - 2.0 1.0 - 3.0 5.0 - 15 2,3,4,6-TeCP * 0.2 0.005 - 0.01 0.5 - 2.0 1.0 - 3.0 5.0 -15 * Analysed in samples from Fraser River only. N/A: Not analysed in suspended sediments. 65 4. RESULTS AND DISCUSSION 4.1. L A B O R A T O R Y S T U D I E S 4.1.1. B i o c o n c e n t r a t i o n a s a F u n c t i o n o f C o n t a m i n a n t C o n c e n t r a t i o n Experiments were carried out in April 1991, in order to determine the effect of the water concentration of chlorinated phenolics on leech bioconcentration. Since pollutant concentration is one of the major determinants of final tissue bioconcentrations (Barron 1990), it is necessary to quantitate the effect of this parameter in order to model chlorinated phenolic bioconcentration. In the context of the Fraser River, water concentrations of chlorinated phenolics vary widely on both temporal and spatial scales as a result of proximity to effluent point sources, changing pulp mill process technology and seasonal variations in water flow. Therefore, an assessment of the effect of variance in water concentration of chlorinated phenolics on leech bioconcentration is important to interpretation of field monitoring data. It was hypothesised that the bioconcentration of chlorinated phenolics would increase with increasing chlorinated phenolic concentration, until water concentrations reached high enough levels to either saturate uptake mechanisms and storage sites or cause toxic effects resulting in altered uptake and elimination kinetics. The elimination rate of a target contaminant from an organism is also of importance in assessing the sensitivity of a biomonitor. Rapid clearance from tissues will result in decreased bioconcentrations and reduced sensitivity to low ambient levels of contaminants. Earlier research conducted by Metcalf et al. (1988) reported linear relationships between leech bioconcentration and elimination time with prolonged half lives exceeding 30 days for tri- and tetrachlorophenols. From this research it was hypothesized that leeches would show similar slow elimination rates for the chloroguaiacols, with relatively little loss of compound after seven day bioassay periods. 66 Leeches (TV. obscura), which had been collected from Black Lake in August 1990, were exposed semi-statically for seven days, to five chlorinated phenolic concentrations ranging from 0.1 - 10 ug/L (n = 5 leeches per test group) at a water temperature of 12.5 °C and a pH of 7.5. The environmental relevance of test concentrations greater than 1.0. ug/L is questionable, since Fraser River studies (Carey 1988) (Hall and Jacob 1988) (Schreier et al. 1991) have shown chlorinated phenolics concentrations to be generally less than 1.0 ug/L. It would be desirable to carry out experiments at levels lower than 0.1 ug/L, however, pilot studies indicted that under laboratory conditions, minimum leech bioconcentration detection limits would be reached after seven days exposure to water concentrations in the area of 0.1 pg/L. Mean leech weights, between test groups ranged from 1.18 - 1.45 g. Elimination of chlorinated phenolics were followed over a 28 day period, by transferring a subsample of leeches to clean water after seven days exposure to a 1.0 ug/L concentration. Water was replaced every 24 hours and samples of leeches (n = 5), were taken at time 0, 72, 168, and 672 h after termination of seven day exposure. Mean leech weights, between sample groups ranged from 0.63 - 1.36 g. The variation in leech weight between sample groups in the elimination experiment resulted in significant differences in bioconcentration due to leech weight. Therefore, it was necessary to normalize elimination data to a leech weight of 1.0 g. Tables detailing normalization calculations are given in Appendix 3 (Table A2 and A3). No mortalities were recorded during either exposure or depuration periods. A strong linear relationship between bioconcentration and ambient contaminant level is of importance in the selection of an effective biomonitoring organism (Phillips 1978). Laboratory bioassays were successful in demonstrating a strong linear relationship between leech bioconcentration of chlorinated phenolic compounds and water concentration (0.1 - 10 ug/L), when transformed to a double log scale (Y = aX b) (Table 4.1; Figure 4.1). Slopes of regression lines, indicated by b in the equation Y = aX*\ varied from 0.77 - 1.0, indicating possible differences in uptake and bioconcentration behaviour. 67 co CO as , . _cz D) CD CO c o £5 c CD on O T3 c CD CO ort o o_ .—' CD CD CO "5 o CO Q. 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E E N N W « a. g- E g-P E co E n co co co co co co • + - + - < £ » tf> CO II ~ c CD co g N O W o - S E n co >• 68 Figure 4.1: The effect of chlorinated phenolic exposure concentration on leech (N. obscura) bioconcentration of 3,4,5-TCG and 3,4,5-TCVer. 10000 T 0.10 1.00 Log Water Concentration (ug/L) + = Standard deviation 3,4,5-TCG; x = Standard deviation 3,4,5-TCVer 69 The observed relationship could be described by the equation: Eqn. 1 Bioconcentration (u.g/kg) = a(Water Concentration (ug/L))b used by Ellgehausen et al. (1980) to describe the bioconcentration of organochlorine pesticides by algae (Scenedesmus acutus), a daphnid (Daphnia magna) and catfish (Ictdlurus melas)at equilibrium water concentrations. This indicates that our semi-static bioassay was a good approximation of steady state water concentrations. Similar linear relationships were found for chlorophenol bioconcentrations in salmonids inhabiting the Fraser River (Carey et al. 1988; Servizi et al. 1988). Leeches were relatively efficient bioconcentrators of chloroguaiacols in laboratory bioassays. Table 4.2 shows the percentage of 4,5,6-TCG found in leeches after seven days exposure. Table 4.2: The average proportion of 4,5,6-TCG found in leeches after seven days of semistatic exposure at various concentrations. Nominal Water Concentration (Ug/L) Total Amount Spiked 7 days(u.g)* Total Amount Detected in Leeches (n = 5) dig) Proportion of 4,5,6-TCG in Leeches (%) 0.1 0.7 0.222 31.8 0.5 3.5 1.00 28.6 1.0 7.0 2.30 33.0 5.0 35.0 5.89 16.8 10.0 70.0 13.2 18.9 Water was changed and freshly spiked at 24 h. intervals. Note that at the two highest water concentrations (5 & 10 ug/L), leeches bioconcentrated a smaller percentage of compound, indicating the approach of a saturation in uptake rate. Since the leeches bioconcentrated such a high proportion of compound in the 70 water, it is difficult to argue that water concentrations were at a steady state, unless most of the uptake took place relatively quickly, reaching nearly steady state bioconcentrations within 24 - 48 h (1 - 2 water changes), with relatively little uptake over the following 120 - 144 hours. In this case water concentrations and leech bioconcentrations would have been relatively stable for most of the bioassay period. The relatively low bioconcentration of 3,4,5-TCVer in leeches was puzzling, since the bioconcentration potential of this compound has been shown to be great. Bioconcentration factors of 3,4,5-TCVer in the zebra fish (B. rerio) were on the order of 3100 (Neilson et al. 1984), an average of 75 times the values observed for leeches in our bioassays. It is possible that the concentrations of 3,4,5-TCVer covered in our experiments caused a toxic reaction, resulting in inhibited uptake of this compound, however this seems unlikely for two reasons: firstly, leeches have been shown to be resistant to high concentrations of chlorinated phenolics (Metcalf et al. 1984) (Hall and Jacob 1988) and secondly sub lethal threshold concentrations for 3,4,5-TCVer to fish embryos and larvae (Brachydanio rerio)weie found, in one study (Neilson et al. 1984), to be 450 ug/L, 45 times the highest exposure concentration tested in this study. However, statements regarding the applicability of toxic threshold concentrations between species (i.e. fish vs. leech) have to be taken with caution, since toxicity can depend on many species specific physiological characteristics (Barron et al. 1990). The bioconcentration factor, defined as the steady state bioconcentration (pg/kg) divided by the water concentration (ug/L) can be derived by rearrangement equation 1, yielding. Eqn. 2 a = Bioconcentration (ug/kg)/Water Concentration (ug/L)'3 where a, the y intercept, is equal to the BCF (Ellgehausen et al. 1980). When the BCF was related back to water concentration there appeared to be very little change over the range of water concentrations tested (Table 4.3). Leeches retained similar relative proportions of compounds in their tissues at each water concentration. 71 co. cr co CC su-es o ir i LL c o '4—• CO •4—' § o c o o o o CD o O CD i _ CO o CL i n 0 T3 c o Q. E o O 7— CO i n i n co 0 ci 0 0 0 CO CO +1 i n CM CM + 1 o + 1 CM GO o CO +1 o (3 0 Q 1 i n + 1 i n CO + 1 CM 00 CO co + 1 CO o m + 1 o m O I— • i n co" , cn CD i n +1 +1 +1 CO m CM co CO CM -,— CO CO CM CM + 1 + 1 + 1 T - CM 00 CO CM o m co CD O H-CD in o i n CM CD > o h-i n co" CM 5 00 + 1 + 1 + 1 m CM CO 03 CM w •*r •si-CO ^ -+1 +1 +1 CO 00 CM CM 00 CM CO O CM O O •sf + 1 + 1 + 1 o 9> 6) o> o m CM 00 CO R ' N O Q 1 " CO O O £ M CO CO g 5 'co 50 0 a CO co -H +1 co S> co £ 1 1 . - u) . . . _ if 10 M m m 1  1  c 1  1  e •= m <= >= Q, CO N CO CO CO CO E E 5 6 N N '55 w N N W WJ _ "j§_ ffi 0) aJ n I i C Q . Q . CD E E § E E aj CO 00 (0 CO CO CO . CO CO 4- (&> X -O 72 E l i m i n a t i o n o f ch lor ina ted phenol ics over a 28 day per iod are ou t l ined i n F igure 4.2. The data were too sparse to either describe the nature o f e l imina t ion kinetics or a l low an accurate estimate o f ha l f l i fe ( T 1 / 2 ) m leeches. Whi l e T 1 / 2 c o u l d not be calculated, i f i t is assumed that the tissue concentration o f compound at T\/2 equals ha l f the measured concentration at T = 0 h, then data presented i n F igure 4.2 indicate that ch lo rogua iaco l ha l f l i fe is approached at T > 28 days (672 h), wh i l e chloroveratrole hal f l i fe is reached at T < 3 days (72 h). The relatively long T 1 / 2 ( > 2 8 days) for the c h l o r o g u a i a c o l s i n leeches is cons i s ten t w i t h h a l f l i v e s o f ch lo ropheno l s i n leeches reported by other researchers. M e t c a l f et a l . (1988) calcula ted the ha l f l ives o f a variety o f d i - , t r i - , and tetra-chlorophenols in three species o f leech (Dina dubia, Erpobdella punctata, Helobdella stagnalis) to range from 25 days to greater than 40 days. In sharp contrast to the s low e l iminat ion rates reported for leeches, chlorinated phenolics appear to be e l iminated rap id ly by other aquatic organisms. Renberg et a l . (1980) reported ha l f l ives of less than 24 h for 4 , 5 , 6 - T C G and T e C G i n the brackish water fish species Alburnus alburnus. T h e T j / 2 for P C P i n two fish species (Salmo gairdneri, Carassius auratus)and one species o f mol lusc (Mytilus edulis) was reported to be under 24 and 72 h respectively. Data in F igure 4.2 indicates that accurate determinat ion o f ha l f l ives o f ch loroguaiaco ls i n leeches (N. obscura) w o u l d require longer depura t ion t imes and more s a m p l i n g p e r i o d s . E l i m i n a t i o n curves for the ch lo rogua i aco l s a l l d i sp l ayed s i m i l a r character . There was an apparent sharp drop i n tissue levels after on ly 72 h , however the higher tissue levels observed at 168 h indicate that the i n i t i a l sharp decrease may s imply have been due to experimental var ia t ion. Fo r a two compartment e l imina t ion mode l , between the water and an aquatic o rgan ism, a s emi log plot ( log tissue concentrat ion vs. t ime) y ie lds a straight l ine re la t ionship. In a three compartment m o d e l , where uptake, d i s t r ibu t ion and e l i m i n a t i o n o f the c o m p o u n d takes place be tween water and two b i o l o g i c a l compar tmen t s , the e l i m i n a t i o n c u r v e is character ised by an i n i t i a l r ap id e l i m i n a t i o n phase f o l l o w e d by a longer , s low e l iminat ion rate phase (Butte 1991). Linear regression o f data i n Figure 4.2, y ie lded 73 Figure 4.2: The elimination of chlorinated phenolics from leeches (N. obscura) over a four week period, after seven days exposure to a concentration of 1.0 ng/L at a water temperature of 12.5 °C and water pH 7.5. Sampling times were at time 0, 72, 168 and 672 h. Horizontal dashed lines represent the concentration of compound at T 1/2, expressed as half the mean tissue concentration at sampling time 0 hours. All tissue concentrations are normalized for a leech weight of 1.0 g.f zi. c q « 2 c CD O c o o CD CO CO 1 0 0 1 00 200 500 300 400 Elimination time (hours) 600 700 800 900 1000 CD c o -*—> CO -t—1 c CD o c O O CD 3 CO CO 10 B — . TeCG + Bp 3,4,5-TCVer 1—1—1— + \ | 1 1 1 1 | 1 1 1 1 I 1 1 1 ' 1 ' ' 1 1 | 1 1 1 1 l l T ' T ' f t | 1 1 1 1 1 1 1 1 1 100 200 300 400 500 600 700 800 900 Elimination time (hours) + = Standard deviation. t Leech sample weights: Sample time 0 h, 1.23 ± 0.44 g; Sample time 72 h, 1.36 ± 0.23 g; Sample time 168 h, 1.17 ± 0.19 g; Sample time 672 h, 0.63 ± 0.15 g. 74 weak correlat ions between tissue concentrat ion wi th t ime: 4 , 5 - D C G , r 2 = 0.37; 3,4,5-T C G , r 2 = 0.41; 4 ,5 ,6 -TCG, r 2 = 0.68; T e C G , r 2 = 0.32; 3 ,4 ,5-TCVer , r 2 = 0.59, indicat ing the p o s s i b i l i t y that e l i m i n a t i o n k ine t i c s f i t a non- l inea r three compar tment m o d e l characterized by an in i t i a l rapid release phase fo l lowed by a prolonged s low release phase. E l i m i n a t i o n curves for chlorophenols i n leeches reported by M e t c a l f et a l . (1988), showed a l inear two compartment depuration model for three leech species. It is l i k e l y that e l i m i n a t i o n o f c h l o r o g u a i a c o l s i n leeches also f o l l o w s a two compartment model and that more sampl ing periods w o u l d have y ie lded a stronger l inea r co r r e l a t i on between l o g t issue concent ra t ion and e l i m i n a t i o n t ime. The e l iminat ion curve for 3 ,4 ,5-TCver was characterized by a rapid depuration rate after T = 168 h , i n d i c a t i n g pos s ib l e d i f ferences i n e l i m i n a t i o n mechan i sms be tween chlorovera t ro le and ch lo rogua iaco l compounds . Since this compound is c h e m i c a l l y dist inct from the chloroguaiacols , due to an O-methylat ion o f the phenol group, i t is poss ib l e that i t is i n a fo rm more eas i ly metabo l i sed (e.g. enzyme media ted conjugation) or d i rec t ly e l iminated (e.g. b i l i a r y excret ion) . A plaus ib le explanat ion for the l ow leech bioconcentrations o f 3 , 4 , 5 - T C V e r is the re la t ively rapid e l imina t ion o f 3 ,4 ,5-TCVer . M e t c a l f et a l . (1984, 1988) attributed h igh leech B C F s for chlorophenols to ex t r eme ly s low e l i m i n a t i o n rates. The apparent s l ow rate o f e l i m i n a t i o n o f ch lorogua iaco ls by N. obscura is o f relevance to their routine use as biomonitors on the Fraser R i v e r . A slow e l imina t ion rate increases the sensi t ivi ty o f a b iomoni tor s ince the amount o f contaminant retained i n tissues increases. Therefore , the threshold l eve l o f water contaminant detectable by the b iomoni tor is lowered. L o w e l im ina t i on rates h igh l igh t exposure to per iod ic h igh pulses o f contaminant , s ince the organism retains cor respondingly h igh body burdens longer after the inc ident . T h e l i n e a r r e l a t i o n s h i p b e t w e e n l e e c h b i o c o n c e n t r a t i o n and wa te r concentra t ion, coup led w i t h the s low rate o f e l i m i n a t i o n o f ch lor ina ted pheno l i c s suggest that ambient chlor inated phenol ic concentrations i n water may be estimated from body burdens i n leeches. However , since ambient chlorinated phenol ic 75 concentrations are often be low the lower l i m i t o f those covered by the laboratory regression relat ionship, estimates may not match measured levels . In addi t ion, other in f luences , such as water temperature, p H , and suspended sediment p a r t i t i o n i n g , w h i c h vary wi th season and loca t ion , must be assessed to create a more complete m o d e l . 76 4.1.2. Bioconcentration as a Function of Leech Weight Experiments were carried out in April 1991, in order to determine the effect of leech weight on bioconcentration of chlorinated phenolics. Animal size and weight are recognized determinants of final tissue bioconcentrations of some aquatic species (Barron 1990). Earlier investigations (Hall and Jacob 1988) concerning bioconcentration of tetra- and pentachlorophenol by leeches, found evidence of weight related bioconcentration differences, indicating the need to pursue the matter in this current study. In the context of routine biomonitoring programs, it is essential to quantify this effect, both to determine if there is an optimal leech size for biomonitoring and to aid in the interpretation of monitoring data from leech weight groups of variable size. It was hypothesised that the bioconcentration would be related to the surface to volume ratio of the leech, which is inversely related to weight. It was anticipated that smaller leeches, with a high surface to volume ratio, would attain higher bioconcentrations. Leeches (N. obscura), which had been collected from Black Lake in August 1990, were exposed semi-statically for seven days to a 1.0 pg/L concentration of chlorinated phenolics at water a temperature of 12.5 °C and a pH of 7.6. Leeches were segregated into four different weight groups (n = 5 leeches per group); mean individual leech weights ranged from 0.148 - 1.77 g. Based on the expected growth rate and life cycle (Davies and Everett 1977) of the leech species N. obscura, this range of leech weights covered individuals born during two separate seasons, late summer of 1989 and spring and early summer of 1990. Leeches under 0.5 g (n = 10) lacked a clitellum, which is an indicator of sexual maturity. Therefore, there may have been age related differences in physiology between the larger and smaller leeches. Results describing the effect of leech weight on bioconcentration of chlorinated phenolics are given in Table 4.4. A strong inverse relationship between bioconcentration of the chloroguaiacols and leech weight was observed (r2 = 0.90 -0.94). The best linear fit was obtained by plotting the variables on a log log scale, in 7 7 c g co ZJ c r UJ c o CO co CD rr CL X CD X (0 II > 3 c o c CD O c o O CD 3 CO CO CO O CO d LO CNJ -H T CM II +1 +1 c co LO CO co ll Q. ZJ 2 CD JZ DJ CD 5 JZ o CD CD O CL e o O CO o 0) 0) CD 0) -t-o o 6 ci X CJ> CM CO o CM O CM CO CM O CM co CO O) CO n=4 +1 +1 n=4 O) o * CO T— CO co CM O CO CO CO +1 +1 +1 +1 CM CO LO CM CO CJ) CM LO CO CO CJ) CO o CJ) o CM CO +1 +1 +l +1 +1 CM CO •tf CM CJ) o CO LO •* o LO CM •tf o co CO CM T— 1— +1 -H -H o CO CM O) CO CO 00 78 which the smallest leeches, represented by the steepest portion of the curve, were extremely efficient bioconcentrators of chloroguaiacols. Larger leeches (1.2 - 1.8 g) showed less change in bioconcentration with equal changes in weight. These data immediately suggest that a trade off in leech size selection must be considered between very small leeches, which are highly efficient and sensitive bioindicators and larger leeches which are less sensitive, but also show less variability with size. As indicated in Figure 4:3, 3,4,5-TCVer displayed a different behaviour; there did not appear to be any correlation between leech weight and bioconcentration of 3,4,5-TCVer. The slopes, b, of the regression equations listed in Table 4.4 indicate that the magnitude of apparent weight effect differed between the chloroguaiacols. 3,4,5-TCG, b = - 0.83, showed the greatest effect, while 4,5,6-TCG, b = - 0.54, displayed less of a weight related difference, implying compound specific differential bioconcentration behaviours between the chloroguaiacols. Bioconcentration models for lipophilic organic compounds often describe the process as a simple partitioning process between the water and organic lipid phase, where bioconcentration is dependent upon the log of the octanol water partition coefficient and the lipid content of the organism (Barron 1990; Connell 1991). Many recent models also incorporate rate limiting physiologically based determinants such as ventilatory volume across respiratory surfaces, blood flow rates and depuration rates, into the bioconcentration process (Barron 1990). For example, in the case of poorly perfused peripheral tissue compartments, blood flow can exert a rate limiting influence on contaminant distribution. Physiological processes, in turn, may be influenced by ambient environmental conditions. A plausible explanation for the observed inverse relationship between bioconcentration and leech weight may be inferred from current knowledge concerning leech physiology. In general, small leeches have been observed to show a greater rate of oxygen consumption. Oxygen consumption in the leech species Mooreobdella microstoma ranged from 1.7 pL/mg in small individuals (20 mg) down to 0.3 uL/mg in large individuals (250 mg) (Sawyer 1989). Davies et al. (1987) found that survival of N. obscura, under anoxic conditions, was greater for larger leeches, 7 9 i m p l y i n g greater oxygen requirements for smal ler s ized i n d i v i d u a l s . Figure 4.3: Bioconcentration of 3,4,5-TCG and 3,4,5-TCVer by leeches (N. obscura) of differing weights. 1000 cn C O s= CD O C O O CO CO CD o 100 + 10 0.1 3,4,5-TCG = O 3,4,5-TCVer = <> + —i 1 1 1—t-1.0 Log Weight (g) 10.0 + = Standard deviation 3,4,5-TCG; x = Standard deviation 3,4,5-TCVer 80 Since the primary uptake site for lipophilic organic chemicals in aquatic organisms is the respiratory surface (Connell 1991), a greater oxygen consumption rate in smaller leeches promotes greater total exposure and bioconcentration. In a similar fashion, rainbow trout, show increased ventilatory volume (Barron 1990) and cardiac output (Barron et al. 1987a) with elevations in water temperature, which is linked to increased bioconcentration of PCBs (Veith et al. 1979). There is evidence that depurative metabolic function in leeches may change with age. Leech carbohydrate and lipid metabolism as well as xenobiotic (DDT) detoxification reactions take place in the betroiydal tissue, a primitive liver homologue, dispersed throughout the body. Betroiydal tissue content has been found to increase with age (Sawyer 1989). While no attempt was made to age leeches in this experiment, life history studies of N. obscura, conducted by Davies and Everett (1976) indicate that there were significant age differences between the two heaviest and lightest leech weight groups. The two heaviest weight groups were sexually mature and probably born 5 - 7 months before the two lightest weight leech groups. Therefore, increased metabolic detoxification capacity could have contributed to the lower bioconcentrations observed in larger leeches. Bioconcentration by smaller leeches may be further enhanced by a greater surface to volume ratio; more specifically, by a greater surface area of respiratory sites relative to the total mass of tissue. Saarikoski et al. (1986) carried out comprehensive investigations of factors affecting the uptake rate of ionizable organic compounds in the guppy (Poecilia reticulata). Equations relating volume and surface area to bioconcentration were developed, based on assumptions of first order uptake and elimination kinetics and steady state tissue concentrations. The rate constant, k, of absorption of several chlorinated phenolic and carboxylic acids was inversely related to the size of the fish. 81 Saarikoski et al. (1986) used Eqn. 3 to calculate the rate constant of absorption: Eqn. 3 k a = C f / (C W * t) = N f / ( V f * C w * t) where k a is the rate constant of absorption, C f is the concentration in the fish after exposure to water concentration Q, for time interval t. In the second half of the equation N f represents the total amount found in the tissues and V f is the volume (= weight) the fish. Eqn. 4 describes the effective permeability of the outer epithelium to organic compounds: Eqn. 4 P = N f / (A f * C w * t) where P is the effective permeability and A f is the surface area of the outer epithelium. Permeability is related to surface area, while the uptake rate constant is related to volume or weight. However, both P and k a remain proportional to each other as long as the surface area to volume ratio remains constant (Eqn. 5). Eqn. 5 P / k = V f / A f or k a = P * A f / V f According to Eqn. 5, if the surface area to volume ratio increases, there will be an increase in the uptake rate constant, ka, and the concentration in the fish. It is likely that a similar mechanism could be operating in the case of the leeches, where there is an increase in the ratio of surface respiratory sites to tissue volume ratio, resulting in the observed inverse relationship between leech size and chlorinated phenolic bioconcentration. The effect of leech weight on bioconcentration was compound specific. A strong inverse relationship was observed for the chlorinated guaiacols, with 3,4,5-TCG showing the greatest effect, while no clear trend was observed for 3,4,5-TCVer. Available evidence suggests that the effect may have been due to a combination of factors, both physiological and physical. Increased respiratory rate and decreased 8 2 depurative enzyme function i n smaller , younger leeches were possible factors. The greater surface area to volume ratio and lower total biomass associated wi th the l ow weight leech groups are other poss ib le con t r ibu t ing factors. M e a s u r i n g enzyme function and respiratory demand i n relat ion to leech size were beyond the scope o f this research. G i v e n the c o m p l e x i t y and v a r i a b i l i t y i n l eech age/s ize re la ted p h y s i o l o g i c a l r e l a t i onsh ip s be tween p o p u l a t i o n s o f N. obscura f r o m d i f f e r en t geographica l locat ions (Davies and Everet t 1977), the best so lu t ion to intraspecies var iab i l i ty i n bioconcentrat ion w o u l d be the use o f a laboratory bred leech stock o f uniform size and age for routine in situ b iomoni tor ing . It was not feasible to achieve this goa l i n the current s tudy, howeve r , the exponen t i a l inverse r e l a t i o n s h i p be tween l eech we igh t and b i o c o n c e n t r a t i o n ind ica tes that any we igh t re la ted differences should be m i n i m i z e d between leeches at the higher end o f the weight scale. Lesser weight related var iab i l i ty should be observed where leech weights are greater that 1.0 g. In the f o l l o w i n g experiments, variations i n leech weight between test groups were kept to a min imum (± 10 - 30 %). 83 4.1.3. B i o c o n c e n t r a t i o n a s a F u n c t i o n o f W a t e r pH Experiments were carried out in September 1991, in order to determine the effect of water pH on bioconcentration of chlorinated phenolics by leeches. Since chlorinated phenolics are weak acids, individual pKa and water pH determine the percent ionization of these compounds. In the context of routine biomonitoring programs, it is essential to quantify this effect for monitoring sites of variable water pH and to aid in the interpretation of monitoring data from regions of differing water chemistry. For the purpose of chlorinated phenolic monitoring on the Fraser River, which ranges in pH from 7.5 - 8.5, the effect of water pH may be of lesser significance for those compounds with a pKa greater than one unit above or below this pH range. However, certain compounds with a pKa falling in this range may show substantial shifts in percent dissociation. For example, 4,5,6-TCG (pKa = 7.4) shifts from 56% to 93% ionization over the above pH range of 7.5 - 8.5. It was hypothesized that leech bioconcentration would decrease with increasing water pH in a relationship roughly parallel to the dissociation curve of the individual compound, making it possible to estimate leech bioconcentration from the water concentration of compound, corrected for the percent ionization. Since, trichloroveratrole is non-ionizable it was hypothesized that bioconcentration would not change with water pH. Leeches (N. obscura), which had been collected from Black Lake in July 1991, were exposed semi-statically for seven days to a 0.1 ug/L concentration of chlorinated phenolics at a water temperature of 12 °C. Three experimental pHs (5.1, 7.1, 9.0) were selected to cover the pH range of most naturally occurring receiving water environments. Leeches were segregated into three test groups (n = 10) and one control group (n = 3); mean individual leech weights .ranged from 0.606 - 0.714 g between test groups. Results showing the effect of water pH on percent ionization and bioconcentration of chlorinated phenolics are given in Table 4.5. A weak inverse 8 4 CO >> co •o c 01 > CD CO CD 4—' ro ro i— r; o co x i O Si CO CD x: C J CD CD ® o Q . C D 1 ro CD : c CD D . o E I E ® °ro § 5 • ;= CO CO ^ i : CO § = * • CJT-5 ° <D O £ C c-S o ro £^ Q _ <D >_ O S§ co cj ^ ro o o t5 CD 1 o CD Q . x: X <D LO •tf 0) XI c » >< (0 Q . LU ^ c —' .9 co CD rr CO CO r--•tf CO in CD 6 6 6 ci b O O O . O - T-,2 ~ l o o c „ ,9 cn O J Z co o>| 8 ^| CO 1 ,2 -I 8 ^ o O CO cn < d oi Z LO CO cn Cci ^ +1 +1 00 CO CO CO CM ^ CO CO cn co co CJ oi tf co +1 +1 +l +1 +1 o> o CO o LO CO CO CO O *— CO O-5 o 3 O a. E o o CN LO CO co tf • t f Cvi +1 +i +i +1 +1 o co CM CO C J O CO , — o tf q CO N CD in o tf T-C9 O Q in •tf O O h- H in co •tf co" D O 1 « CO , •H E '5 at g> 17 E « ^ cu .,_ >-t t : 8 5 l inear re la t ionship between b ioconcen t ra t ion o f the ch lo rogua iaco l s and water p H was observed (r 2 = 0.36 - 0.67). Cor rec t ion o f bioconcentrat ion data for between sample differences i n mean leech weight d id not s ign i f i can t ly alter results. The strongest l inear correla t ion was observed for a semi l og plot o f bioconcentrat ion vs. water p H (Figure 4.4). The non-dissociable 3 , 4 , 5 - T C V e r showed a trend towards increas ing b ioconcent ra t ion at higher water p H . S ince the chemica l character o f the veratrole remains unchanged wi th water p H , the increase i n b ioconcentra t ion is an ind ica t ion that p H may alter the membrane permeabi l i ty or uptake phys io logy o f leeches. Figure 4.4: The effect of water pH on the bioconcentration of 3,4,5-TCG and 3,4,5-TCVer by leeches (N. obscura). + = Standard deviation 3,4,5-TCG; x = Standard deviation 3,4,5-TCVer. 86 Both the fish toxicity and bioconcentration of chlorophenols have been shown to decrease with increasing water pH as a result of dissociation of the neutral phenol to the phenolate ion form (Saarikoski and Viluksela 1981, 1982). As hypothesized, an inverse relationship between leech bioconcentration and water pH was observed for the ionizable chloroguaiacols, however the pattern of the effect did not follow the dissociation curve for the compounds tested. Figure 4.5 plots the dissociation curves for the ionizable chlorinated guaiacols, with an overlay of the bioconcentration of each compound at the three experimental pHs. The change in the percent ionized fraction is generally insignificant at pHs greater than ±1 unit of the pKa. Major changes in the bioavailable fraction take place at pHs near the pKa. For example, the percent dissociation of 3,4,5-TCG (pKa = 8.0) ranges from < 1 - 9.1% over the range of pH 5 - 7. Above pH 9.0, 3,4,5-TCG remains > 90% ionized. This behaviour indicates that if the major bioavailable form is the non-ionized species, then the relationship between bioconcentration and water pH in Figure 4.5 should be a mirror image of the dissociation curve of that compound. According to this theory, the bioconcentration of 4,5-DCG should change little from pH 5.1 to 7.1, since there is only a 1.2% change in percent ionization. From pH 7.1 to 9.0, we should see a dramatic decrease in the bioconcentration of 4,5-DCG, since the percent ionization increases from 1.2% to 50%. In the case of 3,4,5-TCG we should have observed a sharp decrease in bioconcentration between pH 7.1 - 9.0, had the unionized form been absorbed with much greater ease. In a leech bioconcentration study by Hall and Jacob (Figure 4.6) a similar phenomenon was observed for 2,4,6-TCP and PCP (pH 7.5 - 9.0). In general, the slopes of the bioconcentration vs. pH curves were much less steep than would be predicted if the non-ionized species were the only form absorbed. In studies using guppies (Poecilia reticulata) Saarikoski et al. (1986) observed that bioconcentration of ionizable chlorophenols was independent of pH, as long as pH was > 1 unit below the pKa. However, as pH approached the pKa of the compound, bioconcentration in guppies decreased, but at a rate less than predicted if only the non-ionized fraction were bioavailable. The water soluble phenolate ion has been shown to be bioconcentrated at a significant rate in fish (Saarikoski et al. 1986). 87 •o CD o w c (6>)/Brl)uonBJ)uaouoo enssy. (6>|/6Tl)uo!iBJiueouoo enssn co CO CO 3 3 O T CJ) CL LO T3 03 9 ° co O x: o CD Q. N _ C CO .2 c _ „_ CD o o E ° .2 " co O) CD CL X CD — CD CD CD §> S -=S- c: x: o ~~ C O CD O i_ x: <D — <0 "55 "K eg co <= > c 8 » .2 c £ co .2 co § x i § O CO o CD .2 0) « 5> I ' =co c — c X CO " CD CD II C D) N O CD c in £ •S . CD x re 2^ o. o o co CO « * 2 CJ co — £ S o Q co co LO CD uoj)ezjuo| % (6>|/6Tl)uo!iBjjueouoo anssii uo|jBZ!UO| % X C L •2 CO X C L UO!)BZjUO| % (6>|/6Tl)uonBJ}ueouoo enssy. UO!JEZ|UO| % 3 CO co 0 0 c co c\i "— •D ai c O •a O o c CO w C L CO ~3 E in TJ o c CO an he TD cf Hall C L ze IO Hall ted oni o E ina "5 Hs o o c C L pa x: o "CO o •«—. CO Q. c CO o CD X3 E < CD CJ c-CD •Q_ _j zi. II CO ex| o> ter ee o "cn zj wa thr o ion the the ion rati tz CO ^ 0 cz CO c c CD g 0 o x: o c CO cz o £ -»-^ o o c o o CD i !Q o U a> T 3 o CO o c o lee sa ue CO co o •o CO an c O an CD o ok 0 tf-II CD o> c ro c CO x: g J C X o tf-BZj C L k_ CD x: CM i— c CD < O o wat TD wat 0 o>- co o OS -1-c c CO dxi o CO CO 0 CO CD CO CO 'o tted _c he: OS tted T> o 0 co o "o 0 b C L CO _ l CD 0 i _ ZJ CD (6>|/6rl)uo!}BJjueouoo enssi i (6>|/6TI)UO!IBJ}U80UOO onssn uo!JBZ!UO| % uouezjuoi % (6>)/6rl)uouBj}uaouoo enssu. (6>)/6Tl)uo!iBj}uaouoo enssy. U0!}BZ|U0| % U0!JBZ|U0| % 89 Further support for the bioavailability of ionized chlorinated phenolics comes from in situ monitoring studies on the Fraser River (pH 7.5 - 8.5), where measurable levels of TeCG have been detected in fish tissues, even though TeCG is 97 - > 99% ionized (Dwernychuk 1991; Schreier et al. 1991). Relationships between chlorinated phenolic bioconcentration and water pH appear to be complex. The relatively shallow slopes exhibited by the bioconcentration vs. pH curves in the present study (Figure 4.5) suggest that the non-ionized form of the chlorinated guaiacols are bioconcentrated preferentially to the ionized form, however the latter chemical species does represent a significant pool of bioavailable compound. The complexity of the results could be due to a pH induced interaction of both physical chemical dissociation of chlorinated phenolics and alterations in leech physiology. Insight into the mechanism of uptake of ionized chlorinated phenolics comes from sediment partitioning studies. Sediment adsorption of lipophilic organic compounds takes place mostly on the organic fraction (Schellenberg et al. 1984; Schwarzenbach 1985). Partitioning of chlorinated phenolics to suspended sediments of fixed organic content is inversely related to water pH, as is the case for bioconcentration, as well being positively correlated to the ionic concentration (K + , L i + , Ca 2 +) of the water compartment (Schellenberg et al. 1984). Schellenberg et al. (1984) showed that sediment sorption of both the neutral chlorophenol and the anionic form occur. Westall (1985) proposed that the major mechanism of sediment phase partitioning of the chlorophenolate ion was sorption of electrostatically neutral ion pairs (e.g. PCP"»K +). Increased octanol partitioning of tetra- and pentachlorophenol was observed as ionic strength (KC1) of the water phase was increased. It is quite likely that partitioning of chloroguaiacol ion pairs accounts for the apparent bioconcentration of ionized compound observed in both laboratory leeches and Fraser River fish. Bioconcentration of non-ionizable lipophilic organic compounds is generally independent of pH as long as physiological function is not altered over the pH range tested. Saarikoski et al. (1986) found no relationship between water pH and 9 0 bioconcentration of neutral organic compounds (tetrachloroveratrole, pentachloroanisole) in guppies '(P. reticulata). The observation that leech bioconcentration of 3,4,5-TCVer showed a positive relationship with water pH, indicates that leech physiology may have also been affected over the range of water pH tested. A comprehensive leech survey conducted by Herrmann (1970) in Colorado, revealed that N. obscura is found in waters ranging from pH 6.3 - 9.8. The absence of the species from waters of pH < 6.3 may not be due to water pH alone, since other factors such as conductivity, dissolved inorganic/organic content, food abundance and presence of competing leech species are also important variables. No literature concerning the effect of water pH on leech respiratory physiology was found, however the lowest experimental pH of 5.1 may have been below the tolerance level for the species. Evidence for a physiological component of the effect of water pH on leech bioconcentration can be inferred from studies of the effects of low water pH on fish physiology. Effects of low water pH (pH < 4) on fish include: acid-base imbalance, ionoregulatory dysfunction and decreased cellular oxygen delivery (Spry et al. 1981). Increased H + concentrations affect blood acid-base balance by titration of blood buffers, such as HC0 3", yielding H 2 0 and C 0 2 . Metabolic acidosis also leads to altered charges on enzymatic proteins, leading to altered cellular function (Spry et al. 1981). Ionoregulatory dysfunction, which occurs in fish at the gill, has been recorded at water pH < 6.0. This effect is dominant in waters of low C a 2 + concentrations, and is manifested by excessive loss of Na + , Cl" and K + ions (Spry et al. 1981). In fish, acid water also decreases uptake of 0 2 at the gill by two mechanisms. Firstly, 0 2 diffusion across the gill is reduced by copious mucous secretions, induced by acid exposure (Spry et al. 1981). Secondly, the affinity of hemoglobin for 0 2 is reduced by elevated blood levels of C 0 2 and H + (Bohr effect). The leech species N. obscurais, a faculative anaerobe (Sawyer 1989), capable of using both glycogen and amino acids as energy sources under anoxic conditions (Reddy and Davies 1993). If the assumption is made that leeches suffer acid-base 91 i m b a l a n c e f rom exposure to increased water concen t ra t ions o f H + and l e e c h hemoglob in responds to increased b lood C 0 2 and H + w i th decreased aff ini ty to 0 2 , then it is plausible that there may have been a compensatory shift to l o w energy g lyco ly t i c pathway at the lowest test p H of 5.1. Since the major end products o f anoxic metabol ism are succinate and alanine i n N. obscura (Reddy and Davies 1993), this hypothesis could be tested by metabolite quantification over the p H range 5 - 9. A phenomenon s imi l a r to the increased mucous secretions observed i n f ish exposed to l ow p H water could account for the the lowered leech bioconcentrations of 3 , 4 , 5 - T C V e r at p H 5.1. Leeches secrete a mucopolysacchar ide based mucous, f rom numerous subcutaneous mucous ce l l s , i n response to i r r i ta t ing external s t imu l i , such as l o w water p H (Sawyer 1989). S ingha l et a l . (1990) reported excess mucous secretions by N. obscura i n response to hype rox ic condi t ions , e f fec t ive ly protec t ing the organisms by reducing oxygen dif fus ion. M u c o u s secretions i n response to l o w p H may have decreased uptake o f 3 , 4 , 5 - T C V e r by two mechanisms: 1) acting as a permeabi l i ty barrier p h y s i c a l l y b l o c k i n g uptake at respiratory sites; 2 ) act ing as a th i rd adsorpt ive compartment , e f fec t ive ly r e m o v i n g c o m p o u n d f rom the b ioassay system. Such an effect c o u l d also have al tered the uptake o f the ch lo r ina ted guaiacols , such that the inverse relat ionship between bioconcentra t ion and water p H appeared to be o f lower magnitude between water p H 5.1 and 7.1. Strong pos i t ive correlat ions between bioconcentra t ion and l o g K o w have been reported for neutral organic compounds i n aquatic organisms (El lgehausen et a l . 1980). However , this relat ionship general ly shows much var ia t ion between different chemica ls when these data are examined c lo se ly . B ioconcen t r a t ion factors o f 45 dif ferent o rganic compounds ( log K o w 1-7) by fathead m i n n o w s (Pimephales promelas) showed variations of greater than two orders of magnitude for s imi la r l o g K o w values. A direct relationship between bioconcentration and l o g K o w of ionizable ch lor ina ted phenol ics is d i f f i cu l t to es tabl ish at neutral water p H , since percent i o n i z a t i o n varies f rom about 1% to greater than 90% for the compounds tested. However at p H 5.1 it is possible to compare bioconcentration to log K o w , since a l l the chloroguaiacols are ionized less than 1%, except T e C G (11.1%). Figure 4.7 plots the 9 2 o b s e r v e d r e l a t i o n s h i p b e t w e e n t h e l o g o f t h e b i o c o n c e n t r a t i o n a n d l o g K o w . N o c l e a r p o s i t i v e c o r r e l a t i o n b e t w e e n l e e c h b i o c o n c e n t r a t i o n a n d l o g K o w w a s o b s e r v e d . F i g u r e 4 . 7 : R e l a t i o n s h i p b e t w e e n l e e c h b i o c o n c e n t r a t i o n a n d l o g K o w o f c h l o r i n a t e d p h e n o l i c s a t p H 5 . 1 . 1 0 0 CD c O 03 C CD o c o o g CQ CD o 1 = 4,5-DCG; 2 = 4,5,6-TCG; 3 = 3,4,5-TCG; 4 = TeCG; 5 = 3,4,5-TCVer. B i o c o n c e n t r a t i o n s o f a l l c o m p o u n d s w e r e s i m i l a r , a l t h o u g h a p o s i t i v e t r e n d w a s o b s e r v e d f o r 4 , 5 - D C G , 4 , 5 , 6 - T C G a n d 3 , 4 , 5 - T C G . L e e c h b i o c o n c e n t r a t i o n o f T e C G w a s p r o b a b l y l o w e r t h a n e x p e c t e d d u e t o p a r t i a l i o n i z a t i o n . 3 , 4 , 5 - T C V e r b i o c o n c e n t r a t i o n s w e r e t h e l o w e s t ' o f a n y c o m p o u n d a t p H 5 . 1 , w h i c h m a y b e e x p e c t e d c o n s i d e r i n g t h e r a p i d i t y w i t h w h i c h i t a p p e a r s t o b e e l i m i n a t e d f r o m t h e l e e c h s p e c i e s N. obscura ( S e c t i o n 4 . 1 . 1 . ; F i g u r e 4 . 2 ) . S i n c e b o t h d i s s o l v e d a n d p a r t i c u l a t e o r g a n i c m a t e r i a l m a y a l t e r t h e r e l a t i o n s h i p b e t w e e n b i o c o n c e n t r a t i o n a n d l o g K o w , 93 through competitive adsorption (Jaffe 1991), mucous adsorption of the highly lipophilic trichloroveratrole may have also contributed to reduced bioconcentrations. Note that at pH 9.0, where the chloroguaiacols are > 50% ionized, 3,4,5-TCVer was bioconcentrated to the greatest level. The relatively small differences in bioconcentration between di- tri- and tetra-chlorinated guaiacols is consistent with the results of Saarikoski et al. (1986), who reported that the rate of bioconcentration in guppies (P. reticulata) increases linearly with lipophilicity until log Kow reaches a value of about 4.0, at which point there is little increase in bioconcentration with increasing lipophilicity. Log Kow values ranged from 3.41 to 5.35 in leech bioconcentration experiments. This phenomenon was attributed to the permeability barrier to highly lipophilic compounds, which exists as a result of unstirred water layers at the respiratory surface (Dainty and House 1966). Initial guppy bioassays were conducted in unstirred bioassay environments. When bioassay water was continuously stirred using a magnetic stirrer, bioconcentration of PCP increased by 75%. Uptake of less lipophilic compounds, such as phenol and butyric acid, was not significantly affected by stirring. Furthermore, the bioconcentration of the PCP anion was not affected by stirring, which suggests that uptake of chlorophenolate ions or ion pairs follows a different mechanism. In leech pH bioassays a stronger relationship between bioconcentration and log Kow may have been found, had the bioassay water been stirred continuously. The effect of water stirring on leech bioconcentration is investigated in Section 4.1.5. Bioconcentration of ionizable chloroguaiacols was inversely related to water pH, however, bioconcentration was not directly related to the concentration of undissociated compound. The phenolate ion form also contributed significantly to the bioavailable contaminant pool. There was evidence of physiological disturbance to the leech species N. obscura, induced by low pH, which was manifested by an opposite trend in bioconcentration of the non-dissociable compound, 3,4,5-TCVer. Regression relationships between bioconcentration and water pH were generally weak and in reality are probably not linear. Therefore, the relationships derived are only useful as a guide to aid in the interpretation of monitoring data. 94 4.1.4. Bioconcentration as a Function of Water Temperature Experiments were carried out in January 1992 to determine the effect of water temperature on bioconcentration of chlorinated phenolics by leeches. Water temperature is recognized as an important determinant of bioconcentration of organic compounds by aquatic organisms (Barron 1990). Within the context of in situ biomonitoring in the Fraser River, it is essential to quantify this effect, since seasonal fluctuations in water temperature range from 1 - 20 °C (Hall & Jacob 1988). It was hypothesized that leech bioconcentration would show a continuous, linear increase with increasing water temperature. Leeches (N. obscura), which had been collected from Black Lake in July 1991, were exposed semi-statically for seven days to a 0.1 pg/L concentration of chlorinated phenolics at water pH 7.0. Three experimental water temperatures between 4 - 20 °C, were selected covering most of the seasonal temperature range recorded on the Fraser River. N. obscura is a cold water stenotherm, preferring cooler lentic waters of Canada and the Northern United States (Sawyer 1972). Herrmann (1970) found N. obscura in waters ranging in summer and early fall temperature from 0.5 - 24 °C. Leeches used in the current study, originate from Black Lake, B.C. which undergoes a similar annual temperature variation. From this information it is assumed that leeches in the current study were not subject to temperatures out of their tolerance range. Leeches were segregated into three test groups (n = 10/group) and one control group (n = 3) at each temperature; mean individual leech weights ranged from 0.540 - 0.635 g between test groups. x i Table 4.6 shows leech bioconcentrations at water temperatures of 4.4, 11.8 and 20.0 °C. Contrary to the hypothesis, a steady increase in bioconcentration with temperature was not observed. Leech bioconcentrations were significantly (Student's t-Test, p = 0.05) lower at 4.4 °C than at 11.8 °C for the tri- and tetrachloroguaiacols, however bioconcentrations levelled off between 11.8 and 20 °C. No significant change in bioconcentration was observed for 4,5-DCG. Unexpectedly, trichloroveratrole bioconcentration decreased initially and then levelled off 9 5 co 13\ "* O) CO 5 CD CO CD 0 CD CD -a — CD >* t .Q o 1 s o co CO c o -a CD ns o ° X x : a . o M- CD ° « It id c o £ 8 CO o CO CO E 2 CD co o a_ X CD CO >. CO -o CD C CD CD > SI CD I— <o CO • t f XI CO g * * 15 cr cL LU X c o CO co CO CD II cn >-CD tr O) )^ c o '•^ CO c CD O c o O a) 3 CO CO o o <D 15 ® CL E CD I -CD 15 T3 C Z) o a . E o O CO t-CO CM X X CL x CD O CL X CD CO O CO CO II > CD O Q LO •tf II >-CD O \— I m co" co co CO o in 1— CO CM +1 +1 +1 +1 CM in o o in CO CO • t f in CM CO 1— +1 +1 +i +1 ,— 1— CJ) CO CO in • t f 1 - CO CM d r-+1 +1 +i +1 o> GO CO CO CO CO CO CD O • CO_ in" •tf" CL x cn o_ in CM | II > O 1^ +1 •tf CO l« S S I 96 between the highest temperatures. These data are i n d i c a t i v e o f a temperature i nduced a l te ra t ion i n the balance between uptake and e l i m i n a t i o n k ine t i c s o f compounds between 4.4 and 11.8 °C , which stabilized between about 12 - 20 °C . For the purpose o f p r e d i c t i v e m o d e l l i n g o f the effect o f water t empera ture on b ioconcen t ra t ion there are two poss ib le methods o f in terpre t ing the data (F igure 4.8): 1) to assume a bi-phasic re la t ionship, where bioconcentra t ion changes between 4.4 and 11.8 ° C , as measured by a l ine j o i n i n g data points between the two lowest temperatures, after wh ich there is no net change i n bioconcentrat ion, 2) to assume a steady change i n bioconcentra t ion over the entire temperature range, model led by a best l ine drawn between al l three data points. Neither model is perfect. The first option assumes that the no effect temperature starts at about 12 ° C , whi le in reality, i t could start at any point above 4.4 °C. The second option is a poor description o f the data, since i t leads to re la t ionships dev ia t ing s ign i f i can t ly from l inear i ty ( r 2 = 0 -0.33). F r o m in fo rmat ion pub l i shed concern ing the effect o f water temperature on leech b i o l o g y i t is l i k e l y that there are a number o f factors w h i c h c o u l d have contributed to the observed effect o f water temperature on bioconcentrat ion. Sawyer (1989) states that there i s an increase i n leech vent i la tory behaviour at e levated wate r t empera tu re , as m e a s u r e d by i n c r e a s i n g u n d u l a t o r y and s w i m m i n g movements, whi le L i n d m a n (1935) observed increased oxygen uptake by leeches over the temperature range 10 - 17 ° C . L i n t o n et a l . (1983) observed a general increase i n me tabo l i c rate o f N. obscura w i t h i n c r e a s i n g water tempera ture . Inc reased venti latory and metabolic act ivi ty could expla in increased bioconcentrat ion (T = 4.4 -11.8 ° C ) through a greater rate o f uptake at r esp i ra to ry sites as s i m i l a r l y hypothesized for f ish, by Ph i l l i p s (1978) and Barron (1990). A t 4.4 °C leeches showed l i t t l e p h y s i c a l ac t iv i ty and probably had re la t ive ly l o w ce l lu l a r O2 r e q u i r e m e n t s . Leeches d id appear to show greater s w i m m i n g ac t iv i ty i n the 11.8 and 20.0 °C b i o a s s a y s , w h i c h are l i k e l y to be co r r e l a t ed to i nc reased 0 2 up take and b i o c o n c e n t r a t i o n . Research by V e i t h et al . (1979) has relevance to the results of the current 97 Figure 4.8: Alternative interpretations of the effect of water temperature on the bioconcentration of 3,4,5-TCG and 3,4,5-TCVer by leeches (N. obscura). 1) Bi-phasic model 100 ut c o c CD O c o O cu at o 10 3,4,5-TCG = O 3,4,5-TCVer = O + 10 15 Water Temperature (° C) 20 2) Linear model; y = a(b)exp.x + = Standard Deviation 3,4,5-TCG; x= Standard Deviation 3,4,5-TCVer 98 leech research, since it shows that bioconcentration of organic contaminants does not necessarily increase constantly with temperature. Veith et al. (1979) found that fish bioconcentrations of PCBs generally increased with water temperature (5 - 25 °C). However the nature and magnitude of the effect was variable depending upon fish species at test temperature. For example, rainbow trout (Salmo gairdneri) bioconcentrations showed no increase between 5 - 1 0 °C, after which a steady increase was observed. Fathead minnow (Pimephales promelas) bioconcentration followed a pattern similar to that observed for leech bioconcentrations of tri- and tetrachloroguaiacol, with sharp increase between 5 - 10 °C, followed by very little change at 10, 15 and 20 °C. These data are also consistent with a multiphasic model to describe the effect of temperature upon bioconcentration as well as with the idea that there are species specific, competing physiological factors influenced by temperature, which affect bioconcentration. Bioconcentration can be seen as a function of the rate of uptake vs. the rate of elimination. Barron (1990) details the importance of physiological processes in xenobiotic uptake and elimination rates and the enhancing effect of increased temperature. Potential rate limiting , physiological and biochemical factors controlling uptake and elimination include: ventilatory rate and volume, blood flow, membrane permeability, lipid composition, rate of enzymatic biotransformation and metabolism (Barron 1990). For example, fish uptake of organic contaminants can be enhanced at elevated water temperatures through increased cardiac output and ventilatory volume. Sijm (1991) reported increases in both uptake and elimination rate constants of ethanol and 4-amino antipyrine in goldfish (Carassius auratus) over the temperature range 1 0 - 3 5 °C. One plausible interpretation of the leech bioconcentration data for the tri- and tetrachloroguaiacol (Table 4.6) is that the rate of uptake was enhanced to greater extent between 4.4 and 11.8 °C, while the rate of elimination took on greater importance at temperatures > 11.8 °C. The anomalous decrease in the bioconcentration of 3,4,5-TCVer could be explained by a greatly increased rate of elimination specific to this compound. In fact, this explanation is supported by data in Section 4.1.1.(Figure 4.2), which show that leeches are able to 99 eliminate 3,4,5-TCVer much more rapidly than the chloroguaiacols. The current leech data are not consistent with experiments conducted by Hall and Jacob (1988), who reported an arithmetic increase in leech (N. obscura) bioconcentrations of chlorophenols over the temperature range 4 - 2 2 °C. In these experiments uptake was dominant over elimination, The discrepancy in the current data may reflect a shift towards increased elimination rates specific to the chlorinated phenolics tested in this study. Metcalf et al. (1988) reported half lives of chlorophenols in leeches of 1 - > 2 months, while those for the chloroguaiacols are estimated to be < 1 month (Section 4.1.1.; Figure 4.2). Also, the relatively high exposure concentration of 10 pg/L used by Hall and Jacob, while not acutely lethal to leeches, may have had a sublethal effect on chlorophenol elimination function, leading to greater increases in bioconcentration than observed in the leeches exposed to 0.1 pg/L chloroguaiacols in the current study. Therefore, the effect of temperature on bioconcentration could be concentration dependent. In order to shed light on the discrepancy between the two leech studies it may be useful to conduct a search for biotransformation products. Common metabolites of chlorinated phenolics in aquatic species include sulfate and glucuronide conjugates of the parent compound (Kennedy 1989). An assessment of the effect of temperature on the ratio of conjugate to parent compound may indicate whether enhanced elimination rate accounts for the sharp levelling off of chloroguaiacol bioconcentrations above 12 °C. Phillips (1978) reported that a similar technique successfully revealed that the temperature induced increase in fish- bioconcentration of DDT was primarily due to increased uptake. These data can also be explained, in part, on the basis of physiochemical theories concerning the partitioning of organic contaminants to phases other than biota. Water temperature influences compound solubility (Phillips 1978) and interactions with other compartments, such as organic and inorganic particulate material (Schwarzenbach 1985). Since the exposure concentration of 0.1 pg/L was well below the solubility limit of the test compounds the first phenomenon should not have played a significant role. There was however, an unquantified amount of 100 particulate and colloidal organic material in the leech bioassay system, resulting from the secretion of mucopolysaccharide (Sawyer 1989) based slime by leeches. It was not possible to reliably quantitate the concentration of compound sorbed to slime. Moreover, the effect of water temperature on slime secretion was not measured, although visual inspection did not reveal any obvious differences between test groups. A histological search for hypertrophy in mucous secreting epithelial cells may clarify the importance of this effect. Increased water temperature should have increased the rate of desorption of compound from organic particulates with a concomitant increase in bioavailable compound (Phillips 1978). However, temperature may also have induced increased slime production, competitively inhibiting bioconcentration. The sharp decrease in bioconcentration of 3,4,5-TCVer is consistent with increased partitioning to organic particulates, since it is both neutral and highly lipophilic. Together, a temperature induced increase in elimination rate and secretion of excess leech slime may explain the inverse relationship observed for 3,4,5-TCVer and the levelling off of chloroguaiacol bioconcentrations at T > 11.8 °C. It is important to note that any such effect may be of lesser importance in a field biomonitoring system, where water currents are likely to wash away excess slime. From this experiment it is concluded that temperature affects bioconcentration of all test compounds except 4,5-DCG, following a pattern best described by a bi-phasic model. There is significant increase in tri- and tetrachloroguaiacol bioconcentration between 4.4 and 11.8 °C, at which point bioconcentrations remain constant. An opposite effect was observed for 3,4,5-TCVer. The results can be explained as the competitive interaction of uptake and elimination processes, with the former dominating over the lowest temperature range and the latter becoming of increasing importance at higher water temperatures. Considering the variability in the effect of temperature on organochlorine bioconcentration described by Phillips (1978) and Veith et al. (1979) along with the differing results obtained by Hall and Jacob (1988), using the same species of leech as the current study, it appears that there are other possible interfering factors to 1 0 1 consider such as: contaminant concentra t ion, species speci f ic p h y s i o l o g i c a l react ion to temperature (e.g. s l ime secretion), age, l i fe stage. Though a clear explanat ion o f the results is d i f f i cu l t to e lucidate , for the purpose o f routine b i o m o n i t o r i n g , a quant i f ied descr ip t ion o f the effect is enough for p rac t ica l app l ica t ion to routine b i o m o n i t o r i n g p r o g r a m s . 102 4.5. Bioconcentration as a Function of Suspended Sediment Load Experiments were carried out i n June 1992, i n order to determine the effect o f water t u rb id i t y on l e e c h b i o c o n c e n t r a t i o n o f c h l o r i n a t e d p h e n o l i c s . It was hypo thes ized that the b i o a v a i l a b i l i t y and therefore b ioconcen t ra t ion o f ch lo r ina ted phenol ics w o u l d be inh ib i t ed through compet i t ive b ind ing to part icles o f suspended sol ids , par t icular ly the organic fraction. W i t h i n the context of the Fraser R i v e r , the inf luence o f suspended sediments upon chlor ina ted phenol ic b i o a v a i l a b i l i t y cou ld be o f impor t ance , s ince suspended sediment concent ra t ions i n the M i d d l e F rase r (Hansard) have been shown to vary, seasonally, by greater than 600% ( H a l l et a l . 1991). Leeches (N obscura), wh i ch had been caught at B l a c k L a k e in Augus t 1991, were exposed semi-statically for seven days, to 0.1 u g / L concentra t ions o f ch lo r ina ted phenol ics at water a temperature o f 12 °C and a p H of 7.1. M e a n leech weights, between test groups ranged from 0.364 - 0.467 g. Bioassay protocol had to be modified by in t roduc ing a mechanism to main ta in sediments i n a suspension. T h i s was accompl ished through the continuous use o f a magnetic s t i r r ing apparatus, i n w h i c h bioassay jars , conta in ing a teflon coated magnetic stir-bar, were placed on magnetic st irr ing plates. In addit ion, one litre square sided bioassay jars were replaced by two l i t re c y l i n d r i c a l jars, since the s t i r r ing apparatus d id not function re l i ab ly when the former jars were used. Leech loading to bioassay jars remained at 5 leeches per l i tre bioassay water. A control bioassay was set up to assess the effect o f cont inuous s t i r r ing on b ioconcen t ra t ion . N o mor ta l i t ies were recorded du r ing the exposure p e r i o d . Based on evidence (Vo ice et a l . 1983) (Schellenberg et al . 1984) indicat ing that the organic and c lay par t ic les i n the 63 pm suspended sediment f rac t ion may represent the greatest effective adsorptive surface for chlor inated phenol ics , i t was decided to focus on the < 63 pm sediment fraction. Analys is o f the < 63 pm fraction o f Fraser R i v e r suspended sediments f rom Stoner B . C . , showed a r e l a t ive ly constant organic content, measured as the percent volat i le fraction, ranging from 5.4 - 7.4% 103 in July 1991, 3.7 - 5.2% in February 1992, 4.2 - 4.8% in March 1992 (Duncan, personal communication) 3 . Qualitative X-ray diffraction analysis revealed the dominant clay species to be chlorite and illite in an approximate ratio of 60% to 40%. Based on this evidence it was decided to use a suspended sediment mixture composed of 95% inorganic clay (50% chlorite : 50% illite) and 5% organic material. Organic material consisted of dried, crushed, green alder leaves, picked from the campus of the University of British Columbia, in Vancouver, B.C. It was decided to design the turbidity experiment to represent the seasonal range of concentrations and character of Fraser River sediments (Hall et al. 1991). Three suspended sediment levels, representing low, medium and high flow periods on the Fraser River were selected: 0.025 g/L ( 5.0 ±0 .5 NTUs), 0.08 g/L (19.0 ± 1.5 NTUs) and 0.15 g/L (35.0 + 3.0 NTUs). Water turbidity (NTU) was found to be linearly correlated (r2 = 0.98) to suspended sediment concentration (g/L). A turbid (0.08 mg/L) bioassay, with 100% inorganic clay was included, in order to assess the effect of the organic content on bioconcentration. In addition, a stirred bioassay with no introduced suspended sediments was run. Also included were control bioassays, with no chlorinated phenolic exposure in clear water and turbid water (0.08 g/L). Results describing the effects of increasing suspended sediment concentration upon bioconcentration, along with the effect of removing the organic content and differences between continuously stirred and unstirred bioassays are described in Table 4.7. Comparing the bioconcentrations of chlorinated phenolics between unstirred (Group 1) and continuously stirred (Group 2) bioassays, it can be seen that continuous stirring of the water significantly (p = 0.01, two tailed Student t-Test) facilitated bioconcentration of the chlorinated guaiacols. The same effect was not observed for 3,4,5-TCVer. Qualitatively, it was observed that leeches in the continuously stirred bioassay jar secreted much more slime, a stress indicator (Sawyer 1989), than leeches in the unstirred bioassay system. However, slime production was difficult to quantitate reliably, since much of it remained stuck to the walls of the bioassay jars and tb the leeches themselves. 3 William Duncan, British Columbia Ministry of Environment, Lands and Parks, 1011 Fourth Avenue. 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E nJ Q CO Z « 5 5 T. cn • D • • , I— D D ' Z (— H ' - « ) ' - ( / ) „ m T -> 11 11 S 5^ ^ to 1 5 - - o d> cn co CM d O if) to r-i O O II co S 2 n » ^ o o o ii it II D O crj co a a r r r „ „ aj cn g) cn ^ " 5 S S | | | « io c c '•5 o » ffl 5 A U 0 0 ) 0 6 £ cn cn N (j N N N « 0) W W w 2 S> ~ci Q> a> <ii ~ E I .5 co — — CLE E E « W <0 rj fn CO CO CO 105 N o clear trend was observed when examin ing the effect o f suspended sol ids c o n c e n t r a t i o n on l e e c h b i o c o n c e n t r a t i o n o f c h l o r i n a t e d p h e n o l i c s ove r the concentrat ion range o f 0 - 0.15 g / L , w h i c h covers the seasonal range reported for the M i d d l e Fraser R i v e r . Fo r example, bioconcentrat ion o f a l l compounds was less w i th no suspended sediment than at 0.025 g / L suspended sediment concentrat ions, wh ich was contrary to what was . hypothesized. Figure 4.9 shows the variable nature of observed bioconcentra t ion wi th increas ing sediment concentrat ion for both 3,4,5-T C G and 3 , 4 , 5 - T C V e r . Bo th the veratrole and the guaiacols fo l lowed a s imi lar trend even though the guaiacols are polar and ion izable compounds. Figure 4.9: The effect of suspended sediment concentration on leech (N. obscura) bioconcentration of 3,4,5-TCGand 3,4,5-TCVer 100 0.00 0.04 0.08 0.12 0.16 Suspended Sediments (g/L) + = Standard Deviation 3,4,5-TCG; x = Standard Deviation 3,4,5-TCVer. 106 Available evidence suggests that the observed increase in leech bioconcentration associated with the continuous stirring o f bioassay water can be explained by a combination of two processes: first, physiological response to increasing water flow and second, alterations in the water micro-environment surrounding the leeches. Some leech species, including the erpobdellid, Mooreobdella microstoma, a species in the same family as N. obscura, display a positive rheotaxic response to water current, tending to swim against the current (Sawyer 1989). A similar response was observed in all continuously stirred bioassays, while leeches in the unstirred bioassay tended to remain at rest. The natural habitat of the leeches used in this bioassay is the littoral zone of a small still water lake; the artificially induced current could have resulted in an elevation in physiological activity due to stress of abnormal environmental Condi t ions . Since ventilation volume and blood flow have been found to be determining factors in uptake of contaminants by aquatic organisms (Barron 1990), it is quite plausible that increased swimming behaviour and stress led to increased ventilatory rates, resulting in an increase in the rate of chlorinated phenolic uptake through subcutaneous capillary nets. Agitation of the water micro-layers surrounding the leeches skin could have exacerbated the effect. It has been found (Dainty and House 1966) that there exists a micro-zone of poorly mixed water layers at the membrane-water interface of aquatic organisms, which is theorized to impede the bioconcentration of lipophilic organic molecules. Saarikoski et al. (1986) compared the bioconcentration of PCP in guppies (Poecilia reticulata in both magnetically stirred and unstirred bioassay environments. A 75 % increase in the rate of uptake was observed in the stirred bioassay, which was attributed to enhanced penetration of PCP, due to the elimination of the unstirred water micro-layer. In these experiments, the increase in cardiac output was not thought to be of major importance, since the less lipophilic compounds, butyric acid and phenol, were bioconcentrated at virtually the same rates in both stirred and unstirred bioassays. These findings indicate that leech bioconcentrations of chlorinated phenolics under field conditions may vary 107 considerably , depending upon the ambient water f l ow. O v e r a l l , our f ind ing that leech b ioconcent ra t ion o f ch lor ina ted phenol ics is not related to suspended sediment concentrat ion is contrary to ava i lab le evidence i n d i c a t i n g that b i o a v a i l a b i l i t y o f ch lor ina ted phenol ics is reduced w i t h i nc reas ing suspended par t icu la te concent ra t ions , p r o v i d e d that sediment o rgan ic content i s greater than 0 .1% (Schel lenberg et a l . 1984). Schwarzenbach (1985) found that sed iment p a r t i t i o n i n g o f i o n i z a b l e h y d r o p h o b i c c o m p o u n d s was dependent on organic content o f sediments, water p H and and the ionic composi t ion of the aqueous phase, w i th hydrophob ic pa r t i t i on ing to the par t iculate organic phase be ing the dominant factor. Organic sediment par t i t ioning o f 2 , 4 , 5 - T C P ( p K a = 6.94) and P C P ( p K a = 4.75) was found to decrease as water p H increased above chlorophenol p K a values (Schel lenberg et a l . 1984). Th i s f ind ing is consistent w i th the preferential adsorption o f the un ion ized form to organic particulates. Sediment par t i t ioning o f the i on i zed fraction o f chlor inated phenol ics has been observed to be negl ig ib le as long as the ambient water p H is no more than one unit above the compound p K a . Since the experimental water p H was constant at 7.1 and the percent organic content was w e l l above 0 .1%, we wou ld expect, from the chloroguaiacol p K a values (section 4 .3 ; Table 4.5) , that direct hydrophobic par t i t ioning to particulate organic mater ia l , should have been the dominant process affecting the b ioava i lab i l i ty o f 4 , 5 - D C G , 3,4,5-T C G , 4 , 5 , 6 - T C G and 3 ,4 ,5 -TCVer i n our bioassay system, whi le T e C G ( p K a = 6.0) should show s ignif icant anion b ind ing to suspended sediments. So rp t i on o f the i o n i z e d f rac t ion o f ch lo r ina ted pheno l i c s occurs on both organic and inorganic surfaces and at a f ixed p H is related to the ionic concentration of the ambient water (Schellenberg. et a l . 1984; von Oepen et a l . 1991). Par t i t ioning o f ion ized chlorinated phenol ics to particulates has been shown to occur by direct i o n -exchange on charged surfaces ( X i n g et a l . 1993) or by electrostatic attraction o f i on -pairs to particulate surfaces (Westal l 1985). Phenolate ions may direct ly b ind by i o n exchange or by electrostatic attraction to pos i t ive ly charged inorganic surfaces such as exposed ca lc ium, magnesium, a luminium or i ron atoms on c lay surfaces ( X i n g et al . 1993). The rmodynamica l ly stable ion -pa i r ing o f chlorophenolate anions wi th K + , 108 Na + , C a 2 + and M g 2 + cations present in water has been demonstrated, with ion-pair formation showing a positive relationship to ionic strength of the ambient water (Schellenberg et al. 1984; Westall 1985). In the presence of suspended sediments of organic content > 0.1%, partitioning of the anion species to organic material is the dominant transfer process (Schwarzenbach 1985). Partitioning of PCP anions into octanol was shown to increase with increasing K + concentration (Westall 1985). Schellenberg et al. (1984) found a strong positive relationship between sediment adsorption of 2,3,4,6-TeCP and PCP anions and the fraction of organic material in suspended sediments. From the available evidence we would have expected to observe a decrease in leech bioconcentration due to reduced chlorinated phenolic bioavailability, with increasing suspended sediment concentrations due to the following phenomena, presented in decreasing order of importance: 1) Hydrophobic adsorption of undissociated compounds to organic particulates. 2) Ion pair adsorption to both organic and inorganic clay surfaces. 3) Ion exchange of free phenolate ions to clay surfaces. Observed results indicate that at suspended sediment levels ranging from 0 -0.15 g/L, with a 95:5 ratio of inorganic clay to organic material, sediment partitioning plays little role in bioconcentration of chlorinated phenolics by aquatic organisms such as leeches. This is a possible description of the behaviour the relatively water soluble, ionizable chloroguaiacols, however, the highly lipophilic chloroveratroles should display a distinctly different behaviour. However, 3,4,5-TCVer bioconcentrations followed the same trend as did the chloroguaiacols. This finding could be indicative of an overwhelming experimental artifact, masking the effect of the introduced suspended sediments. It is plausible that such an effect could be related to the profuse leech slime secretions observed in the continuously stirred bioassays. Leeches secrete mucous from various secretory cells; the two most common being the pear shaped mucous cells located just below the integument and 109 the tubular mucous cells located deeper in the musculature (Sawyer 1989). Mucous secretion serves physiological functions such as osmoregulation and excretion, as well as in species recognition and defence against physical danger. The excess leech mucous secretion observed in the continuously stirred bioassays may have been due to stress from the constant water current and skin irritation from the pounding of suspended sediments. Chemically different types of mucous may be secreted in response to various stimuli. In general leech mucous is a mucopolysaccharide complex with protein content in the area of 10% (Sawyer 1989). Prominent carbohydrate monomers can include: glucose, galactose, glucosamine, galactosamine and fucose. Within the bioassay jars mucous was visibly present in large and small pieces both free and attached to leeches. In addition, it was likely that there was a significant mucous surface area present in the form of colloidal material. This abundant pool of high surface area organic material is likely to have had a significant effect on chlorinated phenolic binding, perhaps masking the effects of the increasing suspended sediment load on bioconcentration. Physically associated with mucous particles were suspended sediments. The close bonding of suspended sediments with mucous coupled with difficulties in collection and isolation of mucous made it unreliable to try to quantify the relative role of this artifact in chlorinated phenolic absorption. To control or reduce this effect it would be desirable to expose leeches in a larger system under flow through conditions, where waste products could be removed and the amount of mucous relative to the bioassay water volume would be insignificant. The effect of removing the organic sediment fraction from turbid bioassay water is shown in Figure 4.10, comparing bioconcentrations between leeches exposed under identical concentrations (0.08 g/L) of suspended sediments, both with and without organic material (Group. 4 vs. Group 6). Increased chloroguaiacol bioconcentrations were observed with the absence of organic material. . These results are consistent with the hypothesis that the organic fraction of suspended sediments act as the dominant adsorptive surface in reducing the bioavailability of organic pollutants. However, an opposite, but not statistically significant (p = 0.05), effect was 110 observed for 3,4,5-TCVer. The differences in bioconcentration of chloroguaiacols were directly proportional to the pKa of these compounds (Figure 4.11). 4,5-DCG (pKa = 9), which was ionized to the least extent at water pH 7.1 also showed the greatest increase in bioconcentration between turbid waters with and without introduced Figure 4.10: The effect of the presence of organic material in suspended sediments on the bioconcentration of chlorinated phenolics. Asterisks indicate statistically significant differences (p = 0.01; Student's t-Test). Numerical figures indicate the ratio of bioconcentration in the bioassay with 100 % inorganic sediments to the bioconcentration in the bioassay with 95% inorganic : 5% organic material CD cz co o c o O 0) 3 CO CO o u 40 c o 20 I 1.68 10 + 1.46 1.42 1.13 0.87 4,5-DCG 3,4,5-TCG 4,5,6-TCG TeCG 3,4,5-TCVer • 95% Inorganic : 5% Organic Material 100% Inorganic Material I l l Figure 4.11: Relationship between chloroguaiacol pKa and the ratio of leech (N.obscura) bioconcentration in turbid (0.08 mg/L) bioassay water of 0 % organic particulate content versus 5 % organic particulate material. to E o ' c CO cn t_ o g ' c CO O) o c cu o c o o g in CO rr 2.0 1.8 1.6 1.4 + 1.2 + 1.0 4.0 4,5-DCG e 3,4,5-TCG e e 4,5,6-TCG TeCG 5.0 6.0 — I — 7.0 pKa 8.0 9.0 10.0 organic material. T e C G (pKa = 6.0), wi th the lowest p K a , showed the lowest increase i n bioconcentrat ion between the two bioassays. These results are exact ly what wou ld be expected i f the dominant mechan i sm for transfer o f ch lo r ina ted pheno l ics i n a turbid system is the par t i t ion ing o f the un ion ized hydrophobic species to organic material . A t f ixed water p H , chlorinated phenolics of lower chlorine substitution and higher p K a tend to be present i n the undissoc ia ted form and thus tend to be cont ro l led more by hydrophobic interact ions, such as b ind ing to part iculate organic mater ia l than the more h i g h l y d i ssoc ia ted t r i - and tetra- ch lo r ina ted pheno l i c s . Other h y d r o p h o b i c in te rac t ions w i t h glass surfaces and l eech s l i m e may have con t inued to dominate ch lo rovera t ro le b ioconcen t ra t ion . 112 E x p e r i m e n t s a imed at assessing the ro le o f suspended s o l i d mater ia l i n affecting seasonal b ioava i l ab i l i t y o f chlor inated phenol ic compounds to leeches used as biomonitors in the Fraser R i v e r showed no evidence o f any effect. However , the f indings are questionable on the basis o f the extraneous input o f organic material i n the form o f leech mucous. The results are consistent wi th predictions made by Carey (1988), who determined sediment pa r t i t i on ing coeff ic ients o f several ch lo rophenols at the p H of the Fraser (pH = 7.8). Carey predicted moderate sediment part i t ioning o f chlorophenols dur ing the l o w f low per iod when the fract ion o f organic mater ial i n suspended sediments was at its peak, but taken together wi th the actual ratio o f water to sediment mass, the total percent associated wi th suspended sediments w o u l d be very low. However , his predictions did not go into any detail about the possible role p l ayed by c o l l o i d a l macromolecu les and d i s s o l v e d organic mate r ia l . Labora to ry exper iments r evea led that the presence o f o rgan ic ma te r i a l i n the suspended sediment fraction can reduce b ioava i l ab i l i ty . W h i l e this may not be an issue i n the Fraser R i v e r , where the f rac t ion o f organic mater ia l i n suspended sediments is r e l a t ive ly constant, this factor must be taken into cons idera t ion when c o m p a r i n g b iomon i to r i ng studies f rom different watersheds. Labora tory studies also ind ica ted that bioassay protocol has a profound effect on the measured results as was indic ted by the s ix to eight fo ld increase i n ch lo rogua iaco l b ioconcentra t ion observed when laboratory bioassay was cont inuous ly st irred. Therefore, var ia t ion i n ambient water f low may also be of importance i n interpretation of f ie ld b iomoni tor ing data. 113 4.2. FIELD STUDIES 4.2.1. Temporal Variations in Contaminant Concentrations in Effluent, Water, Suspended Sediments and Leeches In situ studies were conducted both to evaluate leeches as biomonitors under a wide range of seasonal environmental conditions and to generate a data base for comparative study with the results of laboratory bioassays. Field monitoring experiments were carried out in July & October 1991 and February 1992. Monitoring took place on the Fraser River at Prince George B.C. A test site was selected at Stoner B.C., located about 40 km downstream of three BKMs, discharging secondary treated effluent into the Fraser River, from a total of two discharge points. A control site was selected at Shelley B.C., about 10 km upstream of the most northerly of the two outfalls (Section 3; Figure 3.1). A comprehensive sampling program was followed, with composite sampling of treated mill effluent, river water and suspended sediments in addition to caged leeches. A full description of monitoring sites and sampling procedures is given in Methodology, section 3.3. Laboratory experiments revealed that leech bioconcentration of chlorinated phenolics is positively correlated to contaminant concentration (0.1 - 10 ug/L) and water temperature (4 - 20 °C). Negative correlations to water pH and leech weight were observed in laboratory bioassays. The quantity of suspended sediments did not appear to be an important determinant of bioconcentration, although the presence of organic material in suspended sediments did seem to competitively inhibit bioconcentration. Introduction of water current, by means of magnetic stirring, to the laboratory bioassay system resulted in increased bioconcentration. In the context of a Fraser River biomonitoring program, it was expected that contaminant concentration, which is governed primarily by BKM discharge into the river and water flow, would be the. major determinant of bioconcentration. Temperatures in the Fraser vary from about 1 - 20 °C (Hall and Jacob 1988), however, 114 temperature - bioconcentration relationships in the laboratory were weak (r2 = < 0.33) and only affected bioconcentration at T < 12 °C. Water pH is relatively constant in the Fraser River, ranging from 7.0 - 7.8 (Hall and Jacob 1988), and was not expected to be a major factor in determining leech bioconcentrations. Difficulties in obtaining a leech stock of uniform size were encountered, therefore leech weight -bioconcentration relationships were expected to be of importance in the analysis of results. Based on laboratory relationships it was hypothesized that leech bioconcentrations could be predicted from measured water concentrations, with correction factors applied to account for variation in water temperature, pH and leech weight. Exposure conditions during monitoring trials are summarized in Table 4.8. Seasonal water flows ranged from high summer flow to low winter flow on the Fraser River. Water temperature decreased progressively over the study, dropping below the lower range of temperatures tested in laboratory bioassays (4.4 °C) between October 1991 and February 1992. Water pH was relatively constant, varying under 1 pH unit over the study. Turbidity remained high over the July and October study periods, but dropped sharply during the winter low flow period. Mill operations were normal in July and February, however periodic shutdowns occurred in all mills in October 1991. Effluent flow and bleaching process parameters are given in Table 4.9. AOX loading data were calculated from measured effluent flow and AOX concentrations. Increasing chlorine dioxide (CIO2) substitution reduces the formation of both AOX (Axegard 1986) and higher chlorine substituted phenolics (Liebergott et al. 1990). All three Prince George mills were in the process of testing higher C10 2 substitution levels, consequently AOX loading generally decreased through the study. From the C10 2 substitution levels and AOX concentrations it was predicted that effluent concentrations of chlorinated phenolics would be greatest in July 1991. Effluent trends in chlorinated phenolic concentrations (Table 4.10) matched these predictions; July effluent concentrations were more than double those of October and February. High C 1 0 2 substitution combined with secondary effluent treatment did not eliminate chlorinated phenolics 1 1 5 from effluents, except in the case o f 2 , 3 , 4 , 6 - T e C P , wh ich became non-detectable February. Decreased use o f elemental chlorine ( C y in bleaching led to shifts in Table 4.8: Field conditions for the July 8-15, October 17-24, 1991 and February 19-26, 1992 at two monitoring stations on the Fraser River at Prince George, B . C . Parameter July 1991 October 1991 February 1992 Shelley Stoner Shelley Stoner Shelley Stoner River Flow (m3/sec)* 1808 21 84 7 5 0 881 4 4 9 574 Temperature ( °C) 1 1 12 6 6 0 0.5 PH 7.8 7.8 6.9 6.9 7.3 7.3 Turbidity (NTUs) 2 6 2 3 1 7 41 4 6 * Water flow data provided by Water Survey of Canada, Environment Canada. Flow at Stoner station calculated as the sum of flows on the Fraser River at Shelley and the Nechako River at Isle Pierre. Table 4.9: Pulp mill process conditions for the July 8-15, October 17-24, 1991 and February 19-26 1992 at Prince George, B.C. Parameter July 1991 October 1991 February 1992 Outfall A* Outfall B " Outfall A* Outfall B** Outfall A* Outfall B " Avg. Effluent Flowt (m3/day) 166 ,025 163 ,513 125 ,522 140 ,475 137,458 137 ,475 Avg. Percent Chlorine Dioxide Substitution 70 7 0 90 81 90 NA Effluent AOXt (mg/L) 16.5 20 5.5 12 8.5 1 6 AOX Loading (kg/day) 2 7 3 9 3 2 7 0 6 9 0 1686 1168 2 2 0 0 NA = not available. * Outfall A = Canadian Forest Products Ltd. (Canfor) outfall. " Outfall B = Northwood Pulp and Paper Co. outfall. t AOX concentration and efliuent flow supplied by Bill Duncan, B.C. Ministry of the Environment, Lands and Parks, Prince George, B.C. branch. 116 Table 4.10: Flow weighted average effluent concentrations of chlorinated phenolics discharged from Prince George bleached kraft mill outfalls during three seven day monitoring periodsf. Compound Concentration (pg/L) July 1991 October 1991 February 1992 4,5-DCG 1.0 0.5 1.4 3,4,5-TCG 17.3 6.0 5.7 4,5,6-TCG 2.8 1.2 1.6 TeCG 7.8 2.7 0.8 3,4,5-TCVR 2.5 1.6 2.9 5,6-DCV 2.8 0.6 5.6 2,4,6-TCP 5.2 1.2 1.5 2,3,4,6-TeCP 1.2 0.2 ND Total 40.6 14.0 19.5 ND = none detected; refer to Table 3.1 for detection limits, t Results from five day composite samples: July 8-12 and October 17, 21-24, 1991, February 19-21, 24, 25, 1992. chlorine substitution of chlorinated phenolics (Liebergott et al. 1990). Levels of 4,5-DCG and 5,6-DCV increased between July and February, while TeCG and 2,3,4,6-TeCP levels dropped sharply. 3,4,5-TCG, 4,5,6-TCG and 2,4,6-TCP all showed decreases with increasing C 1 0 2 substitution. 3,4,5-TCG was the most prominent marker compound in all effluent samples, which represents a shift in dominance from TeCG, which the was prominent compound from Prince George effluents of the late 1980's (Schreier et al. 1991). Future implementation of chlorine reducing technology, such as 100 % CIO2 and 0 2 delignification may lead to shifts towards dominance of di- and monochlorinated species. 117 Despite high CIO2 substi tution, Fraser R i v e r water samples from the Stoner test station consistently showed measurable levels o f 3 , 4 , 5 - T C G , 4 , 5 , 6 - T C G and T e C G , in s imi lar proportions to those observed i n effluent samples (Table 4.11). A s in the case of the effluents, 3 , 4 , 5 - T C G appeared to be the most prominent marker o f B K M derived chlor inated phenol ic discharge to the Fraser R i v e r . Chlor ina ted phenol ic compounds were not detected at the upst ream S h e l l e y con t ro l s ta t ion . R e l a t i v e l y h i g h concentrat ions o f 5 , 6 - D C V were detected i n water samples, but i n much greater re la t ive propor t ion to what was detected i n effluent samples, a poss ible i nd i ca t i on that there were a n a l y t i c a l in terferences unreso lvab le by dua l c a p i l l a r y c o l u m n ana lys i s . A l t e r n a t i v e l y , a p r e v i o u s l y u n k n o w n source o f 5 , 6 - D C V , such as degradation o f A O X discharged to the r iver , cou ld expla in the observations. Water c o n c e n t r a t i o n o f c h l o r o g u a i a c o l s f o l l o w e d the d e c r e a s i n g t rend i n e f f luen t concentrations between Ju ly and October. A sharp increase i n both the number o f chemica l species detected and water concentrat ion was observed i n February, w h i c h can probably be attributed to less d i lu t ion o f the l ow river f low during this period. Table 4.11: Seven day mean water concentrations of. chlorinated phenolics detected in the Fraser River at Stoner, B.C. No target compounds detected in water samples from the control station located at Shelley, B.C. Compound July 1991* Concentration (ug/L) October 1991* February 1992f 4,5-DCG ND ND 0.073 ± 0.009 3,4,5-TCG 0.052 ± 0.005 0.043 ± 0.005 0.067 ± 0.016 4,5,6-TCG 0.002 + 0.001 0.006 ± 0.003 0.015 ± 0.003 TeCG 0.024 + 0.001 0.010 + 0.002 0.012 ± 0.003 3,4,5-TCVer ND ND 0.024 ± 0.003 5,6-DCV 0.036 ± 0.004 0.071 + 0.021 0.467 + 0.018 2,4,6-TCP ND ND 0.017 ± 0.002 2,3,4,6-TeCP ND ND ND ND = none detected; refer to Table 3.1 for detection limits. •July and October 1991 results from analysis of daily composite samples, n = 7 ( 3 samples composited per day). tFebruary 1992 results from analysis of four grab water samples; Feb.19, 21, 24, 26. 118 High concentrations of chlorinated phenolics present in water from the February low flow period carried over to. the suspended sediment samples (Table 4.12). February suspended sediments concentrations of 3,4,5-TCG and TeCG increased by 68 and 36 times respectively over July levels. As in the case of water samples, 5,6-DCV appeared in concentrations out of proportion to effluent levels. Note that 3,4,5-TCVer was not recovered from sediment spiked material by the analytical procedures followed and therefore, could not be monitored in suspended sediments. The neutral lipophilic veratrole family of compounds may not be easily separable from the sediment phase, using the alkaline buffer extraction employed. Further testing, using organic solvent extraction failed to recover spiked compound, indicating that veratroles may bind by a different mechanism to suspended sediments. Table 4.12: Suspended sediment concentrations of chlorinated phenolics and percent organic contant in the Fraser River; July 8-15, 1991 and February 19-26, 1992. Results reported on a dry weight basis. Compound Concentration (pg/kg) July 1991* February 1992** 4,5-DCG ND 95 3,4,5-TCG 0.91 62 4,5,6-TCG ND 8.4 TeCG 0.36 1 3 3,4,5-TCVer NR NR 5,6-DCV ND 1 76 2,4,6-TCP ND 3.0 2,3,4,6-TeCP ND ND Organic Content (%)tt 3,1 7.8 ND = none detected; refer to Table 3.1 tor detection limits. NR = not recovered in sediments by extraction procedure used. * July 1991 study; no control samples taken. ** February 1992 study; 2,4,6-TCP (0.63 ug/kg) found in control sample, tt Measured as loss on ignition. 119 Size fractionation of July suspended sediment material revealed a composition of 4.3% coarse sand (> 250 u-m), 78.5% fine sand (> 63 pm < 250 pm) and 17.2% silt (< 63 um). February suspended sediments were not size fractioned due to a lack of enough material. Qualitative X-ray diffraction analysis was carried out on the silt fraction of July suspended sediments to better characterize clay components. Prominent clay minerals were chlorite and illite in an approximately 3:2 ratio. Also prominent were quartz, feldspar and horneblend components. The percent volatile organic material, as measured by the loss on ignition, was-3.1% in July 1991 and 7.8% in February 1992, which is consistent for the 1.6% and 6.0% organic fractions predicted by Carey (1988) for spring freshet and winter low flow periods on the Fraser. A lower organic particulate content during the high flow period is related to the increased input of low carbon mineral components eroded from bed sediments and river banks during freshet periods (Carey 1988). Sediment partitioning of both neutral and ionizable hydrophobic compounds is dependent primarily on the organic content of sediments (Schwarzenbach 1985). Schellenberg et al. (1984) found a strong positive relationship between sediment adsorption of 2,3,4,6-TeCP and PCP and the fraction of organic material in suspended sediments. In the case of ionizable hydrophobic compounds, such as the chlorinated phenolics, sediment partitioning is also dependent upon water pH and ionic concentration (Schellenberg et al. 1984; von Oepen et al. 1991). The unionized fraction preferentially adsorbs to suspended particulates (Schellenberg et al. 1984), although stable ion-pairs (e.g. phenolate-K+ pair) can also partition to the sediment phase (Westall 1985; Xing et al. 1993). Preferential partitioning of the unionized fraction is consistent with the observation that the ratio of 4,5-DCG (pKa = 9.0) to 3,4,5-TCG (pKa = 8.0), 4,5,6-TCG (pKa = 7.4) and TeCG is greater in sediment than in water samples from February 1992 (i.e. there is an increased rate of sediment partitioning as pKa increases). The increase in suspended sediment bound compound between July 1991 and February 1992 was likely associated with the increase in the percent organic material in suspended sediments. Using the following equations developed by 120 Karickhoff et al.(1979) to describe sediment partitioning of hydrophobic pollutants to sediments: Eqn. 6 Ksw = Koc* foe Eqn. 7 Koc = Kow* 0.63 Where: Ksw = sediment water partition coefficient, Koc = carbon normalized partition coefficient, foe = fraction of organic carbon and Kow = octanol water partition coefficient, increased Ksw values are predicted for 3,4,5-TCG and TeCG during the February low flow Table 4.13. The equations were developed to describe the behavior of neutral organic compounds, however they give an idea of the possible seasonal differences in sediment partitioning behavior of chlorinated phenolics. From Table 4.13 there should be a 2.5 fold increase in Ksw between July and February, considerably less than the measured 68 and 36 fold increase in sediment concentrations of 3,4,5-TCG and TeCG. Table 4.13: Predicted sediment-water partition coefficients (Ksw) for 3,4,5-TCG and TeCG in the Fraser River at Stoner, BC. in July 1991 and February 1992. Sampling Period Fraction Organic Ksw Carbon (foe) 3,4,5-TCG TeCG Jul. 1991 0.031 550 2700 Feb. 1992 0.078 1380 6800 Kow 3,4,5-TCG = 4.45; Kow TeCG = 5.14. 121 Sediment concentration o f hydrophobic organic compounds is related to part icle size, wi th the greatest part i t ioning associated with the fine silt fraction (< 63 pm) (Vo ice et a l . 1983; Schellenberg et a l . 1984). Differences i n size composi t ion of the suspended sediment load probably accounted for the remain ing increases i n February sediment concentra t ions . A l t h o u g h February suspended sediments qua l i t a t ive ly appeared to be much finer than Ju ly sediments, there was not enough sample to permit size fractionation, so this hypothesis could not be tested. To ta l suspended sol ids concentrations i n the Fraser R i v e r were not measured i n this study, however both Carey (1988) and H a l l et a l . (1991) report the seasonal range o f suspended sediments to be from about 15 - 150 m g / L . If i t is assumed that The Ju ly 1991 and February 1992 mon i to r i ng tr ials corresponded rough ly to the above range i n suspended sediment concen t ra t ions , i t is es t imated, u s i n g the suspended sediment concentrations obtained for 3 , 4 , 5 - T C G i n July 1991 and February 1992, that from 0.3 - 1.4% of 3 , 4 , 5 - T C G is adsorbed to the sediment load, whi le the remain ing 99.7 - 98 .6% is carr ied i n the water co lumn. Therefore, the effect o f s ed imen t p a r t i t i o n i n g on l e e c h b i o c o n c e n t r a t i o n m a y have been n e g l i g i b l e . Labora tory investigations (section 4.1.5; Table 4.7) also support this conc lus ion , since no c o r r e l a t i o n was observed between b i o c o n c e n t r a t i o n and suspended sediment concentration (25 - 150 m g / L ; 5% organic content). However , o f the quantity carried i n the water c o l u m n , a poo r ly def ined fract ion is p robably adsorbed to organic c o l l o i d s (e.g. b ioso l ids ) and macromolecu les (Derksen personal communica t ion ) to smal l to be col lected i n the sediment traps used i n the current study 4 . This remains an unquanti tated poo l o f compound , non-ava i l ab le to aquatic organisms such as leeches. L e e c h e s b i o c o n c e n t r a t e d r e l a t i v e l y h i g h l e v e l s o f c h l o r i n a t e d p h e n o l i c compounds dur ing a l l three moni tor ing trials (Table 4.14). B ioconcen t ra t ion factors ( B C F ) ranged from 465 - 6000 times water concentrations. A s expected from effluent and water prof i les , the t r i - and tetrachloroguaiacols were detected most consis tent ly by leeches at the Stoner site. In addit ion the observed proportions o f chloroguaiacols present in leeches were similar to levels detected in fish tissues by Rogers et a l . 4 George Derksen, Environment Canada, 1992. 122 c o CO t o C L CD CD O c CL OS I— CO > CO CC 3 >_ 0) CD 0) co DC co LL c\i CD S .b: CO -a w CD ' 8 CO C L X CO 3 CD U _ "O C CO 8 I x: fl) CM cn O XI C O JB TJ Q. O "8 £ "co . CO o >, 3 O M — o "55 3, CO o c CD O £Z o o g CO C L CO -a c CD > CD CO CO i— o § 9 S tz CO CO XI ca cn CD 5 CD 5 CM O) i— CO 3 h-X3 CD CO co o co CM r~ o 00 < CO CO o m z o i— i— c o CD O o O CD 3 co CO CD .Q O o O >> - 3 C O c CD O c o O CD 3 co co CO CO + 1 + 1 + 1 o •* CM CO CM T— CO CO CM -H CO "> ° < < < < ID <D W 5 5 2 7 O) LO c o c CD O c o O CD 3 CO CO "O c 3 O a. E o O in O) CO r1 d + 1 +i + 1 +1 o CM 00 t^" in CD ^— Q CO eo in o o o CO in eo < < < < tv. CM in + 1 + 1 +i CO CM T— CM CM 00 + 1 a o a in a o in co" (3 O H co m" •*" CD > O h-in •<t co" > O Q co in O h-co t" CM" D_ O CO I -I CO co" CM" CL o ~ a. 2 P \% r~ •* ^ CM a. O a co 6 a co C5 - 5 J= c OJ CO '3 fi 3 t Co CO E 11 . - c °> © II N E Ss "a CL O H cc • a a. co co 3. 2!. Q. O © CD CO Q CO UJ c o » » 5 « | -a o 3 to t CO © 8 I 2 c CO © s -g » $ CO .© II ft II S- » S « m M CO £ © •c •£ -c CO S co 1 if 2 c 3 c S T3 T -- = o a -s O) o. ^ u. 1 1 » f O O CO z .£ o> .E T> 2 -O O CD to £" T3 oi Ji £ 2 B >. to O 123 (1988a) and Dwernychuk et al. (1991). No chlorinated guaiacols were detected in control leeches. However, Shelley leeches were successful in detecting upstream contamination of tri- and tetrachlorophenols during all three monitoring trials, after water testing had revealed no detectable traces. Note that measurable levels of neither 4,5-DCG nor 5,6-DCV were bioconcentrated by leeches in any field trial, even though these compounds were detected in effluents, at all times. The apparent high water concentration of 5,6-DCV coupled with its absence from leeches can be interpreted as either evidence that 5,6-DCV is transported in the Fraser in a non-bioavailable form or it is simply not bioconcentrated effectively by leeches. The latter explanation is favoured by laboratory studies where the compound was difficult to detect in leeches at water concentrations < 5 pg/L. The relatively low GC-ECD response of 5,6-DCV also inhibits the study of this compound; detection limits (20 -50 pg/kg) in leeches range to twenty times above those for tri- and tetrachloroguaiacols ( 1 - 3 |xg/kg). 4,5-DCG has a relatively lower detection limit (3 -5 pg/kg), however, its lower log Kow (3.41) make it less bioavailable to aquatic organisms. Both in situ monitoring and laboratory bioassays indicate that leeches are not effective biomonitors of either 5,6-DCV or 4,5-DCG at present levels of analytical sensitivity. One of the criteria of a good biomonitor is the ability to concentrate target compounds in proportions reflective of ambient levels (Phillips 1978). Leeches bioconcentrated tri- and tetrachloroguaiacols and 3,4,5-TCVer in proportions similar to both measured effluent and Fraser River water concentrations (Figure 4.12). Leech bioconcentrations followed decreasing proportions of TeCG and increasing proportion of 4,5,6-TCG associated with increased C10 2 substitution employed by the mills. Chlorinated phenolics were detected in water, suspended sediments and leeches throughout our monitoring study. The results indicate that the winter low flow period on the Fraser River may be the critical exposure time for many native organisms, since the highest concentrations of compounds were found during this period. In support of our findings, monitoring of juvenile chinook salmon on the 124 Figure 4.12: Relative concentrations of chlorinated phenolics in leeches and water samples from Stoner, B C . and effluents from two Prince George B K M outfalls, over three seasonal monitoring periods. July 1991 October 1991 February 1992 0.08 ^ 0.07 ~ 0.06 + B 0.05 co *" 0.04 0.03 0.02 0.01 0.00 co o c o O 1_ CD CO July 1991 Water n October 1991 February 1992 zt CZ o c CO o c o o cl CD _ 3 3= LU July 1991 Effluent n October 1991 February 1992 125 Fraser R i v e r by Rogers et a l . (1988a) ind ica ted that body burdens o f ch lor ina ted phenol ics were higher dur ing the winter l o w f low per iod than dur ing med ium f low periods i n the fa l l . v Some of the chlorinated phenolics ( T e C P and P C P ) are known to be uncouplers o f o x i d a t i v e p h o s p h o r y l a t i o n ( K e n n e d y 1989) , r a i s i n g the main tenance energy demand upon f ish. The effect on juven i le sockeye salmon (Oncorhynchus nerka) growth rate was found to be m i n i m a l at very l o w (0.7 °C) temperatures, however increas ing water temperature may result i n measurable changes i n metabol i sm and fish growth rates (Webb and Brett 1973). Therefore, impacts o f higher winter body burdens o f chlor ina ted phenol ics may not be felt un t i l the spr ing runof f pe r iod , when water temperatures and f i sh metabol ic rates are elevated. Ev idence suggests that year round moni tor ing o f such dynamic systems as the Fraser R i v e r is required, i n order to identify and isolate cr i t ica l exposure periods. 4 . 2 . 2 . In Situ B i o c o n c e n t r a t i o n a s a F u n c t i o n o f C o n t a m i n a n t C o n c e n t r a t i o n W h i l e i n l abo ra to ry b ioassays , b i o c o n c e n t r a t i o n inc reased l o g a r i t h m i c a l l y w i t h water c o n c e n t r a t i o n , o n l y a s l i g h t decrease i n B C F (the ra t io o f b ioconcent ra t ion to water concentrat ion) was observed over the exposure range o f 0.1 - 10 p g / L ; B C F was re la t ive ly constant over a l l water concentrations Table 4.3). However , there appeared to be a trend towards decreasing B C F wi th increasing water concentrat ion in f ie ld studies. Together laboratory and f ie ld data show a re la t ive ly s t rong r e l a t i o n s h i p ( r 2 = 0.78) between B C F and water concentration o f 4 , 5 , 6 - T C G (Figure 4 .13) . A t the lower water concentrat ions moni tored i n the f i e l d , there appeared to be an increase i n the B C F , w h i c h is an ind i ca t i on that there is a departure from l inear i ty o f the bioconcentra t ion vs. water concentrat ion 126 relationships derived in laboratory studies (Figure 4.14). As water concentration decreases below the lowest level tested in laboratory bioassays (0.1 p g / L ) , b ioconcentra t ion predic t ions w i l l increas ing ly underest imate f ie ld bioconcentrations. ~ —• • i Figure 4.13: Leech (N. obscura) bioconcentration factor (BCF) as a function of water concentration for 4,5,6-TCG detected in field studies (x) for July, October 1991 and February 1992 and . laboratory studiest(O). 10000 o CO q 1000 Jul. 1991 100 • 0.001 Oct. 1991 o o -0 0.010 0.100 1 000 Log Water Concentration (u.g/L) t Laboratory leeches exposed lor seven days at T = 12.5 °C, pH = 7.5. Figure 4.14: Leech (N. obscura) bioconcentration as a function of water concentration for 4,5,6-TCG detected in field studies (x) for July, October 1991 and February 1992 and laboratory studiesf (O). 10000 0.001 0.010 0.100 1.000 Log Water Concentration (p,g/L) t Laboratory leeches exposed tor seven days at T = 12.5 °C, pH = 7.5. 127 In the case o f 4 , 5 , 6 - T C G , the magnitude o f error is not accurately depicted in Figure 4 .14, s ince f i e l d condi t ions var ied marked ly both between f i e ld t r ials and f rom laboratory condi t ions . Fo r example, low water temperatures i n October 1991 (6 °C) and February 1992 (0.5 °C) probably resulted in lower bioconcentrat ions than what was observed at the laboratory temperature o f 12.5 °C . In the case o f 3 , 4 , 5 - T C G , water concentrations (0.043 - 0.067 p g / L ) d id not deviate greatly from the lowest laboratory assayed levels . However , higher than expected in situ bioconcentrat ions o f 3 , 4 , 5 - T C G were s t i l l observed (Figure 4.15). It appears that there are in situ e n v i r o n m e n t a l factors unaccounted for i n laboratory studies, wh ich result i n higher than predicted b i o c o n c e n t r a t i o n s . Figure 4.15: Leech (JN. obscura) bioconcentration as a function of water concentration for 3,4,5-TCG detected in field studies (x) for July, October 1991 and February 1992 and laboratory studiest(O). t Laboratory leeches exposed for seven days at T = 12.5 °C, pH = 7.5. 128 Insight into the hypo thes ized u n k n o w n var iab les r e su l t ing i n greater than expected in situ bioconcentrat ions can be gained from the testing conducted to ver i fy that leech bioconcentrat ion should be equal across the breadth o f the r iver . B K M effluent streams do not instantaneously m i x wi th the bu lk r iver f l o w . A e r i a l photographs o f the Nor thwood Pu lp and Paper Company outfal l indicate that the effluent plume remains v i s i b l y dis t inct from the Fraser R i v e r for at least 1 k m ( D w e r n y c h u k and L e v y 1994). The Stoner moni to r ing station, located about 40 k m downstream of the Pr ince George B K M outfalls was selected on the assumption that effluents w o u l d be complete ly m i x e d wi th the Fraser R i v e r at that point. Indeed, s o d i u m tracer data ind ica te that comple te eff luent m i x i n g occurs about 7 k m downstream of the confluence o f the Nechako and Fraser R i v e r s (Dwernychuk and L e v y 1994). A l s o the presence o f a strong mix ing zone at R e d R o c k Canyon , about 10 k m upstream of Stoner, reduces the probabi l i ty o f incomplete effluent m i x i n g , even at low r iver f low. Based upon concern for the security of water sampl ing equipment, the lack o f a second automatic water sampler and logis t ica l problems it was decided to conduct comprehensive water sampling on on ly one side o f the Fraser. T o test the assumption o f homogeneous effluent m i x i n g leech b iomon i to r i ng was conducted on both sides o f the Fraser i n July and October 1991. It was hypothesized that leech bioconcentra t ions should show no s igni f icant differences across the breadth o f the r i v e r . S t a t i s t i ca l ly s ign i f i can t differences i n l eech b ioconcen t ra t ion were observed between moni tor ing stations at opposite sides o f the Fraser R i v e r at Stoner i n both Ju ly and October 1991 (Table 4.15). Leech bioconcentrations o f chloroguaiacols were highest on the east side of the river i n Ju ly , while the opposite trend was observed i n October . Ju ly b ioconcentra t ions o f 2 , 4 , 6 - T C P d id not f o l l o w the trend o f other compounds, being s l igh t ly greater i n west side leeches. The cross channel differences observed i n leech b ioconcent ra t ion cannot be accounted for by in te r sample d i f ferences i n l e e c h we igh t . F r o m reg res s ion equat ions r e l a t ing l eech weigh t to b i o co n cen t r a t i o n , i t was found that l igh te r leeches are more effective bioconcehtrators. However , the l ighter west side leeches 129 bioconcentrated less than the heavier east side leeches in July 1991. October leech sample groups did not differ enough in weight to cause any notable differences in bioconcentration. Higher bioconcentrations of chlorinated phenolics in leeches on one side of the river could be explained by incomplete effluent mixing resulting in cross channel differences in water concentrations. This theory cannot be verified, since water sampling was only conducted on one side of the river. Table 4.15: Cross channel differences in tissue concentrations of chlorinated phenolics in leeches (NL obscura) exposed for seven day periods in the Fraser River at Stoner, BC. Results reported on a wet weight basis. Results in bold text indicate significant differences (two-tailed Student's t-Test; p = 0.05) between east and west side test groups. Compound Tissue Concentration (pg/kg) July 1991 October 1991 East side* West side** East sidef West s ide f t 4,5-DCG ND ND ND ND 3,4,5-TCG 49 ± 9 28 ± 7 16 + 4 20 ± 3 4,5,6-TCG 20 + 3 12 ± 2 3.7 + 1.8 5.8 + 1.5 TeCG 30 + 5 21 ± 5 4.2 ± 1.9 6.4 ± 1.9 3,4,5-TCVer ND ND 3.0 + 1.7 1.9 + 0.9 5,6-DCV ND ND ND ND 2,4,6-TCP 17 ± 3 19 + 8 ND ND 2,3,4,6-TeCP 4.7 ± 0.7 3.4 + 0.6 9.8 + 1.6 11+2 ND = None detected; refer to Table 3.1 for dectection limits. * Sample size n = 19; mean weight ± SD = 1.18 + 0.14 g. ** Sample size n = 9; mean weight ± SD = 1.07 + 0.14 g. t Sample size n = 10; mean weight ± SD = 0.85 + 0.09 g. t t Sample size n = 10; mean weight ± SD = 0.872 + 0.15 g Alternatively, the differences, could also be explained by micro-environmental differences between monitoring stations, leading to enhanced bioconcentration at one station over the other. While it is not likely that sampling sites differed significantly in either temperature or pH, field notes indicated notable differences in water current. No attempts were made to measure this variable, since it had not 130 b e e n o r i g i n a l l y h y p o t h e s i z e d that w a t e r c u r r e n t w o u l d a f fec t l e e c h b ioconcent ra t ion . Howeve r , there is evidence from laboratory bioassays from this and other invest igat ions , w h i c h support the theory that water f low cou ld have had substantial effect on the uptake rate o f chlor inated phenol ics . In July 1991 lower bioconcentrations were observed from the west side, where leeches were stationed i n a l ow f low backwater area. East side leeches had been stationed i n a h igh f low zone. Ev idence from laboratory invest igat ions , conducted after comple t ion o f f ie ld t r ials , showed that when bioassays were conducted under c o n t i n u o u s s t i r r i n g c o n d i t i o n s ( i . e . p r e s e n c e o f w a t e r c u r r e n t ) l e e c h bioconcentrat ions o f t r i - and tetrachlorinated phenol ics ranged from 5.2 - 7.8 times greater than those f rom uns t i r red b ioassay water ( S e c t i o n 4 .1 .5 ; T a b l e 4 .7 ) . Inves t iga t ions i n to b i o c o n c e n t r a t i o n o f P C P by guppies (Poecilia reticulata) (Saar ikosk i et a l . 1986) revealed s imi la r b ioconcentra t ion differences between st irred and unstirred laboratory bioassays; P C P bioconcentrat ion increased by 75 % wi th the presence o f a current. A mic ro -zone o f poor ly m i x e d water layers exists at the membrane-water interface o f aquat ic o rgan i sms , w h i c h is may impede the b i o c o n c e n t r a t i o n o f l i p o p h i l i c organic molecules (Dainty and House 1966). The enhancing effect o f water current can be attributed to the agitation o f the water micro- layers surrounding the outer epi thel ium, increasing the rate of uptake (Saar ikoski et a l . 1986). Since leeches take up o x y g e n and xenob io t i c through "pores" scattered across the e p i t h e l i u m (Sawyer 1989), a disrupt ion o f surface diffusion barriers cou ld result i n a substantial increase i n b ioconcen t r a t i on . Enhanced uptake rates i n areas o f stronger water f low cou ld also have been exacerbated by stress, related phys io log i ca l response to water current. Some leech species, i n c l u d i n g the e rpobde l l id , Mooreobdella microstoma, a species i n the same family as N. obscura, d isplay a posi t ive rheotaxic response to water current, tending to s w i m against the current (Sawyer 1989). A s imi l a r response was observed i n labora tory bioassays. Leeches i n con t inuous ly s t i rred bioassays tended to s w i m act ively, whi le leeches in the unstirred bioassay tended to remain at rest. 131 The natural habitat of the leeches used in this bioassay is the littoral zone of a small still water lake; the artificially induced current could have resulted in an elevation in physiological activity due to stress of abnormal environmental conditions. Since ventilation volume and blood flow have been found to be one of the determining factors in uptake of contaminants by aquatic organisms (Barron 1990), it is quite plausible that increased swimming behaviour and stress led to increased ventilatory rates, resulting in an increase in the rate of chlorinated phenolic uptake through subcutaneous capillary nets. Both laboratory investigations and field observations support the theory that water flow may have a significant effect on leech bioconcentration due to a combination a physiological response to increasing water flow and alterations in the water micro-environment surrounding the leeches. This factor alone may account for the discrepancy between leech bioconcentrations observed in the field and those measured in laboratory bioassays (Figures 4.14 & 4.15). 132 4 . 2 . 3 . In Situ B i o c o n c e n t r a t i o n a s a F u n c t i o n o f L e e c h W e i g h t In order to val idate the inverse re la t ionships between b ioconcent ra t ion and leech weight documented i n laboratory bioassays, this re la t ionship was invest igated during the July 1991 f ie ld study. Due to a lack o f ava i lab i l i ty of small leeches, the f ie ld invest igat ion only tested leech samples ranging in mean weight from 0.63 - 1.94 g, wh i l e laboratory invest igat ions had ranged down to a weight o f about 0.15 g. Inverse r e l a t i o n s h i p s be tween c h l o r i n a t e d p h e n o l i c b i o c o n c e n t r a t i o n and l e e c h weight , s imi l a r to those reported for the laboratory studies, were observed under f i e l d cond i t ions (Table 4 .16) . Labo ra to ry inves t iga t ions revea led no s ign i f i can t re la t ionsh ip between b ioconcen t ra t ion and leech weight for 3 , 4 , 5 - T C V e r , however this could not be confi rmed by f ie ld moni tor ing , since no detectable quantities of the veratrole were detected i n July f ie ld exposed leeches. A s expected from their s imi lar c h e m i c a l s t ruc tu res , t r i - and t e t r a c h l o r o p h e n o l s b e h a v e d s i m i l a r l y to the ch loroguaiacols , al though 2 , 4 , 6 - T C P showed a shal lower slope and lower corre la t ion coefficient than other test compounds. In general, in situ b i o c o n c e n t r a t i o n s s h o w e d less sens i t iv i ty to leech weight , as manifested by shal lower regression slopes, than d i d l abora tory b ioconcen t ra t ions (F igu re 4 .16) . The di f ference was e s p e c i a l l y apparent for T e C G . The decreased slopes may not reflect true differences between f i e ld and laboratory relat ionships, since the f ie ld studies encompassed less than ha l f the weight range examined i n the laboratory. Compar i son o f the laboratory and f ie ld relat ionships, i n the form o f the predicted percent decrease i n bioconcentra t ion over the weight range 0.5 - 1.0 g is provided i n Table 4.17. Direc t comparison, i n terms o f absolute b ioconcen t r a t i ons were not p o s s i b l e , s ince the l abo ra to ry and f i e l d condit ions varied markedly i n terms o f water concentration (i.e. y intercept a o f y = a(X)exp.b) . M a r k e d differences are only apparent for T e C G , over the weight range 0.5 - 1.0 g. 133 CO >. * d CQ CD o 55 to CD > be CD CO CO 1 Li. CD +^ C CD CO i_ co Z3 CO CO X i O C L £ ex cn ex CD ys CO T 3 CD C CD CO > CD c CO o I" g co 13 cr LU co co CD CO CD rr O) ~c5) c o -t—» CO 4—' c CD CJ c o O CD W W TJ C o C L E o o X x co o o LO o d CD O Q CO LO tf CO • CO d d d d d CD d tf +1 CD U O X c c CO tf ^ CVJ LO +1 +1 +1 CO CM , -CM eg tf m +i +i +i tf r- in IO CM CD Q > o h-uS tf" co" Q 9 i > o o tf d T- tf CL O ^ CM r, "S £ 134 Figure 4.16: Bioconcentration of 3 ,4 ,5-TCG (a) and T e C G (b) by leeches (N. obscura) of differing weights, exposed under both laboratory (water concentration, 1.0 ug/L; temperature = 12.5 °C; pH = 7.5; n = 19) and field conditions (water concentration = 0.052 pg/L; temperature = 12.0 °C; pH =7.8; n = 28). a) "9) JZ O CD o o O CD ZJ co co O 1 0 0 0 1 0 0 t 1 0 3,4,5-TCG - LAB = O 3,4,5-TCG - FIELD = • 0.1 1.0 Log Leech Weight (g) 10 .0 : Standard Deviation 3,4,5-TCG - LAB; x = Standard Deviation 3,4,5-TCG - FIELD. b) C o CD CJ c o O CD co co o 1 0 0 0 1 ^ + 1 0 0 + 1 0 0.1 1.0 Log Leech Weight (g) 10 .0 + = Standard Deviation TeCG - LAB; x = Standard Deviation TeCG - FIELD. 135 Table 4.17: Predicted decreases in bioconcentration of chlorinated phenolics by leeches (N_. obscura) between 0.5 g and 1.0 g under both laboratory* and field conditions**. Predicted bioconcentrations calculated from regression relationships developed for laboratory bioassys (Table 4.5) and Fraser River field bioassys (Table 4.18). Laboratory Study Field Study Compound Predicted Bioconcentration (u.g/kg) Decrease Predicted Bioconcentration (u.g/kg) Decrease Leech Wt. = 0.5 g Leech Wt. = 1.0 g (%) Leech Wt. = 0.5 g Leech Wt. = 1.0 g (%) 3,4 ,5-TCG 121 68.0 - 43.8 53.2 3 3 - 40.0 4,5 ,6 -TCG 446 3 0 7 - 31.2 18.2 12 - 34.1 TeCG 3 3 3 202 - 39.3 27.5 20 - 27.3 * Semi-static seven day exposure; water concentration = 1.0 ug/L, temperature = 12.5 °C, pH = 7.5. " Seven day caged field exposure at Stoner BC. July 8 - 15, 1991; water concentrations: 3,4,5-TCG = 0.052 ug/L, 4,5,6-TCG = 0.002 ug/L, TeCG = 0.024 ug/L, temperature = 12.0 °C, pH = 7.8. The combined effect of the following factors were argued to be responsible for the enhanced bioconcentration observed for small leeches observed in laboratory bioassays: 1) greater rate of 0 2 consumption (Davies et al. 1987; Sawyer 1989) leading to an increased rate of xenobio'tic uptake at surface respiratory sites (Barron 1990; Connell 1991), 2) decreased content of detoxifying betroiydal tissue (primitive liver homologue), leading to a lower rate of depuration (Sawyer 1989), 3) greater uptake surface area to volume ratio (Saarikoski et al. 1986), 4) lower biomass to contaminant ratio for bioassays containing the lighter leech weight groups. There is no reason to believe that the first three phenomena do not occur in field exposed leeches. The fourth factor is an artifact of the laboratory bioassay protocol and is not a factor in the continuous flow conditions of the Fraser River. In laboratory bioassays, leeches were segregated into four weight groups (n = 5 per group) of mean weights, 0.148 g, 0.303 g, 1.23 g and 1.77 g. This corresponds to biomass loadings ranging from 0.74 -8.85 g/L of bioassay water. Therefore, there was more compound per unit biomass available to leeches in the light weight groups in effect, increasing the apparent exposure concentration to lighter leech groups. The magnitude of the the effect is uncertain, however it may have been of little significance, since water was 136 completely changed every 24 h and field exposures showed similar relationships. A biomass effect could account for the slightly greater rate of change in bioconcentration per unit change in leech weight observed in laboratory relative to in situ bioassays (Figure 4.16), although the limited range of leech weight examined in the field make it difficult to compare the two studies. 4 . 2 . 4 . I n t e r s p e c i e s D i f f e r e n c e s i n B i o c o n c e n t r a t i o n Criteria for the selection of an appropriate leech species for routine biomonitoring of water contaminants should include: sensitivity to the presence of target compounds, with respect to both toxicity and bioconcentration, suitability for routine laboratory analysis, maintenance in the laboratory and availability of uncontaminated stock. Nephelopsis obscura was selected as the primary biomonitoring organism for this study, because early pilot investigations indicted that all the above criteria were met. For comparative purposes, a second leech species, Percymoorensis marmorata, was evaluated against N. obscura during the July 1991 field study. Refer to section 2.4 for a comparative description of leech biology between leech species. No clear interspecies trend was observed. Table 4.18, indicates that N. obscura was a more effective bioconcentrator of 4,5,6-TCG, TeCG and 2,3,4,6-TeCP, while P. marmorata attained greater concentrations of 3,4,5-TCG and equal levels of 2,4,6-TCP. Laboratory bioassays, conducted at a water concentrations of 1.0 pg/L, revealed a different pattern (Table 4.19), in which P. marmorataa.tta.ined greater bioconcentrations of all chloroguaiacols. The dominance of P. marmorata under laboratory conditions cannot be explained by differences in leech weight, since the mean weight of N. obscura (1.23 g) was about half of that for P. marmorata (2.66 g). Furthermore, leeches used in both laboratory and field studies were maintained under identical conditions and appeared to be in similar physical condition prior to the experiments. Therefore, differences between laboratory and 137 f i e ld results can be expla ined by the effect o f differ ing environmental condi t ions on uptake and depurat ion p h y s i o l o g y . There are many b i o l o g i c a l factors thought to exert affect i n t e r spec ies d i f ferences i n b i o c o n c e n t r a t i o n i n c l u d i n g : surface area to v o l u m e ra t io , l i p i d content, ac t iv i ty o f and mechanisms o f depurative metabol ism ( C o n n e l l 1991), b lood f low rate and d i s t r ibu t ion pattern, and respiratory rate (Bar ron 1990). Ce r t a in factors, such as respiratory rate and b lood f low are important rate l i m i t i n g factors, w h i c h are inf luenced, i n the short term, by environmental condi t ions . F o r example , increased vent i la tory vo lume and b l o o d f low rate to the g i l l , brought about by warmer water temperature, may lead to increased bioconcentra t ions (Bar ron 1990). The effect o f water temperature on respiratory phys io logy may exp la in interspecies differences in bioconcentrat ion o f P C B s , observed by V e i t h et al . (1979), between the fathead m i n n o w (Pimephales promelas) and ra inbow trout (Salmo gairdneri). N. obscura a t t a ined greater b i o c o n c e n t r a t i o n s o f more c o n t a m i n a n t s than P. marmorata i n f i e l d exposures . In a s i m i l a r fashion, to that descr ibed above, env i ronmen ta l l y i nduced stress reac t ion , manifes ted by increased respi ra tory and metabolic rate and cou ld have induced greater uptake of chlor inated phenolics i n the Fraser R i v e r , by the species N. obscura. A basis for this explanation comes from studies o f the known habitat and tolerance ranges for N. obscura and P. marmorata Herrmann (1970) studied the habitat preferences o f Colorado leech species and found P. marmorata to be dis t r ibuted re la t ive ly evenly between lent ic and lo t i c habitats and was tolerant o f h igh suspended sediment l eve l s , wh i l e N. obscura showed a dis t inct preference for clear water lent ic habitats. H i g h in situ water f l o w and suspended sediment levels cou ld have led to the theorized stress induced increase i n b ioconcent ra t ion by N. obscura. 1 3 8 Table 4.18: Bioconcentrations (ng/kg) of chlorinated phenolics in two leech species, Nephelopsis. obscura and percymoorensis marmorata. exposed in the Fraser River at Stoner, B.C.*, for a seven day period, July 8 - 15, 1991. Results reported on a wet weight basis. Compound Tissue Concentration (pg/kg) N. obscura f P. marmorataft 4,5-DCG ND ND 3,4,5-TCG 2 8 + 7 159 ± 63 4,5,6-TCG 12 ± 2 3.0 ± 1.0 TeCG 21 ± 5 5.0 ± 1.0 3,4,5-TCVer ND ND 5,6-DCV ND ND 2,4,6-TCP 19 ± 8 19 ± 6 2,3,4,6-TeCP 3.4 ± 0.6 ND ND = None detected; refer to Table 3.1 for dectection limits. * Leeches exposed on the West side of the Fraser River, t Sample size n = 9; mean weight + SD = 1.07 + 0.14 g. f t Sample size n = 10; mean weight ± SD = 3.07 ± 0.52 g. Table 4.19: Bioconcentration of chlorinated phenolics by two species of leeches, Nephelopsis Qb£Cjjia.and Percymoorensis marmorata. after seven days exposure to a water concentration of 1.0 ug/ L at a water temperature of 12.5 °C and water pH 7.5. Compound Tissue Concentration (pg/kg) N. obscuraf P- marmorataft 4,5-DCG 182 ± 44 311 ± 105 3,4,5-TCG 82 ± 18 303 ± 96 4,5,6-TCG 325 ± 27 354 ± 119 TeCG 219 ± 41 256 + 65 3,4,5-TCVer 45 ± 18 45 ± 8 t Sample size n = 5; mean weight ± SD = 1.23 ± 0.43 g. t t Sample size n = 4; mean weight + SD = 2.66 ± 0.49 g. Both N. obscura and P. marmorata are suitable biomonitors of chlorinated phenolics and have been useful in situ riverine biomonitors in past investigations (Hall and Jacob 1988; Metcalf and Hayton 1989). N. obscura is available throughout British Columbia as well as all the rest of Canada (Sawyer 1974), is easy to maintain in 139 laboratory and provides clean, easy to analyse extracts. While P. marmorata may be more suitable for Fraser River monitoring due to its riverine habitat preferences, N. obscura has shown a greater tolerance to lower water temperature (0 - 20 °C) (Herrmann 1970; Sawyer 1974) and may be the better choice for cold water monitoring, characteristic of the Fraser River during spring, fall and winter seasons. 4 . 2 . 5 . E s t i m a t i o n o f In Situ W a t e r C o n t a m i n a n t C o n c e n t r a t i o n s U s i n g L a b o r a t o r y B i o c o n c e n t r a t i o n R e l a t i o n s h i p s Leech bioconcentration data provides a measure of the bioavailability of water contaminants. However, it would be useful to relate leech tissue concentrations back to in situ exposure levels (i.e. derive water contaminant concentration from bioconcentration data). Laboratory relationships between bioconcentration, water concentration and environmental (temperature, pH, water current) and biotic (leech weight) factors were used to calculate the original exposure concentration during July, October, 1991 and February 1992 field trials (Table 4.20). Sample calculations are given in appendix 5. In each case the relationship between leech bioconcentration and water concentration (Section 4.11; Table 4.1) was rearranged to give an initial prediction of in situ water concentration from measured in situ bioconcentration (Section 4.2.1; Table 4.14). Since this laboratory relationship was derived for a mean leech weight of 1.17 g and water temperature 12.5 °C and pH 7.5, corrections were made to normalize the prediction to the corresponding field conditions. Thus the resulting predicted water concentration took into account exposure concentration, water temperature, pH and leech weight. The resulting value, termed the static water predicted water concentration is given in column 4 of Table 4.20. 140 E J 8? o — CO = co ^ X LU a Q - E £ £ j= 3 2 ~ 2 ffl o CD -a 8 I I CD CO o 0 c CO CD Q) •4—1 i . CO 5 -o CD g ^ O -B 0 , ' i to "> C * • CD " c o ffl OT .2 co o co c > 2 P ffl ca p 2 >_ *• o co eg 1 0 •o X 12 ^  0 a$ CO c o S E £ c m CD O c o o co I -6 o CO Q - i _ JZ JD co -9 c £ .2 o C L Q . co 2 ~ w _ 3 1 o S co - 1 CC " D C 1 0 -5 >. >, c CD > CD CO T3 CD CO ~c CD O c o o o 3 CO s ^ E x: o c a 0 co b 2 O CO c o CD 0 cu > 00 CO -r-•g O - -ffl TS £ 2 § • O O) C £ H fl) o 5 « 5 E o CM 3 ! c o co c 0 o c o o 0 c 0 k_ 3 o CO 3 o 3 p C c o o 0 5 -a 0 k_ 3 CO co 0 0 as o » ffl CO c o co k_ •*—' c 0 o c o O T3 C 3 o C L E o O 00 00 CO CM i n T— CM O o o O O O o q 6 b c i b d 6 LO CO CM O CM i n o CM O •a o 0 0-O) C "a. E co co CD O H I i n co" OT OT i n oo 3 ~3 r -CO T -T ~ O CO m 2 "& =1 * 5 5! O O o CD I- I -- OI c n ^ =£. > "a 3> i i =-00 CO CM CD I  II CO o o o CO o CM T— o 6 CM " O (5 (5 U o £ ^ " T io ii "5. -a- , f CO £Z ^ ! I § CD O r— I i n •>* co" CD O 1-I m co" n CO tr ci n 2 .2> .2> oi ffl | | OT OT •0-CM O O CM OT OT CO CM OT XI 0 co cn 1^  CO I X 9 o O O csi CO CD ffl ffl CO CO <B ffl CL CL E E 0) ffl fl) ffl ffl I I I 141 Note that i n a l l cases the predicted water concentrat ion for an unstirred seven day b ioassay overes t imated the measured va lue . The effect o f water current on b ioconcen t r a t i on was never f o r m a l l y e x a m i n e d , but was acc iden t a l l y d i s c o v e r e d when bioassays were magne t ica l ly st irred to retain sediments i n suspension. T o prov ide a more accurate estimate o f water concentra t ion under f i e ld condi t ions a c o r r e c t i o n factor , equ iva len t to the re la t ive increase i n l eech b i o c o n c e n t r a t i o n observed i n the bioassays wi th magnetic s t i r r ing (i.e. water current), was d iv ided by the static water p red ic t ions i n c o l u m n 4. The water f l ow co r r ec t i on factor underest imated the in situ measured levels , except for 3 , 4 , 5 - T C G i n February 1992. The accuracy o f the predic t ions are quest ionable since the water f l ow cor rec t ion factor was derived from only a single bioassay. Nei ther method o f water contaminant concentra t ion es t imat ion was accurate, but together the two estimates provided a useful range o f exposure levels . Further cha rac t e r i za t ion o f the re l a t ionsh ip be tween water current and b i o c o n c e n t r a t i o n may provide more accurate estimates o f in situ water concent ra t ions f rom leeches exposed i n lo t ic environments. The laboratory bioassay relat ionships der ived i n the present study may be more suitable for pred ic t ing water concentrat ions i n len t ic water habitats . 142 5 . SUMMARY, RECOMMENDATIONS AND APPLICATIONS Leeches were assessed as biomoni tors o f chlorinated phenol ic type compounds. The leech species Nephelopsis obscura was used as the pr imary test organism, whi le Percymoorensis marmorata was used as secondary test species , for compara t ive purposes. A n integrated set o f labora tory and f i e l d inves t iga t ions assessed the importance o f both environmental and b io t ic factors i n the bioconcentra t ion o f four bleach kraft m i l l der ived chlorinated guaiacol compounds ( 4 , 5 - D C G , 3 , 4 , 5 - T C G , 4,5,6-T C G , T e C G ) and one chlorinated guaiacol biotransformation product, 3 , 4 , 5 - T C V e r . A n attempt was made to investigate a th i rd group o f compounds , the c h l o r o v a n i l l i n s , represented by 5 , 6 - D C V , but rel iable analysis and quantif icat ion o f this compound i n water and leech tissue samples was not achieved due to poor analyt ical sensi t ivi ty by G C - e lec t ron capture detect ion and apparent l o w leech b ioconcent ra t ions at the contaminant levels assayed. In f ie ld studies, two chlorophenol compounds, 2 , 4 , 6 - T C P and 2 ,3 ,4 ,6-TeCP were also monitored. In the l a b o r a t o r y , the effects o f the f o l l o w i n g fac tors on l e e c h b i o c o n c e n t r a t i o n were inves t iga t ed : E n v i r o n m e n t a l Fac to r s : - Contaminant concentration (0.1 - 10 p g / L ) - Water p H (5.1 - 9.0) - Water temperature (4.4 - 20.0 °C) - Suspended sediment load and composi t ion (0 - 0.15 g /L) B i o t i c Factors: - Leech weight (0.15 - 1.80 g) - Leech species (N. obscura, P. marmorata) 143 Three f ie ld moni tor ing trials, cover ing summer (July 1991), f a l l (October 1991) and winter (February 1992) seasonal condi t ions , were conducted to evaluate leeches under va ry ing in situ c o n d i t i o n s . Unde r f i e ld condi t ions leeches proved to be effective moni tors o f t r i - and tetrachlorinated phenol ic compounds under diverse seasonal condi t ions (Table 4.14) . Leech bioconcentration factors ( B C F ) ranged from 465 - 6000. Leeches were also able to bioconcentrate measurable levels of the compounds ( 3 , 4 , 5 - T C V e r , 2 , 4 , 6 - T C P , 2,3,4,6-T e C P ) , after water analysis had failed to detect any trace of the compound. Leeches were not effective b iomoni tors o f d ichlor ina ted phenol ics over a seven day per iod . Leeches were accurate indicators o f the relat ive proportions o f contaminants i n both B K M E effluents and r e c e i v i n g water (F igure 4 .12) . L e e c h b i o m o n i t o r i n g data ref lected the decrease i n m i l l discharge o f T e C G , resu l t ing f rom an increase i n chlor ine d ioxide substitution i n bleaching. L e e c h moni tor ing data also indicated that h igh contaminant body burdens for aquatic organisms may occur dur ing the winter l ow flow period, when water contaminants are diluted the least by river f low. The p r ima ry determinant o f l eech b ioconcen t ra t ion i n the labora tory was contaminant concent ra t ion . A l l contaminants showed a strong l inear co r re l a t ion be tween the l o g o f the c o n t a m i n a n t c o n c e n t r a t i o n and the l o g o f the bioconcentrat ion (Table 4.1). However , a sl ight trend towards increased B C F at l o w contaminant l eve l s suggested that there may be a departure f rom l inea r i ty ( i .e . increased rate o f uptake) at water concentrations < 0.1 pg /L . In situ studies, at l o w con taminan t concen t r a t i ons (< 0.06 p g / L ) also suggested elevated uptake rates at l o w e r water concentra t ions (F igure 4 .13) . The resu l t ing h igher than p red ic ted bioconcentra t ions w o u l d lead to overestimates i n water concentrat ion der ived f rom the laboratory regression model descr ib ing b ioconcent ra t ion as a funct ion o f water contaminant concentra t ion. R e m o v a l o f leeches from contaminated water revealed that leeches e l imina t ed body burdens o f ch lo r ina ted gua iaco l s r e l a t i v e l y s l o w l y (estimated T 1 / 2 ^ 2 8 days) (Figure 4.2), suggesting that loss by e l iminat ion w o u l d not be a c r i t i ca l factor i n the estimation o f ambient water concentrations from seven day leech body burdens. Tr ichlorovera t ro le showed a d is t inc t ly different depuration 144 prof i le , wi th constant, rapid e l imina t ion (estimated T 1 / 2 < 7 days). B ioconcen t r a t i on o f ion izab le ch lo rogua iaco l s was inve r se ly related to water p H , however , b ioconcen t ra t ion was not d i r ec t ly related to the concent ra t ion o f und issoc ia ted c o m p o u n d and regress ion re la t ionsh ips were genera l ly weak (Table 4 .5) . The phenolate i o n form also cont r ibuted s i gn i f i c an t l y to the b ioava i l ab l e contaminant p o o l . There was evidence o f p h y s i o l o g i c a l disturbance to the leech species N. obscura, induced by low p H , wh ich was manifested by a posi t ive trend i n b ioconcentra t ion o f the non-dissociable compound , 3 , 4 , 5 - T C V e r . Water temperature affected bioconcentrat ion o f a l l test compounds except 4 ,5-D C G (Table 4.6), in a pattern best described by a bi-phasic model (Figure 4.8). There was s ign i f i can t increase i n t r i - and t e t rach lo rogua iaco l b ioconcen t ra t ion between 4.4 and 11.8 ° C , at w h i c h point bioconcentrat ions d id not change s igni f icant ly . A n opposite effect was observed for 3 , 4 , 5 - T C V e r . The results can be expla ined as the c o m p e t i t i v e i n t e r ac t i on o f uptake and e l i m i n a t i o n processes , w i t h the former dominat ing over the lowest temperature range and the latter becoming o f increas ing importance at higher water temperatures. Expe r imen t s a imed at assessing the ro le o f suspended s o l i d mater ia l i n affecting seasonal b ioava i l ab i l i t y o f ch lorophenol ic compounds to leeches showed no evidence o f any effect (Table 4.7) , when us ing a suspended sediment mixture o f u n i f o r m c o m p o s i t i o n (< 63 pm diameter; 9 5 % inorgan ic : 5% organic mater ia l ) . L a b o r a t o r y exper iments revea led that the presence o f o rgan ic ma te r i a l i n the suspended sediment f rac t ion reduced b i o a v a i l a b i l i t y , s ince b ioconcen t r a t i ons o f chlor inated guaiacols were greater i n the absence o f organic particulates. However , the results were compl ica ted by unquantif ied, extraneous input o f suspended organic material , i n the form of leech mucous secretions. Suspended sediment studies also indicated that bioassay protocol has a profound effect on the measured results as was i n d i c t e d by the s ix to e ight f o l d increase i n c h l o r o g u a i a c o l b i o c o n c e n t r a t i o n observed when laboratory bioassays were cont inuous ly st irred. Therefore, va r ia t ion i n ambient water current may also be an important envi ronmenta l determinant o f chlor inated phenol ic bioconcentrat ion. The f ind ing that Fraser R i v e r exposed 145 leeches attained greater bioconcentrations of certain compounds, even though water contaminant concentrations were significantly lower than in laboratory bioassays supports this theory (Figure 4.15). Further corroborating evidence was obtained from the apparent cross channel differences in Fraser River leech bioconcentrations (Table 4.15). Strong river water mixing reduced the possibility of heterogeneous water contaminant concentrations at the Stoner site. However, higher leech bioconcentrations were related to positioning in areas of swifter water flow. The effect of leech weight on bioconcentration was compound specific. A strong inverse relationship was observed for the chlorinated guaiacols, while no clear trend was observed for 3,4,5-TCVer (Table 4.4). Similar relationships were found for field exposed leeches (Table 4.16). Available evidence suggested that the effect may have been due to a combination of age and size dependent factors of both a physiological and physical nature. Increased respiratory rate and decreased depurative enzyme function and . a greater surface area to volume ratio in smaller, younger leeches were possible contributing factors. No clear trend in interspecies differences in bioconcentration was observed between the leech species Nephelopsis obscura and Percymoorensis marmorata . Under laboratory conditions P. marmorata attained greater tissue concentrations of chlorinated guaiacols than N. obscura (Table 4.19). Both species bioconcentrated similar levels of 3,4,5-TCVer. Differences in tissue concentration of chloroguaiacols could not be explained by leech weight differences, since the smaller species, N. obscura, attained the lower bioconcentrations. Under in situ conditions, N. obscura attained greater bioconcentrations of 4,5,6-TCG, TeCG and 2,3,4,6-TeCP, while P. marmorata bioconcentrated greater levels of 3,4,5-TCG (Table 4.18). Both leech species proved to be suitable for biomonitoring chlorinated phenolic compounds. However, N. obscura may be better suited for monitoring in cool northern waters, such as the Fraser River, since it shows a greater temperature tolerance range that P. marmorata. In addition, N. obscura provides cleaner tissue extracts for chlorinated phenolic analysis, allowing for lower limits of detection. 146 The re la t ive importance o f the envi ronmenta l and b io t ic factors inves t igated i n re la t ion to the magnitude o f changes occur r ing i n the natural environment can be seen in Table 5.1. Two compounds, 3 , 4 , 5 - T C G and 3 ,4 ,5 -TCVer are used to illustrate the effects on each o f these classes o f compounds. Leeches were very sensitive to changes i n contaminant concentrat ions, w h i c h is useful for detecting smal l changes i n ambient l eve l s . Water p H s ign i f i can t ly affected b ioconcent ra t ion over the p H range 5 - 9 . In reality, the effect in any single aquatic system of relatively stable p H , such as the Fraser R i v e r (pH 7.5 - 8.5), would probably be min imal . Temperature was not f o u n d to s i g n i f i c a n t l y affect b i o c o n c e n t r a t i o n above 12 ° C , h o w e v e r bioconcentration appeared to be affected between 0 - 1 2 °C , wh ich corresponds to the temperature range exhib i ted by many northern aquatic systems from fa l l through to the sp r ing . A l t h o u g h not quant i ta t ive ly invest igated i n the present study, the effect o f water cur ren t on b i o c o n c e n t r a t i o n m a y be an i m p o r t a n t fac tor i n l e e c h b iomon i to r i ng o f r iver ine systems, such as the Fraser R i v e r . The effect may be cont ro l led , to some extent, by conduct ing a l l moni to r ing i n areas o f s imi l a r water f low rate. L e e c h weight was a powerful determinant o f bioconcentrat ion. Use o f leeches at the smaller end o f the weight scale wou ld promote m a x i m u m sensi t iv i ty . At tempts to use de r ived labora tory re la t ionsh ips , f rom semi-s ta t ic b ioassays , to estimate water contaminant concentrat ions f rom leech body burdens o f ch lor ina ted pheno l i c s after exposure i n the Fraser R i v e r (Table 4 .20) overes t imated water concentrat ions. If a cor rec t ion factor, cor responding to the observed increase i n b ioconcen t ra t ion w i t h con t inuous ly s t i r red b ioassay water, was app l i ed , p red ic ted water concen t r a t i ons unde res t ima ted measured l e v e l s . T h e r e f o r e , l a b o r a t o r y s i m u l a t i o n s , w h i l e s u c c e s s f u l i n i d e n t i f y i n g i m p o r t a n t d e t e r m i n a n t s o f b ioconcen t ra t ion , fa i l ed to dupl ica te the c o m p l e x and dynamic condi t ions i n the Fraser R i v e r . 147 O c p co k_ "cz 0) o c: o o o JQ as o CO o -Q o pe 21 >. co x: TJ o c 0 0 CD > C CO o co co 1 CD o o AO CO 0 o 3 ' o SO Q_ Ja X 73 0 c c CO o ed c co CD CO E CQ c o > Ve C o CD i -O i IT) w o co" ffe nd CD co O CD O h-ma lO E ^ ~ n CO CO LO _0 CO SZ 2> CD SI CJ CD CD CO CO 00 CO 2 CD t o 1 CD c o CO c 0 o c o o g CQ - - I O CO •tf LO LO X Q_ CD o c 0 CJ c o O 4—' c CO c E o O "D C ZI o CL E o O • t f CO LO 00 CD LO CD CM • t f o CD O H Lf) •tf co" LU UJ UJ z UJ LO CM CM 05 CVJ o CO 00 • t f o •tf o •tf o CD > O H-in •tf" co" =^  «=! §> D) O 1 i i n c o •2 1 I J 1 I I » o a S -5 n •= <S IS -2 •3 M 148 S e m i - s t a t i c l a b o r a t o r y b i o a s s a y s d i d not p r o v i d e a c c u r a t e l e e c h bioconcentra t ion data at water contaminant concentrations be low about 0.1 pg /L . It is recommended that future l ow level (0.1 - 0.001 p g / L ) studies be carr ied out under f l ow through cond i t ions . Fur ther , i t i s necessary to accurate ly quantitate the r e l a t ionsh ip between water current f l o w rate and b ioconcen t r a t ion , i n order to generate a more real is t ic laboratory model and evaluate in situ data f rom d i f f e r ing f low regimes. Di rec t appl icat ion o f leech bioconcentrat ion data to b ioava i l ab i l i t y to other aquatic organisms cou ld be achieved through comparat ive uptake rate studies between leeches and representative aquatic species. Fo r routine appl ica t ion o f leech b i o m o n i t o r i n g programs i t w o u l d be desirable to es tabl ish a laboratory breed ing program, u t i l i z i n g ind iv idua ls o f standard age and size. Ev idence from the present study indicates that leeches are useful for gauging seasonal b i o a v a i l a b i l i t y and body burden changes i n B K M der ived contaminants . Leeches also gave an indicat ion o f effectiveness o f m i l l process changes aimed at A O X r e d u c t i o n , such as reduced c h l o r i n e use. L e e c h e s p r o v i d e d t ime in tegra ted in format ion on the relat ive propor t ion o f a variety o f contaminants i n the Fraser R i v e r . The potential exists for a more comprehensive in situ b i o m o n i t o r i n g p r o g r a m , i n v o l v i n g a series o f test s ta t ions , genera t ing i n f o r m a t i o n c o n c e r n i n g spa t ia l changes i n d i s t r ibu t ion and propor t ion o f contaminants throughout an entire r ive r system. Further, this strategy could be applied i n a national moni tor ing program. In the case o f Canadian B K M effluents, a federal government sponsored r isk assessment o f the effects o f bleached pulp m i l l effluent on the r e c e i v i n g waters concluded that "effluents from pulp m i l l s us ing bleach is entering the environment i n a quanti ty or concentra t ion under condi t ions h a v i n g immedia te and long- t e rm harmful effects on the environment. Effluents. . . . are therefore considered to be toxic as defined under Paragraph 11(a) o f the Canad ian E n v i r o n m e n t a l P ro tec t ion A c t " ( C E P A ) (Env i ronmen t Canada 1991a). Amendment s to P u l p and Paper Ef f luen t Regu la t ions , under the F i sher ies A c t i n M a y 1992, were enacted to reduce the environmental r i sk associated wi th bleached pulp m i l l effluent discharge 149 (Envi ronment Canada 1992). However , extensive knowledge gaps were ident i f ied i n the area o f acute and chron ic effects o f effluent (Env i ronmen t Canada 1991a), spawning the need for a mechan i sm to assess the effectiveness o f the amended regulat ions i n protect ion o f aquatic ecosystems. The Federal Env i ronmen ta l Effects M o n i t o r i n g Program ( E E M ) for pulp and paper m i l l effluents provides a nat ion-wide p rog ram to mon i to r the effects o f pu lp and paper eff luents on the aquat ic envi ronment (Env i ronment Canada 1992). N a t i o n a l coord ina t ion and guidance are p rov ided by a Na t iona l E E M Off ice at Env i ronment Canada, regional adminis t ra t ion and guidance is the respons ib i l i ty o f both p r o v i n c i a l and regional federal off ices . Each pulp and paper m i l l is responsible for conducting the E E M . Under the E E M program, a set o f both s i te-specif ic and core variables are monitored i n order to determine: the magnitude, spatial extent and temporal changes i n effects. The E E M program, is to be carried out by ind iv idua l pulp & paper companies under federal and p rov inc i a l supervis ion. M o n i t o r i n g focuses on the effects of both w h o l e e f f luen t and s p e c i f i c c o m p o n e n t s , i n c l u d i n g the c h l o r i n a t e d p h e n o l i c c o m p o u n d s ( E n v i r o n m e n t C a n a d a 1991b) . S p e c i f i c t e s t ing i n c l u d e s b o t h environmental chemica l moni tor ing and tox ic i ty bioassay tests. It is proposed here that leeches could be an integral part the E E M program. Since leeches proved to be sensi t ive in situ moni tors o f ch lor ina ted phenol ic contamina t ion , they w o u l d make good E E M ind ica to r s for measur ing spat ia l and tempora l contaminant changes . L e e c h e s may be e s p e c i a l l y usefu l after i m p l e m e n t a t i o n o f c h l o r i n e r e d u c i n g t e chno log i e s , s ince they can b ioconcen t r a t e measurab le l e v e l s o f c h l o r i n a t e d phenol ics , when water concentrations are too l o w for convent ional analysis . W i t h further research in to the re la t ionsh ip between l eech b ioconcen t ra t ion and effects on other aquat ic o rgan i sms (eg. benth ic inver tebra te c o m m u n i t y impac t s , f i sh de tox i f ica t ion enzyme ac t iv i ty , f i sh l a rva t ox i c i t y ) , leech b ioconcent ra t ion cou ld be correlated wi th other observed impacts . Since in situ leech bioassays are r e l a t ive ly s imple to conduct, leeches could be used as routine indicators of trouble areas, wh ich may be fo l lowed up by specific toxic i ty testing. 1 5 0 The E E M program represents a single application of leech biomonitoring. 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Influence of p H and ionic strength on the aqueous-nonaqueous dis tr ibut ion o f chlorinated phenols. E n v i r o n . S c i . Technol . 19, 193-198. X i e , T . (1983). Determination o f trace amounts o f chlorophenols and chloroguaiacols i n sediment. Chemosphere 12, 1183-1191. X i n g , B . , W . B . M c G i l l and M . J . Dudas (1993). Thermodynamic parameters for pentachlorophenol sorption on montmor i l lon i te i n aqueous suspensions. Chemosphere 26, 1311-1324. 167 P A R T 7 : A P P E N D I C E S 168 Appendix 1 Leech Identification Distinguishing features useful in the identification of the leech species Nephelopsis obscura and Percymoorensis marmorata. Information from North American Freshwater Leeches Exclusive of the Piscicolidae: with a Key to all Species (Sawyer 1972). 1) Nephelopsis obscuraNemW 1872 (Family Erpobdellidae) Description: Length: 5 - 6 cm. Weight: 0.1 - 2.0 g. Colouration: light brown to dark brown with irregularly scattered black splotches dorsally and less commonly ventrally. Distinguishing anatomical features: a) External features: sub-division of most mid-body annuli. Male and female gonopores separated by two annuli, with pores usually located in furrows (Figure Al) . b) Internal features: spirally coiled atrial cornua, pre-atrial loop of the vas deferans, extending to ganglion XI. Similar species: Erpobdella punctata, Dina dubia Figure A l : Relative positions of male and female gonopores for: 1) Nephelopsis obscura 2) Erpobdella punctata 3) Dina dubia 169 1) Percymoorensis (Haemopsis) marmorata. ( F a m i l y H i rud in idae ) . Note that this species is ve ry d i f f i c u l t to d i s t i n g u i s h f rom Haemopsis grandis by e x t e r n a l examinat ion on ly . Ident i f icat ion by dissect ion and internal examinat ion is the on ly reliable method o f identif ication and is best conducted by a recognized expert. Descript ion: Length: 2 - 8 cm. Weight: 1 - 5 g. Coloura t ion : H i g h l y variable. Rang ing from l ight o l ive green wi th sparse dark splotches to heavi ly mottled wi th almost b lack colourat ion. M o t t l i n g occurs on both dorsal and ventral surfaces. D i s t i n g u i s h i n g ana tomica l features: a) External features: M a l e and female gonopores separated by five to five and ha l f annuli , with pores usually located on the anterior third o f their respective annul i , however actual pos i t ion ing may vary. b) Internal features: extraordinari ly long and slender penis. Ante r io r part o f prostate gland located between gangl ion XIIIi/4 and X V i / 2 (usually X I V ) . Center o f the ovaries located between ganglion X I I and XIV1/2 (usual ly XIII3/4 ). The posterior extension o f the vaginal system located between ganglion X V and X V I I (usually X V I ) . The intestine usual ly begins at gangl ion X I X . D i s t i ngu i sh ing behaviour: amphibious habits, often found under part ial ly submerged rocks and logs at the waters edge. Ventures on to shore in search o f prey at night. B o d y extremely f l acc id ; drapes l i m p l y over fingers when p i cked u p . S imi l a r species: Haemopsis grandis 170 A P P E N D I X . 2 CALCULATION OF REPORTED CHLORINATED PHENOLICS CONCENTRATIONS C a l c u l a t i o n o f ch lo r ina ted pheno l i c concentra t ions were based on the f o l l o w i n g equations adapted f rom M o r a l e s et a l . 1992. Target c o m p o u n d refers to any chlor ina ted phenol ic analyzed dur ing this project. a) Response factors (RF) o f the target compound standard and the internal standard : ( 1 ) R F S T D = Peak Area of S T D / [STD] (2) R F I S = Peak Area of IS / [IS] W h e r e : ^ S T D = r e s P o n s e factor o f the target compound or surrogate [STD] = concentration o f target compound or surrogate i n the G C reference s tandard R F I S = response factor o f internal standard [IS] = concentration o f the internal standard i n the G C reference s t a n d a r d b) Relat ive response factor ( R R F ) o f the target compound to the internal standard (3) R R F = R F S T D / R F I S W h e r e : R R F = relative response factor of test compound to internal standard c) Concentrat ion o f target compound [A] i n the or ig ina l sample ( 4 ) [A] = ([IS] / RRF)*(Peak Area o f A / Peak Area of IS) * (F ina l Extract V o l u m e / Or ig ina l Sample Weigh t or Vo lume) 171 APPENDIX 3 Raw Data Laboratory Bioassays 172 Table A1: Bioconcentration of chlorinated phenolics by leeches (N. obscura) as a function of contaminant concentration at T= 12.5 °C and pH = 7.5. w a t e r Tissue Concentation (ng/kg) Cone. Leech wt. (ug/L) (9) 45-DCG 345-TCG 4,5,6-TCG TeCG 345-TCVR 1 .460 ND 9 2 3 1 6 2 0.1 1 .430 ND 1 3 3 0 2 2 1 1 .240 ND 14 2 7 2 3 9 0 . 7 4 9 ND 1 5 4 2 2 8 7 0 . 7 2 1 ND 2 4 5 5 3 7 ND Mean 1 .12 ND 1 5 3 5 2 5 5 S i d Dev. 0 . 3 6 5 1 2 7 4 1 .290 5 7 3 0 1 2 8 7 0 1 8 0 .5 1 .050 6 3 31 1 3 3 7 9 1 5 0 . 5 9 0 1 2 3 7 4 2 1 7 1 7 4 1 3 1 .580 4 8 2 4 1 1 3 5 9 1 7 1 .360 5 9 3 0 1 2 8 7 2 1 5 Mean 1.17 7 0 3 8 1 4 4 91 1 6 Std Dev. 0 . 3 8 3 0 1 8 3 7 4 2 2 0 .861 1 7 9 9 5 3 1 0 2 4 7 2 1 1.0 0 . 9 6 2 2 1 0 9 9 3 3 7 2 5 4 4 9 1 .770 1 3 0 5 7 2 9 3 1 6 5 4 7 1 .650 1 4 2 6 3 3 1 5 1 7 2 6 3 0 . 9 1 2 2 4 9 9 8 3 7 1 2 5 7 ND Mean 1 .23 1 8 2 8 2 3 2 5 2 1 9 4 5 Std Dev. 0 . 4 3 4 4 1 8 2 7 4 1 1 8 1 .230 4 4 1 1 9 4 8 4 2 6 4 6 2 2 3 5 .0 1 .220 4 9 3 2 4 9 9 6 4 8 0 7 1 1 8 0 . 9 2 0 6 0 3 2 5 7 8 5 6 7 2 4 2 4 4 1 .750 3 0 4 1 3 5 6 1 5 4 1 9 2 5 2 1 .300 4 3 0 1 9 9 7 8 5 5 9 4 1 9 4 Mean 1.28 4 5 4 2 0 7 8 1 2 6 3 8 2 0 6 S i d Dev. 0 . 3 0 9 7 4 4 1 14 131 4 8 1 .220 8 7 2 4 4 9 1 7 1 0 1 4 2 0 3 4 0 10 .0 1 .020 1 1 5 0 5 7 0 2 2 4 0 1 8 8 0 4 0 3 0 . 9 5 6 1 6 8 0 8 9 2 3 3 4 0 2 9 3 0 3 1 9 0 . 8 0 3 1 2 8 0 5 9 6 2 2 1 0 1 8 1 0 3 8 3 1 .150 7 5 6 3 3 5 1 6 2 0 1 2 4 0 3 0 2 Mean 1.03 1 1 5 0 5 6 8 2 2 4 0 1 8 6 0 3 4 9 Std Dev. 0 . 1 6 3 3 0 1 8 7 6 1 2 5 9 0 3 8 173 Table A 2 : The depuration of chlorinated phenolics from leeches (N. obscura) over a 28 day period after seven days exposure to a concentration of 1.0 ug/L at a water temperature of 12° C and water pH 7.5. Samples taken on day 0, 3 (72 h ) , 7 (168 h.), and 28 (672 h.) after seven day exposure. Results reported on a wet weight basis. Depuration Raw Tissue Concentration (ng/kg) Time Leech Wt. (h) (g) 4,5-DCG 3,4,5-TCG 4,5,6-TCG TeCG 3,4,5-TCVR 0.861 179 95 3 1 0 2 4 7 21 0 0 .962 2 1 0 9 9 3 3 7 2 5 4 4 9 1 .77 1 30 57 2 9 3 1 65 4 7 1.65 142 6 3 3 1 5 172 6 3 0 .912 2 4 9 9 8 371 2 5 7 ND Mean 1 .231 182 8 2 3 2 5 2 1 9 4 5 Std. Dev. 0.441 44 1 8 27 41 1 8 1.350 9 3 4 0 244 128 1 1 7 2 1 .540 7 5 31 2 1 6 9 8 1 4 1 .450 91 34 2 2 2 107 1 5 1.480 91 38 2 3 8 1 1 6 1 2 0 .970 147 5 9 2 9 7 179 ND Mean 1 .358 99 4 0 2 4 3 126 1 3 Std. Dev. 0 .228 28 1 1 32 32 2 1.240 1 1 5 5 0 2 2 8 146 4 1 68 1 .350 135 61 3 0 0 184 1 9 0 .890 1 99 6 6 2 9 9 2 1 2 1 3 1.290 108 4 5 2 4 4 149 8 1 .080 1 3 6 58 301 198 7 Mean 1 .170 139 5 6 274 178 1 0 Std. Dev. 0 .186 3 6 8 3 6 2 9 6 0 .468 2 2 7 87 3 5 8 2 8 7 ND 6 7 2 0 .616 147 6 5 2 6 2 2 0 2 ND 0 .864 1 25 4 9 2 3 6 168 ND 0 .538 174 78 2 8 3 2 1 6 ND 0.66 161 64 2 5 2 1 7 3 ND Mean 0 .629 167 69 2 7 8 2 0 9 ND Std. Dev. 0 .150 3 8 1 5 4 8 4 8 174 O "a + ^ +- CJ) i i CO A 3 ra £ TI « o co 3 ~ 3^) < co o »- z a. cc > o Tt~ CO i? "a 1= c o "5 C3 c o o H o c CO o in" o cu Z3 V) "lis G TJ O <1) H J= in £ •tf m" o S-DCC CO CM lO O CM ^ CO O < - CO CJ CD ^ N CO (D CM Tf Tf ""fr Tf CM CNJ CM CM CM O O ) CO CO CO 1 00 CD O o CO COi LO T LO O O T f— CM CM CM CM I •et LO CD CO 1 -LO i - CO if) CM CD CM CO CM O) CO CM CM CM CM CM CO CO CM . LO LO CM CD Tf Tf LO CO T - O CO CM CO LO 00 CM O) N T - CM i -(!) W CD IO N LO CM O C> CO iD lO (D N CM CO CM CV] CO r*- CM co LO co LO LO CO CO CD 00 CO CD io CO CM CO CO CO O T - o CM CM CM CM CM LO CD CO CO CO CD Tf Tf Tf Tf LO 2 2 "a» • H O <t E 8 o ^ cd tj C (D * " 0) "D * 8 CO 2? o Q o •= a> 5 § -s FT--5 « t "> TJ «-.fi c x: _ CO CJ) t i l C U) .2 £ § "5 .2> -g 3 9> c o s • (0 r O o co ^1 s CM r - co LO w o n CM CM -r-CO Tf CM CO LO CM CO CD -tf , i - CM CO CM I CO CM CM CO CD CO CM 00 T -cb ai CM T} N ID t N T - CM CD CD CO O ) CM r»- LO Tf T— r— T— T— CM If) S S CO o ^ CO CO CM CD Tf LO Tf i -CM CM CM CM CO 00 ^ CO o t W N CO O) CO LO -sr co (fj Cll S CO O) CO -r- CD ' TJ o o o o o If) Tf Ifl CO I s -io " t ' " f o Tf CO CM CO cn LO CM i -CO CM CD CO CO CO CM t- -r-CD LO N O) CD T - CD h-CD CM CD CD CM CO CM CM CO U) CD U) CM Tr Tf co lO IS-. LO CD o o o o o Q Q Q Q Q r*- CD CM IS-LO Tf Tf CO CM T - N CO CM CM CO CM CO CO O N t O r - CM r-. co CO t f IO N N T - (O N co co r*- LO CM CM T- CM CM CO (D Tf CO CD i — CD CO t f CD CO LO d o d o s co s s co O) V o - ° b o-1 CD Q_ co B Ife 8 t S rs » « I- 11 o "S CM . 2 II "&> f ft •a "& ) CD CD . o o •H C3 ^ o o> O : Q O I • Ul 111 I LU CC CC •: CC CL 0- Q 18 •<t CO -t H CO 175 Table A4 : Bioconcentration of chlorinated phenolics by leeches (N. obscura) as a function of leech weight at water temperature 12 °C, pH 7.5 and a contaminant concentration of 1 ug/L. Tissue Concentation (ng/kg) Leech wt. Group (9) 45-DCG 345-TCG 4,5,6-TCG TeCG 345-TCVR 0.162 705 325 992 810 50 1 0.123 805 395 1 120 920 29 0.168 573 246 669 592 27 0.139 673 273 779 741 21 Mean 0.148 689 310 890 766 32 Std. Dev. 0.021 96 66 204 137 1 3 0.207 488 218 644 542 76 2 0.282 ; 391 186 504 424 30 0.380 301 130 435 377 52 0.322 471 197 628 559 113 0.322 51 1 231 691 616 90 Mean 0.303 432 192 580 504 72 Std. Dev. 0.064 86 39 107 99 32 0.861 179 95 310 247 21 3 0.962 210 99 337 254 49 0.912 249 98 371 257 ND 1.770 130 57 293 165 47 1.650 142 63 315 172 63 Mean 1.231 182 82 325 219 45 Std. Dev. 0.441 49 21 30 46 18 1.740 85 33 219 120 34 4 1.550 92 32 200 116 42 2.010 76 32 207 110 88 2.010 62 23 1 73 90 36 1.300 102 56 223 162 33 Mean 1.722 83 35 204 120 47 Std. Dev. 0.306 1 5 1 2 20 26 26 1 7 6 Table AS- Bioconcentration of chlorinated phenolics by leeches (N. gacup) as a function of water pH after seven days exposure at water temperature 12 - C and contaminant concentration 0.1 u.g/L. Water PH 5.1 ± 0.2 Mean Std Dev. 7.1 ± 0.1 Mean Std Dev. 9.0 ± 0.2 Mean Std Dev. Leech wt. (g) 0.575 0.491 0.592 0.428 0 .535 0.718 0.690 0.766 0.655 0.606 0.111 0 .743 0.859 0.514 0.707 0.772 0.995 0.451 0.670 0.714 0 .175 0.696 0.682 0.677 0.585 0 .815 0.630 0.801 0.622 0.567 0.726 0.680 0.084 45-DCG 345-TCG 4,5 ,6-TCG 11.0 13.0 11.0 19.0 8.6 7.6 8.9 13.0 7.4 11.1 3.7 6.6 6.9 12.0 5.0 5.0 5.5 9.6 4.2 6.9 2.7 8.0 5.1 4.2 5.6 3.0 4.3 4.0 6.0 8.2 4.5 5.3 1.7 Tissue Concentation (u.g/kg) TeCG 9.7 15.0 22.0 16.0 14.0 6.9 12.0 8.1 9.5 12.0 17.0 12.0 9.1 8.1 15.0 7.9 13.0 4.9 11.3 3.3 8.7 7.5 7.5 6.0 12.0 9.4 11.0 5.7 9.8 10.0 5.8 2.7 2.5 3.5 2.1 2.5 4.0 3.4 5.9 3.2 8.8 2.1 3.6 1.3 4.3 6.9 7.7 7.5 6.4 3.3 6.8 4.7 6.0 1.6 4.7 2.4 2.4 2.9 1.8 2.1 2.3 2.0 3.0 3.7 2.7 0.9 345-TCVR 13.0 12.0 9.7 11.0 16.0 13.0 11.0 9.3 13.0 9.8 11.0 7.8 15.0 13.0 14.0 10.0 21.0 13.0 10.0 7.1 11.0 7.0 4.1 4.9 8.9 6.0 4.3 9.2 24.0 16.0 9.9 5.2 22.0 21.0 18.0 8.0 16.0 12.3 10.2 8.1 5.2 4.5 4.3 2.1 25.0 6.7 5.7 3.7 6.4 13.0 16.0 7.1 10.5 7.2 32.0 6.4 13.0 19.0 6.7 10.0 11.0 33.0 15.0 9.6 15.6 9.7 177 Table A6 • Bioconcentration of chlorinated phenolics by leeches obscura) as a function of water temperature after seven days exposure at pH 7.0 and contaminant concentration.0.1 ug/L. Water Tissue Concentation (u.g/kg) T p m r i Leech wt. 1 C l 1 l\J-(° C) (g) 45-DCG 345-TCG 4,5 ,6-TCG TeCG 345-TCVR 0.664 6.5 8.6 6.8 4.4 4.7 20.0 0.826 6.0 16.0 4.5 3.8 5.0 0.678 6.5 8.1 4.0 2.7 15.0 0.559 8.3 9.3 4.7 3.3 12.0 0.554 6.8 4.1 3.4 2.4 12.0 0.494 10.0 6.1 7.4 3.1 13.0 0.986 7.3 9.1 9.1 7.3 18.0 0.602 5.4 3.7 4.1 2.7 29.0 0.479 8.4 4.9 7.1 4.9 16.0 0.507 6.3 5.5 8.3 4.7 9.5 Mean 0 .635 7.2 7.5 5.9 3.9 13.4 Std Dev. 0 .162 1.4 3.6 2.0 1.5 7.0 0.447 ND 4.3 4.1 2.9 9.0 11.8 0.635 5.0 5.9 5.3 4.8 22 .3 0 .755 5.8 8.3 6.0 5.7 13.9 0.426 8.8 5.7 5.8 4.4 11.6 0.664 7.3 4.9 4.4 4.1 9.0 0.672 10.3 7.7 9.5 7.2 18.6 0.759 3.6 6.2 5.5 4.2 7.8 0.502 9.5 7.1 6.5 5.4 10.5 0 .683 8.3 7.1 6.5 5.7 11.4 0.436 2.9 4.1 5.4 4.6 12.7 Mean 0.598 6.1 6.1 5.9 4.9 12.7 RtiH nou 0.132 3.3 1.4 1.5 1.2 4.6 0 .633 6.2 2.9 3.8 3.8 24.0 4.4 0 .535 7.3 3.3 3.5 3.5 48.0 0.627 5.8 3.9 2.3 2.5 31.0 0 .563 7.8 6.0 2.4 2.3 13.0 0.410 8.0 4.2 5.0 4.7 26.0 0.421 6.3 2.4 2.9 2.2 35.0 0.493 7.6 4.4 4.4 4.8 34.0 0.600 5.3 4.3 4.3 4.3 21.0 0.494 6.4 2.8 2.9 2.8 24.0 0 .623 6.4 5.1 6.4 5.4 29.0 Mean 0.540 6.7 3.9 3.8 3.6 28 .5 Std Dev. 0.083 0.9 1.1 1.3 1.2 9.4 178 Table A 7 : Bioconcentration of chlorinated phenolics by leeches (N. obscura) as a function of water temperature after seven days exposure at pH 7.0 and contaminant concentration 0.01 ^g/L. Water Tissue Concentation (ug/kg) Temp. Leech wt. (° C) (g) 45-DCG 345-TCG 4,5 ,6-TCG TeCG 3 4 5 - T C V R 0.6 ND 4.1 ND 0.5 3.3 20.0 0.6 ND 3.3 1.8 1.3 1.4 0.7 ND 2.5 0.9 1.0 1.6 0.5 ND 3.7 2.6 2.0 3.3 0.7 ND 2.4 1.1 0.9 1 .5 0.6 ND 1.7 1 1.8 0.7 2.2 0.8 ND 2.0 1.1 0.8 1.1 0.6 ND 1.7 1.1 0.9 2.0 0.6 ND 1.5 1.4 0.8 1.4 0.7 ND 3.5 3.5 1.7 1.6 Mean 0.640 0.0 2.6 1.5 1.1 1.9 Std Dev. 0 .100 0.0 0.9 1.0 0.5 0.8 0.770 ND 1.6 1.0 1.3 ND 12.6 0.546 ND 5.1 1.5 1.7 ND 0.417 ND 3.2 2.9 2.3 ND 0.799 ND 2.8 0.8 2.0 ND 0.666 ND 3.7 0.9 0.9 ND 0.692 ND 4.6 0.8 1.4 ND 0.893 ND 6.4 1.1 ND ND 0.460 ND 6.8 1.0 1.2 ND 0.631 ND 4.6 3.0 1.9 ND Mean 0 .653 ND 4.3 1.4 1.4 ND Std Dev. 0 .158 1.7 0.9 0.7 0 .533 ND ND ND ND 3.3 4.4 0.647 ND 0.9 ND 0.4 4.6 0.421 ND 1.9 ND 0.6 3.5 0 .365 ND 0.9 ND 0.8 5.8 0.689 ND 0.3 ND 0.2 2.6 0 .533 ND ND 0.8 0.9 4.3 0.58 ND 0.3 0.5 0.4 2.9 0 .525 ND ND ND 0.3 2.4 0.62 ND 1.4 0.7 0.4 2.6 0.469 ND 1.0 0.9 1.1 3.2 Mean 0.538 ND 0.7 0.3 0.5 3.5 Std Dev. 0.101 0.7 0.4 0.3 1.1 179 Table A8a : Bioconcentration of chlorinated phenolics by leeches (J±. obscura) as a function of bioassay water stirring, suspended sediment concentration and composition, at water temperature 12 °C, pH 7.1 and contaminant concentration 0.1 iig/L. Suspended Tissue Concentation (|xg/kg) Sediment Leech wt. Cone. (g) 45-DCG 345-TCG 4,5,6-TCG TeCG 345-TCVR (g/L) 0 .486 ND 3.08 2 .50 2.61 35 .0 0 0 .482 ND 3.71 2.54 3.67 26 .7 (Unstirred) 0 .394 ND 4 .56 4 .67 3.91 17.3 0 .629 ND 3 .47 3.38 2 .83 9.5 0 .346 ND 5.16 2 .45 4.51 9.0 Mean 0 .467 ND 4 . 0 0 3.11 3.51 19.5 Std Dev. 0 .108 0 .85 0 .95 0 .78 11 .3 0 .425 7.91 17.4 17.6 15.1 15 .5 0 0 .475 10.4 17 .3 • 17.7 11.8 15.8 (Stirred) 0 .285 12.5 27 .7 26 .0 22 .6 14 .5 0 .364 8 .55 21 .7 21 .2 18.7 14.5 0 .448 9 .70 18.0 18.4 15.4 14,9 0 .440 10.0 23 .0 25.1 18.5 11 .5 0 .293 16.9 22 .6 25 .9 17.4 15 .3 0 .476 15.9 26 .2 28 .9 20.1 13.4 0 .235 ND 27.4 30 .9 21 .9 12.9 0 .384 19.2 30.4 3 1 . 3 22.1 13.2 Mean 0 .383 12.3 23 .2 2 4 . 3 18.4 14.1 Std Dev. 0 .086 4 .9 4 .7 5.3 3.5 1.3 0 .278 10.2 20 .3 25 .0 2 1 . 3 10.0 0.025 g/L* 0 .443 15.4 22.1 28 .2 16.7 53.1 (Stirred) 0.381 20 .9 3 8 . 5 38 .3 28 .8 15.1 0 .477 20.1 34.4 39 .2 31.1 16.1 0 .379 22 .3 4 1 . 7 50 .0 38 .2 7 .98 0 .414 21 .9 4 5 . 7 51.4 34 .3 11.7 0 .342 22 .6 42 .8 48 .9 34.1 10.9 0 .403 25 .2 45 .7 4 8 . 8 38 .5 8.32 0 .438 18.9 35.1 39 .6 24 .0 7 .45 Mean 0 .395 19.7 36 .2 4 1 . 0 29 .7 15.6 Std Dev. 0 .059 4 .5 9.5 9.7 7.6 14.4 0 .406 . 6 . 5 7 23 .7 27 ,6 16.1 6 .0 0.08 g/L* 0 .543 11.5 21 .8 24 .2 16.6 22.1 (Stirred) 0 .296 ND 15.7 24 .6 18 .3 8 .07 0 .380 13.20 26.1 32 .7 26 .7 8 .18 0 .384 6.81 22 .8 26 .8 18.5 14.8 0 .492 8 .16 24 ,0 29.4 19.4 25 .6 0 .614 9 .23 27 ,2 29 .9 20 .6 7.44 0 .345 10.1 22 .9 27 .7 2 3 . 5 10.1 0 .448 9.38 2 2 . 3 26 .8 19.6 8 .40 0 .415 6.21 22 .2 26.1 19.5 8 .27 Mean 0 .432 9.0 22 .9 27 .6 19.9 11.9 Std Dev. 0 .095 2.4 2.9 2.4 3.0 6.4 180 Table A8b : Bioconcentration of chlorinated phenolics by leeches (N_. obscura) as a function of bioassay water stirring, suspended sediment concentration and composition, at water temperature 12 °C, pH 7.1 and contaminant concentration 0.1 iig/L. Suspended Tissue Concentation (|xg/kg) Sediment Leech wt. Cone. (g) 45-DCG 345-TCG 4,5,6-TCG TeCG 345-TCVR (g/L) 0.401 21.2 36.6 40.7 21.8 8.43 0.08 g/Lt 0.351 15.1 28.2 35.1 19.5 7.77 Inorganic 0.418 12.8 30.4 38.4 19.3 23.2 (Stirred) 0.302 14.9 32.5 37.6 20.2 9.81 . 0.414 14.6 36.2 43.1 26.1 6.56 0.468 18.6 35.9 45.4 24.6 12.3 0.438 12.8 33.8 38.7 26.5 7.57 0.342 14.3 37.1 48.6 27.1 9.66 0.370 11.3 23.0 34.0 17.8 8.33 Mean 0.389 15.1 32.6 40.2 22.5 10.4 Std Dev. 0.052 3.1 4.7 4.8 3.6 5.1 0.451 26.0 43.1 47.4 36.1 14.79 0.15 g/L* 0.302 11.0 20.2 24.8 17.8 7.21 (Stirred) 0.325 13.2 29.8 37.0 21.6 9.9 0.444 15.5 35.0 43.8 33.3 8.54 0.305 19.5 34.2 43.7 28.7 11.01 0.367 11.8 28.0 31.7 19.6 7.1 0.367 12.5 33.5 35.0 25.4 39.38 0.376 6.6 25.0 28.8 20.3 10.69 0.341 9.0 29.5 32.7 24.4 8.35 Mean 0.364 13.9 30.9 36.1 25.2 13.0 Std Dev. 0.054 5.9 6.6 7.6 6.3 10.1 * Suspended material: size < 63 pm, 5% organic : 95% inorganic material, t Suspended material: size < 63 pm, 100% inorganic material. 181 APPENDIX 4 Raw Data Field Monitoring Trials 182 183 O cri i— CD ^ C CM O CD oo ? CO co i_ CM CD > tr CM i_ CD 1 — CO <N CO . LL °2 CD sz n — CD C LL TJ " ° £ a o CD T -S| CO -.CJ -tf o C ' CD » o T _ c CO CJ) c 'co 3 _co co £ CD c 00 c CD co O . 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S 2 S tf f- r CM Q Q Q D Q z z o CM O CO CM O O co in CO CM CO CM O CO co in q cd in co 0 Oi CO CM CO CM CO CM CO To o £ > co O CD 5 -d 55 191 Table A18: Bioconcentration of chlorinated phenolics by the leech species Percvmoorensis marmorata after seven days laboratory exposure to various contaminant concentrations at T = 12.5 °C and pH = 7.5. Water Tissue Concentation (p.g/kg) Cone. Leech wt. (ng/L) (g) 45-DCG 345-TCG 4,5 ,6-TCG TeCG 345-TCVR 2.200 110 190 1 55 128 3 2 0.5 2 .230 7 5 122 1 1 1 9 7 2 0 2 .410 128 151 138 131 58 2 .290 1 11 139 142 122 4 6 Mean 2 .283 106 151 137 120 39 Std Dev. 0 .093 22 2 9 1 8 1 5 1 7 2 .230 288 287 319 240 39 1.0 3.070 466 443 529 350 4 3 2.240 242 235 266 228 4 0 3.090 249 247 302 204 5 6 Mean 2.658 311 3 0 3 354 256 4 5 Std Dev. 0.488 105 96 119 65 8 1.660 1300 1070 1440 673 143 5.0 3.780 1290 1290 1270 8 1 3 163 0.779 1520 1070 2130 1630 2 5 7 2 .050 168 1380 1870 1 180 191 Mean 2.067 1070 1203 1678 1074 189 Std Dev. 1.260 610 157 393 428 4 8 1.370 3 9 9 0 2360 4 9 3 0 3370 2 8 9 10.0 3.270 2 2 5 0 1820 2 7 4 0 1230 452 1.560 3060 2 1 0 0 3410 2 6 8 0 2 0 3 2 .980 2 5 1 0 1960 3 0 7 0 1330 94 Mean 2.295 2 9 5 3 2060 3 5 3 8 2 1 5 3 2 6 0 Std Dev. 0.969 770 230 968 1047 151 192 Appendix 5 Predictions of field water contaminant concentrations using measured leech bioconcentration and laboratory regression relationships Sample Calculation Predicted water concentration, ug/L 3,4,5-TCG; July 1991. Predicted water concentrations were calculated from the laboratory relationship between water contaminant concentration and bioconcentration, with rearrangement of the equation to yield water concentration from observed bioconcentration. Correction factors for differences in water temperature, pH, leech weight and water current between measured field and laboratory experiments were multiplied by the initial predicted concentration. No formal regression relationship water was developed for the effect of water current on bioconentration. The correction factor for the effect of water current divided by the water concentration predictions was calculated as the inverse of the percent increase observed between bioconcentration in static and stirred bioassays. Measured Field Values Average Measured leech bioconcentration = 28 fig/kg (Table 4.14) Average measured water concentration = 0.052 ug/L (Table 4.11) Water temperature = 12.0 °C (Table 4.8) pH = 7.8 (Table 4.8) Average leech weight = 1.07 g (Table 4.14). Predicted Water Concentration, |xg/L. 1) Water concentration as a function of leech bioconcentration at water temperature 12 °C, pH 7.5, average leech weight 1.17 g (n = 25). a) 3,4,5-TCG (ng/kg) = 74(water cone. ng/L)expt.0.77 = 74(0.052)expt.0.77 (Table 4.1) Rearranged to: Water cone. (iig/L) = 0.77 Bioconc (ng/kg)/74 Water cone (ug/L) = 0.77 28/74 = 0.283 2) Correction factors used for adjustment of calculated value for differences in laboratory and field conditions. Correction factor = Predicted lab value/Predicted field value b) Water temperature: no significant effect 11.8 - 20 °C c) Water pH: 3,4,5-TCG (ng/kg) = 33.1 (0.86)expt. pH (Table 4.3) Field water pH = 7.8: 33.1(0 86)expt.7.8 = 10.21 Laboratory water pH = 7.5: 33.1(0.86)expt.7.5 = 10.67 Correction factor = 10.67/10.21 = 1.045 d) Leech weight: 3,4,5-TCG (jig/kg) = 68(weight)expt.-0.83 (Table 4.4) Field leech weight = 1.07 g: 68(1 07)expt -0.83 = 64.28 Laboratory leech weight = 1.17 g: 68(1 17)expt.-0.83 = 59.69 Correction factor = 59.69/64.28 = 0.929 3) Predicted water concentration (static bioassay water) 3,4,5-TCG (ug/L) = 0.283 ug/L * 1.045 * 0.929 = 0.275 ug/L. 4) Water current: no laboratory regression relationship: Correction factor for presence of water current in laboratory = 5.80 * bioconcentration without current (Table 4.7) Inverse of correction factor = 1/5.80 = 0.172 5) Predicted water concentration (continuous current) 3,4,5-TCG (ug/L) = 0.275 ug/L * (1/5.80) = 0.047 jig/L. (Table 4.20). 

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