UBC Theses and Dissertations

UBC Theses Logo

UBC Theses and Dissertations

All dredged up and no place to go : the disposal of contaminated dredged material from greater Vancouver,… Gorham, Richard Arthur 1985

Your browser doesn't seem to have a PDF viewer, please download the PDF to view this item.

Item Metadata


831-UBC_1985_A6_7 G67_3.pdf [ 10.49MB ]
JSON: 831-1.0100294.json
JSON-LD: 831-1.0100294-ld.json
RDF/XML (Pretty): 831-1.0100294-rdf.xml
RDF/JSON: 831-1.0100294-rdf.json
Turtle: 831-1.0100294-turtle.txt
N-Triples: 831-1.0100294-rdf-ntriples.txt
Original Record: 831-1.0100294-source.json
Full Text

Full Text

A L L D R E D G E D UP A N D NO P L A C E T O GO: T H E DISPOSAL OF CONTAM INATED D R E D G E D MATER IAL F R O M G R E A T E R V A N C O U V E R , BRITISH COLUMBIA, INTO T H E NEIGHBOURING STRAIT OF GEORGIA By R ICHARD A R T H U R G O R H A M LL.B., The University of Western Australia, 1974 B.Sc, The University of British Columbia, 1979 A THESIS SUBMITTED IN PART IAL F U L F I L L M E N T OF T H E REQUIREMENTS FOR T H E D E G R E E OF MASTER OF SCIENCE in T H E F A C U L T Y OF G R A D U A T E STUDIES (Resource Management Science) . We accept this thesis as conforming to the required standard T H E UNIVERSITY OF BRITISH COLUMBIA September 1985 © R i c h a r d Arthur Gorham, 1985 In presenting this thesis in partial fulfilment of the requirements for an advanced degree at the University of British Columbia, I agree that the Library shall make it freely available for reference and study. I further agree that permission for extensive copying of this thesis for scholarly purposes may be granted by the head of my department or by his or her representatives. It is understood that copying or publication of this thesis for financial gain shall not be allowed without my written permission. Department of The University of British Columbia 1956 Main Mall Vancouver, Canada V6T 1Y3 DE-6(3/81) i i A B S T R A C T The thesis provides an integrated review and analysis of strategic scientific information from which management procedures for the environmentally acceptable disposal of dredged materials from Greater Vancouver into the Strait of Georgia are determined. An analysis of reported and suspected contamination of Vancouver's waterways identifies trace metals and hydrophobic organic chemicals that warrant concern by authorities responsible for the management of dredged material disposal. The processes, transformations and pathways of these contaminants in the marine environment subsequent to dredged material disposal are reviewed. It is argued that areally confined degradation of suitable disposal sites is of negligible concern, but that release of contaminants from the disposal area, were it to occur, could have unpredictable and perhaps substantial environmental conse-quences. The major potential pathway of contaminant release from dredged material during or subsequent to disposal is via the resuspension and transport of fine particulate material. Biological accumulation of contaminants and their trans-port through the food chain is a potentially significant release pathway for hydrophobic organic contaminants. Biological transformation to more soluble metabolites is also potentially significant for mercury and some of the less chlorinated and lower molecular weight organic compounds of concern. Desorptive release of contaminants from the disposed sediments into solution is usually negligible, with the possible exceptions of cadmium and mercury. i i i A review of the physical factors that promote contaminated sediment erosion and dispersion during or subsequent to dredged material disposal identifies oceanographic characteristics where such release will be minimal. There are only four areas within the Strait of Georgia that exhibit these characteristics. Two of these areas are adjacent to valuable and sensitive biological resources and are consequently unsuitable as ocean dumping sites. Dredged material disposal at the other two sites, one near Smelt Bay, Cortes Island, and the other off McNaughton Point, Sechelt Peninsula, should result in minimal adverse environmental impact. Available methodologies to identify dredged materials with contaminant levels that preclude their environmentally acceptable disposal at these recommended sites are discussed. Title Page i v T A B L E OF CONTENTS Page No. Acknowledgements ix 1.0 INTRODUCTION I 1.1 Study Purpose 8 2.0 A SCIENTIFIC PERSPECTIVE I I 2.1 Containment or Dispersion II 2.2 Release 14 2.3 The Chemistry of Contaminant Behaviour 15 2.4 Chemicals of Environmental Concern 19 3.0 DREDGING AND D R E D G E D MATERIALS IN T H E LOWER MA INLAND 22 3.1 Dredging in the Vancouver Region 22 3.2 Sediment Contaminants in the Vancouver Region 25 3.2.1 Trace Metals 36 3.2.2 Organic Contaminants 42 3.3 Conclusion 56 4.0 T R A C E M E T A L PATHWAYS AND PROCESSES IN T H E MARINE ENVIRONMENT 59 4.1 Introduction 59 4.1.1 Trace Metal Speciation 62 4.2 Trace Metal Release During Water Column Passage 66 4.2.1 Salinity Alteration 67 4.2.2 Alteration of Redox and pH Conditions 70 4.2.3 Field Monitoring 75 4.3 Trace Metal Release from Settled Sediments 77 4.3.1 Chemical Release Processes 77 4.3.2 Biological Release Processes 81 4.3.3 Conclusion 1 92 4.4 Environmental Effects of Released Trace Metals 96 4.4.1 Cadmium 96 4.4.2 Mercury 99 V Page No. 5.0 ORGANIC P O L L U T A N T PATHWAYS AND PROCESSES IN T H E MARINE ENVIRONMENT 104 5.1 Introduction 104 5.2 Release of Organic Contaminants to the Water Column 106 5.3 Biological Uptake of Organic Contaminants 117 5.4 Biological Transformation 127 5.4.1 Microbial Degradation 128 5.4.2 Metabolism by Higher Organisms 129 5.5 Properties, Bioaccumulation and Metabolism of Organic Contaminants of Concern 131 5.5.1 Polychlorinated Biphenyls (PCBs) 5.5.2 Hexachlorobutadiene (HCBD) 136 5.5.3 Polychlorinated Dibenzofurans (PCDF) 138 5.5.4 Pentachlorophenol (PCP) 140 5.5.5 Chlorinated Benzenes 141 5.5.6 Polyaromatic Hydrocarbons (PAHs) 144 5.5.7 Phthalate Esters (PAEs) 150 5.6 Conclusion 152 6.0 • PROCESSES E F F E C T I N G T H E PHYSICAL DISTRIBUTION OF DISPOSED D R E D G E D MATER IAL 157 6.1 Introduction 157 6.2 During Water Column Passage 158 6.3 Subsequent to Settling 161 6.3.1 Physical Factors Effecting Sediment Resuspension 161 6.3.2 Biological Factors Effecting Sediment Resuspension 165 6.4 Conclusion 167 7.0 DISPOSAL SITE OPPORTUNITIES WITHIN GEORGIA STRAIT 170 7.1 Recommended Disposal Site Characteristics 170 7.2 Identification of Potential Disposal Sites 173 8.0 RESOURCES IN CONFL ICT 182 8.1 Introduction 182 8.2 Groundfish 185 8.3 Bent hie Crustaceans > 189 8.4 Molluscs 191 8.5 Pelagic Fish 196 8.6 Marine Mammals 202 8.7 Marine Birds 204 8.8 Human Recreation 209 8.9 Parks and Reserves 21 I 8.10 Cone lusion 213 Page No. 9.0 CONCLUSIONS 221 9.1 Introduction 221 9.2 Dredged Sediment Criteria 225 9.3 The Need for Field Assessment 230 10.0 L I T E R A T U R E CITED 233 VI 1 LIST OF FIGURES Figure Page No. 1. Location map - Strait of Georgia 4 2. Location map - Greater Vancouver 5 3. Schematic representation of contaminant transport and fate 16 k. Freundlich adsorption isotherm 108 5. Bioaccumulation of six chlorinated benzenes in guppies 121 6. Relationship between current, speed, particle diameter, and sediment erosion, transport and deposition. 163 7. Georgia Strait - Location of proposed dumpsites 175 8. Location of proposed North Stuart Channel dumpsite (Area I) 176 9. Location of proposed McNaughton Point dumpsite (Area 2) 177 10. Location of proposed Comox and Denman Island dumpsites (Area 3) 178 11. Location of proposed Smelt Bay dumpsite (Area k) 179 v i i i LIST OF TABLES Table Page No. I. Comparative towing costs for Point Grey and ocean disposal sites. 6 II. Quantity of dredged materials disposed at dumpsites adjacent to the lower mainland, 1976-79. 24 III. Contaminant concentrations in False Creek and Burrard Inlet sediments. 26 IV. Potential anthropogenic sources of trace metal contamination of Vancouver's water-ways. 29 V. Canadian Environmental Contaminants Act list of priority chemicals. 31 VI. EPA priority pollutant list showing classi-fication by extraction group. 33 VII. Priority ratings for industrial chemicals in British Columbia and the Yukon. 34 VIII. Contaminants of concern identified or sus-pected within Vancouver sediments. 37 IX. Trace metals in sediments of the Fraser River. 38 X. Levels of metals in sediments from the Fraser River and the Columbia River, B.C. 40 XI. Summary of incidents of human poisoning by toxic metals in the aguatic environment of Japan. 61 XII. Mechanisms for trace mineral bonding. 64 XIII. Potential dredged material disposal areas in Georgia Strait. 174 XIV. Species names of organisms cited in Section 8.0. 184 XV. Quantification of resource conflicts at the proposed ocean dumping sites. 214 XVI. Comparative towing costs for recommended disposal sites. 218 i x A C K N O W L E D G E M E N T S The successful completion of this thesis is in no small part due to my wife, Catherine, who has given me encouragement and support (as well as a charming daughter) during its preparation. I also owe extreme gratitude to Maria Mees for her uncomplaining and extremely efficient word processing of the manu-script. 1 INTRODUCTION The environmental management of dredged materials is faced with the residue of past neglect. Urban, agricultural and industrial activities adjacent to many areas of our coast have resulted in accumulations of contaminants within the sediments. The sediments of protected bays and estuarine areas are especially prone to interrment of contami-nants. Sediments that have been contaminated by chemicals from surface runoff and groundwater inflow, or from industrial and domestic waste discharges, present a difficult problem when dredging is required. The removal and relocation of coastal and harbour sediments is often essential to establish and maintain navigational waterways and port facilities. However it can also promote the release of potentially deleterious materials into the pelagic and benthic environment. If this release tends to favour uptake of the contaminants by marine organisms, then either short- or long-term deleterious effects may occur. Dredging and dredged material disposal involves the mechanical agitation of the sediments, their exposure to the water column and their relocation into a physical, chemical and biological environment that may differ substantially from the dredged site. Contaminant release may result from each of these processes. 2 Water quality perturbations at the dredge site generally subside rapidly upon cessation of dredging activities (Lee, 1976; Fulk, Gruber and Wullschleger, 1975; Wakeman, 1976; Hoos, 1976). Mitigative measures (e.g. techniques, timing and rate of dredging) may be necessary to reduce potential impacts upon particularly sensitive biota, but our dependence upon navigable waterways will otherwise generally override our concern for the short-term environmental effects at the dredge site. For the environmental manager the primary issue concerning con-taminated dredged material is related to disposal. The options are limited: on land, into the coastal sea, or into the deep ocean. The most economical of these is the coastal sea but this option requires careful and critical examination so that we may be assured that the valuable resources of this region are preserved and protected for future generations. Vancouver, British Columbia, is the largest urban-industrial centre and the major port in western Canada. Dredging activities are crucial to the maintenance of the port facilities (Willams, 1976). Between three million and six million tonnes of sediments are annually dredged from the region (H. Nelson, EPS, pers. comm., 1985). Although some of these materials can be disposed at foreshore sites, .lack of available sites, concerns for the environmental effects of upland disposal, and the high engineering costs of this disposal methodology, require that substantial quantities of dredged material are marine disposed. 3 Vancouver is located on the south-eastern edge of the Strait of Georgia, which comprises a dynamic estuarine system supporting a variety of valuable fisheries and commercial and recreational act i -vities (Figures I and 2). Due to its limited municipal and industrial development, classical pollution impacts have rarely been observed and the region has generally been perceived as nearly pristine (Waldichuk, 1983) . The disposal of dredged materials from Vancouver into the Strait of Georgia has recently induced considerable and increasing controversy. The nearest marine alternative to disposing of dredged material in the Strait of Georgia is in the offshore deep ocean to the west of Vancouver Island. The use of this alternative was recently required by the Environmental Protection Service for the disposal of contaminated sediments from False Creek (Brothers and Sullivan, 1984) . Although the use of the deep ocean may lessen the potential environmental costs of dredged material disposal, it adds substantially to the engineering costs of a dredging program (Table I) and may not always be preferred. Dredged sediments from the Greater Vancouver area , vary from uncontaminated natural sediments to deposits that contain substantial levels of toxic pollutants. The environmental hazard associated with the appropriately managed disposal of natural sandy sediments in Georgia Strait is negligible (Levings, 1982). Short term turbidity, possible substrate alteration and burial of benthic organisms may result, but these effects are transient and rarely cause measurable 4 125° 124* 123° ' FIGURE 1. Location Map 0 C o m p i l e d f r o m C a n a d i a n E n e r g y , M i n e s and R e s o u r c e s Map N M - 9 / 1 0 6 TABLE I . Comparative Towing C o s t s f o r P o i n t Grey and Ocean D i s p o s a l S i t e s . 3 Component D i s p o s a l S i t e s P o i n t Grey Ocean Tug r e q u i r e m e n t s 500 HP 2400 HP + 1200 HP Tug h i r e $170/hr $8000/day + $6000/day Barge r e q u i r e m e n t s 1000 m3 7000 m3 i n t o t a l 6 Barge h i r e $300/day $1500/day b T r i p d u r a t i o n 12 hrs 4 days T r i p c o s t $2340 $ 6 2 , 0 0 0 Cost per m3 $ 2 . 3 4 $ 8 . 8 6 C a . S o u r c e : P e r s o n a l communication w i t h Mr K. Longbottom, Seaspan I n t . , 1984. C o s t s f o r ocean d i s p o s a l are based on the use o f a dumpsite r e c e n t l y used by B . C . P l a c e C o r p . to d i s p o s e o f c o n t a m i n a t e d F a l s e Creek s e d i m e n t . I t i s 50 n a u t i c a l m i l e s west o f B a r k l e y Sound. b. The two 2000m and one 3000m barges used by B . C . P l a c e C o r p . were l e a s e d from San D i e g o , none b e i n g a v a i l a b l e i n t h i s s i z e range i n B r i t i s h C o l u m b i a . H i r e r a t e s used here are quoted w i t h t h e a s s u m p t i o n t h a t , i f ocean d i s p o s a l became t h e common p r a c t i c e , l o c a l o p e r a t o r s would p r o v i d e t h i s s i z e barge under c o m p e t i t i v e c i r c u m s t a n c e s . c . T h i s e s t i m a t e r e f l e c t s c o s t s f o r a r o u t i n e o p e r a t i o n . B . C . P l a c e ' s r e c e n t d i s p o s a l o f 3 2 5 , 0 0 0 m3 o f c o n t a m i n a t e d s u r f a c e s e d i m e n t s from F a l s e Creek i n t o deep ocean water c o s t $ 3 . 2 m i l l i o n ( .pers. comm., R. W a t e r s , B . C . P l a c e C o r p . , 1984). T h i s equates to a t o w i n g c o s t o f $ 9 . 8 2 / m 3 . 7 long term degradation of the environmental quality at appropriate disposal sites (Peddicord, 1980). Upward migration of buried infauna and recolonization of the disposal mound from adjacent areas will usually restore the disposal site to its original or equivalent produc-tivity within the short term (Levings, 1982). The fact that clean sedimentary material can be disposed cheaply within Georgia Strait without significant ecological impact justifies the continued use of the Strait for this purpose. The guestion at issue thus becomes one of appropriate management and degree. Between natural and grossly polluted sediments there exists a continuum of degrees of contamination. Up to some level of contamination, disposal sites and methodologies exist that will allow disposal within Georgia Strait without significant impact. However the selection of inappropriate disposal sites, the use of unsuitable disposal methodo-logies, or the disposal of material that exceeds a particular level of contamination, may result in sufficient environmental degradation within the Strait that alternatives should be sought. The objectives for management of dredged material disposal into the Strait are therefore to identify disposal sites and methodologies that will ensure minimal ecological impact, and to determine the limits to sediment guality that may be appropriately disposed under such conditions. 8 I.I Study Purpose This study analyses and integrates strategic information pertaining to the disposal of dredged material from Greater Vancouver into the neighbouring Strait of Georgia. The purpose is to use existing knowledge to develop ecologically appropriate procedures for the coastal marine disposal of Vancouver's dredged materials. Consideration of available strategies for disposal is confined to the use of the Strait of Georgia, which is the least expensive and most commonly used of the land/ocean/coastal sea disposal options. At issue is whether, how and where dredged material disposal into this resource rich coastal zone can be accomplished with minimal adverse environmental impact. The study aims to provide a rational scientific basis to decision making in the management of dredged material. The ultimate selec-tion of an appropriate management strategy depends not only on rational scientific factors, but also upon multifarious socio-economic factors, including the tangible and intangible values to society of maintaining the resources of the Strait of Georgia in their present bountiful state. Although environmental science cannot, and should not, presume authority to adjudicate upon the intangible socio-economic values of alternate resource uses, it has the task to reduce the myths and misunderstandings which are too often an influence on their reckoning. To accomplish this reguires a delicate balance between interpreting and articulating the scientific knowledge in a practical language, whilst recognizing and avoiding the imposition of hidden subjective judgements. I therefore attempt to remain cog-9 nizant of, and to describe, the socio-economic factors that impinge upon this review, but I do not deliberate upon them at length. Their determination is properly left to the political system. Accordingly, the objectives of this study are: o To define the scientific criteria for assessing the environmental hazard associated with disposal of dredged materials and from these, to identify the critical environmental objectives for which strategies for dredged material disposal should aim. o To determine the nature and distribution of contamination of Vancouver's sediments and to identify the contaminants of concern to the management of dredged material disposal. This will be accomplished by compiling and interpreting relevant existing information, and reviewing the occupation of Vancouver's shoreline areas to identify the potential sources of contaminant release. o To review the current state of knowledge describing the pathways, transformations and effects of contaminants in the marine environment, taking into account the quality of sediments shown or estimated to exist in the Vancouver region. This review will identify and discuss the environmental processes that induce potentially adverse impacts from dredged material disposal, arid will illuminate the uncertainties that constrain definitive assessment of these impacts. 10 o To identify environmental characteristics that minimize the potential for adverse impacts from dredged material disposal, and to delineate areas of the Strait of Georgia which correspond to these characteristics. o To describe and compare the resource uses near to these delineated areas with which dredged material disposal might conflict, and to evaluate the dependence of these areas upon unperturbed environmental conditions. From this analysis, to identify sites in the Strait of Georgia where dredged material disposal will result in minimal environmental degradation. o To determine the limits to disposal of contaminated sediments at the recommended sites. Some sediments in the Vancouver region may be too highly contaminated for disposal within coastal waters. How do we distinguish between materials that will be accommodated at the recommended disposal sites without signi-ficant adverse ecological impact and those where the environ-mental risks associated with coastal zone disposal remain unacceptable? Although this study pertains directly to the disposal of dredged materials from Vancouver into the Strait of Georgia, the analytical framework and many of the principles applied are equally applicable elsewhere. 11 2 . 0 A SCIENTIFIC PERSPECTIVE 2 . 1 Containment or Dispersion An initial and fundamental argument that is developed within this thesis is that for disposal of dredged materials in coastal waters to have minimal environmental impact, the disposal strategy must maximize containment of the materials within the selected disposal site. This is in contrast to the strategy presently adopted at off-shore deep ocean dredged material disposal sites, where the dispersion of the disposed materials is usually preferred. In this latter case, given the homogeneity and low biological productivity of the deep ocean ecosystem, dilution offers itself as the answer to pollution. However, if contaminated sediments are to be disposed into coastal waters, then dispersion of the contaminants imposes unpredictable and perhaps substantial environmental risk. Even where coastal waters are well flushed, contaminants that disperse may be transported into fragile or highly valued environments before dilution processes ameliorate the environmental hazard. The complexity of nearshore current regimes will frequently preclude adeguate monitoring of the environmental effects of a dispersed contaminant loqding. Dispersion induced dilution of the contaminant loading from a disposal operation may not reduce the associated environmental hazard, although it will probably spread it beyond the detection limits of a practicable monitoring program. 12 A management preference towards confinement of the disposed materials to the dump-site implies a preference for more intense but areally confined environmental degradation rather than diluted degradation over a more extensive area. This may seem to run counter to the common regulatory allowance for mixing and dilution factors in assessing the environmental risks associated with most anthropogenic discharges to the marine system (e.g. water quality criteria enforcement, see U.S. EPA, 1976). But for most marine discharges confinement of the pollution to the discharge area is not possible, so in the absence of alternatives, rapid dilution of the environmental effect becomes important. Other factors (e.g. bio-availability) being equal, the bioaccumulation and toxicity of pollutants are essentially dose dependent. If ocean dumped dredged materials can be effectively contained within the disposal site, a more relevant analogy to the ecological objectives of ocean dumping management is with management strategies for waste disposal at terrestrial sites. The management objectives in these instances are generally towards containment, minimizing dis-persion of the pollution through the atmospheric, aquatic or biotic pathways. Periodic disposal of contaminated dredged materials onto the seafloor will likely result in severe depauperation of the biological community at the disposal site. However, environmental degradation at an areally constrained site is of little concern if the following conditions are met: 13 1. The disposal site is neither highly productive nor ecologically unique. 2. Dredged sediment disposal at the site does not conflict with other beneficial resource uses of the environment. 3. The detrimental alteration of the environment will not extend beyond specified site boundries. k. Upon cessation of disposal activities, the site can be returned, or will naturally return over time, to an "acceptable" environmental condition.' Fulfillment of these conditions has justified the use of many areas of our terrestrial sphere for the disposal of anthropogenic wastes. Indeed, it is on this basis that upland disposal of contaminated dredged materials has often been recommended2 (e.g. U.S. E P A / C O E , 1977). Our reluctance to similarly recommend disposal of wastes into the marine sphere stems primarily from concern for the uncontrolled redistribution of waste components subsequent to their disposal into this fluid medium. However, this concern applies primarily to contaminant loadings that breach one or all of the four conditions nominated above. If the marine disposal of dredged sediments can be managed to comply with the above conditions, then there is no obvious reason why it should not be preferred over upland disposal. The reluctance to use the ocean in this instance may be said to place greater weight on preserving the 1. Because a disposal site may be recolonized by a substantially different biological community than was present before disposal, the "acceptability" of the rehabilitated condition includes both objective (e.g. biomass, productivity, species diversity) and "subjective (e.g. equivalence of the replacement community) judgements. 2. This is despite the high likelihood of trace metal leaching and dispersion that accompanies this method (see Lee et al., 1975). 14 marine environment than on preserving the terrestrial environment, avoiding health risks due to groundwater contamination and other social goals. Release Given an acceptance of areally constrained detriment at a suitably located disposal site, the critical factor in determining the potential hazard of contaminated dredged material disposal is the degree to which the contaminants may be remobilized or released from the disposed sediments to impact adjacent environments. Without con-taminant release the guestion is merely whether we relocate our inshore pollution. But if contaminant release occurs, the act of relocation accentuates the potential harm, adding unpredictable and perhaps substantial environmental cost to the dredging program. When dredged materials are disposed into the marine environment, they are mixed with the oxygenated seawater during their descent to the bottom. Dissolution may occur, and a fraction of the particles will remain suspended within the water column. Pore water will also be partially released to the water column. The disposed sediments which settle to the bottom will be subject to physical and biological resuspension processes at the disposal site. In addition, the chemical environment in which they have settled may be substantially different from where they were dredged. Chemical 15 transformations may occur that permit long term release to the pore water, and chemical flux from the sediment system to the water system may thereby be promoted. The questions at issue are the extent to which these physical, chemical and biological processes at the disposal site cause release of contaminants and the degree to which the contaminants released due to the disposal operation effect the biological system. The Chemistry of Contaminant Behaviour By definition, contaminants of concern in dredged sediments have a strong association with the sediment phase. They are relatively insoluble in the water phase, so their introduction to the aguatic ecosystem has resulted in an eguilibrium condition where they have sorbed (adsorbed or absorbed) to suspended particulate matter and have been deposited on the bottom when the suspended matter settled. Although the sediments act as a "sink" for these contaminants in aquatic ecosystems, the water and biological systems can also become enriched. Figure 3 describes a schematic representation of the transport and fate of contaminants introduced to the marine environment. The contaminants occur in both the sediment system and the water system as dissolved or sorbed species, with two directional change taking place across the solid-water interface. Physical (sedimentation and resuspension) and chemical exchange 16 FIGURE 3. Schematic Representation of Contaminant Transport and Fate in the Marine Environment. INTERNAL DISTRIBUTION EXTERNAL LOSS WATER DISSOLVED Sorption Partitioning SUSPENDED PARTICLES SEDIMENTS Direct Uptake Indirect Uptake BIOTA <« Depuration, Excretion PELAG » C/EPIBENTHIC. , Foodcha i n processes BENTHIC PARTICULATE Sorption Parti t i o n i ng INTERSTITIAL WATER Indirect Uptake Direct Uptake AOVECTIVE TRANSPORT CONSUMPTION IRREVERSIBLE BURIAL 17 occurs between the sediments and the overlying water column. Biological organisms take up the contaminants from either the water column or from the sediments. When contaminants are released into the ecosystem, a new equilibrium between the respective concentrations in the sediments, the water column and the biota has to be established. This new equilibrium may tend to favour greater dissolution, or more sorption of the con-taminants by biological organisms. An understanding is required of the mechanisms which control the forms and distributions of the contaminants between environmental compartments and the responses of these mechanisms to environmental alteration. The partitioning behaviour of pollutants in the marine environment may be described in terms of their relative chemodynamics in the sediment, water and biotic phases. Spontaneous transfer of chemical and thermal energies occurs across the interfaces between these three phases until an equilibrium condition is reached. If we assume that the net chemical transfer is the sum of reversible transfers between the environmental phases, thermodynamic considerations require that at equilibrium, the chemical "escaping tendencies" of pollutant molecules in each phase will be egual (Thibodeaux, 1979). Alteration of the equilibrium position results from changes in the relative "escaping tendencies" of the pollutant, which is dependent upon the chemical reactivity of the pollutant molecule with each phase. 18 An important distinction exists in this regard between the chemistry of trace metals and hydrophobic organic compounds. Trace metals are simple cations that exist in a number of chemical forms within the environment, and are essentially indestructable. They may cycle repeatedly through the different chemical forms dependent upon the chemistry of their environment. Metals react as electron-pair acceptors (Lewis acids) with electron-pair donors (Lewis bases) to form various chemical groups such as an ion pair, a metal complex, a coordination compound or a donor-acceptor complex. Inorganic and organic complexes are formed, and their distribution between the sediment, water and biotic phases is influenced by the diverse and dynamic chemical interactions of these many associations. In contrast, toxic organic compounds of concern to ocean dumping management are large non-polar molecules which exhibit minimal variation in their speciation. Chemical reactions involving these compounds in the natural environment are generally very slow and are uni-directional. They result in the formation of new compounds within a series of reactions leading ultimately to either the complete destruction of the compound (metabolized to CO2 and H2O) or to the incorporation of fragments into other organic polymers (e.g. humic acids). This difference in the environmental chemistry of trace metals and organic contaminants necessitates their separate consideration when discussing their availability to the biological system. The reactivity of trace metals in the environment requires analysis of the processes 19 which alter the equilibrium condition of trace metal speciation, and hence effect bioavailability. Toxic organic compounds are essentially mono-specific: their availability to the biological system is simply a guestion of their exposure to, and their tendency to incorporate with, the biological system. Chemicals of Environmental Concern Trace metals are ubiquitous in nature. They are natural components of soils and sediments and most are required nutrients for biological organisms. As implied by their nomenclature however, these elements normally occur only at very low levels in natural systems (parts per million or less). At higher concentrations, many are toxic. Although the term "trace metals" usually includes all metals found in trace amounts in the marine environment, my concern is with trace metals that are highly toxic to marine biota. These include mercury, cadmium, lead, chromium, copper, zinc, arsenic, nickel, silver and selenium. While considerable quantities of organic matter of biogenic origin are natural components of all ecosystems, the organic compounds of concern to ocean dumping management are principally of anthro-pogenic origin, either due to limited natural production or because the compounds are totally man-made. Chlorinated pesticides such as DDT, and other organohalogen compounds such as polychlorinated biphenyls (PCBs), terphenyls (PCTs), dibenzofurans (PCDFs), benzenes (e.g. HCB), phenols (e.g. PCP) and butadienes (CBDs), do not exist 20 naturally and have become notorious environmental contaminants of worldwide significance. Although petroleum and related hydrocarbons are ubiquitous in nature, industrial and domestic activities have resulted in substantially elevated loadings of these compounds in nearshore marine sediments. For example, Hites, LaFlame and Farrington (1977) measured polyaromatic hydrocarbon (PAH) levels in core samples from Buzzard's Bay, Massachusetts, and found that phenanthrene, flouranthrene and pyrene started to increase signifi-cantly in sediment layers indicative of the period 1850 to 1900. The persistence and high toxicity (and carcinogenicity) of many petroleum hydrocarbon derivatives such as these give cause for appreciable concern about their distribution, fate and effects in the marine environment. The organic pollutants of primary concern to dredged material management are the relatively insoluble compounds that accumulate and persist within the sediments. More soluble organic pollutants in the marine environment are not necessarily of less environmental concern, but because they are more prone to advective transport, hydrolysis and/or biodegradation, they are not of such direct concern to the management of dredged material disposal. High levels of more soluble contaminants within dredged sediments is indicative of a recent spill or discharge. This warrants concern, but is essentially outside the scope of this thesis. 21 Highly halogenated organic compounds, as well as the higher molecular weight non-polar petroleum hydrocarbons, are generally not soluble in surface waters at concentrations higher than a few parts per billion (ug/L). Most water column concentrations are found associated with either biological organisms or suspended sediments (Poirrier, Bordelon and Laseter, 1972). Their tendency to associate with the solid phase is a direct product of hydrophobic interactions, so is an inverse function of their solubility (Chiou, Peters and Freed, 1979). The environmental concerns for sediment-bound organic pollutants pertain to their persistence in the environment, their toxicity and/or carcinogenicity, and their tendency to accumulate to high levels in animal tissues. As will be outlined in Section 5.0 below, the per-sistence and bioaccumulation potential of water-insoluble organic pollutants are largely related to their water solubility. This relationship is so significant that Lu, Metcalfe and Cole (1978) suggested a classification of organic compounds as: 1. Water solubility less than 0.5 mg/L, likely to be environmentally hazardous. 2. Water solubility greater than 50 mg/L, likely to be environ-mentally nonhazardous. 3. Water solubility from 0.5 to 50 mg/L to be used, with caution. 22 DREDGING A N D D R E D G E D MATERIALS IN T H E LOWER MAINLAND Dredging in the Vancouver Region The vast majority of dredging activities in British Columbia's Lower Mainland occur in the Fraser River. Public Works Canada annually removes approximately two million tonnes of freshet sediment from the Fraser River navigation channels (H. Nelson, EPS, pers. comm., 1985). In addition, substantial quantities of sediment (the amount varies from less than I million to approximately 3 million tonnes) are annually dredged from the river by private commercial interests for industrial and urban developments. Dredging activities elsewhere in the Lower Mainland are more sporadic. Occasional dredging adjacent to harbour wharf areas in Burrard Inlet is required to maintain or improve berthing facilities. Isolated dredging activities have also taken place for foreshore developments, although these are often cut and fill operations that do not involve ocean dumping. The most recent requirement for major dredging outside of the Fraser River has been for the B.C. Place and Expo '86 developments in False Creek. Excavations for the entire False Creek rehabilitation have involved the dredging of approximately 2.26 million m3 (B.C. Place et al., 1983). 23 Much of the clean sediment dredged from Vancouver's waterways are in-river dumped or disposed at designated foreshore sites. For example, during 1985 the Department of Public Works plans to dredge 2.35 million m3 of sediments from the lower Fraser River. An estimated 850,000 m3 will be dredged using suction-pipeline tech-niques, of which 180,000 m3 will be disposed in-river and the remainder on the foreshore in the Duck-Woodward-Barber Island area and at Steveston Island (see Figure 2). A further 1.5 million m3 will be dredged by suction-hopper. Of this, some 700,000 m3 are planned for disposal at the Sandheads dumpsite at the mouth of the Main Channel: the other 800,000 m3 will be disposed in-river (H. Nelson, pers. comm., 1985). An average of approximately 1.5 million m3 of dredged materials have been annually disposed at nominated Dredge Material Disposal Sites adjacent to the Lower Mainland (Table II). This comprises a range of materials including sand, mud and silt, woodwastes, and rock, gravel and construction debris. The Sandheads and Fraser South Arm dumpsites primarily receive clean riverine sediments with particle size distribution or organic content which deem them unsuitable for use as aggregate or landfill. Dredged material from the Fraser River North Arm, Burrard Inlet and False creek, as well as woodwastes, con-struction debris and contaminated sediments dredged from throughout Vancouver's waterways, are predominantly disposed at the Point Grey dumpsite. The Point Grey dumpsite is four nautical miles west of Point Grey in 240 m of water. TABLE I I . Q u a n t i t y o f Dredged M a t e r i a l s Disposed at Dumpsites A d j a c e n t to the Lower M a i n l a n d , 1976-1979. (Compiled from Ward and S u l l i v a n , 1980). Dumpsite Volume Dumped 1976 1977 (thousands o f 1978 1979 m ) Average Predominant M a t e r i a l s Dumped P o i n t Grey 192 737 181 58 292 S i l t , s a n d , mud, r o c k , g r a v e l , c o n s t r u c t i o n d e b r i s and woodwastes. Sandheads 1 ,408 1 ,083 14 986 873 Sand and s i l t from maintenance d r e d g i n g o f the F r a s e r R i v e r Main Arm. B u r r a r d I n l e t 24 154 0 0 44 Mud, s i l t , sand and c o n s t r u c t i o n d e b r i s ( i n a c t i v e dumpsite s i n c e 1977). F r a s e r R i v e r - 169 503 256 50 244 S a n d , s i l t , mud and c l a y . South Arm T o t a l 1 ,793 2,477 451 1,094 1 ,454 25 This review is mainly concerned with dredged materials currently disposed at Point Grey, particularly the woodwastes and fine organic materials that accumulate adjacent to Vancouver's sawmills, and the organic "oozes" from quiscent areas of the Fraser River, Burrard Inlet and False Creek. It is normally the fine grained and organic materials, and not the coarse grained, that warrant environmental concern related to disposal (Boyd et al., 1972). These fine sediments include natural and anthropogenic deposits, and frequently contain high concentrations of organic and inorganic contaminants. Table III summarizes the elevated concentrations of various metals and toxic PCB compounds that have been found in surface sediments of False Creek and adjacent to Vancouver Wharves Ltd.'s terminal in Burrard Inlet. Sediment Contaminants in the Vancouver Region Data describing the degree of contamination of riverine and marine sediments in the Vancouver region mostly date from the 1970's and are very limited. A review of the available data describing the distribution of trace metals in the Fraser River estuary was compiled by Stancil (1980). Garret (1980) similarly reviewed the distribution through the lower Fraser of chlorinated pesticides, PCB's, hexa-chlorobenzene (HCB) and chlorinated phenols. Garrett (1983) more recently published an overview of PCB's in British Columbia, which presents additional data on the aquatic distribution of these com-pounds in the Vancouver region. Intensive sampling data describing 26 TABLE I I I . Contaminant C o n c e n t r a t i o n s i n F a l s e Creek and B u r r a r d I n l e t S e d i m e n t s . C o n c e n t r a t i o n i n S u r f a c e Sediments (mg.kg d r y wt) Background! , Vancouver Wharves Element Sediments F a l s e Creek T e r m i n a l , B u r r a r d I n l e t Copper 25 2 3 7 . 0 17 , 9 0 3 u Lead 6 5 6 2 . 0 141 d Cadmium 0 . 2 3 . 5 1 8 . 7 C Mercury 0 . 0 5 0 . 9 6 2 . 4 7 c Z i n c 80 8 0 8 . 0 l , 6 3 5 d N i c k e l 20 25 6 1 d A r s e n i c 6 11 .3 9 5 . 3 C Chromi urn 32 6 0 . 2 3 8 . 8 C PCBs 0 . 2 0 . 3 7 e 1 4 . 4 f a . For uncontaminated Puget Sound s e d i m e n t s , r e p o r t e d by Mai i n s e t a l . , 1980. b. From N i x and Chapman, 1984. V a l u e s shown a r e average f o r 11 s t a t i o n s . c . From Chapman and B a r l o w , 1984. V a l u e s shown a r e from a c o m p o s i t e o f 2 s a m p l e s . d . From Chapman, Munday and V i g e r s , 1982. V a l u e s shown are average f o r 13 s t a t i o n s . e . From B r o t h e r s and S u l l i v a n , 1984. V a l u e shown i s average f o r 13 s t a t i o n s i n F a l s e Creek E a s t B a s i n . f . From G a r r e t t , 1983. V a l u e shown i s f o r s i n g l e sample i n A u g u s t , 1980. P r e v i o u s sample i n J u l y , 1976 showed 0 . 1 8 m g . k g - ^ . -27 the distributions of polychlorinated phenols, PCB's and trace metals in False Creek sediments are summarized in Brothers and Sullivan (1984). Garrett, MacLeod and Sneddon's (1980) summary of data on mercury contamination in the British Columbian and Yukon environments provide additional limited data with respect to Burrard Inlet. Sediment quality data on the common parameters used in evaluating ocean dumping permits in Canada were, for a few years, compiled on computer file in the System 2000 Ocean Dumping Data Base by the Environmental Protection Service (EPS) of Environment Canada. The parameters listed for storage in the data base comprised a wide range of chemical and physical characteristics of dredged material. The data file thus provided a potential vehicle for the accumulation of extensive information on coastal sediment quality in Canada. However infrequent monitoring of these parameters within the ocean dumping permitting process has substantially degraded this opportunity. Only Hg, Cd, PCB's, COD, T O C , moisture and particle size have been routinely monitored in ocean dumping permit applications on the Pacif ic coast (McGreer and Konasewich, 1981). The System 2000 data base has not been maintained since 1983 (H. Nelson, EPS, pers. comm., I 985). Substantial data gaps, particularly with respect to a wide range of organic contaminants of concern, thus frustrate our ability to fully document the nature and distribution of contaminants in Vancouver's sediments. Most previous studies have been limited in their analyses to trace metals, chlorinated organopesticides, and PCB's, with 28 occasional monitoring of chlorinated phenols and benzenes. Data on these substances is sporadic, and virtually no data exists describing the local distribution of other persistent organic chemicals such as chlorinated butadienes, polybrominated biphenyls and polyaromatic hydrocarbons. The U.S. EPA Priority Pollutant list includes I 16 organic compounds: there are inadequate data describing the distri-bution of most of these compounds within the Vancouver region. A prohibitively expensive sedimentological survey would be required to fully document the nature and extent of contamination of Vancouver's sediments. However an integration of the available data with information describing past and present activities along Vancouver's waterfront can assist compilation of a list of "suspected" contaminants of local concern. For example, the lower Fraser River and estuary receives wastes from approximately five metal finishing and fabri-cating plants; thirteen pulp and paper and lumber mills, twenty wood preserving plants; six sewage treatment plants; five active municipal landfills and numerous smaller wood waste and industrial disposal sites; one hundred and eighty storm sewers; a coal bulk loading terminal; and unknown numbers of chemical plants, cement producers, and numerous other facilities (Garrett, 1982). Table IV examples the broad spectrum of sources and activities that may contribute nomi-nated trace metals to the environment. A comparison of the list in Table IV with the range of activities discharging into the river indicates a significant potential for trace metal contamination of the river sediments. 29 TABLE IV. P o t e n t i a l A n t h r o p o g e n i c Sources o f T r a c e Metal C o n t a m i n a t i o n o f V a n c o u v e r ' s Waterways ( c o m p i l e d from G a r r e t t , 1982; S w a i n e , 1980; H a l l , 1976; S t a n c i l , 1980; and G a r r e t t , MacLeod and Sneddon, 1980) Metal P o t e n t i a l Sources Mercury M a n u f a c t u r e o f e l e c t r i c a l a p p a r a t u s , c h l o r i n e , c a u s t i c s o d a , p a i n t , d r y w a l l s , p h a r m a c e u t i c a l s , c h e m i c a l s , p u l p and p a p e r , f u n g i c i d e s , amalgams and cement. A l s o from l e a d and z i n c o r e s . Cadmium M e t a l l u r g y ( a l l o y s , s o l d e r s ) , e l e c t r o p l a t i n g and metal f i n i s h i n g , p a i n t and v a r n i s h , b a t t e r y m a n u f a c t u r e , f u n g i c i d e s , f e r t i l i z e r s , rubber t i r e s , motor o i l s , z i n c , copper and l e a d o r e s . Lead P r i n c i p a l l y from motor v e h i c l e s . A l s o from base metal o r e s and metal f o u n d r i e s , p a i n t s and t h e d i s p o s a l o f b a t t e r i e s . Chromium P r i n c i p a l l y e l e c t r o p l a t i n g and metal f i n i s h i n g . A l s o from p a i n t s , p h o t o g r a p h i c p r o c e s s i n g , the c o n c r e t e i n d u s t r y and wood p r e s e r v a t i v e s Copper Metal f i n i s h i n g , a n t i - f o u l i n g p a i n t s , a l g i c i d e s , p e s t i c i d e s , base metal o r e s , c o r r o s i o n o f household p l u m b i n g , wood p r e s e r v a t i v e s . Z i n c E l e c t r o p l a t i n g and metal f i n i s h i n g , c o r r o s i o n o f h o u s e h o l d p l u m b i n g , c o n c r e t e p r o d u c t i o n , wood p r e s e r v a t i v e s . A r s e n i c L e a d , z i n c and copper o r e s , i n s e c t i c i d e s , h e r b i c i d e s , wood p r e s e r v a t i v e s , c e r a m i c and g l a s s m a n u f a c t u r e , m a n u f a c t u r e o f s e m i c o n d u c t o r s . N i c k e l E l e c t r o p l a t i n g and metal f i n i s h i n g , base metal o r e s . S i l v e r P h o t o g r a p h i c m a t e r i a l s , e l e c t r o p l a t i n g , e l e c t r i c a l equipment d e n t a l a l l o y s , p a i n t s , g l a s s m a n u f a c t u r e . S e l e n i u m C o p p e r , l e a d and z i n c o r e s , p h o s p h a t e s , m a n u f a c t u r e o f g l a s s and c e r a m i c s , s e m i c o n d u c t o r s , metal f i n i s h i n g , p a i n t s , p h o t o c o p y i n g , p h o t o c e l1s . 30 Although the existence of particular urban or industrial activities adjacent to a waterway does not automatically indict those activities for contaminating the sediments, it does provide, through reference to documented cases of pollution elsewhere, presumptive evidence suggesting the existence of nominated contaminants. In the face of a very limited data base, ocean dumping management must often presume the presence of suspected contaminants until contrary evidence is provided. This presumption implies a requirement to account for the identified contaminants within the management scheme. Many of the compounds that are presumed to enter our aquatic system do not constitute an environmental concern to ocean dumping management as they are relatively non-toxic, are not likely to associate with the sediments in amounts great enough to cause harm, or are rapidly degraded by natural mechanisms. However trace metals and many man-made organic compounds that may enter the environment in significant amounts associate predominantly with sediments, are extremely stable and exhibit a high toxicity to various organisms. Numerous documents exist which purport to nominate the environ-mental contaminants that warrant priority concern by regulatory authorities. The Departments of Environment and National Health and Welfare (1979) have identified an assorted list of priority contaminants requiring further study under the Environmental Contaminants Act (Table V). The U.S. EPA (1977) List of Priority Pollutants includes 31 * TABLE V. Canadian E n v i r o n m e n t a l Contaminants A c t L i s t o f P r i o r i t y C h e m i c a l s . C a t e g o r y I : Those s u b s t a n c e s which t h e government i s s a t i s f i e d pose a s i g n i f i c a n t danger to human h e a l t h o r t h e environment and f o r which r e g u l a t i o n s a r e b e i n g d e v e l o p e d . P o l y c h l o r i n a t e d B i p h e n y l s (PCBs) C h l o r o f l u o r o m e t h a n e s C a t e g o r y I I : Those s u b s t a n c e s which a r e b e i n g i n v e s t i g a t e d to d e t e r m i n e the n a t u r e and the e x t e n t o f the danger t o human h e a l t h o r the environment and the a p p r o p r i a t e means t o a l l e v i a t e t h a t d a n g e r . Cadmium C h l o r o p h e n o l s Mercury C a t e g o r y I I I : Those s u b s t a n c e s w h i c h may pose a s i g n i f i c a n t danger t o human h e a l t h or the e n v i r o n m e n t and about w h i c h f u r t h e r d e t a i l e d i n f o r m a t i o n ( f o r example t o x i c o l o g y and amounts used) i s r e q u i r e d . C h l o r o b e n z e n e s H e x a c h l o r o b u t a d i e n e H e x a c h l o r o c y c l o p e n t a d i e n e and i t s Adducts O r g a n o t i ns T r i a r y l phosphates and r e l a t e d s u b s t a n c e s Compounds f o r which r e g u l a t i o n s a r e a l r e a d y i n f o r c e under t h e A c t . P o l y c h l o r i n a t e d B i p h e n y l s (PCBs) Mi r e x P o l y b r o m i n a t e d B i p h e n y l s P o l y c h l o r i n a t e d T e r p h e n y l s Canada G a z e t t e , 1979 ( l a ) 32 129 substances (Table VI). On a more regional scale, Garrett (1982) priorized industrial chemicals that warrant concern in British Columbia and the Yukon (Table VII). However, the environmental concerns supporting the inclusion of substances in these lists are multifarious, including air, water, land, biota and man. Environmental management of ocean dumping is essentially concerned with substances from these lists which meet the following criteria: o they are known to be toxic and/or mutagenic to marine biota, or to bioaccumulate in marine biota to an extent that they pose a health concern to subsequent consumers (including humans). o they tend to associate with the sediments where they are not readily degraded under normal environmental conditions o they have documented or potential distribution in the sediments of the Lower Mainland. For the purposes of this study, the definition of "toxic substance" includes all substances which can cause death, disease, behavioural abnormalities, cancer, genetic mutations, physiological or reproductive malfunctions, or physical deformities in any organism or its offspring, or which can become poisonous after its concentration in the food chain or in combination with other substances. This definition was jointly derived by the Governments of Canada and the United States for the 1978 Great Lakes Water Quality Agreement (IJC, 1978). 33 T A B L E VI- U . S . E . P . A . P r i o r i t y P o l l u t a n t L i s t Showing C l a s s i f i c a t i o n by E x t r a c t i o n Group. PURGEABLE GROUP PESTICIDE GROUP ACID GROUP B e n z e n e B r o m o d i c h l o r o m e t h a n e B r o m o f o r m Bromomethane C a r b o n T e t r a c h l o r i d e C h l o r o b e n z e n e C h l o r o e t h a n e 2 - C h l o r o e t h y l v i n y l e t h e r C h l o r o f o r m C h l o r o m e t h a n e b i s - C h l o r o m e t h y l e t h e r Di b r o m o c h l o r o m e t h a n e D i c h l o r o d i f l u o r o m e t h a n e 1 , 1 , - D i c h l o r o e t h a n e 1 , 2 , - D i c h l o r o e t h a n e 1 . 1 , - D i c h l o r o e t h y l ene t r a n s - 1 , 2 , - D i c h l o r o e t h y l ene 1 . 2 , - D i c h l o r o p r o p a n e 1 , 2 , - D i c h l o r o p r o p y l e n e E t h y l b e n z e n e M e t h y l e n e C h l o r i d e 1 , 1 , 2 , 2 , - T e t r a c h l o r o e t h a n e 1 , 1 , 2 , 2 , - T e t r a c h l o r o e t h e n e T o l u e n e 1 , 1 , 1 , - T r i c h l o r o e t h a n e 1 , 1 , 2 , - T r i c h l o r o e t h a n e T r i c h l o r o e t h y l ene T r i c h l o r o f l u o r o m e t h a n e V i n y l C h l o r i d e A l d r i n a l p h a - B H C b e t a - B H C gamma-BHC ( L i n d a n e ) d e l t a - B H C C h l o r d a n e 4 , 4 1 - O D D 4 , 4 * - D D T a l p h a - E n d o s u l f a n b e t a - E n d o s u l f a n E n d o s u l f a n s u l p h a t e E n d r i n E n d r i n A l d e h y d e H e p t a c h l o r H e p t a c h l o r E p o x i d e PCB-1016 PCB-1221 P C B - 1 2 3 2 P C B - 1 2 4 2 P C B - 1 2 4 8 P C B - 1 2 5 4 P C B - 1 2 6 0 T o x a p h e n e p - C h l o r o - m - c r e s o l 2 - C h l o r o p h e n o l 2 , 4 , - D i c h l o r o p h e n o l 2 , 4 , - D i m e t h y l p h e n o l 4 , 6 , - D i n i t r o - o - C r e s o l 2 , 4 , - D i n i t r o p h e n o l 2 - N i t r o p h e n o l 4 - N i t r o p h e n o l P e n t a c h l o r o p h e n o l P h e n o l 2 , 3 , 6 , - T r i c h l o r o p h e n o l BASE NEUTRAL GROUP MISCELLANEOUS GROUP P o l y n u c l e a r A r o m a t i c H y d r o c a r b o n s A c e n a p h t h e n e A c e n a p h t h y l e n e A n t h r a c e n e B e n z o ( a ) a n t h r a c e n e B e n z o ( b ) f l u o r a n t h e n e B e n z o ( k ) f l u o r a n t h e n e B e n z o ( g h i ) p e r y l e n e B e n z o ( a ) p y r e n e C h r y s e n e O i b e n z o ( a h ) a n t h r a c e n e F l u o r a n t h e n e F l u o r e n e I n d e n o ( 1 , 1 2 , - c d ) p y r e n e N a p h t h a l e n e P h e n a n t h r e n e P y r e n e C h l o r i n a t e d B e n z e n e s 1 . 2 - 0 i c h l o r o b e n z e n e 1 . 3 - D i c h l o r o b e n z e n e 1 . 4 - D i c h l o r o b e n z e n e 1 , 2 , 4 - T r i c h l o r o b e n z e n e H e x a c h l o r o b e n z e n e N i t r o s a m i n e s N - n i t r o s o d i m e t h y l amine N - n i t r o s o d i p h e n y l a m i n e N - n i t r o s o d i - n - p r o p y l a m i ne P h t h a l a t e E s t e r s B u t y l B e n z y l p h t h a l a t e D i - n - b u t y l p h t h a l a t e D i e t h y l p h t h a l a t e D i m e t h y l p h t h a l a t e D i - n - o c t y l p h t h a l a t e b i s - ( 2 - e t h y l h e x y l ) p h t h a l a t e H a l o e t h e r s 4 - B r o m o p h e n y l p h e n y l e t h e r b i s ( 2 - C h l o r o e t h o x y ) methane b i s ( 2 - C h l o r o e t h y l ) e t h e r 4 - C h l o r o p h e n y l p h e n y l e t h e r O t h e r compounds B e n z i d i n e 2 - C h l o r o n a p h t h a l e n e 3 , 3 ' - D i c h l o r o b e n z i d i n e 2 , 4 , - D i n i t r o t o l u e n e 2 , 6 , - D i n i t r o t o l u e n e 1 , 2 , - D i p h e n y l h y d r a z i ne H e x a c h l o r o b u t a d i e n e H e x a c h l o r o c y c l o p e n t a d i e n e H e x a c h l o r o e t h a n e I s o p h o r o n e N i t r o b e n z e n e A c r o l e i n A c r y l o n i t r i l e 2 , 3 , 7 , 8 , - t e t r a c h l o r o -d i b e n z o - p - d i o x i n A s b e s t o s C y a n i d e T o t a l P h e n o l s A n t i m o n y A r s e n i c B e r y l 1ium Cadmi urn Chromium C o p p e r L e a d M e r c u r y N i c k e l S e l e n i u m S i l v e r T h a l l i u m Z i n c 34 TABLE V I I . P r i o r i t y R a t i n g s f o r I n d u s t r i a l C h e m i c a l s i n B r i t i s h Columbia and the Yukon. (From G a r r e t t , 1982). Chemical T o x i c i t y , P e r s i s t e n c e Q u a n t i t y Used and P o t e n t i a l & B i o a c c u m u l a t i o n f o r R e l e a s e C h l o r i n a t e d Phenols Other Wood P r e s e r v a t i v e s ( c r e o s o t e , a r s e n i c compounds) Heavy M e t a l s C h i o r i ne Cyanide PCBs P l a s t i c s A d d i t i v e s ( t r i a r y l p h o s p h a t e s , o r g a n o t i n s , p h t h a l a t e e s t e r s , c h l o r i n a t e d p a r a f f i n s ) C h l o r i n a t e d Hydrocarbons ( s o l v e n t s ) C h l o r i n a t e d Benzenes PAHs C h l o r i n a t e d D i b e n z o d i o x i n s and D i b e n z o f u r a n s F l o u r i d e s 4 3 3 4 3-4 4 2-3 3 3 4 2 4 4 4 3 3 2 - 2 . 5 2 - 3 2 2 R e l a t i v e  R e g i o n a l Concern 16 12 12 12 9-12 8-10 6-9 6-9 6 6 ? f a . low = 1 ; moderate = 2 ; h i g h = 3 ; v e r y h i g h = 4 b. low = 1 t o 4 ; moderate = 5 t o 8 ; h i g h = 9 to 1 2 ; v e r y h i g h = 13 t o 16 35 Uncertainty and controversy have surrounded the application of this definition to quantify concentrations at which sedimentary contami-nants are considered toxic. The problem is that the science of measuring, much less predicting, the effects of contaminants in the marine environment is still in its infancy. Each of the compounds and elements listed in Tables V, VI and VII are known to induce toxic responses at some level (hence their inclusion in the lists), but levels at which they may be considered to be environmentally "safe" continue to defy certain definition. For the purposes of this study therefore, all of the organic contaminants listed within the Tables are considered to be toxic at current levels of detection. Trace metals are considered to be toxic if they are in excess of average concen-trations in deep sea sediments, as reported by Bryan (1976). This approach refrains from the controversial evaluation of environmental risk i.e. whether "no effect" levels exist or whether carcinogenic risk can be extrapolated to all levels of exposure (Maugh, 1978). Although all of the priority pollutants listed in Tables V, VI and VII thus fall within the criterion of toxicity, many fall outside the criterion of persistent association within the sediments. For example chloroflouromethanes and other purgeable chlorinated hydrocarbons introduced to agueous systems will mostly volatilize to the atmosphere (Callahan and Slimak, 1979). The major fate process of other substances in the aquatic system (e.g. nitrosamines, cyanide, fluorine) is dispersion in solution rather than adsorption to the sediments. 36 Hexachlorocyclopentadiene, triaryl phosphates, chlorinated paraffins and chlorine tend to associate with sediment material, but are not environmentally persistent (Callahan and Slimak, 1979). Contaminants that do meet the first two criteria justifying priority concern in the management of dredged materials from the Vancouver region are listed in Table VIII. Summary information supporting their inclusion in the list and describing our knowledge of their distribution in the sediments of the Lower Mainland follows. 3.2.1 Trace Metals Trace metal discharges to the Fraser River occur throughout the lower reaches below the Pattullo Bridge. Ferguson and Hall (1979), Swaine (1980) and Stancil (1980) describe the sources and distribution of metal loadings to the river, particularly for lead, copper, nickel and zinc. The distribution of trace metals in sediments is determined by input location and hydrodynamic relocation. Approximately 60 percent of the lead and nickel loading to the Fraser River comes from storm-water discharges, and is dispersed over a wide area from more than 180 outfalls (Hall, 1976). In contrast, the lona Sewage Treatment Plant discharge clearly effects trace metal levels in sediments of Sturgeon Bank (Table IX, c f . Roberts Bank). 37 TABLE V I I I . Contaminants o f Concern I d e n t i f i e d o r Suspected w i t h i n Sediments o f G r e a t e r V a n c o u v e r . T r a c e M e t a l s Mercury Cadmium Lead Chromium Copper Z i n c A r s e n i c N i c k e l S i 1ver S e l e n i urn Halogenated O r g a n i c s O r g a n o c h l o r i n e P e s t i c i d e s P o l y c h l o r i n a t e d B i p h e n y l s C h l o r i n a t e d B u t a d i e n e s P o l y c h l o r i n a t e d D i b e n z o f u r a n s P o l y c h l o r i n a t e d Benzenes P o l y c h l o r i n a t e d Phenols H a l o g e n a t e d P o l y a r o m a t i c Hydrocarbons N o n - h a l o g e n a t e d O r g a n i c s P o l y a r o m a t i c Hydrocarbons P h t h a l a t e E s t e r s O r g a n o t i ns 38 TABLE IX. T r a c e M e t a l s i n Sediments o f t h e F r a s e r R i v e r (From H a l l , 1976). ( C o n c e n t r a t i o n s , i n mg/kg, d r y w t . ) Area Copper Z i n c Lead North Arm 34 72 16 S t u r g e o n Bank 183 170 157 Roberts Bank 28 61 4 Main Arm 17-36 25-60 4-11 Ladner Slough 43-57 95-115 14-23 39 Hydrodynamic considerations result in a tendancy for trace metals to especially accumulate in backwater areas, which are sinks for the fine sediments to which metals preferentially sorb (Boyd et al., 1972). For example, sediment metal loadings are higher in Ladner Slough than at adjacent sites in the main channel (Table IX). Table X examples data compiled by Stancil (1980) describing trace metal concentrations in Fraser River sediments, with comparative figures from a relatively unpolluted area of the Columbia River, B.C. For each of the trace metals except mercury, sediment concentrations in one or more reaches of the river are substantially above natural levels. Of particular interest is the very wide variability shown in the data. For example, although mean copper and nickel concentra-tions did not vary substantially among most stations, the sediments from the Ladner Side Channel showed high variability in each of these metals: some sediment samples contained up to four times the river's average concentration. Select samples from New Westminster similarly showed extremely elevated concentrations of lead, zinc and cadmium. This variability demonstrates the difficulty of broad-scale sediment characterizations. The site specificity of trace metal concentrations requires appropriate site specificity in sediment characterization analyses during ocean dumping assessment. Trace metal levels in surface sediments of False Creek and at Vancouver Wharves terminal in Burrard Inlet have been previously documented in Table III. Both areas showed extremely elevated concentrations of copper, lead, cadmium and zinc. Concentrations of TABLE X. L e v e l s o f M e t a l s i n Sediments from the F r a s e r R i v e r and the Columbia R i v e r , B . C . (from S t a n c i l , 1980). Metal C o n c e n t r a t i o n s i n mg/kg ( d r y w t ) . Hg Cd Pb Cu Zn Ni Mean Mean Mean Mean Mean Mean Range Range Range Range Range Range FRASER RIVER 1. MAINSTEM B a r n s t o n I s l a n d 0.07 1.4 12.2 2 6 . 4 5 4 . 6 4 1 . 8 0 . 0 5 - 0 . 1 1 < 1 . 0 - 3 . 0 6 . 0 - 2 0 . 0 2 4 . 0 - 2 8 . 0 4 8 . 0 - 6 2 . 0 3 5 . 0 - 5 0 . 0 New W e s t m i n s t e r 0.10 <2.2 142.0 3 0 . 6 211.0 3 0 . 9 0 . 0 5 - 0 . 3 6 < 1 . 0 - 1 9 . 0 1 8 . 0 - 3 7 0 2 1 . 0 - 4 2 . 0 5 5 . 0 - 4 7 0 2 3 . 0 - 3 9 . 0 2 . MAIN ARM D/S A n n a c i s I s l a n d 0 . 0 8 <1.5 11.5 2 8 . 8 6 8 . 8 4 5 . 0 0 . 0 6 - 0 . 1 0 < 1 . 0 - 3 . 0 9 . 0 - 2 0 . 0 2 3 . 0 - 3 2 . 0 5 8 . 0 - 9 5 . 0 3 9 . 0 - 6 0 . 0 Ladner S i d e Channel - <0.3 2 5 . 7 6 3 . 0 115.0 7 9 . 3 ND - 0 . 6 1 6 . 0 - 5 1 . 0 4 1 . 0 - 1 1 8 8 3 . 0 - 2 0 0 6 3 . 0 - 1 7 0 3 . NORTH ARM M i t c h e l l I s l a n d 0 . 0 9 1.7 3 9 . 0 2 9 . 0 117.0 4 0 . 0 0 . 0 6 - 0 . 1 2 1 . 0 - 3 . 0 3 2 . 0 - 4 5 . 0 2 5 . 0 - 3 3 . 0 100-136 3 1 . 0 - 5 0 . 0 Oak S t . B r i d g e 0 . 0 6 <1.6 4 4 . 1 3 2 . 8 114.0 3 2 . 8 0 . 0 4 - 0 . 0 9 <1.0-<2.0 2 7 . 0 - 7 6 . 0 2 7 . 0 - 4 3 . 0 8 0 . 0 - 1 7 0 2 6 . 0 - 4 6 . 0 COLUMBIA RIVER U/S from T r a i l S m e l t e r 0 . 3 8 <0.5 1 9 . 0 2 9 . 0 4 4 . 3 41 arsenic, chromium and mercury were also substantially above natural levels in False Creek, as were arsenic, nickel and mercury in sediments adjacent to Vancouver Wharve's Bulk Terminal. The False Creek data described in Table III was determined prior to the major dredging program recently conducted for the B.C. Place and Expo '86 projects. Up to 325,000 m3 of contaminated surface sediments dredged for B.C. Place Corporation have now been removed from False Creek and disposed in oceanic water to the west of Vancouver Island. However, approximately 25,000 m3 of contaminated surface sediments dredged from the south-eastern area of False Creek for Expo '86 were disposed within the central area of the False Creek East Basin (Nix and Chapman, 1984). Also, not all areas of False Creek have been dredged. The dredging program has certainly reduced the contaminant loading of the waterway from the very high levels described in Table III, but localized areas will still exist with high trace metals. There is a paucity of data describing trace metal levels in the dredgable areas of Burrard Inlet. The extreme pollution of sediments adjacent to Vancouver Wharves Bulk Terminal (Table III) has resulted from spillage and washdown during ore (copper, lead, nickel, zinc) loading operations (Chapman, Munday and Vigers, 1982). Chapman and Barlow (1984) reported similarly elevated levels of cadmium (LI mg/kg dry wt), chromium (29.8 mg/kg), copper (199 mg/kg), nickel 238 mg/kg), lead (59 mg/kg), and zinc (145 mg/kg) in sediments of Neptune Terminals in Burrard Inlet. It is considered likely that each 42 of the bulk loading terminals surrounding Burrard Inlet, and also the major shipyards (Burrard Yarrows, Vancouver Shipyards), would show high contaminant loadings were the sediments analysed. Areas adjacent to boat marinas and to sewage and stormwater outfalls also warrant monitoring. In conclusion, existing data has documented areas within Vancouver's waterways with substantially elevated sediment concentrations of Hg, Cd, Pb, Cu, Zn, As and Ni. The failure to document pollution by Cr, Ag and Se is more likely due to the paucity of data rather than their absence from the sediments. Given their high toxicity and their likely occurrence in our sediments, it is recommended that trace metal analyses of contaminated sediments should also routinely include these metals. Concern for their presence in the sediments can only be confirmed or denied by routine measurement. 3.2.2 Organic Contaminants Existing data relating to organic contaminant levels in the sediments of Vancouver's waterways are confined primarily to organochlorine pesticides and PCB's, with occasional data for chlorinated phenols, chlorinated benzenes, polyaromatic hydrocarbons and phthalate esters. Other persistent and toxic organic contaminants that are identified in Table VIII are the chlorinated butadienes and dibenzofurans, and organotins. 43 Orqanochlorine Pesticides The common representative of the organochlorine pesticides is DDT, but this pesticide was never used extensively in British Columbia either in forestry or agriculture (Waldichuk, 1983). Hall and Weins (1976) analysed 37 sediment samples from the agricultural Salmon and Sumas River basins, but only one contained detectable DDT. Sediment samples from the lower Fraser River that were collected by the Inland Waters Directorate (IWD) in 1977 were analysed for a range of organochlorine pesticides including DDT and its metabolites, aldrin, dieldrum, endrin, lindane, heptachlor and its epoxide, methoxychlor, alpha- and beta-endosulphan and alpha- and gamma-chlordane. These compunds were not present in measureable amounts in any sample (Garrett, 1980). Only two reports of organochlorine pesticide contamination of sediments in the Lower Mainland were found in the literature. Hall, Yesaki and Chan (1976) tested sediments of the urban/industrial Brunette River basin for the presence of organochlorine pesticides. The range of analyses was the same as was tested in the IWD study. Of the 30 sediment samples analysed, 12 of them contained measureable guantities of DDT and its metabolites. Alpha- and gamma-chlordanes were found at four stations i n , a reach of Still Creek. Total DDT levels did not exceed 0.19 mg/kg, while total chlordane levels were below 0.09 mg/kg. It was speculated that these low levels were the result of recent illegal usage rather than historic residuals (Hall and Weins, 1976). No other pesticides were detected. 44 Garrett (1980) reported data collected by the EPS and Pesticide Control Branch from soils and sediment of the abandoned site of Later Chemicals manufacturing plant in Richmond. Particularly high levels of several organochlorine pesticides were detected with sediments in the drainage ditch containing 3200 mg/kg total DDT, 1020 mg/kg total chlordane, 120 mg/kg aldrin, 110 mg/kg heptachlor, 15 mg/kg methoxychlor, 21 mg/kg dieldrin and 13 mg/kg lindane. Despite the local concern that is warranted by this latter case, it is concluded that organochlorine pesticides are not a significant concern to the management of dredged sediments in the Lower Mainland. DDT was never extensively used in British Columbia, and its use was banned in Canada in the early 1970's. Other organochlorine pesticides are more than an order of magnitude more soluble and less persistent in the environment. Most pesticides in current use are organo-phosphates, which have very low association tendencies with sediments. Polychlorinated Biphenyls (PCBs) Polychlorinated biphenyls have been detected in low but measurable guantities in the sediments of almost all industrially active areas of the Fraser River downstream from the Port Mann Bridge, and in higher concentrations in False Creek and Burrard Inlet (Garrett, 1983). Although the manufacture of PCBs was terminated in North America 45 in the early 1970's and the allowable uses have been severely restricted, PCBs remain a ubiquitous pollutant deserving substantial environmental concern. Garrett (1983) conducted a comprehensive review of the potential sources and distribution of PCBs in the British Columbia environment. High sediment concentrations, often exceeding 1.0 mg/kg dry weight, were reported off certain industrial facilities in the Lower Mainland. These include levels of up to 4.46 mg/kg in the east basin of False Creek and up to 6.9 mg/kg under the roller bearings of the Granville Street Bridge. Concentrations of up to 16.8 mg/kg were reported in sediments of Coal Harbour. Very high levels were also found in the vicinity of the various shipyards and wharves of Burrard Inlet. The sediments of the Fraser River generally have low levels of PCBs, except in localized areas adjacent to paper and paperboard mills and certain of Delta's land drainage pump stations (Garrett, 1983). High levels of PCBs in these sediments are probably a reflection of local contamination. Concentrations away from point sources are generally low. The extreme toxicity and persistence (up to 30 years; Horn et al., 1974) of PCBs in the environment, together with .the high concen-trations found in the sediments of areas of Vancouver's waterways, justify very high concern for these compounds within the management of dredged materials from the Lower Mainland. 46 Polychlorinated terphenyls (PCTs) and polybrominated biphenyls (PBBs) are related compounds which are also listed as priority pollutants under the Canadian Environmental Contaminants Act . Each of these closely resemble PCBs in chemical structure and their environmental effects may be similar. However no information was found describing their usage or possible distribution in the Lower Mainland. The use, manufacture and importation of PBBs and PCTs has been banned in Canada under the Environmental Contaminants Act (1979). However PBBs were previously widely used as flame retarders (Jamieson, 1977a), while PCTs served a variety of uses, including plasticizers, adhesives, lubricants, caulking compounds, in printing inks and as fire retarders (Jamieson, 1977b). There is conseguently a need to determine the levels and extent of distribution of these compounds within Vancouver sediments. Chlorinated Phenols Chlorinated phenols are used extensively in the Lower Mainland, primarily as wood preservatives and as slimicides in paper production. Other uses include germicides, mildewcides, insecticides and herbi-cides. The most environmentally persistent chlorophenol is penta-chlorophenol (PCP). Both PCP and tetrachlorophenol (TCP) are used in wood preserving operations at numerous foreshore sites in the Vancouver region. There are more than twenty mills using chlorinated phenols for wood protection located along the foreshore of the lower Fraser River alone (Garrett, 1982). 47 Maynard and Vigers (1979) analysed sediments from adjacent to a number of chlorophenol users in coastal areas of B.C., including three wood preserving operations in the Fraser River. Sediment samples from all sites contained detectable concentrations of tetra- and pentachlorophenol. Maximum sediment levels of up to 70 ug/kg dry weight of P C P and up to 60 ug/kg T C P were found adjacent to the Crown Zellerbach Mill. Garrett (1980) reported data from Waste Management Branch moni-toring of sediments off Koppers International pole treating facility that showed high levels of PCP (525 mg/kg) and T C P (90 mg/kg). Garrett (1980) also described the results of tissue analyses of fish from the Fraser River, which found that chlorinated phenol compounds were identified with greater freguency and at higher concentrations in fish from the industrial areas of the lower Fraser. PCP and T C P concentrations did not exceed 125 ug/kg and 62.0 ug/kg wet weight, respectively. The high usage of pentachlorophenol by foreshore industries in the Lower Mainland justifies continued concern for sediment contamina-tion. Relative to the other halogenated organic contaminants of concern to ocean dumping management, PCP (and especially TCP) is less persistent in the sediments. However the likelihood that there are regular and substantial discharges or spillages of this compound to the coastal environment requires routine monitoring of PCP levels in sediments proposed for ocean dumping. 48 Chlorinated Benzenes Chlorinated benzenes are listed in the Canadian Environmental Contaminants Act list of priority chemicals. Hexachlorobenzene (HCB) is the most persistent compound of this class, and is the only chlorobenzene that has been found in significant concentrations in sediments of the Lower Mainland (Garrett, 1980). HCB is among the U.S. EPA's 129 Priority Pollutants (U.S. EPA, 1977). HCB was used extensively in agriculture for prevention of fungus disease until restrictions were placed on such use in 1971 (Garrett, 1980). Current sources of release of HCB to Vancouver's aguatic environment include tire-rubber residues within storm water run-off and industrial discharges, e.g., from producers of chlorinated solvents (Safe et al., 1979). HCB may also be a component of waste dis-charges from electrolytic chlorine production facilities (Quinlivan, Ghassemi and Teshendok, 1977), and may be formed by the chlori-nation of sewage effluents (Kopperman, Juehl and Glass, 1976). There is virtually no data describing HCB levels in Vancouver sediments. During their survey of contamination from the wood preserving industry, Maynard and Vigers (1979) found trace levels of HCB in sediments of the Fraser River adjacent to.Domtar Ltd. and Later Chemicals Ltd., and concentrations of up to 1.9 ug/kg (dry weight) in the vicinity of MacMillan Bloedel White Pine Division. Johnson et al. (1975) found HCB in tissues of a wide range of fishes 49 collected from throughout the Fraser River downstream from Hope. Concentrations were low however (mean = 10 ug/kg wet wt.), and there was no apparent variation with location. No other information was found describing HCB levels in sediments of the Fraser River, and there was a total absence of data pertaining to HCB levels in False Creek and Burrard Inlet. Given the lack of evidence to the contrary, and with regard to the numerous potential sources of HCB contamination along Vancouver's shoreline, this compound must be assumed to be a contaminant of concern to ocean dumping management. Polyaromatic Hydrocarbons (PAHs) Reference to PAHs includes a large class of compounds which consist of two or more fused benzene rings in linear, angular or cluster arrangements. Other terms commonly used for these compounds are polycyclic hydrocarbons and polynuclear aromatic hydrocarbons. PAHs are ubiquitous in nature, but high anthropogenic inputs to coastal environments warrant appreciable concern. Chlorinated PAHs, which are widely used industrial compounds with physical properties similar to PCBs, are not naturally occurring compounds and are particularly persistent in the environment. PAHs are priority chemicals under the Canadian Environmental Contaminants Act . Of the U.S. EPA's 129 Priority Pollutants, 16 are PAHs. 50 There are few data pertaining to P A H concentrations in Vancouver's sediments. Hall, Parkinson and Ma (1983) reported high total P A H concentrations among sediment samples analysed from unidentified locations in False Creek and Burrard Inlet. Total P A H levels in nine samples from False Creek ranged from 3.86 mg/kg (dry weight) to 71.69 mg/kg, with a mean of 21.36 mg/kg. In the nine sediment samples analysed from Burrard Inlet, concentrations ranged from 1.62 mg/kg to 1551.99 mg/kg, with a mean value of 314.96 mg/kg. Individual PAHs identified in particularly high concentrations included naphthalene, acenaphthene, fluorene, phenanthracene, anthracene, fluoranthene, pyrene and benzo(a)anthracene. Lower concentrations of PAHs have been reported in Fraser River sediments. Dunn and Stich (1976) reported that benzo(a)pyrene concentrations on Sturgeon Bank increased with proximity to the lona Island sewage treatment plant outfall. Concentration maxima were 120 ug/kg (dry wt.). Garrett (1980) reported Department of Fisheries and Oceans data showing naphthalene concentrations of up to 22 ug/kg in the same area. Low concentrations of benzo(a)pyrene have also been found in sediments adjacent to creosoted pilings in the Fraser Estuary (Garrett, 1980). The largest source of PAHs to the environment is the combustion of coal, wood and oils. However point source contamination is generally responsible for more substantially elevated sediment concentrations within localized areas, so is of more interest to ocean dumping management. 51 Contamination of the aquatic environment by crude and refined oils can be a significant source of PAHs, particularly naphthalenes (Lee, 1977). Other significant sources of naphthalenes include plastic and resin manufacturers, timber products processors, paint and ink formulators and iron foundries. A U.S. EPA survey found naphthalene levels in timber products processing wastewaters of up to 45 mg/L (U.S. EPA, 1978). The U.S. EPA (1980) Treatability Manual lists timber products pro-cessing and automobile and other laundries as major sources of benzo(a)pyrene. Man-made sources of flouranthenes include shipping and harbour oil discharges and industrial effluents from oil refineries, timber products processors, coke production industries, plastic and dyestuff industries and industries using high temperature furnaces. Borneff and Kunte (1965) indicated that road runoff can also be a source of flouranthenes as a result of bituminous road surfaces, car tire wear and vehicle exhausts. Matins et al. (1980) reported 26 halogenated PAHs in sediment samples from Puget Sound. Although there are very limited data on sources of halogenated PAHs to the aquatic environment, possible sources appear to be discharges from motor vehicle washing facilities (Halowaxes), leaching from landfill sites, discarded capacitors and cable insulation. Historical usage of chlorinated PAHs included rubber substitutes, cutting oil additives and fabric dyes. 52 Polyaromatic hydrocarbons have been found in very high concentra-tions in sediments of urban and industrialized embayments in Puget Sound, where they have been linked to tissue abnormalities in resident fish and crabs (Malins et al., 1980). Until evidence to the contrary can be established for sediments from the Vancouver region, PAHs must be assumed to be contaminants of significant concern to ocean dumping management in Canada. Phthalate Acid Esters (PAEs) Phthalate esters have a variety of uses ranging from antifoam agents in the paper industry to perfume vehicles in cosmetic production. However they are mainly used as plasticizers in the production of polyvinyl chloride. Within some plastic formulations, PAEs can comprise up to 60% of the total weight of plastic, depending on the degree of plasticity required. The market for PAEs is therefore extremely diversified, including use in clothing, home furnishings and packaging industries. Contamination of the environment by PAEs can occur through discharge of effluents from such industries, or by leaching of P A E containing substances from land disposal sites (Johnson et al., 1977). As a result, there are many potential sources of PAEs in the Vancouver region. Phthalate acid esters are priority chemicals under the Canadian Environmental Contaminants Act and the U.S. EPA's Priority Pollutants List. Of the wide variety of PAE's manufactured, di-2-53 ethylhexylphthalate (DEHP), diethyl phthalate (DEP), dibutylphthalate (DBP), butyloctylphthalate (BOP) and di-isononylphthalate appear to be of the most significant environmental concern (Garrett, 1980). Limited data were found from the literature review describing PAE levels in Vancouver sediments. Hall, Parkinson and Ma (1983) analysed for PAEs in sediment samples from unidentified locations in False Creek and Burrard Inlet. Total phthalate concentrations ranged from below detection up to 33.10 mg/kg (dry weight) in False Creek and up to 20.96 mg/kg in Burrard Inlet. Individual PAEs found in the highest concentrations were DEHP (up to 25.2 mg/kg in False Creek and 7.82 mg/kg in Burrard Inlet) and DBP (7.9 mg/kg and 4.27 mg/kg respec-tively). Dimethyl phthalate concentrations were found to be 8.7 mg/kg in two of the samples from Burrard Inlet, but were undetect-able or in trace amounts at all other sites. Garret (1980) reported that an extensive survey of contaminant levels in biota in the lona Island mudflats region found PAEs in all samples analysed. DEHP was present in highest concentrations (up to 696 ug/kg wet weight in flounder). Cain, Clark and Zorkin (1979) found low but guantifiable concentrations of a range of PAEs in the effluents from the Annacis Island, lona Island and Lulu Island primary sewage treatment plants. Total P A E concentrations ranged from 2.0 -50.0 ug/L. 54 PAEs have lower recorded toxicity than the other nominated con-taminants of concern to ocean dumping management. The concern for PAEs in the environment is a result of their high production and subsequent release into the environment, and their persistence within aquatic sediments. The high concentrations of PAEs found in sediments of False Creek and Burrard Inlet by Hall, Parkinson and Ma (1983), and the probability that other areas of Vancouver's waterways also contain substantial concentrations of PAE's, justify their inclusion as contaminants of concern to ocean dumping management. Chlorinated Butadienes No data was found describing chlorinated butadiene contamination of Vancouver's sediments. However these compounds, particularly hexachlorobutadiene, are significant by-products of wastes from the production of chlorine, hexachlorobenzene, chlorinated ethylenes and methanes, and other chemical industries (U.S. EPA, 1975). Hexa-chlorobutadiene (HCBD) is also used as a solvent for natural and synthetic rubber and other polymers, as a washing liquor for removing hydrocarbons and as a hydraulic fluid (Verschueren, 1983). Chlorinated butadienes have been found in high concentrations in sediments of urban and industrialized embayments in Puget Sound (up to 3.3 mg/kg dry weight of HCBD; Malins et al., 1980; Riley et al., 1980). The likelihood of comparable levels in Vancouver's sediments 55 cannot be discounted. The high probability of chronic effects in biota that are exposed to very low levels of HCBD has justified its inclusion in the Canadian Environmenal Act list of priority chemicals. Polychlorinated Dibenzafurans (PCDFs) No data was found describing polychlorinated dibenzafuran contami-nation of Vancouver's sediments. PCDFs are not produced inten-tionally for any purpose, but are common impurities in PCBs and chlorinated phenols (Norstrom et al., 1976). Other possible sources include coal tar used as an anticorrosion coating on sub-aguatic pipes and in fly ash from municipal or industrial incinerators. The primary cause for suspicion of P C D F contamination in Vancouver's waterways is its presence as impurities in pentachlorophenol (PCP), which is used widely in local sawmill facilities, lumber shipping terminals and pulp and paper mills. PCDFs may also be formed during the burning of PCP contaminated hogfuel from sawmills (Norstrom et al., 1976). Although PCDFs are not nominated as priority chemicals under the Canadian Environmental Contaminants Act , the identification of three PCDFs in Puget Sound sediments (Malins et a l „ 1980) justifies appreciable concern in British Columbia. PCDFs are structurally very similar to chlorinated dibenzodioxins, which are among the most toxic chemicals known. Recent findings suggest that small amounts of 56 PCDFs found along with PCBs in contaminated rice oil may have been the principle factor causing mass food poisoning in Yusho, Japan in 1968 (Koroki and Masuda, 1978). Organotins The major concern for t r i - and tetra-organotins in Vancouver's sediments derives from its use in antifouling paints. There is also minor local usage in wood preservation (Garrett, 1982). Di-organotins are used extensively as stabilizers in PVC plastics. Information on environmental levels of organotins is lacking as analytical methodology is still in the developmental stage. However high levels are suspected for sediments underlying boat and ship repair facilities, where concentrations have accumulated via leaching and during the removal of antifouling paint and the application of new paint. Recent innovations in analytical techniques have greatly improved our ability to monitor organotins in sediments (C. Garrett, EPS, pers. comm., 1985). Conclusion It is concluded that each of the classes of toxic and persistent chemicals identified in Table VIII, except for the organochlorine pesticides, warrant concern in the management of dredged material from the Vancouver region. Organochlorine pesticides have never been extensively used in British Columbia, and have only been found at 57 significant levels in the aquatic sediments of the Lower Mainland on one isolated occasion. However the toxic trace metals, PCBs, PCP, HCB, HCBD, PCDFs, PAHs and PAEs each have documented or suspected distribution within Vancouvers sediments so each demand specific consideration in the formulation of a scientifically rational dredged material management program. Implicit within this conclusion is a requirement that routine analyses of sediments proposed for ocean dumping should be extended from the present cursory examination to cover this broad range of substances. In practice however, analysis costs (especially when considered in the light of requirements for sample replication necessary to provide adequate sediment characterization over an area such as False Creek) will prohibit the conduct of this breadth of analyses on a routine basis. A tiered structure of sediment analyses may provide some reduction in the suggested analytical requirements. The majority of sediments dredged from Vancouver's waterways are relatively unpolluted, so should not require detailed characterization. For example, coarse sediments from the Fraser River Main Arm do not warrant the detailed assessment that should be required of "ooze" sediments from backwater areas. If an initial assessment indicates that sediments proposed for ocean dumping may be significantly contaminated, then a requirement for scientifically informed management practices implies a responsibility to demand substantive and comprehensive sediment characterization data before the issuance of a permit. 58 I will return to this question subsequently. For the present purpose however, it is concluded that a significant portion of ocean dumped dredged sediments from the Vancouver region may contain any or all of the nominated contaminants of concern. Therefore, the environmental processes effecting the release of any of these contaminants from disposed dredged material, and the environmental effects of the contaminants that may be released, must be accounted for in a comprehensive management scheme. These tasks comprise the objectives of the following two sections. 59 4 . 0 T R A C E M E T A L PATHWAYS A N D PROCESSES IN T H E MARINE ENVIRONMENT 4 . 1 Introduction Trace metals occur naturally in the marine environment. Their distribution in seawater was first compared and discussed by Fabricand et al. (1962) and Shutz and Turekian (1965). They showed that the considerable fluctuation observed in coastal waters can be attributed partly to man's influence and partly to natural biological cycles. Only in the past few years however, have reliable experimental data become available. Natural processes in riverine and marine ecosystems tend to con-centrate trace metals in bottom sediments. Trace metals are not very soluble in water under the conditions that normally occur in oxygenated uncontaminated surface waters, so the introduction of high concentrations of metals into aquatic ecosystems will generally result in a equilibrium condition where most of the metals will be sorbed (adsorbed or absorbed) by suspended particulate matter and then deposited on the bottom when the suspended material settles. The time necessary to achieve this equilibrium condition depends on the physiochemical conditions of the aquatic system and the quality and duration of contaminant introduction. 60 Society's awareness of trace metal contamination in the marine environment is a recent phenomenon. It was first brought to prominence in the early 1950's by a mysterious neurological illness and resultant human fatalities among the subsistence fisherfolk of Minamata Bay in Japan. Since this disease also prevailed among local seabirds and household cats, investigations led to the discovery that the consumption of high concentrations of mercury compounds accumulated in fish and shellfish had evoked disasterous end effects in the nutritional food chain. Society became suddenly aware of the existence of toxic metals in the environment. Four other tragedies of human poisoning due to toxic metals have been documented in Japan since Minamata (Table XI). In each case the concentration of the metals in a sediment phase and subsequent release and transfer to the food chain was a factor (Kester et al., 1983). Consequently there is good cause to understand the behaviour of metals in the disposal of contaminated dredge spoil. The release of trace metals from disposed dredged materials to the water column is complex, being influenced by a number of physical, chemical and biological factors. Although the fundamental concepts relating to the transformation and release or retention of trace metals in the aquatic environment have long been developed, it is only in the last decade that they have been applied . to dredging and disposal operations. The Dredged Material Research Program, conducted by the U.S. Army Corps of Engineers, have studied the release of chemical contaminants from a wide variety of dredged sediments. Although the magnitude of metal release is largely 61 TABLE X I . Summary o f I n c i d e n t s o f Human P o i s o n i n g by T o x i c M e t a l s i n t h e A q u a t i c Environment o f Japan ( F o r s t n e r and W i t t m a n , 1981). L o c a t i o n Year Metal Minamata Bay Ni i gata Goshonoura Fuchu Tokyo 1953-1960 1964-1965 1973 1947-1965 1975 M e r c u r y Mercury M e r c u r y Cadmi urn Chromi um 62 specific to the dredged material and the receiving environment, general principles have been derived which should be applied to environmental regulation. 4.1.1 Trace Metal Speciation The action of chemical processes effecting the release of trace metals is highly dependent upon the speciation of the metals within the sediments. Most trace metals in the natural sedimentary "sink" are incorporated into the crystalline lattice of their mineral carrier (Brannon et al., 1976). Because this incorporation is by metal bonding and is predominantly in inert positions, these trace metals are effectively removed from the aquatic and biological systems. Trace metals associated with parts of the dredged material other than the mineral crystalline lattice can also be essentially immobile and biologically unavailable. Metals associated with crystalline Fe and Mn oxides in dredged sediments are a case in point. For example, most sediment As is usually associated with these highly crystalline Fe and Mn oxides and is chemically immobile and biologically unavailable (Brannon et al., 1976). However there is a diverse range of other metal associations in aquatic sediment, some of which are only weakly bound to the mineral lattice. Metals in sediment interstitial waters or adsorbed to the cation exchange complex, and metals associated with poorly crystalline, amorphous Fe and Mn oxides, are the most mobile and 63 potentially available contaminants in dredged material (Burks and Engler, 1978). Forstner and Patchineelam (1976) characterized metal associations in both natural and polluted aquatic systems by the bonding processes (Table XII). A number of studies have been conducted to characterize the ability of a sediment to release material to the overlying water. Procedures used in these studies have ranged from filtration (to determine interstitial water) to total acid digestion. In addition, several intermediate operationally defined sediment extraction procedures have been used to assess sediment constituent availability. These latter procedures are similar to soil extraction procedures that have been used for many years to define exchangeable, available and extractable constituents in soils (Jackson, 1958; Black, 1965). No single extract has been shown to be universally successful in defining chemical exchangeability (Jenne and Luoma, 1977). Con-seguently, a variety of extraction procedures ranging in severity from distilled water or a dilute salt solution to strong oxidizing agents such as hydrogen peroxide to a strong acid solution have been used to define sediment geochemical phases. The elutriate test, developed jointly by the U.S. Army Corps of Engineers and the U.S. Environmental Protection Agency to evaluate .proposed dredged material disposal operations, is similar to the weaker extractant used in elemental partitioning studies except that disposal site water rather than distilled water or a well defined salt solution is used as the leaching agent. 64 TABLE X I I . Mechanisms o f T r a c e Metal Bonding (from F o r s t n e r and P a t c h i n e e l a m , 1976) C a r r i e r Bonding P r o c e s s C r y s t a l l i n e m i n e r a l Metal bonding p r i m a r i l y a t i n e r t p o s i t i o n s H y d r o x i d e s - , Carbonates S u l f i d e s H y d r o x i d e s and Oxides o f Fe/Mn B i t u m e n , L i p i d s , Humic s u b s t a n c e s , R e s i d u a l o r g a n i c s C a l c i u m c a r b o n a t e pH dependent pH dependent pH dependent P r e c i p i t a t i o n as a r e s u l t o f e x c e e d i n g s o l u b i l i t y p r o d u c t i n s i t u P h y s i c o - s o r p t i o n Chemical s o r p t i o n (exchange o f H + i n f i x e d p o s i t i o n s ) C o p r e c i p i t a t i o n as a r e s u l t o f e x c e e d i n g s o l u b i l i t y p r o d u c t P h y s i c o - s o r p t i o n Chemical s o r p t i o n (exchange o f H + i n COOH", OH" groups) Complexes P h y s i c o - s o r p t i o n Pseudomorphosus C o p r e c i pi t a t i o n ( i n c o r p o r a t i o n by e x c e e d i n g t h e s o l u b i l i t y p r o d u c t ) 65 Another commonly adopted strategy of selective leaching was developed by Brannon et al. (1976). This procedure partitions trace metals within the sediments into six phases as follows: 1. Interstitial water, which is subject to diffusive exchange with benthic organisms or the overlying water column. 2. Easily exchangeable phase, which may release dissolved cations in response to pH and salinity changes. 3. Easily reduceable phase, which may be released with a slight decrease in redox potential. 4. Organic plus sulfide phase, which could be mobilized by exposure to oxygenated conditions. 5. Moderately reduceable phase, which could be mobilized if the redox potential substantially decreases. 6. Residual phase, which is tightly bound and unlikely to be released to the marine environment or biota upon dredging and disposal. However neither the elutriate test nor selective leaching strategies provide complete information describing the lability of trace metals in disposed dredged material. This failing is perhaps more the result of the complexity of _in situ transformation processes than inappropriate analytical procedures. The receiving environment includes not only 66 the bulk seawater column, but also such complex and variable micro-environments as the interstitial water of the sediment, the interfacial transition zone between solid and solution, and the digestive tracts of particulate feeding benthic animals. In the following discussion the emphasis is placed on the environmental conditions which induce metal transformations and remobilization. The phase partitioning of trace metals within individual sediments effects the susceptability of those sediments to the described transformation processes. Trace Metal Release During Water Column Passage The remobilization of trace metals from dredged material is mainly caused by four types of chemical changes. 1. Elevated salt concentrations, whereby the alkali and alkaline earth cations can compete with metal ions sorbed onto solid particles. 2. Changes in redox conditions. Under more reduced conditions iron and manganese hydroxides partly or completely dissolve thus releasing part of the incorporated or sorbed trace metal load. Under more oxidized conditions, metal sulfides may be oxidized to more soluble forms. 3. Lowering of pH which leads to a dissolution of carbonates and hydroxides, as well as to increased desorption of metal cations due to competition with H+ ions. 67 4. Increased concentrations of complexing agents in solution, which can compete for trace metals adsorbed to solid particles and form soluble metal complexes, sometimes with high stability. 4.2.1 Salinity Alteration Concern that salinity alteration may cause trace metals desorption from riverine sediments disposed into the marine environment is primarily based upon research on trace metal behaviour in estuaries. However estuarine chemistry displays a very complex system of interactions that are still not completely understood, especially with respect to the behaviour of trace metals. Turekian (1977) noted that in an estuarine system there is continuous movement of some metals in and out of solution, but little is actually lost out of the system. The loss that does occur is primarily in the form of fine particles. Experiments conducted by Kharkar, Turekian and Bertine (1968) on the desorption of absorbed cobalt, silver, selenium, chromium and molybdenum show that where the trace metal was absorbed from freshwater it was always released to a greater or lesser extent upon contact with seawater. They concluded that the mechanism involved was competitive displacement of adsorbed ions by magnesium and sodium ions. Sholkovitz (1976) argued that the use • by these authors of standard clays and distilled water, the absence of dissolved and solid organic matter and the failure to monitor pH restricts the 68 conclusions that can be drawn from these artificial test conditions. Nevertheless, desorption did seem to occur and this has been confirmed in other studies. When Van der Weiden, Arnoldus and Meurs (1977) resuspended Rhine River sediments in seawater they demonstrated significant desorption, primarily due to chloride and sulfate complex formation. The order of decreasing desorption of metals in the experiments was Cd > Zn > Mn > Ni > Co > Cu == Cr; for Fe and Pb no desorption was found. However, in a more recent series of experiments on suspended Rhine River sediments in seawater, Patchineelam and Calmano (1981) concluded that none of the investigated elements (Zn, Co, Cr and Fe) released more than 10% of the original metal concentration in the seawater. In the Columbia River estuary, similar studies by Evans and Cutshell (1973) observed minor desorption of Mn and Zn due to salinity changes, but saw no loss of Cr, Sb or Sc. Trefry and Presley (1976) conducted extensive studies in the Mississippi River estuary, comparing river suspended matter to that in seawater just outside the river mouth. They showed that individual particulate metals behave differently upon mixing with Gulf of Mexico water. Fe, A l , Co, Ni and Cr concentrations were very similar in river and gulf suspended matter whereas Mn content generally decreased (i.e. was released) seaward and Zn, Pb, Cu and Cd concentrations were either similar or higher in gulf samples. These observations argue against extensive desorption of any of these metals except for Mn. The decreased Mn 69 concentrations in gulf suspended matter, which were up to 40% lower than those in the river, suggest desorption of Mn similar to Evans and Cutshell (1973). In a number of instances, the Zn, Pb, Cu and Cd concentrations were higher in the gulf suspended matter, suggesting that under certain conditions (e.g. following pH changes from less than 8.0 to 8.5, or during phytoplankton blooms) uptake of these metals may occur. Similar studies of the effects of the sediment exchange reactions in the Fraser River estuary on zinc and copper distributions in the Strait of Georgia concluded that both these metals desorbed from suspended sediments when the river water mixed with seawater (Thomas and Gril l, 1977). De Mora (1981) showed that this also occurred for Mn, which exhibited a maximum dissolved concentration in the Fraser River estuary at salinities ranging from 4 to 12 ppt. Bindra (1983) used static and agitated water column tests to inves-tigate the effects of salinity alteration on desorption of copper, iron, manganese, lead and zinc from sediments of the Lower Fraser water shed. Variable results were achieved. There was minor release of copper to a static, aerated water column due to a salinity increase from 0 ppt to 14.5 ppt, but much of this was removed from the water column with further increases in salinity. In contrast, the agitated test measured negligible release of copper due to salinity increases. Analyses of zinc concentrations in a static, aerated water column showed release of zinc into solution following a salinity increase to 14.5 ppt, subseguent readsorption after a salinity increase 70 to 21.5 ppt, then slow release after salinity was further increased to 29.5 ppt. In the agitated test, high organic sediments released more zinc at higher salinities, but low organic sediments showed high release due to a very slight salinity increase (to I ppt), and removal from solution at higher salinities. Iron and lead were generally removed from solution due to increased salinity, whereas manganese was released. It is perhaps not surprising that so much seemingly conflicting data have been published on marine adsorbtion-desorption phenomena in view of the likely importance of such variables as the nature and con-centration of both dissolved and suspended matter, pH, time of equilibrium, temperature, etc. These matters are discussed in some detail in O'Connor and Kester (1975) and Parks (1975). It may be then, that each dredged material and disposal environment will have to be considered individually if there is concern for minor changes in the chemistry of the material in response to salinity changes. However it seems unlikely that major chemical changes can be expected as a response to salinity changes alone. Changes in redox conditions and pH are of much greater significance. 4.2.2 Alteration of Redox and pH Conditions The redox potential (Eh) and the pH are two of the primary variables that control the chemical behaviour of metals in marine sediments. The redox potential of the water column or sediment is a measure of the availability of electrons and thus the intensity of oxidation or 71 reduction of a sediment-water system. Similarly, the pH indicates the activity of the hydrogen ion in the aqueous phase and is thus a measure of the degree of acidity or alkalinity. Redox potential and pH act in combination to influence processes affecting the mobili-zation or immobilization of metals in sediment-water systems. Some of these processes can be described from thermodynamic considera-tions (e.g. valence state), but many processes do not lend themselves to mathematical description. In situ, the fine-ground "ooze" sediments that are of primary concern to this review are generally reduced, containing organic matter and other substances that did not completely oxidize before being cut off from the overlying water by burial. Microbial degradation of the organics in anaerobic conditions results in the formation of sulfides, ammonia and reduced forms of iron and manganese. Trace metals are normally stabilized as insoluble metal sulfides. Typical chemical conditions are a low redox potential (Eh <-50 mV) and pH (usually in the range of 6.8 - 7.5). The waters in Georgia Strait are generally oxidized throughout (Waldichuk, 1957). Marine environments that contain dissolved oxygen usually have Eh values greater than 400 mV with pH values generally in the range of 7.5 - 8.4. The sediments within low energy deposi-tional areas of the Strait, such as might be suitable for dredged material disposal, may have an oxidized surface veneer (to a depth of a few centimetres or less), underlayed by anaerobic sediments. 72 The degree to which dredged sediments become oxidized upon disposal depends upon the extent of contact between the material and oxygenated water. Clam-shell dredged material which is disposed from a bottom dumping scow tends to remain cohesive during its passage through the water column, so contact will be minimal. In contrast, hydraulically dredged material entrains large volumes of water during both the dredging and the disposal operation, so may be substantially oxidized. When reduced sediments are mixed with oxygenated water, the metal sulfides are oxidized. The oxidized forms of the metals, with the exception of manganese and iron, are somewhat more soluble than the sulfide compounds. In addition, the oxidation of sulfides to sulfate may reduce the pH, resulting in increased dissolution of exchangeable trace metal cations. However the release of trace metals is counteracted by sorption of the dissolved cations onto the surface of poorly soluble iron and manganese hydrous oxides. The hydrous oxides of iron and manganese form colloidal particles that precipitate and exhibit a large, active surface area that scavenges other trace metals from the water column (Burks and Engler, 1978). The extent to which this scavenging process compensates the release of soluble metal cations depends upon the degree of oxidation of the sediments, the pH conditions and the relative phase concentrations of the various metals. Freshly preci-pitated colloidal hydrous oxides are especially efficient in scavenging for trace metals (Lee, 1973). 73 The above theoretical framework is substantially complexed in natural sediments by the activity of organic material within the dredged sediments. Chelation with soluble organic compounds may enhance heavy metal solubility to levels considerably greater than the con-centration of soluble free ions. Conversely, complexation with insoluble organics is an important sink for many metals. Redox potential, and to a lesser extent pH, influence both of these processes by guantitatively and qualitatively affecting the organic compounds present (Gambrell et al., 1976). The structurally complex, large molecular weight organic compounds characteristic of humic materials in reduced environments are reported to be altered to smaller, less complex molecules with less metal binding capacity as a sediment is oxidized (Patrick and Mikkel sen, 1971). Though metal ions may be released from insoluble humic materials as a consequence of oxidation or acidification, it is likely that such release would occur only where a reduced dredged material was subjected to an oxidizing environment for a considerable period of time (Gambrell, Khalid and Patrick, 1976). Numerous researchers have investigated trace metal release under varying Eh and pH conditions using the elutriate test described by Lee (1975). This test involves the mixing of one volume of sediment with four volumes of seawater for a 30 minute shaking period. Lee (1975) considered that this will induce greater mobilization of trace metals than will occur under actual dredged material disposal conditions. 74 Bindra (1983) investigated the effects of varying pH conditions upon the release of copper, lead and zinc from sediments of the Lower Fraser watershed. Generally a high release of all metals was observed at extreme pH conditions (pH of 5 and 10). Under normal pH (between 6.5 and 8) and both oxic and anoxic conditions, release of copper and lead was very low and release of zinc was slight. Physical characteristics of the sediments (organic matter, particle size distribution) appeared to affect the release pattern. Chen, Lu and Sycup (1976) tested release of Ag, Cd, Cr, Cu, Fe, Hg, Mn, Ni, Pb and Zn from sediments taken from seven areas of Los Angeles Harbor and released into oxidized (DO = 5 - 8 mg/L), slightly oxidized (DO = 0 - 1 mg/L), and reduced (DO = 0 mg/L) conditions. Neutral pH conditions were maintained. Under all conditions there was an immediate release of metal to the seawater, followed by subsequent removal from solution, either gradually in reducing environments, or immediately in slightly oxidized or oxidized environments. The metals that were significantly released in this study were Fe, Mn and Ni, with Cr, Cu, Pb and Zn being released slightly. The release phenomena behaved similarly for each of these metals, with release in reducing conditions > in slightly oxidizing > oxidizing. The concentrations of metals in the equilibrated water were very low, in the parts per billion range or less. There was no release of Ag, Cd or Hg. 75 Other investigations testing a variety of sediments from U.S. rivers and coastal inlets have shown that manganese is the only metal released in substantial quantities during the Elutriate Test conducted under fully aerated conditions (Brannon et al., 1976; Lee et al., 1975; Lee et al., 1978; Brannon, Plumb and Smith, 1978; Shubel, 1978; Wright, 1978). Transitory release (a matter of minutes) of mercury (0.01 - 0.05 ug/L), lead (< 40 ug/L), cadmium (0.08 - 2.5 ug/L) and nickel (5 - 20 ug/L) have been observed on occasion in the field (Lee et al., 1978, Wright, 1978). Iron is usually released initially in much higher concentrations than metals other than Mn, but is subject to very rapid oxidation and precipitation in the water column (Wright, 1979). The iron oxide precipitates have been generally found to result in the removal of most other soluble metal constituents from the water column (Lee et al., 1975, Brannon et al., 1976). Brannon et al. (1976) observed large releases of zinc during elutriate tests on Mobile Bay sediments run under anaerobic acidic conditions. However these releases have not been observed by others conducting aerated elutriate tests on similar sediments. 4.2.3 Field Monitoring In field monitoring during disposal by barges and hopper dredges, releases of trace metals have been found to be minor and of limited duration. Such releases should not exert even short-term adverse effects on water quality or marine organisms at the disposal site. Even in continuous discharge pipeline disposal in areas where 76 sediments contained elevated levels of most metals, Shubel, (1978) found no plumes of dissolved metals significantly greater than background levels. Trace metals in the disposal plume were asso-ciated with particulate matter and were rapidly removed from the water column. It thus seems unlikely that metal release to the water column during the disposal operation would have significant biological effects. Clam-shell dredged materials that are disposed by bottom opening scow remain relatively cohesive as they fall through the water column (discussed subsequently in Section 6.0). Because seawater entrainment within the falling mass is small, the release of trace metals dissolved within the pore water should be minimal. The limited extent and duration of contact between the disposed sediments and the seawater should also mitigate against oxidative dissolution of trace metal contaminants. Concern for metal release during dredging and disposal has largely been based on the extended and vigorous mixing which occurs during hydraulic dredging operations. In recognition of the reduced concern attached to clam-shell dredging, Lee (1977) proposed a short-contact "plop test" as a more reasonable simulation of trace metal release under these conditions than the elutriate test. 77 4.3 Trace Metal Release from Settled Sediments 4.3.1 Chemical Release Processes After the sediment settles it normally returns to an anaerobic and near neutral pH condition. The surface sediment will probably be oxidized to a shallow depth. The vertical extent of the oxic zone is a function of the natural sediment accumulation rate (Shokes, 1976), the concentration of metabolizable organic matter (Berner, 1971), and the nature and extent of biological activity. If sizeable amounts of anoxic dredged material are deposited in oxygenated water there will be a limited oxygen supply except in the surface centimetres of sediment where biological mixing occurs. With freguent disposal of high organic content waste material, it may be possible to generate a completely reducing, anoxic sediment column. Hoos' (1976) study of the Point Grey disposal site evidenced low organic content in the disposed sediments (maximum value from 50 core samples was 1.61%), despite the high volume of disposed wood debris and fibrous material at the site. It is assumed that appropriate dredge spoil sites in Georgia Strait will be generally anoxic, overlayed by a variable but shallow oxic zone. In the anoxic region, some chemical species may be released to the interstitial water whereas others may be rendered immobile. Field data has shown that partitioning within settled sediments is complex and variable. On the basis of solubility product constants for metal 78 sulfide compounds, trace metals should remain essentially fixed in reducing sediments. Thomson, Turekian and McCaffrey (1975) ca l -culated that trace metals should remain immobile in pore waters with a sulfide ion concentration of IO -9 mole per litre. However Manheim (1976) reported that trace metal concentrations in pore water were very high; up to 1000 times higher than the overlying seawater. Hallberg (1974) ascribed the high solubility of trace chemicals measured in pore waters of reducing sediments to the stabilization of metals leached from silicates and oxides onto soluble organic complexes. Nissenbaum and Swaine (1976) analyzed the metals bound to dissolved organic complexes from interstitial water of reducing sediments in Saanich Inlet, B.C. The results showed that all of the Zn, almost all of the Cu and a major part of the Fe, Ni and Co are bound to humic material. Experimental data of Lu and Chen (1977) suggests that under reducing conditions the concentrations of Cd, Hg and Pb are controlled by sulfide complexes with organic complexes controlling Fe, Ni and Zn, chloride complexes for Mn and hydroxide complexes for Cr. High pore water concentrations of trace metals will establish a concentration gradient between the sediments and overlying seawater, which will promote upward diffusive migration. Dissolved trace metal complexes that migrate upward will reoxidize in the aerobic surface sediments, so release to the overlying seawater will generally be controlled by the solubility of the oxidized species. 79 Large molecular weight organic molecules (e.g. humic compounds) slowly degrade to simpler fulvic compounds in an oxic environment. From a colloid chemistry point of view, metal ions in large molecular weight molecules may be included in central hydrophilic "cavities" of the molecule, and would then be prohibited from leaching. Only the higher molecular weight fractions exhibit colloidal properties, and these metals may be released following oxidation to simpler com-pounds (Reddy and Patrick, 1975; Chen et al., 1976). However this release process will be countered by the higher complexing capacity of many simpler organic molecules (e.g. fulvic compounds) which have more functional groups (i.e. OH, COOH) with which the released metals can attach. Also, the very slow rate of oxidation of large organic molecules will generally mean that the absorption of any released trace metals onto iron and manganese hydroxides within the surface sediments will substantially counter release to the overlying water column. Long-term incubation tests (3 and 5 months) conducted by Lu and Chen (1977) showed that, with the exception of Cr and Hg most trace metals were found to be released relative to original seawater concentrations. However total release remained in the sub-ppb to ppb levels. The explanation offered for the negligible release was the scavenging effects of Fe/Mn hydroxides or clay minerals and the formation of insoluble humic complexes. Studies by Windom (1976) on the exchange of metals between sediment and water from marshes and estuaries in the Gulf of Mexico similarly showed no significant increases of heavy metals such as Pb, Cu and Hg above ambient levels. It was concluded that the slowness 80 of the diffusion processes and the rapid absorption of released metals by Fe/Mn hydroxides resulted in insignificant contribution to the water column. In contrast to these results, studies by Gambrell et al. (1977) with Mississippi River dredged sediment indicated that Cd release was heavily pH-redox dependent, and that significant release may occur in time under low pH, oxygen rich conditions. Gilbert, Clay and Leighty (1976) found that prolonged oxidation of reduced polluted sediments can result in significant Cd release. Due to the high toxicity of Cd, it was concluded that care must be exercised with polluted dredged material. Khalid, Patrick and Gambrell (1978) subjected reduced sediment suspensions containing cadmium, sulfides and other inorganics to streams of oxygenated gas mixtures. They found that gas mixtures with high and low oxygen contents solubilized more Pb, Cd and Cu than gases with an intermediate oxygen content. Their explanation was that a low oxygen regime favoured metal solubilization by organics; a high oxygen regime supported conversion of insoluble sulfides to soluble sulfates as Eh rose from -170 mV to +600 mV, and the concomitant decrease in pH promoted the formation of more soluble cadmium species. Khalid et al. (1978) suggested that hydrous oxides of iron and manganese acted as scavengers for dissolved cadmium at moderate Eh, but could release absorbed soluble cadmium as sediments became slightly acidic and oxidizing. 81 4.3.2 Biological Release Processes The laboratory studies cited above provide no measure of the processes by which the biological community that colonizes a dredged material disposal site may effect trace metal release. With time, biologically induced release of trace metals to the water column may be significant. Wood (1974) and Petr (1977) suggested that of the several mechanisms responsible for the movement of trace metals across the sediment-water interface, diffusion is probably the least important. The movement of certain biota and release of gases from microbial activity were considered of greater significance. Macrofaunal and microbial activity in the seafloor sediments can each significantly effect trace metal release. Bioturbation Burrowing and tunnel building influence trace metal release rates by active and passive regulation of pore water exchange with the water column. Although little is known about the mechanisms, rates or effects of biologically induced pore water exchange, Swartz and Lee (1980) contend that it is more important than particle transport in controlling the flux of many dissolved compounds. Biogenic activities resulting in pore water exchange include burrowing by vagile organisms and irrigation by vagile and stationary fauna. Irrigation is the transport of water through a tube or tunnel for respiratory or feeding purposes. 82 Unfortunately few studies have investigated the influence of organisms on the rate of exchange of contaminants from marine sediments to the water column. Tube irrigation by a polychaete was shown to enhance ammonia flux to overlying water (Aller and Yingst, 1978). Burrow irrigation was also shown to influence the benthic flux of methane (Martens, 1976) and manganese (McCaffrey et al., 1980). Irrigation by Arenicola marina (Gordon, Dale and Kerzer, 1976) and Uca pugnax (Pemberton, Risk and Buckley, 1976) may have contributed to the dissipation of buried oil. Bryan (1976) suggested that burrowing organisms would release heavy metals dissolved in interstitial water, but provided no substantiation. The only study that could be found to support this belief was by Renfro (1973), where the addition of the polychaete (Nereis diversi- color) to contaminated sediments increased the loss rate of &$Zn three to seven times compared with sediments without worms. It is considered that the scavenging properties of Fe/Mn hydroxides probably plays a significant role in counteracting biogenic trace metal release due to mechanical processes. Burrowing and tube construction do increase the depth and surface area for passive or active exchange of pore water and seawater. However the sediment-water interface of the burrows or tubes will generally be oxidized.. Dissolved trace metals within the pore water that diffuses or is pumped into this aerated layer will generally adsorb and coprecipitate with the Fe/Mn hydroxides. 83 Aller (1978) determined that sediments with the polychaete Clymenella  torquata had higher pore water concentrations of Fe and Mn in the surface 2 cm, but lower concentrations from 2 - 1 2 cm. There was an increased flux of Fe, Mn, NH4 and HPO4 from the pore waters due to diffusive exchange of ions through the polychaete tubes. Aller (1978) reasoned that closely spaced tubes increased the surface area of the sediment-water column interface and decreased the average distance these ions had to travel to diffuse into the water column. However the coprecipitation of dissolved trace metals with precipi-tated Fe/Mn hydroxides was believed to have counteracted the expected release of trace metals to the water column. Although trace metals release to the water column may not be increased by mechanical processes of biological activities, it is noted that these processes do result in substantially elevated trace metal concentrations in the aerated surface sediments. In areas of high reworking, surface sediments may be ingested several times a year. Passage through the gut of invertebrate exposes particulate and sorbed trace metals to an acidic environment (pH < 5), which may release organically and inorganically sorbed metals. Pryor (1975) demonstrated that the structure of some clay minerals is modified by digestive processes. The nature and extent of chemical alteration of trace metal species through digestive processes, and the degree to which trace metal elimination in feces effects trace metal release, remains unknown. High concentrations of organic pollutants have been measured in 84 invertebrate feces (Lee and Swartz, 1980), and fecal deposition by benthic filter feeding populations can be a significant source of biodeposition of water borne trace metals (Kraueter, 1976). Presumably, digestion chemically alters trace metal speciation in feces vis a v[s the sediments, but the nature and relative importance of this process is not known. Of possibly greater significance, irrigation and fecal deposition by macrofaunal organisms that colonize the disposal mound can sub-stantially increase microbial density in the sediments (Rhoads, McCal l and Yingst, 1978). The importance of microbial processes in the remobilization of metal components from the sedimentary reservoir is frequently stressed in the literature, but the kinetics are poorly understood. Microbial Processes There are two major processes leading to the mobilization of metals from marine sediments by microbial activity: 1. The destruction of organic matter. 2. The conversion of inorganic metals into soluble or volatile metal-organic compounds. 85 The first of these is essentially the acceleration of organic matter decomposition in aerobic environments, described in Section 4.3.1. The structurally complex, large molecular weight organic compounds characteristic of humic materials in reduced environments are altered to smaller, less complex molecules (e.g. fulvic compounds) by microbial degradation (Reddy and Patrick, 1975). Nissenbaum and Kaplan (1972) showed that humic compounds are a major component of the organic matter reservoir in recent marine sediments, comprising up to 70 percent of the organic material. From a colloid chemistry point of view, metal ions may be included in the central hydrophilic "cavities" of humic molecules, and would then be prevented from leaching. Only the higher molecular weight fractions exhibit colloidal properties, and the metals may be released following microbial oxidation (Chen et al., 1976). However this process will be substan-tially countered by the higher metal complexing capacity of fulvic compounds, which contain more functional groups (i.e. OH, COOH) than humic compounds. The net rate and significance of metal ion release from the sediments to the water column as a conseguence of these counteractive processes is highly dependent on Jn situ conditions and is largely unstudied. However it is generally considered that microbial degradation of humic substances in marine sediments is a relatively slow process and that the resulting desorption of metals to the water column, should it occur, will be in very small guantities (Chen et al., 1976). 86 Considerably more research effort has been dedicated to bacterial methylation of inorganic metals. Bacterial methylation is a form of detoxification within the bacteria, but the products are released from the sediments and are readily partitioned into organisms. This is particularly true for the bacterial production of methylmercury, which is highly toxic to higher life forms. The inorganic forms of mercury may be methylated by at least two mechanisms (Ladner, 1971; Wood, Kennedy and Rosen, 1968). The extent and rates of methylation are effected by many factors, including concentration and availability of mercury ions; abundance, growth rate and metabolic activity of the methylating organisms; temperature and pH (Bisogni and Lawrence, 1973; Baker et al., 1981). Methylmercury may also be demethylated by bacteria in sediments (Spangler et al., 1973). Thus the amount of methylmercury released from the sediments is dependent on the combined reaction kinetics of the methylating and demethylating processes. Methylation of mercury involves the conversion of H g + + to methyl-and dimethyl-mercury. The reactions and enzymes required to make these conversions are described in Wood (1974). Dimethylmercury is synthesized from methylmercury, with the rate of synthesis of the former being about 6000 times slower than the latter. Windom et al. (1976) studied the transfer of mercury in marine sediments in Georgia. Mercury was primarily present as an insoluble sulfide, strongly sorbed to the sediments. Bacterial methylation in 87 sediments increased mercury mobility but the rate of methylation was very slow: annual production of methylmercury was calculated to be 50 ug for each gram of total mercury. The relationship of bacterial numbers and activity to the production of methylmercury has not been studied, but may be an important consideration in areas receiving sewage effluent (e.g. Point Grey) where the density of bacteria in the sediments is extremely high. The ability of micoorganisms to detoxify their environment through organometallic transformation is not restricted to mercury. It has been established that the mechanism is also effective in the formation of volatile metal-alky I compounds of As, Pb, Sn and Se. Relatively high concentrations of mono- and di-methyl arsenic acid have been found in seawater overlying contaminated sediments in Southern California (Andreae, 1978) and Florida (Braman, 1976). The results of Chau et al. (1976) showed that biomethylation of selenium may also be a significant process. However Thompson and Crerar (1980) reported that methylation of lead may not be significant in marine sediments. Only 0.03% of lead added as lead nitrate underwent methylation in that study. Thompson (1981) subsequently confirmed this finding in the field at Saanich Inlet and Al ice Arm, B.C. 88 Biological Uptake and Transfer The initial uptake of metals by marine biota can occur from water through the respiratory or digestive surfaces, or from ingested food or particles. Absorption of metals occurs by passive diffusion and for many metals, rates of absorption are directly proportional to the levels of availability in the environment (Bryan, 1979). Although marine animals readily absorb trace metals, they are also able to regulate the concentration of many metals in the body to normal levels through excretion via the gills, gut, feces, urine or molts (Bryan, 1979). This occurs more commonly with the essential and relatively abundant metals such as Cu, Zn and Fe rather than the non-essential metals such as Hg and Cd (Connell and Miller, 1984). When uptake of essential metals exceeds the levels necessary to maintain biochemical functions, excretion, induction of metallothionein synthesis and storage in the kidneys or liver, or storage at inactive sites (e.g. bone, exoskeleton), can provide homeostatic control of tissue concentrations. However, with essential and non-essential metals, if uptake is excessive the homeostatic mechanisms are inhibited and bioaccumulation in soft tissues and body fluids proceeds as uptake exceeds the loss rate (Connell and Miller, 1984). The ability of many organisms to control metal concentrations in tissues means that bioaccumulation is species dependent. For example, Home (1969) quoted a copper bioconcentration factor of 80 89 for fish, and 3,000 and 4,300 for bottom dwelling scallops and nudibranchs, respectively. Mussels and oysters had concentration factors of 3,000 and 13,700. Bryan and Hummerstone's (1973) classical study of the response of the polychaete Nereis diversicolor to metal contamination in estuaries of southwestern England concluded that adaptation of this species to zinc and copper are different, appear to be genetically determined, and have probably evolved separately. Similarly, Brown (1976) showed that lead tolerance of the isopod Asellus meridianus may be genetically determined. Physico-chemical sediment characteristics and the geochemistry of the trace metals also substantially affect metal accumulation in orga-nisms. Hall and Bindra (1979) examined the relationships between sediment trace metal geochemistry and accumulation of copper, lead and zinc in benthic invertebrates from four areas of the Lower Fraser watershed. Lead was the only metal to show some linear correlation between the total concentration in the sediment and organism accumulation, and this varied with the sites and the organisms studied. Copper levels in organisms were correlated with copper in the exchangeable and easily reducible phases, demonstrating that these phases were apparently more available or retained by the organism. No single phase of zinc appeared to regulate organism concentrations. 90 Despite this apparent complexity and confusion, the net process of biological uptake appears to fit within a simple model described by Thomann (1978). The accumulation of metals by marine organisms is a balance between the intake rate, the storage capacity of the organism and the elimination rate. For smaller organisms that comprise the base of the food chain, net metal accumulation is essentially the result of diffusive exchange between the organism and the water column. Residues may be accumulated from food or directly from the sediments, but diffusive exchange results in the maintenance of equilibrium levels in organisms relative to levels in the surrounding water. Because most forms of metals have weak sorptive tendencies for biological tissue, preferencial sorption within biota is low. If a metal is present in solution simply as an uncomplexed ion, it will not tend to be lipid soluble and (in the case of fish) may accumulate in the gills rather than in internal organs. Relative to sediment concentrations, bioaccumulation factors are less than unity for most metals. Aquatic food chain enrichment of trace metals has been frequently postulated in the literature. However it is now generally accepted that for most metals this mechanism has been confused by over-simplification (see Prosi, 1979) and that biomagnification of metals does not occur (Kay, 1984; Biddinger and Gloss, 1984). The exception to this is trace metals that occur in lipophilic organic forms. If a metal is present in lipid soluble form (e.g. methyl mercury), then accumulation in internal organs is enhanced. The high 91 sorptive tendency to organic tissue results in very high storage capacities within organisms, and low elimination rates. For example, although Young et al. (1980) found no evidence for biomagnification of Ag, Cd, Cr, Cu, Fe, Mn, Zn, Ni and Pb in several ecosystems in southern California, mercury, notably the organic forms, generally increased with trophic levels. The bioaccumulation of lipophilic substances has been subject to extensive research over the last decade. The high mercury concen-trations in several species of fish may be more a function of time and body weight rather than trophic level (Phillips, 1980). Larger organisms have a small exchange surface relative to their body volume, so exchange with the surrounding water is less important than intake from food and elimination with excretia. Because elimination via the latter route is relatively inefficient, concentrations in animals resident in contaminated areas tend to increase with residence time. The role of biological uptake and transfer in the transport of trace metals from a dredged material disposal mound is accordingly insignificant. For smaller organisms, bioaccumulation is essentially dependent upon dissolved metals. Biological uptake by epibenthic organisms affects a net removal of trace metals from the water column. Benthic infauna accumulate metals from trie interstitial and easily exchangeable fractions in the sediment system (Hall and Bindra, 1979) but bioconcentration factors rarely exceed unity. Renfro (1973) estimated that a population of Nereis diversicolor at a density of 50 worms/m2 would accumulate only 0.08% of the zinc content of the 92 upper 2 cm of the sediment layer. The fact that this fraction is within biological tissue is insignificant to the transport of trace metals from the dredged material disposal mound. Biomagnification of trace metals is essentially a misnomer, and accumulation in larger predator species is dependent upon the duration of exposure. Appropriate selection of disposal grounds in areas of low productivity, augmented by the expected reduction in disposal site biomass of prey species due to dumping activities, will ensure that disposal sites are unattractive to migrating predators. It is thus expected that dietary uptake of contaminants by predatory species will remain insignificant. 4.3.3 Conclusion It is concluded that for the purpose of minimizing trace metal release, dredged material disposal sites in Georgia Strait will preferably be within a low energy, depositional environment charac-terized by fine, reduced sediments and overlayed by a permanently oxygenated water column. Erosion and dispersion of particulate-bound trace metals will be minimized in a low energy environment. The chemically and biologically induced release of trace metals will be most effectively controlled by disposal onto reduced sediments overlayed by oxygenated water. Maintenance of a low redox potential in the disposal mound will tend to maintain the stability of sulfide and humic bound trace metals. The diffusive migration of dissolved trace metals from pore water will be mostly counteracted by oxidation and Fe/Mn hydroxide co-precipitation within the aerobic surface veneer. 93 Although appropriate disposal site selection will substantially reduce the potential for trace metal release, the extent of release over the long term defies accurate prediction. A large number of studies have been conducted to elucidate the geochemical processes effecting trace metal transport and fate, yet a high degree of uncertainty remains. The data are often conflicting, and the dominant processes appear to be largely site specific, each responding to the particular charac-teristics of different regimes e.g. differences in minerology of particulate phases and in the amounts and/or types of organic matter. Thus it is not possible, even in laboratory testing, to determine a priori which of these processes are applicable to Vancouver sediments disposed in Georgia Strait. Further uncertainty abounds when we attempt to account for the influences of biological activities upon these geochemical processes. Numerous biogenic processes that promote the release of trace metals have been identified, but the kinetics and significance of these processes remain poorly understood. Recent investigations in controlled microcosms provide valuable information on the chemical and biological behaviour of trace metals in sediments, but have also served to confirm the complexity of the system. Hunt and Smith (1981) presented the results of a study conducted in cylindrical tanks (2 m diameter and 5 m deep) containing a 30 cm layer of sediments and associated benthic organisms. These tanks were filled with seawater from lower Narragansett Bay, Rhode Island, and they simulate many features of the planktonic, benthic and 94 bacterial components of the bay ecosystem. The sediment was collected from three locations along a pollution gradient: highly-, slightly-, and uncontaminated. Hunt and Smith's (1981) study shows release of some metals from the sediments and in some instances, trapping of metals. Metal concen-tration in the water column varied during the 17 month experiment, and during much of the experiment the three types of sediment showed similar results. During the summer there was a substantial flux of copper and cadmium from the highly contaminated sediment that was greater than those observed from the other two sediments. The study showed that the release of metals from sediments is not a continual process, but is episodic and perhaps seasonal. The causes of the episodic variation in release rates could not be determined. However the release rate of trace metals from the sediments was, except for cadmium, very slow (sub-umole m _2 day - ' ) . Hunt and Smith (1981) estimated the times required to purge the upper I cm of the sediments to be decades or centuries. When dredged material is disposed into a depositional area of Georgia Strait, burial by naturally deposited sediments or by capping the disposal mound with clean fill may counteract this release rate. I I. Sediment deposition rates at the current Point Grey dumpsite and elsewhere in the Strait of Georgia are currently unknown. Near-sediment oceanographic studies are a priority topic for 1985-86 R.O.D.A.C. (Pacific Region) research projects (R. Kussat, EPS, pers. comm., 1985). 95 Thus, the conclusion drawn by Engler (1981) in his review of research conducted by the U.S. Army Corps of Engineers, that "contaminant releases (from dredged material) are usually limited to nutrients with negligible releases of toxic metals", is essentially confirmed by Hunt and Smith's (1981) work. The exception is cadmium. Similar results by Gambrell et al. (1976), Gilbert, Clay and Leighty (1976) and Hunt and Smith (1981) have shown that prolonged oxidation of polluted sediments can induce significant cadmium release. In light of the high toxicity of cadmium, this aspect requires very careful evaluation in the local context. The other significant concern in the British Columbia context pertains to the methylation of mercury. The present Point Grey disposal site is within five nautical miles of the lona Island sewage outfall. The high mercury content of some of our dredged sediments, combined with the very high bacterial densities that might be expected within Point Grey sediments, provides cause for substantial concern. Parsons, Bawden and Heath (1973) found mercury levels in crab from Sturgeon Bank ranged up to 3.7 ppm dry weight (0.74 ppm wet weight). Forrester, Ketchen and Wong (1972) similarly recorded their highest mercury levels in dogfish from specimens collected off Point Grey. Methylmercury content was not measured in either of these two studies. 96 4 . 4 Environmental Effects of Released Trace Metals There is an extremely abundant and diverse literature describing the environmental effects of trace metals released to the water column. It is well beyond the scope of this review to attempt to define the current state of knowledge pertaining to each of the trace metals known to be toxic. Wood (1974) listed 19 metals that are relatively available to marine biota in dissolved forms and which are known to be highly toxic. Connell and Miller (1984) have conducted a thorough review of the principles of ecotoxicology. This review has shown that principal concerns for metal release subsequent to the disposal of contaminated dredged materials in Georgia Strait pertain to cadmium and mercury. This section reviews the known toxicity data for these metals in the marine ecosystem. Through reviewing the toxicity of these metals, insight is also provided on the difficulty of linking toxicology and environmental effects. 4.4.1 Cadmium There is a general lack of information on the acute toxicity of cadmium to marine biota. The 96-hour LC50 values for marine fish are generally quite high, i.e., for larval Atlantic silversides, the LC50 was 1600 ug/L and for juvenile mummichog, the LC50 was 114,000 ug/L (U.S. EPA, 1978). Gastropod molluscs are generally quite insensitive with LC50 values above 1,500 ug/L. However, other 97 invertebrates are more sensitive, with the LC50 for the mysid shrimp being 15.5 ug/L. Inhibition of growth rate of the alga, Skeletonema  costatum, was observed at 175 ug/L (U.S. EPA, 1978). Low levels of cadmium are capable of causing various chronic effects. Nimmo et al. (1977) reported that mysid shrimp exposed to 6.4 ug/L cadmium exhibited 48-hour delays in brood formation and a 57% decrease in the number of young born per female. No effects were observed at 4.8 ug/L. Calabrese et al. (1975) demonstrated significant alterations in gill-tissue respiration rates measured jn vitro after 60 days exposure to 5 ug/L cadmium. The significance of this observation has not been evaluated. It has been suggested that because of its chemical kinship to zinc, cadmium may displace zinc in certain enzymes, thereby disrupting normal metabolic functions (Callahan and Slimak, 1979). Bioaccumulation is an important fate process for cadmium, which is actively accumulated by marine organisms. There are, however, significant differences in the literature concerning the degree of accumulation. Chapman, Fisher and Pratt (1968) reported bio-accumulation factors of cadmium within marine invertebrates as high as 250,000. The U.S. EPA (1978) indicated that bivalve molluscs most actively accumulated cadmium and a bioaccumulation factor of 2600 was observed after 280 days exposure. Vernberg et al. (1977) reported 700-fold bioaccumulation of cadmium by grass shrimp and Nimmo et al. (1977) reported a 57-fold bioaccumulation by pink shrimp. 98 Cadmium uptake studies require long-term experimentation. Frazier (1979) in his studies with the American oyster, Crassostrea virginica, found continuing cadmium accumulation even after 280 days exposure. Uptake is affected by speciation; an inverse relationship exists between tissue level accumulation and water salinity (Frazier, 1979; Engel and Fowler, 1979). Uptake is also affected by ambient water temperature with increased uptake at higher temperatures. Depuration rates are slow. Greig and Wenzloff (1978) and Zaroogian (1979) found no decrease in cadmium concentration to C. virginica when transferred from metal impacted waters to cleaner waters or in a declining temperature regime. However, the latter worker found that both concentration and weight of the oysters declined during depuration. Although the cadmium concentration stayed the same, the total amount of cadmium in the organism decreased. In a study of the relationship between toxicity and bioaccumulation in aquatic invertebrates, Spehar, Anderson and Fiandt (1978) found that the 28-day LC50 values for cadmium-exposed snails and lead-exposed amphipods were I I and 4 times lower than the 7- and 4-day LC50 values for these metals, respectively. The results indicate that effects could occur at lower concentrations during longer exposure periods. Data are not available on the relationship between toxicity and accumulated body burdens of metals in aquatic, invertebrates and research into this area is required for marine species. Since concerns over ecological impacts of heavy metal contamination are commonly developed from environmental surveys that identify contaminant levels in biological tissues, relationships between toxic effects and metal 99 burdens in major body organs must be established. Such information would provide much needed fundamental data for the development of sediment criteria. 4.4.2 Mercury Mercury is one of the most hazardous trace metals present in the environment. The toxicity of mercury varies greatly with its chemical form (Luckey, Venngopal and Hucheson, 1975). Monovalent mercury is relatively non-toxic due to the low availability of its salts. However divalent mercury and organic mercurials are highly toxic. Alkyl or methylmercury poisoning in higher animals differ somewhat from inorganic mercury poisoning. Symptoms may be dormant for weeks or months after acute exposure, and if brain damage occurs, effects may be irreversible. Alkyl mercury poisonings have caused permanent neurological damage resulting in impaired vision and hearing, sensory loss in limbs, ataxia, and tremor. Neurological disorders such as mental retardation and convulsive cerebral palsy have occurred in infants whose mothers were exposed to methyl-mercury during pregnancy. Fetal nerve tissue may be especially sensitive to methylmercury (Grant, 1971). The U.S. EPA (1978) Ambient Water Quality Criteria recommended an upper limit of 0.19 ug/L inorganic mercury (mercuric chloride) as a 24-hour average and a maximum concentration of 1.0 ug/L at any time for the protection of saltwater life. The criterion for 100 methylmercury is more stringent with 0.025 ug/L as a 24-hour average and a maximum concentration of 0.26 ug/L at any time. Invertebrate animal species, especially at early life stages, are noted to be more sensitive to inorganic mercury than are fish species. The EPA recommended a maximum permissible tissue concentration of I ug/g for protection of wildlife. The present U.S. Food and Drug Administration and Canadian Food and Drug Directorate administrative guidelines for protection of human health are 0.5 ug/g (wet weight) mercury in edible portions of fish (I.J.C, 1977). Concentrations of mercury in fish that have been killed by chronic exposure to methylmercury ranged from 9.5 to 23.5 ug/g (McKim et al., 1975). Fish have been shown to accumulate mercury compounds more than other aquatic organisms, both directly from seawater and indirectly through the food chain. Concentrations of mercury up to 9.8 ug/g have been recorded from fish taken from contaminated environments (Keckes and Miettinen, 1972). The position of the fish in the food chain was found to be an important factor in relation to its mercury content. Ratkowsky, Dix and Wilson (1975) found that in the Derwent estuary, Tasmania, approximately 50% of the fish species whose diet consisted predominantly of other fish had mercury concentrations in excess of 0.5 ug/g. wet weight. In contrast, 24.% of invertebrate predators and only 7% of individuals of herbivorous habit had mercury concentrations above the 0.5 ug/g level. A similar correlation between the mercury concentrations and feeding habits was reported for fishes in the lower Fraser River (Northcote, Johnston and Tsumura 101 1975). Mercury levels in piscivorous fishes (i.e., northern squawfish and prickly sculpins) were greater than mercury levels in benthic feeders, which in turn were greater than levels found in planktonic and surface feeders. Methylmercury is the form of mercury most readily bioconcentrated by estuarine fish. It has a long biological half life, i.e., 400-700 days for flounder in brackish water (Keckes and Miettinen, 1972). Hartung (1976) calculated that elimination half lives of methylmercury in pike and eel were in the order of 600 to 1,030 days. Uptake and bioconcentration of mercury is dependent upon various factors. Studies on rainbow trout (MacLeod and Pessah, 1973; Uthe, Atton and Royer, 1973) have shown that temperature affects the uptake rate of mercury into fish muscle. Increases in temperature from I 0 ° C to 20°C increased the biological magnification factor (mercury concentration in fish divided by the mercury concentration in water) from 10 to 22 times (MacLeod and Pessah, 1973). Rapid uptake of mercury occurred during the first warm summer months (Uthe et al., 1973). Luoma (1977) showed that the total mercury concentrations in biota from a small Hawaiin estuary, fluctuated over nearly two orders of magnitude during one year. Shrimp (Palaemon  debilis) rapidly concentrated soluble mercury which periodically entered the estuary from storm runoff. Between rainstorms, minimal mercury remaining in the estuary was available. The net loss of mercury in 102 tissues was slow relative to the rate of uptake, and long periods of time were necessary to lower the levels of mercury accumulated during the short rain storms. Luoma (1976) studied the uptake and inter-organ distribution of mercury in the carnivorous crab (Thalamita crenata). The rate of increase of mercury in muscle vs. uptake in the viscera was used as an indication of the relative importance of the source of mercury (i.e., from water or food) to the crab. High mercury levels in crab gills were observed after rainstorms. The body muscle-chela muscle portioning of mercury changed seasonally from the wet to the dry season, which suggested a considerable lag time in translocation of mercury from the environment to some slowly exchanging tissues. Luoma (1976) suggested that because of the lag in metal trans-location, short-term laboratory experiments may underestimate the potential for contamination of such tissues. McGreer (1981) reported seasonal variation in concentration of mercury from crabs on Roberts Bank (Strait of Georgia). The lowest concentrations were reported from winter and spring sampling occasions, while higher concentrations were reported in samples taken in late summer to fall. Based on studies of uptake with mercury, it can be seen that all field studies on contaminants should assure standardization of the sex, size, and age of the organism, period of collection, and analytical methods to assess the rates of accumulation and changes in body burdens over 103 a period of time. It is also important to recognize the mechanisms of excretion and regulation of heavy metals in various marine organisms (Bryan, 1976). For example, in studies of the role of metal regulation in the accumulation of mercury in body tissues of Dungeness crab, the destruction or blocking of enzyme pathways in the excretory organs was suggested as a mechanism leading to progressively reduced ability to eliminate free mercury in the crabs (Sloan, Thompson and Larkin, 1974). Deterioration of excretory capability after chronic exposure to mercury would lead to higher concentrations in the oldest, largest animals. Marine invertebrates have some protection from low level chronic heavy metal pollution by a detoxification mechanism involving metal binding proteins. The low molecular weight protein metallonthionein is present in several species of marine invertebrates from Sturgeon Bank (Brown et al., 1977) and is instrumental in providing protection against high body burdens of metals. The term biomagnification refers to a specific phenomenon in which a compound is concentrated through consumption by progressively higher food chain organisms, and results in a net increase in the tissue concentration at each successive trophic level. Biomagnification of mercury occurs, especially, in long-lived predatory fish species such as swordfish, tuna, and pike. In this way, even relatively low concen-trations of mercury in the environment can be built up through the food chain to mercury levels exceeding the 0.5 ug/g wet weight concentration considered safe for human consumption. 104 5.0 ORGAN IC P O L L U T A N T PATHWAYS A N D PROCESSES IN T H E MARINE ENVIRONMENT 5.1 Introduction Rachel Carson's "The Silent Spring" (1971) provided a major stimulus for research into the fate and effects of organic chemicals in the environment. That this popular work was specifically concerned with DDT led to much of the initial research effort in this field being directed towards organochlorine pesticides. In more recent years this effort has expanded rapidly to include a very wide range of organic chemicals. The work is a formidable and complex task: there are literally hundreds, if not thousands, of anthropogenic organic compounds that can reasonably be considered likely components of the present and future environment. For example, in excess of 200 organic compounds considered to be toxic have been identified in the water, sediment and biota of Puget Sound and/or in effluents discharging into the Sound (E. Long, NOAA, pers. comm., 1985). Neither convenient, inexpensive analytical techniques for routinely monitoring this host of contami-nants nor the toxicological information to even estimate their potential hazards to biota and their consumers are available. 105 Despite this apparent complexity and uncertainty, scientific investi-gation has determined that many shared physical and chemical properties of organic pollutants permit a broad understanding of the processes and pathways that effect their fate in the environment. Of particular interest to the present thesis is an understanding of the processes that effect water column release and biological uptake of sediment bound contaminants. Cognisance and avoidance of environmental conditions which promote contaminant release to the biological system is a prerequisite to appropriate environmental management. The schematic representation of contaminant transport and fate in the marine environment that was described above in Figure 3 (see Section 2.3) applies equally to organic contaminants as to trace metals. The sediments act as a "sink" to the organic contaminants of concern, but the water column and biota can also become enriched. When con-taminated dredged materials are disposed into a new environment, a new equilibrium between the respective concentrations in the sedi-ments, the water and the biota may be established. As with trace metals, the important question is whether dredged material disposal operations shift organic pollutant equilibria enough to release bio-logically harmful concentrations of those pollutants. As was outlined in Section 2.4, the contaminants of interest to ocean dumping management are the relatively insoluble organic compounds which preferentially accumulate and persist in the sediments. Section 3.2 reviewed information describing the documented and suspected 106 contamination of Vancouver's sediments to compile a list of con-taminants of concern in the present instance. These included polychlorinated biphenyls, hexachlorinated butadiene, polychlorinated dibenzofurans, pentachlorophenol, hexachlorobenzene, halogenated and non-halogenated polyaromatic hydrocarbons and phthalate esters. The present section reviews the tendency of these contaminants to be released to the water column or to be accumulated by marine biota subsequent to the marine disposal of dredged sediments. Release of Organic Contaminants to the Water Column The water solubility of an organic pollutant is an important charac-teristic not only for determining its susceptibility to dissolution, but for establishing its overall potential environmental distribution and fate (Hague et al., 1980). Substances with higher water solubility are less likely to adsorb and more likely to desorb from marine sedi-ments. Water solubility can also effect possible transformation by hydrolysis, photolysis, oxidation, reduction and biodegradation in water: hydrophobic organic compounds are less susceptible to degradation than hydrophilic compounds. The partitioning tendency of an organic pollutant -between the adsorbed and solution phases is described by the equilibrium coeff i -cient (K), as in the expression: 107 K = Cone, of pollutant in sorbent Cone, of pollutant in water The Freundlich equation describes adsorption equilibria and is fre-quently used in discussions of adsorption and desorption phenomena. This equation: C s = K (Cw)'/n enables a prediction of the adsorption tendencies of a pollutant in water. C s is the amount of adsorbed chemical per unit of adsorbent, Cw is the equilibrium solution concentration of the chemical and l/n is a constant describing the degree of non-linearity (see Figure 4). In dilute systems, adsorption isotherms approach linearity (n—1 = 1). How dilute the system must be to show this limiting behaviour varies from system to system, depending on the nature of the solute and sorbent, and the type of sorption interaction. Karickoff, Brown and Scott (1979) determined that for hydrophobic organics, sorption isotherms for natural sediments are linear if the eguilibrium aqueous phase pollutant concentration is below I0"5 mole/litre or less than half of the solute water solubility (whichever is lower). Other works substantiate low loading isotherm linearity for a range of hydrophobic organics (Hassett et al., 1980; Mearns et al., 1982; Rogers, McFarlane and Cross, 1980). This low loading limit is typically met in most 108 FIGURE 4. Freundl ich Adsorption Isotherm. Equ i l ibr ium Concentration of Solute in So lut ion (Cw). 109 environmental situations. Thus, the tendency of hydrophobic organic pollutants to enter the adsorbed state is inversely proportional to its solubility. From pesticide-soil studies, numerous soil properties (organic carbon content, particle size distribution, clay mineral composition, pH, cation exchange capacity) have been identified as affecting the mechanism and/or degree of pesticide sorption (Mingelgrin and Gerstl, 1982). However, the predominance of organic carbon in "controlling" the sorption of uncharged organic compounds has been extensively documented for soils (e.g. Kenaga and Goring, 1978) and more recently confirmed for sediments (Karickoff, Brown and Scott, 1979; Brown and Flagg, 1981). This suggests that the equilibrium partition coefficient for a sediment consisting of multiple sorbent components can be reduced to a form containing only sorption to organic matter. The experimental distribution of an organic compound between the organic solvent n-octanol and water is now widely used to predict the adsorption of organic compounds to the organic portion of sediments. The log of the distribution coefficient is referred to as "log P". It is emphasized however, that the sorbed state concentrations and the thermodynamic variables derived therefrom are operationally defined. Seldom can the actual distributional volume of the sorbate be defined. The actual degree of adsorption to sediments in any instance is dependent not only on the characteristics of the pollutant (e.g. 110 polarity or solubility) and sorbent (e.g. surface area and organic content), but also on the characteristics of. the water, most notably temperature, salinity, pH and redox potential. Under static environmental conditions, hydrophobic organic con-taminants in undisturbed sediments have a strong tendency to remain associated with solids so are not free to volatilize into the air or be dispersed as a solute. Their ultimate environmental fate is essentially limited to microbial degradation and/or burial within the sediments. However if the environmental character of the surrounding medium alters, or more particularly if it is altered by relocation of the sediments into a different medium, then the kinetics of adsorption and desorption need to be considered. A great deal of confusion exists in the sorption literature regarding adsorption-desorption kinetics. Sorption is often described as rapid and readily reversible (times to reach equilibrium of minutes to a few hours). Yet as Karickhoff (1985) has pointed out, analytical chemists commonly encounter extreme difficulty extracting pollutants from field soils or sediments. Highly sorbed hydrophobic chemicals frequently require lengthy extraction periods (days), multiple solvents and abrasive mixing techniques for quantitative chemical recovery. A concept long held by soil chemists (see Rao et al., 1982) and recently revamped for pollutant sorption to sediments (DiToro and Horzempa, 1982) distinguishes the relative rates of the adsorption and desorption process. DiToro and Horzempa (1982) found that for the I l l sorption of hexachlorobiphenyl to various sorbents (sediments, clays, silica), the adsorption process was rapid (minutes to hours) but for the reverse process most of the sorbed chemical was "highly resistant" to release. Hiraizumi, Takahashi and Nishimura (1979) conducted extensive laboratory and field studies of PCB adsorption and desorption from bed sediments. PCB adsorption was found to occur rapidly, and could be approximately described by a Freundlich isotherm. Desorption rates were much slower: seabed mud released only 1% of total adsorbed PCB during the first 24 hours, and a further 0.8% during the next five days. Karickoff (1980) reported a similar finding for polynuclear aromatic hydrocarbons. Also, he freguently observed a drastic change in the ease of extraction of sorbed chemical with increased incubation time. After short incubation periods (less than 5 minutes) essentially all (greater than 90%) of the sorbed chemicals could be easily extracted from the sediment by direct extraction of the suspensions with hexane for approximately 3 minutes. The fraction of sorbed chemical recovered by this brief extraction decreased to 50% after 3-5 hours of incubation and continued to decrease slowly over the next few days, levelling out in the range of 20-40%. Karickoff (1985) concluded that these findings suggested a diffusion controlled process in the sorbed state where sorbed chemical is slowly incorporated into release resistent components of the sorbent. 112 Accordingly, it has been generally assumed that desorption of hydrophobic contaminants from resuspended dredged material is a relatively minor pathway of contaminant release. The likelihood of insoluble organic chemicals desorbing from riverine sediments disposed into the marine environment is further reduced by the "salting out" effects of increased salinity upon hydrophobic reactions. Wildish et al. (1980) studied the adsorption-desorption of the PCB mixture, Aroclor 1254, by sediment over a range of salinities. Adsorption was found to increase with increasing salinity. Concentration factors (i.e. sediment versus water concentrations) for I gram of silt-clay exposed to 0.2 ug/L Aroclor 1254 were: 810 for deionized water; 2710 for half strength seawater; and 4240 for full strength seawater. The concentration factors for desorption were greater than for adsorption, and sediment adsorption of PCBs in seawater was considered to be essentially irreversible. Field confirmation of these observations as they pertain to disposal of dredged sediments has been difficult due to the difficulty in obtaining a water sample for analysis that contains only soluble organics. Insoluble material cannot be removed by filtration because of the probability of soluble organic contaminants adsorbing to the filter. A preferable method is high speed centrifugation, but the results of this methodology can be substantially exaggerated by contamination with suspended particles and oil that have specific gravity equal to or less than the water phase (Hartung and Klinger, 1970). 113 Fulk, Gruber and Wullschleger (1975) conducted extensive laboratory studies on the desorption of a variety of organohalogens from dredged materials sampled from throughout the United States. PCBs were present in 59 of the 64 samples. Dieldrin and DDT were also commonly present and aldrin and 2-4-T were each detected in two samples. The objectives of the study were to determine the transfer of these materials to the water column during simulated disposal operations. A significant finding of this study was that pore water concentrations of hydrophobic organics in the dredged sediments were extremely low: the maximum concentrations were 0.15 ug/L for PCBs and DDT. With such low concentrations, the release of some of the pore water to the surrounding environment during disposal operations would result in negligible contaminant release. Analysis of the water column after simulated disposal found no soluble pesticides or PCBs. Pesticides and PCBs which were released were all associated with suspended particulate material and were in low concentrations. The maximum concentrations determined were 0.04 ug/L for PCBs and 0.004 ug/L for dieldrin. The relative release of PCBs ranged from 0.4 - 1.8 percent of the total PCB concentrations in the sediments. These results strongly support the assumption that hydrophobic organohalogens are rapidly and strongly sorbed from agueous solution and tend to remain associated with the sediments upon disposal. 114 In their review of research studies conducted by the U.S. Army Corps of Engineers Dredged Material Research Program, Burks and Engler (1978) concluded that desorption of chlorinated pesticides and PCBs was negligible, and that any desorbed material would be rapidly resorbed by suspended particles. Lee et al. (1978) showed that this behaviour is generally the case for chlorinated pesticides during the Elutriate Test. However releases of PCBs ranging from 1.3 to 6.9 times the concentrations in the receiving waters were found by these authors in some tests. Lee et al. (1978) found that the release was inversely related to the oil and grease content of the test sediments, i.e. sediments low in oil and grease released the largest quantities of PCBs. In field studies by Lee et al. (1975), no significant release of organohalogen pesticides or PCBs was detected during disposal of highly contaminated dredged sediments. The authors concluded that the PCB release shown in the Elutriate Test was not detected in the field due to rapid mixing and dilution of the very small quantities which may have been released. Blazevich et al. (1977) reported the results of the U.S. EPA's moni-toring of the disposal of PCB contaminated sediments into Puget Sound. In 1974 there was a spill of almost 1000 litres of PCBs into the maintenance dredging area of the Duwamish River. The extremely contaminated sediments were hydraulically dredged and placed into an impervious upland containment area, while I 14,000 m3 115 of the remaining contaminated material was removed by a clamshell dredge and disposed from barges at an experimental disposal site in Ell iott Bay, Puget Sound. Water quality monitoring showed that there were very slight increases in water column PCB concentrations associated with the disposal operation. These changes occurred with increases in suspended particulate matter and, when the particulate matter decreased, so did the PCB concentration. The increases were of short duration, usually less than 30 minutes. There was no significant uptake of PCBs by fish or most invertebrates analysed during or after the disposal operations. In addition, caged animals were held at the disposal site for up to 3 weeks. Mussels in cages at the disposal site accumulated PCBs to levels above background, but the increase was not statis-tically significant. It was pointed out however, that some of the animals collected from Elliott Bay prior to disposal contained sub-stantial amounts of PCBs, so an appreciable uptake would have been necessary for statistical significance. Similar studies by Arimoto and Feng (1983) in Long Island Sound were confounded by PCB levels in the water column due to discharge from an adjacent river. Although significant uptake of PCBs by mussels occurred coincident with the disposal operation, multiple regression analysis could account for only 20 - 40% of the observed variance in concentrations. The authors concluded that dumping was only a minor influence on the PCB concentrations of the monitoring animals, even for those living on or adjacent to the dumpsite. 116 Boehm and Fiest (1983) conducted extensive analyses of water samples from New York Bight prior to, during and after a dredged material disposal operation. The dredged material contained total PCB concentrations ranging from 3700 - 6900 ug/kg. Individual P A H concentrations ranged from 800 - 7000 ug/kg. Sampling at 15 minutes after the dump detected negligible concen-trations of PCBs or PAHs in the water column, except at one station within the particulate plume near the seafloor. Within this plume, the concentration of PCBs was 0.058 ug/L and total PAHs were 1.8 ug/L. After 2.5 hours, concentrations at this station had decreased but remained detectable: PCB levels were 0.038 ug/L and total PAHs were 0.45 ug/L. For perspective, the U.S. EPA (1978) Ambient Water Quality Criteria recommends that for the protection of marine life, PCB concentra-tions should not exceed a 24 hour average of 0.024 ug/L nor 0.20 ug/L at any one time. Similarly for P A H levels: of the individual P A H compounds analysed by Boehm and Fiest (1983), flouranthrene was prominent in both the source material (5.8 - 6.9 mg/kg) and in the water column sediment plume 15 minutes after disposal (0.1 ug/L). The U.S. EPA (1978) criteria for flouranthrene suggest a 24 hour average seawater concentration of 0.30 ug/L with a maximum of 0.69 ug/L at any time. 117 It is concluded that available evidence concurs with the general theory that hydrophobic organic pollutants are not readily released into solution subsequent to dredged material disposal. Field monitoring of ocean dumping operations have not detected environmentally signi-ficant release of these pollutants in either the sorbed or desorbed state. However the confidence that can be expressed in these results is low considering the background interference that was often present. Laboratory tests indicate that there is a significant potential for environmental impact from disposing contaminated sediments due to resuspension of sediments. It is therefore strongly recommended that the disposal of dredged materials contaminated with organic pollutants should be managed to minimize losses of suspended or resuspended particulates to the water column. Biological Uptake of Organic Contaminants Accumulation within biotic tissue is a significant fate of many hydrophobic organic contaminants within the aquatic and marine environment. Relative to the very low concentrations that may be released into solution, the absorption of hydrophobic chemicals into biotic tissue, and their subsequent transport from the disposal site via the food chain, may be an important source of their release from disposed sediments. The fact that contaminants released in this manner are incorporated within animal tissue adds to the environ-mental significance of this release process. 118 Contaminants may enter marine organisms from the aqueous phase, the sediment, or their food. Whatever the source of entry, a number of fundamental events are involved in the accumulation process (MacKay, 1982): o partitioning of the pollutant compound hetween the source and some surface of the organism o diffusive transport of the pollutant compound across cell membranes. (Klaassen (1980) also describes various active transport mechanisms which have been postulated), o transport mediated by body fluids, such as exchange between blood vessels and serum lipoproteins o concentration of the compound in various tissues, depending on its affinity for certain biomolecules, such as nerve lipids o biodegradation of the compound. The bioaccumulation process is thus seen to be a result of both kinetic (diffusional, transport and biodegradation) and equilibrium (partitioning) processes. For persistent organic pollutants in the marine environment, biological uptake and accumulation is largely determined by their relative chemical activities in water and in tissues (MacKay, 1982). The time required to approach steady state tissue concentrations is determined by kinetic processes, and can range from a few hours to several weeks (Southward et al., 1978; Branson et al., 1975). However, the more hydrophobic the compound, the greater is its tendency to move through biological membranes. 119 The rate and the extent of bioaccumulation of organic molecules is thus greatest for lipophilic (or hydrophobic) compounds. Although the pharmaceutical industry first drew attention to the link between lipid solubility and biological effectiveness of organics at about the turn of the century (see Tute, 1971 for a review of this topic), the environmental sciences have only turned their attention to this topic in the last fifteen years. The uptake and location of persistent organic compounds in the marine fauna has now been shown to depend upon the amount and distribution of lipid content (McFarland, Gibson and Meade, 1984). Stout (1980) and Schneider (1982) observed that tissue concentrations of hydrophobic chemicals in field sampled fish could be directly correlated with lipid content. Schneider (1982) also concluded that differences in biological concentration factors of chemicals in different organisms were due to differences in the lipid pool. McFarland, Gibson and Meade (1984) found that the equilibrium levels of PCBs in fish and clams were equivalent when expressed relative to lipid content. The common approach to a quantification of lipid solubility is to determine the experimental octanol/water partition coefficient (P), the alcohol being the model for biological lipid tissue. Metcalf et al. (1975) correlated the bioaccumulation of a number of organic com-pounds (tri- tetra- and pentachlorobiphenyls, DDT, chlorobenzene, benzoic acid, anisole, nitrobenzene, aniline) from the fish of model ecosystems with the octanol/water partition coefficient. For the limited number of compounds included the correlation was excellent. Konemann (1979) reported a non-linear relation between log P and the 120 log of the bioconcentration factor for chlorobenzenes in fat of guppies. Bioaccumulation approached a maximum value when log P = 6.5, further bioaccumulation being limited by uptake kinetics (Figure 5). Mackay (1982) considered that for high log P values (i.e. greater than 6) bioaccumulation potential data is uncertain. Reasons given were that: o direct equilibrium measurement of partition coefficients are difficult due to the very low water phase concentrations o time to reach steady state is extremely long due to membrane permeability resistance to large molecules. This means that growth and lipid deposition need to be considered in the analysis. For compounds with log P less than 6, Konemann's (1979) results evidenced an essentially linear relationship between bioaccumulation and P values (Figure 5). This relationship has since been confirmed by Mackay (1982) and Esser and Moser (1982). It is now generally accepted that for lipophilic organic compounds with log P values between 2 and 6, bioaccumulation potential may be reasonably pre-dicted from the octanol/water partition coefficient. The American Institute of Biological Sciences (1978) Aquatic Hazards of Pesticides Task Group has recommended that compounds with log P greater than 3 be considered to have high bioaccumulation potential. Currently the state of the art is such that theoretical calculations of log P can be made solely from consideration of the structure of the organic molecule. The derived log P is expressed in terms of additive 121 FIGURE 5 . . B i o a c c u m u l a t i o n o f S i x C h l o r i n a t e d Benzenes i n G u p p i e s . (from Konemann, 1979.) A. R e l a t i o n s h i p between l o g b i o a c c u m u l a t i o n and l o g P. Log P B. R e l a t i o n s h i p between uptake r a t e c o n s t a n t and l o g P. 122 structural contributions from different parts of the molecule, using earlier experimental data as sources of log P contributions for these different molecular parts (Verschueren, 1983). Similarly, the behaviour of organic compounds in the aquatic environment can increasingly be predicted through theoretical application of established structure-reactivity relationships (Wolfe et al., 1980). The overall performance of such theoretical estimates has been encouraging, but the application of experimentally or theoretically derived log P values for predicting contaminant concentrations in biota in field situations is not straightforward. The uptake of lipophilic contaminants from the environment correlates with the partitioning characteristics of the chemical, but the resulting tissue concentrations in biota are strongly influenced by characteristics of the individual organisms. Thomann (1978) proposed a mechanistic model which allows inter-pretation of the variability between different organisms within a single coherent framework. According to this model, the accumulation of contaminants in organisms is a balance between the intake rate, the storage capacity of the organisms and the elimination rate. For smaller organisms such as zooplankton, the data indicate that con-taminant residues are capable of facile exchange between the sur-rounding water and the internal sorption sites, primarily lipids. Residues may be accumulated from food but the facile exchange results in the equilibration of the residue levels in the organisms with 123 the levels in the surrounding water. The ultimate body burden thus reflects essentially a partitioning of the residues between the lipid and water phases (i.e. log P for the contaminant). At the other end of the scale, larger organisms do not have as great an exchange surface relative to their body volume, so maintain an internal environment which has greater independence from the external medium. This is particularly true for the marine mammals, which do not have gills, an effective exchange organ in fish. For these organisms, exchange with the surrounding water is considerably less important than is intake of contaminants from food and el imi-nation primarily with excretia. Since elimination via the latter route is relatively inefficient (e.g., seal scat were found to have about the same PCB levels as observed in fish food; Calambodikis et al., 1978), some portion of the residues in food is retained. The concentrations in these animals consequently tend to increase with time, reflecting the consumption of greater guantities of food. Studies in Puget Sound appear to confirm Thomann's (1978) model, especially for the two extremes of equilibration and food chain transfer. In one study, the concentrations of PCBs in zooplankton (copepods, euphausiids and ctenophores) were analysed as functions of a variety of habitat and physiological variables (Clayton et al., 1977). The lipid weight normalized data yielded rough correlations between the concentrations of PCBs in the water and in the organisms. The relative accumulation was greater for more chlorinated PCBs (i.e. higher log P). 124 Similar behaviour was noted in studies on the depuration of PCBs by the mussel Mytilus edulis (Calambodikis et al., 1979). Mussels co l -lected from a heavily polluted area of Puget Sound were caged and placed in a relatively pristine area of the Sound. The high levels of PCBs originally present in the mussels decreased logarithmically with time of exposure to the cleaner environment. More highly chlorinated PCB isomers were retained longer, with depuration half lives esti-mated to range from 3 days for d i - and tri-chloro compounds to 50 days for hexa- and hepta-chlorobiphenyls. At the other end of the scale, the concentrations of PCBs in the blubber of Harbour seals from Puget Sound and the Washington coast were found to increase with the age of the individual (Calambodikis et al., 1978). The concentrations in the blubber were greater than observed in the fish species known to provide food for the seals, even when the latter levels were normalized to a lipid weight basis. Burns and Teal (1979) similarly found more oil residues in the herring gull than was contained in their food after exposure to the West Falmouth oil spill, but no evidence for this food chain magnification among invertebrates or fish in the study area. According to Thomann (1978), between the two extremes of equili-bration and food chain transfer there exists a continuum of rela-tionships, primarily dependent on the size of the organisms. Thomann (1978) considered that the continuum shifts as a function of increasing 125 size of the organisms from the dominance of facile sorptive/desorptive equilibrium exchange to dominance by food intake and reduced elimination rates relative to the rate of accumulation. Although recent data has generally confirmed Thomann's model, they have shown that he over-estimated the role of dietary intake and under-estimated the role of facile desorptive equilibrium exchange in large water breathing (i.e. non-mammal) organisms. Relative to air breathing animals, all water breathing organisms have large respiratory surfaces in proportion to their body size. The solubility of oxygen in water, especially seawater, is low. Therefore large quantities of water must be passed over their gill surfaces to provide adeguate oxygen for respiration, simultaneously increasing the opportunity for facile eguilibrium exchange of contaminants with the surrounding medium. Even when dietary intake of contaminants is high, the opportunity for desorptive equilibrium exchange via the gills will also be high. In addition to this, the consistent dietary intake of highly conta-minated prey items that results in food-chain biomagnification of contaminants in terrestrial systems (where contaminant levels in prey items are successively magnified at successive trophic levels) rarely occurs in the marine ecosystem. Food-chain biomagnification reguires a food pathway that is essentially linear and highly structured, with the predominant energy flow from lower to higher trophic levels. Most marine food webs are rather weakly structured and do not have trophic levels as clearly defined as those of terrestrial systems. One species may occupy several trophic levels during its lifetime due to 126 different feeding habits at different stages of its life cycle. Energy flow in marine food webs is multidirectional with, for example, crabs being both prey to and scavengers of fish; a large component of the energy in marine systems is bound with the sediments. The combination of intimate physical contact with the external medium and a complexly interactive trophic web is why trace contaminants do not increase nearly as much with trophic levels in aquatic systems as in nonaquatic systems. Kay (1984) and Biddinger and Gloss (1984) provided comprehensive reviews of the potential for persistent organic compounds to biomagnify up the food chain. They concluded that where biomagnification did occur amongst gill bearing organisms, it produced concentrations in the order of only one to ten times higher in the upper trophic levels than in the lower ones, in contrast to the tens or hundreds of times higher as has occurred with DDT in fish eating birds. The application of Thomann's (1978) model to different organic compounds in the marine system therefore mostly involves con-sideration of contaminant transfer processes that are dominated by equilibrum partitioning processes. These involve the same energetic considerations discussed for sorption onto the sediments and parti-culate matter. Irrespective of the source of intake, the potential for contaminants to concentrate in biological tissue will essentially reflect preferential accumulation of less soluble compounds. For the hydrophobic pollutants that are of concern to dredged material management this potential will be high. Information describing the 127 partitioning characteristics and exampling the bioaccumulation potential of each of the organic contaminants identified in Section 3.2 is therefore summarized below in Section 5.5. Biological Transformation Abiotic transformations of organic compounds (e.g. hydrolysis and photolysis) are generally high energy reactions involving heat or light (Larson, Blankenship and Hunt, 1976). For sediment sorbed hydro-phobic compounds in the aguatic environment, these processes are generally considered to be negligible (Verschueren, 1983). However biological transformation and biodegradation of organic pollutants by microorganisms and higher marine organisms are potentially more significant modification and release processes. Recent research has shown that most organic contaminants can be transformed in organisms by several major reaction pathways in which they are oxidized, hydrolyzed or reduced (Phase I reactions), or conjugated (Phase II reactions), or both Phase I and Phase II reactions occur (Connell and Miller, 1984). When a pollutant undergoes chemical change, the reaction products differ from the parent compound in toxicity, bioaccumulation potential, solubility, volatility, etc., so such changes may be of great importance. 128 5.4.1 Microbial Degradation Many marine and estuarine microorganisms rapidly oxidize the less complex fractions of petroleum hydrocarbon mixtures (e.g. n-alkanes), eventually to carbon dioxide and water. But while the n-alkanes are oxidized quite readily, the rate of this process decreases markedly with increasing branching. GESAMP (1977) conducted extensive research into the microbial degradation of petroleum hydrocarbons, and found that the rate of degradation of complex fractions (e.g. cyclo-alkanes and PAHs) can be exceptionally slow. Gibson (1976) and Furukawa and Matsumura (1976) similarly determined that bacteria have limited capacity to degrade complex arenes and more highly chlorinated PCBs. Steen, Paris and Baughman (1980) found that the sorption of phthalate esters and other nonhalogenated industrial compounds to suspended sediments rendered them unavailable for degradation in the adsorbed state. Concentration gradients of organohalogen contaminants in sediment cores from polluted embayments of Puget Sound have been shown to generally correspond with the probable long term historical input rates, i.e. increasing levels in more recent sediment (Horn, 1978; Dexter et al., 1979). The same profiles would tend to be produced by more complete degradation with time of burial, but the residues in the cores did not indicate the preferential reduction in the relative concentrations of the less chlorinated components which should be induced by metabolism. 129 The concentrations of non-halogenated PAHs in Puget Sound sediment cores showed similar general decreases to low levels at depths in the sediments corresponding to pre-1900 in dated cores (Carpenter and Fairhall, 1979). However many of the individual P A H components varied with depth, some showing subsurface maxima. This was not readily explanable from input concentrations, and the investigators felt it could not be clearly established that P A H levels had not been significantly altered by metabolism. No data could be found on the microbial metabolism of many of the contaminants of concern to ocean dumping management. However, studies of the kinetics of microbial degradation of organics in natural water samples have shown significant correlations between degradation rates and chemical structure (Paris et al., 1975; Wolfe et al., 1980). These correlations tend to confirm the generally held belief that the more highly chlorinated or more complex organic compounds (which are of concern to ocean dumping management) are not susceptible to appreciable degradation by microbial organisms. 5.4.2 Metabolism by Higher Organisms Many higher organisms, particularly vertebrates, possess metabolic pathways capable of substantial alteration of organic contaminants. The general mechanism for biotransformation of hydrophobic organics consists of two phases: formation of metabolites, usually by mixed-function oxidations, and transformations into more hydrophilic conjugates (Connell and Miller, 1984). Although the majority of these 130 reactions result in detoxification of the contaminants, some can form reactive metabolites that have been implicated as primary carcino-gens, mutagens and cytotoxins. The formation and excretion of biologically active, hydrophilic metabolites by biota that inhabit a dredged material disposal site may conseguently be of ecological significance. Metabolism and degradation of organic contaminants by invertebrate and vertebrate biota is highly complex, and appears to be species specific. Although our understanding of the multiplicity of processes that occur is far from comprehensive, it is generally understood that the rate of metabolism of organic contaminants in any organism is related to the chemical structure of the contaminant. While there is probably no compound totally resistant to metabolic alteration, the more highly chlorinated and more complex organic contaminants appear to be less prone to metabolism. For example, while many marine organisms, particularly vertebrates, possess metabolic pathways capable of rapidly altering simple petroleum hydrocarbons (Varanasi and Malins, 1977), PCBs containing more than four chlorine atoms are essentially resistant to metabolism (Sundstrom, Hutzinger and Safe, 1976). Sundstrom et al. (1976) also suggest that for less chlorinated PCBs, the stereochemistry of the individual molecule has a significant impact on the metabolism process. Circumstantial evidence indicates that the biological degradation of several of the hydrophobic contaminants of concern to ocean dumping may be significant. For example, studies by Malins et al. (1980) 131 showed that while tetrachlorobutadienes often dominated in sediments of Puget Sound, hexachlorobutadiene generally dominated in biotic tissue. This latter component was often the only detected chloro-butadiene in groundfish tissue. These data were interpreted to show selective metabolism of the less chlorinated butadienes. It is concluded that the release of metabolites of the more highly halogenated and/or complex organic contaminants of concern is unlikely to be a major fate process. However the fact that meta-bolites which are formed will be released to the water column as a more hydrophilic conjugate substantially increases the environmental significance of this release. Information describing the metabolic and degradation products of each of the organic contaminants identified in Section 3.2 is therefore summarized in the following section. Properties, Bioaccumulation and Metabolism of Organic Contaminants of Concern The foregoing review of environmental processes which promote the release of organic contaminants from disposed dredged material has identified three potentially significant release pathways: o the erosion and/or resuspension of particulates to which the contaminants are sorbed o accumulation of contaminants in biological tissue and release via the food chain 132 o biological transformation and release of soluble metabolites to the water column. The most substantial of these pathways is potentially the foremost, and it is recommended in Section 5.2 that disposal operations should be managed to minimize these losses. The degree to which this is possible is considered in Section 6. The significance of the latter two release processes is highly dependent upon the chemistry of the individual contaminants. This section therefore summarizes information describing the relevant properties, bioaccumulation potential and metabolic/depuration pathways of each of the contaminants that were identified in Section 3.2 to be of concern to ocean dumping management. 5.5.1 Polychlorinated Biphenyls (PCBs) Properties There are 209 possible compounds and isomers resulting from the chlorination of biphenyl with I to 10 chlorine atoms. In commercial PCB mixtures, about 100 individual compounds and isomers have been detected (Hutzinger et al., 1974). PCB is no longer manufactured in North America. The major producer of PCB mixtures was the Monsanto Company, which marketed PCB mixtures for closed electrical system applications under 133 the Aroclor trademark. A numbering code was used to designate the Aroclor products. The first two digits, 12, indicate that the parent compound is biphenyl. The last two digits specified the percentage by weight of chlorine. Solubility data for PCBs vary considerably for individual isomers. For example, a review by Tulp and Hutzinger (1978) found published solubility values for 2-chlorobiphenyl that varied between 900 and 5,800 ug/L and for 2,2',3,3',4,4',5,5 ,,6,6'-decachlorobiphenyl the variation was from 0.016 to 15 ug/L. Tulp and Hutzinger (1978) expressed their concerns about some of the methodologies used for determination of PCB solubilities. They stated that most of the published solubility data on Aroclors is based on selective solubilization of the lower chlorinated components from the mixtures. If the solubilities of individual components of a mixture are not identical, the solubility of such a mixture in water cannot be defined as a physical constant. Nonetheless, the reported solubility values of PCB mixtures and their components illustrate their hydrophobic tendencies. The experimental solubilities of PCB mixtures in fresh water, as reported by various investigators, vary between 200 - 340 ug/L for Aroclor 1242 and 2.7 -25 ug/L for Aroclor 1260 (MacKay and Leinonen, 1975). Solubilities in seawater would be less than those reported for freshwater. The partition coefficients of PCBs indicate that the components are also extremely lipophilic. Reported log P values range from 4.0 for monochlorinated isomers to as high as 6.85 for pentachlorinated 134 isomers (Callahan and Slimak, 1979). Where partition coefficients are reported for Aroclors, it should be noted that the more highly chlorinated compounds will more selectively partition within the n-octanol phase; therefore, partitioning will differ for each of the components of Aroclor mixtures. As a result, Tulp and Hutzinger (1978) concluded that log P values of PCB mixtures cannot be considered as physical constants. Bioaccumulation Bioaccumulation of PCBs in pelagic organisms appears to occur primarily through direct uptake of PCBs from water and is strongly dependent upon lipid levels in organisms. Zitko (1974) showed that with juvenile Atlantic salmon, the accumulation ratios of Aroclor 1254 from water and food were 950:1. Clayton et al. (1977) recorded a 106 amplification of chlorinated biphenyls by zooplankton above the levels observed in Puget Sound waters. McDermott, Young and Heesen (1976), reported a 10^ bioconcentration of Aroclor 1254 in the mussel (Mytilus californianus). This amplification was considered a function of the lipid fraction in the organisms. Food chain bio-magnification was determined not to be a controlling factor. There is uncertainty concerning the uptake mechanism for PCBs in benthic species such as clams, crabs, and polychaetes. Stainken and Rollwagen (1979) studied the uptake of PCBs by three species of bivalves (C. virginica, M. arenaria, and M. mercenaria) from sediments of Lower New York Bay. No valid relationship could be found 135 between sediment silt-clay content and sediment PCB content nor between tissue PCB content and sediment PCB content. Generally, tissue levels at all sites were greater than the residue values in sediments. However, Malins et al. (1980) found that bottom-dwelling fish and benthic invertebrates sampled from areas in Puget Sound with the highest PCB levels in sediments had the highest tissue levels of PCBs. Uptake obviously does occur, but it is uncertain whether uptake occurred via absorption from interstitial waters or from ingestion of solid materials. Although dietary intake may not be important in the uptake of PCBs by fish, marine mammals can readily accumulate PCBs contained within diet. Herman and Calambodikis (1978) reported a level of 38 mg/kg PCBs in blubber of a killer whale which was found washed ashore in 1977 in the Strait of Georgia. Few, if any, direct, acute effects on aquatic organisms have been recorded as a result of PCB contamination in North America. However PCBs do have sublethal effects on higher food chain organisms. Such effects include eggshell thinning of fish-consuming birds; behavioural alterations in birds; and reproductive failure of such wildlife as mink (Stendell, 1976). 136 Metabolism and Depuration Sundstrom, Hutzinger and Safe (1976) prepared an extensive review on biological metabolism of chlorobiphenyls. Generally speaking, biphenyls with four or fewer chlorine substitutions can be metabolized by a number of test organisms. Metabolites of PCBs by mammalian species are predominantly monohydroxy and di hydroxy compounds. Highly chlorinated compounds do not appear to be metabolized. Carey and Harvey (1978) found no evidence for metabolism of chloro-biphenyls in anaerobic marine mud, indicating that anaerobic environments may serve as long-term sinks for PCBs. Eder (1976) found that the percentage of lower-chlorinated PCBs in anaerobic sediments of the Norwegian Depression was higher than in well-aerated marine sand suggesting a greater degree of biological degradation in the aerated sands, if the input ratios were identical for both areas. 5.5.2 Hexachlorobutadiene (HCBD) Properties Callahan and Slimak (1979) reported that HCBD has low water solubility (2 mg/L) and high lipophilicity (log p = 3.74). Studies in Puget Sound showed that HCBD tends to be associated with suspended particles at concentrations about four orders of magnitude higher than the surrounding water (Riley et al., 1980). Pearson and McConnell (1975) described HCBD as being "tenaciously adsorbed" on sediment. 137 Bioaccumulation There does not appear to be a pattern of extensive bioaccumulation of HCBD in marine food chains (Pearson and McConnell, 1975). Confined uptake studies, such as laboratory or jn situ studies in contaminated areas, have generally shown erratic bioconcentration factors. For example, Laseter et al. (1976) exposed crayfish in cages to waters and sediments contaminated with HCBD, and found that the bioconcentration factor varied from 7.8 to 300 among different test animals during the same time period. No explanations were suggested by the investigators. The log P of chlorinated butadienes suggests moderate retention of the compounds by biota in the marine environment. Pearson and McConnell (1975) exposed dab and plaice to 1.6 ug/L HCBD over 39 days and 106 days, respectively. They recorded HCBD accumulation in the livers of dab by a factor of 7,000, and in plaice by a factor of 10,000. Negligible uptake was observed via contaminated food over a period of 88 days. The lack of bioaccumulation of HCBD via the food chain was also observed by Goldbach et al. (1976) and Leeuwangh et al. (1975). Pearson and McConnell (1975) found that mussels exposed to 1.6 ug/L HCBD bioconcentrated the compound by factors of 900 - 2,000 within a period of up to 50 days. The largest accumulation of H C B D occurred in the digestive gland and relatively less accumulation occurred in the foot of the mussel. 138 Laseter et al. (1976) showed that crayfish exposed to 2.97 ug/L HCBD concentrated the compound primarily in brain tissue (11,875 times) and the green gland (737 times). Concentration factors in hepatopancreas were 58; gills, 56; and muscle, 15. Bass exposed to 33 ug/L HCBD had highest concentrations of HCBD in its gut (1610 times) with lower concentrations in liver, kidney and brain. It is uncertain if the lower concentrations of HCBD in liver of the test organisms was due to metabolism by the liver. Metabolism and Depuration No experimental studies regarding biological metabolism of chlorinated butadienes were found in the literature. However as was described in Section 5.4.2, circumstantial evidence by Malins et al. (1980) indicated selective metabolism by fish of less chlorinated butadienes. 5.5.3 Polychlorinated Dibenzofurans (PCDF) Properties There are 135 possible isomers of chlorinated dibenzofurans (Rappe et al. 1978). No information was found on the solubilities of any chlorinated dibenzofurans in water. However the solubilities are expected to be lower than 3.1 mg/L, the solubility of dibenzofuran in freshwater at 25oC (Lu et dl., 1978). On the basis of the log 139 n-octanol/water partition coefficient of 4.12 for dibenzofuran (Leo et al., 1971), the tetra- and pentachlorodibenzofurans are theoretically very lipophilic. Bioaccumulation Few studies were found in the literature on bioaccumulation of PCDFs. Lu et al. (1978) followed the passage of dibenzofuran through a model aquatic system by applying this compound to sorghum leaves which, when eaten by salt marsh caterpillars, resulted in its release into water containing plankton, algae, Daphnia, snails, mosquito larvae, and mosquito fish. They found that the dibenzofuran was relatively water soluble, was not degraded, and exhibited substantial accumulation. An ecological magnification factor of 947 was determined for fish. The authors noted a good fit between log octanol/water partition coefficients and bioaccumulation by fish. A higher ecological magnification factor would be expected for chlorinated dibenzofurans. Lake et al. (1981) exposed caged mussels (Mytilus edulis) in various areas of Narragansett Bay, Rhode Island, and found from 643 to 1,180 ug/kg (dry wt) of 2,4,8-trichlorodibenzofuran in the mussel tissues from the areas of the Bay described as highly polluted. • Lobster hepato-pancreas from the lesser polluted areas where mussels had a 52 ug/kg level of trichlorodibenzofuran, contained 1,260 ppb trichloro-dibenzofuran. The authors concluded that "the high toxicity generally 140 associated with PCDFs calls for futher documentation of their dis-tributions throughout the coastal United States and for rapid identi-fication and abatement of their sources". Metabolism and Depuration Information on metabolism of PCDFs was sparse. On the basis of structural similarity to dibenzofurans, it appears that chlorinated dibenzofurans would not be appreciably biologically degraded in the environment (Lu et al., 1978). 5.5.4 Pentachlorophenol (PCP) Properties The fresh water solubility of PCP is 20 mg/L at 25oC (Bailey and White, 1965). The log P for pentachlorophenol has been reported to be 3.4 (Lu and Metcalf, 1975). Bioaccumulation Lu, Metcalf and Cole (1978) used a model aquatic ecosystem con-sisting of a 3 liter flask containing Daphnia, mosquito larvae, snails, mosquito fish and green algae, to study the uptake of radiolabeled PCP. After 48 hours exposure to 0.01 to 0.1 mg/L levels of PCP, all organisms were found to contain the tracer. The bioaccumulation factor in the fish was calculated to be 296. The other organisms 141 took up less quantities and had bioaccumulation factors of Daphnia, 165; mosquito larvae, 16; snail, 121; and algae, 1.5. Kobayashi (1978) reported that goldfish exposed to 0.1 mg/L of PCP for 120 hours showed a bioaccumulation factor of approximately 1000. Metabolism and Depuration Kobayashi (1978) showed that PCP absorbed by goldfish, and also by clams, was guickly excreted into the surrounding water, mostly as pentachlorophenyl-sulfate. Tissue concentrations dropped to half after 10 hours, and approximately one-fifth after 20 hours depuration in clean water. The author concluded that PCP will not remain at high concentrations in fish and shellfish for a long time without causing death. Schimmel, Patrick and Faas (1978) similarly found that oysters purged themselves of PCP in tissues within four days after the cessation of PCP delivery. Microbial degradation of PCP has also been shown to be relatively rapid (K. Hall, Westwater, pers. comm., 1985). 5.5.5 Chlorinated Benzenes Properties No solubility values for chlorinated benzenes in marine waters were found in the literature, but they are expected to be less than in freshwater where they range from approximately 140 mg/L for dichlorobenzene down to 6 ug/L for hexachlorobenzene (Callahan and 142 Slimak, 1979). Increased chlorination results in decreased solubility and increased lipophilicity. The log P for hexachlorobenzene has been reported to be 6.18 (Callahan and Slimak, 1979). Bioaccumulation Few data for bioaccumulation of chlorinated benzenes by marine biota were found in the literature. Veith et al. (1980) determined bio-concentration factors from 60 to 89 for three isomers of dichloro-benzene in bluegill sunfish during a 14-day test. Macek et al. (1977) investigated the relative significance of aqueous and dietary uptake of 1,2,4-trichlorobenzene using the water flea, Daphnia magna, as a food source for the bluegill sunfish, Lepomis macrochirus. D. magna exposed to 3 ug/L in water accumulated up to 0.44 mg/kg in tissues. L. macrochirus accumulated to 0.53 mg/kg from water alone, 0.03 mg/kg from food alone, and 0.57 mg/kg from water and food. Thus, the authors concluded that the dietary contribution was insignificant. Similar results were shown by Zitko (1977) in a freshwater study on juvenile Atlantic salmon (Salmo salar). Amounts adsorbed from water and from food were in the proportion 82.6%/17.4% indicating the importance of the water medium to bioaccumulation of HCB. Zitko (1977) reported an accumulation coefficient for HCB of 753 by salmon. 143 The highest bioaccumulation value found in the literature was 44,000, which was reported by Laseter et al. (1976) for juvenile bass. In field experiments, Laseter et al. (1976) found that resident crayfish which were exposed to 44 ug/L of HCB in natural waters had whole body levels of 1164 times this level. Metabolism and Depuration Depuration rates of bioaccumulated chlorinated benzenes are dependent upon the degree of substitution on the benzene ring. For example, Veith et al. (1980) showed that the half lives of dichloro-benzenes in tissue of bluegill sunfish were less than one day. Tetrachlorobenzene had a half life between 2 and 4 days and pentachlorobenzene had a half life greater than 7 days. Lu and Metcalf (1975) suggested that chlorobenzene is a persistent chemical which is not readily biodegraded by microorganisms unless they are already growing on another hydrocarbon source. Thorn and Agg (1975) found 1,2- and 1,3-dichlorobenzene to be resistant to microbial biodegradation. Safe et al. (1979) found that in higher organisms, chlorinated benzenes are metabolically degraded to hydroxylated and conjugated metabo-lites. These authors found that tetrachlorobenzene can form tetrd-chlorophenol in frogs. Pentachlorophenol is produced by rats as a metabolite of HCB (Koss et al., 1977). 144 5.5.6 Polyaromatic Hydrocarbons (PAHs) There is a very diverse range of compounds within this class of chemicals, not all of which could be reasonably reviewed within the scope of this study. Therefore, following some observations regarding the general properties of PAHs, the properties and bioaccumulation potential of several common PAHs, benzo(a)anthracenes, flouranthenes and benzo(a)pyrene, and the commonly detected polychlorinated naphthalenes, are reviewed as examples. General Properties Solubilities and partition coefficients of PAHs in freshwater and seawater are generally dependent upon molecular weight. Alkylation of a parent aromatic compound will decrease solubility and increase lipophilic properties. For example, the addition of two methyl groups decreases the solubility of naphthalene in seawater from 20 mg/L to 2.4 mg/L (Rossi and Neff, 1978). Addition of a third methyl group decreased the solubility to 1.7 mg/L. Positions of the added groups may also affect solubility. Hollifield (1979) showed that 9 methyl benzo(a)anthracene was approximately three times more soluble than 10 methyl benzo(a)anthracene. Eganhouse and Calder (1976) found that mixtures of hydrocarbons may affect the solubilities of various PAHs. For example, naphthalene and phenanthrene enhance the solubility of acenaphthene. On the other hand, the solubility of naphthalene is essentially unaffected by the 145 presence of other PAHs. Aromatic hydrocarbons are "salted out" by increasing concentrations of inorganic salts, with the result that solubilities in saltwater are lower than in freshwater. Bioaccumulation Higher molecular weight PAHs are generally bioaccumulated to a greater extent by aquatic biota and are retained for longer periods of time. For example, Roesijadi et al. (1978) found that the benthic deposit feeding clam Macoma inguinata concentrated phenanthrene from water by a factor of 10.3; chrysene by 694; dimethyl benzo(a)-anthracene by 1349, and benzo(a)pyrene by 861. Southworth et al. (1978) observed similar trends for P A H bioconcentration with Daphnia  pulex. Uptake and elimination rate constants varied considerably with no discrete pattern. Naphthalene, however, had the lowest uptake rate and the highest elimination rate. Metabolism Microbes in water-sediment interfaces effectively degraded PAHs. However, the higher molecular weight PAHs were degraded more slowly (Lee, 1977). Gibson (1976) stated that for PAHs with more than three rings, microbial degradation is difficult. • 146 Benzo(a)anthracene Callahan and Slimak (1979) reported log octanol water/partition coefficients for benzoanthracenes between 5.6 and 6.4. Oysters which were suspended in enclosures for 8 days were found by Lee et al. (1978) to have accumulated benzo(a)anthracene at a level 18,000 times greater than the ambient water concentration. When the osyters were transferred to uncontaminated waters, benzo(a)anthracene was slowly released and its half life in the oysters was 9 days. Roesijadi et al. (1978) exposed the benthic species Macoma inquinata to marine sediments contaminated with Prudhoe Bay crude oil and found that dimethyl benzo(a)anthracene was bioconcentrated by a factor of 1349 above concentrations in water. Dimethyl b e n z o l -anthracene had the highest bioconcentration factor of the PAHs studied, which included phenanthrene, chrysene, and benzo(a)pyrene. Gardner et al. (1979) used a laboratory (20°C) flowing seawater system to study the effects of microbes and polychaete worms on the degradation of benzo(a)anthracene. Microbial degradation rates of benzo(a)anthracene were 1.5 - 1.8% per week in coastal sand sedi-ments and 1.4% per week in coastal marsh sediments. In the presence of the benthic polychaete worm, Capitella capitata, which is associated with areas of high oil input, degradation rates were increased to 2.7 - 3.0% per week in sand sediments and 1.8% in the marsh sediments. Two reasons were postulated for the increased 147 degradation in the presence of the polychaete: I) the worms improved conditions in the sediments for microbial degradation (i.e., mixing the sediment distributes nutrients to subsurface areas); and/or, 2) the worms metabolically degraded the compounds. Short-term field tests with benzo(a)anthracene have not shown detectable microbial degradation. For example, Lee et al. (1978), in experiments with C-14 labeled benzo(a)anthracene, found no I4CC>2 produced after 72 hours exposure in the waters of "controlled eco-system enclosures" located in Saanich Inlet. Studies of the bioaccumulation of the benzanthracenes in higher aguatic organisms are complicated by the presence of metabolic elimination/detoxification mechanisms. For example, Varanasi et al. (described in Malins et al., 1980) found that dietary radio-labelled naphthalene can be readily converted by flatfish into a variety of oxygenated metabolites that remained in tissues for extended periods. Fluoranthenes Fluoranthenes are characterized by low solubility and high partition coefficients. Hollifield (1979) reported a solubility of fluoranthene of 0.12 mg/L in distilled water. Rossi and Neff (1978) reported a solubility of fluoranthene in saltwater as 0.1 _+ 0.06 mg/L. Callahan and Slimak (1979) reported log octanol/water partition coefficients of 5.33 and 6.57 for fluoranthene and methyl fluoranthene respectively. 148 Fluoranthene was one of the PAHs added to the controlled ecosystem enclosures by Lee et al. (1978). Oysters in the enclosure bioaccu-mulated 5 mg/kg fluoranthene in 2 days from water containing 7.2 ug/L fluoranthene. After 23 days depuration time, 0.4 mg/kg fluor-anthene remained in the tissue. A half life of 5 days was calculated. Benzo(a)pyrene Benzo(a)pyrene has a solubility in freshwater of approximately 4 ug/L (MacKay and Shui, 1977) and a log P value of 6.04 (Radding et al., 1976), so is highly susceptible to adsorption to organic sediments and to biological uptake. Biological uptake studies of benzo(a)pyrene have predominantly utilized invertebrate test species. Vertebrate species such as fish metabolize polyaromatic hydrocarbons, hence complicating interpretation of uptake data. Roesijadi et al. (1978) exposed the bivalve Macoma inquinata for 7 days to oil contaminated sediments and found that the concentration factor for benzo(a)pyrene uptake from sediment was 0.09 compared to an uptake concentration factor from water of 861. Long-term exposure of M. inquinata to contaminated sediment^ showed continuing increases of benzo(a)pyrene within the bivalves, despite decreasing quantities of BaP in the sediments, indicating that equilibrium was not quickly reached with benzo(a)pyrene and that benzo(a)pyrene would be relatively persistent in tissue of exposed bivalves. 149 Freshwater studies by Lu et al. (1977) found that fish metabolize benzo(a)pyrene about "as fast as it is adsorbed and converts it to unextractable products". However, the authors noted minimal degradation of BaP in snails, with a 2177-fold concentration observed after 3 days. Therefore, studies of uptake of benzo(a)pyrene indicate that it is rapidly metabolized in fish livers, explaining the low values of BaP found in fish livers during the Puget Sound studies by Mai ins et al. (1980). Nonetheless, these authors considered that metabolites may remain in fish tissue and possibly affect physiological processes. The implications of the metabolites are unknown. Chlorinated Naphthalenes Callahan and Slimak (1979) calculated the solubility of 2-chloro-naphthalene to be 0.007 ug/L in freshwater at 25°C and the log P to be 4.12. Increased chlorination will result in decreased solubility and increased log P. Bioaccumulation of chlorinated naphthalenes, particularly those with four or more chlorine substitutes, is probably a significant environmental fate. Green and Neff (1977) exposed adult grass shrimp, Palaemonetes puqio, to 40 ppb (ug/L) levels of each of Halowax 1099, 1013, and 1000 for periods of up to 16 days. Halowax 1000 (monochloro- and dichloronaphthalene) was accumulated by a factor of 63; Halowax 1013 (tetrachloro- and pentachloronaphthalene) by a factor of 187; and Halowax 1099 (trichloro- and tetrachloro-150 naphthalene) by a factor of 257. In each case accumulation was rapid in the first 3 days. Depuration was rapid in clean water; tissue levels were generally reduced to I mg/kg or less within 5 days. The bioconcentration factors reported by Green and Neff (1977) for grass shrimp are much lower than the 2300 factor reported by the U.S. EPA (1976) for brown shrimp. Halowax 1014, a mixture of tetra-, penta-and hexa- chloronaphthalene, was used for the U.S. EPA (1976) studies and the increased number of substituted chlorines may have contributed to increased bioaccumulation levels in these studies. The U.S. EPA (1978) in its review of chlorinated naphthalenes reported that metabolism decreases with increasing chlorine levels. No metabolism of naphthalenes with five or more substituted chlorines was observed. 5.5.7 Phthalate Esters (PAEs) Properties Individual PAEs vary considerably in solubility and partitioning tendency, dependent upon molecular weight and substituted groups. The most persistent PAEs include di-butyl phthalates (DBPs), butyl benzyl phthalates (BBPs) and di-ethyl hexyl phthalates (DEHPs). Callahan and Slimak (1979) determined log P's for these compounds as 5.2, 4.8 - 5.8 and 5.3 - 8.7 respectively. 151 Bioaccumulation There have been few systematic studies of bioaccumulation of PAEs by marine biota. Giam et al. (1978) attributed the lack of data to difficulties in analytical capabilities - some of which are due to contamination of dilution waters, reagents, and possibly equipment. Kinetic uptake studies of phthalate esters found during this literature review all used freshwater aquatic test organisms. Sanders et al. (1973) showed that I4C-labeled DEHP was taken up rapidly and bio-accumulation factors for six invertebrate species varied from 1900 to 6500 after equilibrium was reached, generally within 7 days. However in another publication by the same authors and for the same test species, much lower bioconcentration factors were reported (Mayer and Sanders, 1973). Accumulation studies were apparently dependent upon the concentration of PAEs in water. For example, DEHP was bio-concentrated within the scud by a factor of 13,400 using a water concentration of 0.1 ug/L and by a factor of 270 using a water concentration of 62.8 ug/L. Metabolism and Depuration Biological elimination of accumulated PAEs was found by Sanders et al. (1973) to be rapid. Scud exposed to 0.1 ug/L of 14C-DEHP accumulated 5.4 mg/kg in 3 days. Residual radioactivity decreased rapidly during 4 days in phthalate-free water to 20% of the initial activity. 152 Rapid depuration rates were also determined within bluegills, by Gledhill et al. (1980) who observed a bioconcentration factor of 663 at equilibrium (21 days) for butylbenzyl phthalate, based on I^C-uptake. The depuration half life was less than 2 days. Giam et al. (1978) found low levels of phthalates in biota in the Gulf of Mexico despite elevated levels of phthalates in water and sediments. This observation was attributed to rapid metabolism by the organisms. Also, concentrations in liver were lower relative to muscle levels indicating the presence of metabolic degradation of PAEs. Conclusion The persistent organic contaminants that accumulate in sediments to concentrations that warrant concern to dredged material management have strong sorption tendencies with the sediments. Available evidence suggests that they will not appreciably desorb from the sediments upon disposal in the marine environment. However, significant contaminant release from the disposal site may occur via dispersion of contaminated particulates that become suspended in the water column. The potential for contaminated particulate material to become suspended into the water column during or subsequent to disposal is dependent upon the characteristics of the dredged material, the disposal methodology, and the oceanographic characteristics of the disposal site. Given that contaminants preferentially sorb to fine 153 grained sediments (Boyd et al., 1972), and that fine grained sediments are the most susceptible to resuspension and transport away from the disposal site, it is of primary importance that dredged material disposal be managed to minimize sediment losses via this pathway. The oceanographic characteristics of disposal sites where suspended sediment losses are minimized, and the extent of suspended sediment losses at such sites, are considered in Section 6. Section 7 then identifies potential disposal areas within the Strait of Georgia where resuspended sediment losses will be minimal. The potential role of biological uptake and transfer in the transport of hydrophobic contaminants from disposed dredged material may also be significant. Each of the reviewed organic contaminants of concern have a high potential for bioaccumulation. Bioaccumulation factors (vis a vis_ water concentrations) range from approximately 103 (e.g. PCP, PAHs) up to 106 (e.g. PCBs). Relative to the very low con-centrations that may be released into solution, the adsorption of hydrophobic chemicals into biotic tissue, and their subsequent release from the disposal site via migration of the organism or via the food chain, may conseguently provide an important pathway for their release from a disposal mound. The fact that contaminants released in this manner are incorporated within animal tissue adds to the environmental significance of this release process. The extent of contaminant uptake from disposed dredged material to the biological system is directly dependent upon the degree of biotic exposure to the material. Demersal fish and benthic invertebrates 154 that reside at a disposal site will be particularly prone to accumu-lating high body burdens. Their relatively sedentary and benthic habits, and in the case of groundfish, their long life span, will result in direct and prolonged exposure to the material. Hydrophobic organic chemicals have been shown to rapidly accumulate to high levels in these organisms. However the contribution of bioaccumulation to the actual transport of contaminants from a disposal area remains at issue. The process of food chain biomagnification, which can result in very high contaminant concentrations in predatory terrestrial animals, is apparently ineffective in the marine environment. Because contaminant concentrations in gill-bearing marine organisms are effectively moderated by equilibrium partitioning exchange, they rarely exceed sediment concentrations by more than an order of magnitude. The concern for bioaccumulative contaminants from a disposal site is thus largely dependent upon the physical migration of contaminated organisms from the area. The extent to which this occurs is specific to the biotic assemblage. Groundfish are more mobile than infaunal vertebrates, so possibly contribute more substantially to contaminant release. Migrating predatory fish that prey at a disposal site may also accumulate then transport contaminants from the area, but their contribution will be limited by their length of residence and feeding habits at the site. 155 It is concluded that the contribution of bioaccumulation as a pathway of possible contaminant release defies definative prediction on present knowledge. It is probably substantially less than release via suspended particulate dispersion, but is potentially much greater than desorptive release to solution. Minimizing the environmental implications of bioaccumulation requires selection of disposal sites where resident biota are relatively sparse and of low ecological and economic value. These reguirements will be applied in a review of appropriate disposal sites in the Strait of Georgia in Section 8. Biological transformation and release of metabolites of organic contaminants is a further fate process of potential significance. Vertebrates are particularly efficient at metabolizing and excreting accumulated contaminants. Although the metabolites are generally less toxic than the parent compounds, they are usually more hydro-philic and in some cases have been implicated with carcinogenic, mutagenic and cytotoxic effects. The rate at which higher organisms metabolize organic contaminants appears to be related to chemical structure: more highly chlorinated and more complex compounds are less susceptible to metabolic degradation. Rapid depuration of contaminants by fish has been shown to occur with less chlorinated phenols and benzenes, and with the low molecular weight PAHs and PAEs. The implications of the metabo-lites are unknown. 156 Ocean dumping management is primarily concerned with the more persistent organic contaminants that are relatively resistant to metabolism. However the uncertainty surrounding our knowledge of metabolic transformations requires that caution be exercised in this regard. Minimizing the environmental implications of biological transformation and release of contaminants requires selection of disposal sites where resident biota are relatively sparse. This requirement is coincidental with that for minimizing bioaccumulation, and is applied in the review of appropriate disposal sites in the Strait of Georgia in Section 8. 157 6.0 PROCESSES E F F E C T I N G T H E PHYS ICAL DISTRIBUTION OF DISPOSED D R E D G E D MATER IAL 6.1 Introduction It has been concluded that the primary pathway of potential release of trace metals and organic compounds from ocean dumped materials is through the dispersion or erosion of the particulates to which the contaminants are bound. Particulate release from ocean dumped material may occur from frictional forces during passage through the water column, or from erosion of the settled disposal mound. In either case, the advective loss of contaminants will extend the area of potential impact of the ocean dumping activities. As was outlined in Section 2.1, this imposes unpredictable and possibly substantial environmental risk. The finer particles are the most susceptible to resuspension and dispersive forces. The environmental risk associated with fine particle dispersion is accentuated by the fact that these fines frequently contain a disproportionately high contaminant loading. This is because, other factors being equal, contaminant adsorption is a surface phenomenon: fine particulates with a high surface area to volume ratio tend to contain higher contaminant concentrations than the coarser fractions (Boyd et al., 1972). For example, Brothers and 158 Sullivan (1984) determined that the more highly contaminated dredged materials from False Creek could be visually distinguished on the basis of colour and texture. Determination of the physical fate of dredged material is thus of critical environmental concern. In the selection process for a disposal site, consideration must be given to the eventual disposition of the dredged material in order that adeguate determination of the zone of impact can be made. The conduct of the disposal operation should then be calculated to maximize the assurance that the impact boundaries do not extent beyond a confined and acceptable area. Nittrouer and Sternberg (1975) investigated the fate of dredged material in Puget Sound, and concluded that the controlling factors of fine grained sediment movement were the degree of water entrainment during dredging; the water depth at the disposal site; and the energy regime of the near bottom environment at the site. The following describes the processes through which these factors are effective, then defines operational procedures by which fine material release may be controlled. During Water Column Passage Dredged material that is released at the sea surface from a scow is deposited on the seafloor in three steps (Gordon, 1974). Upon release, the dredged sediment descends rapidly (> 100 m/min) through the water column as a high density jet. If the disposed sediment has 159 a moisture content of less than 50%, then the bulk of material will fall essentially intact rapidly through the water column with little if any entrainment of seawater. If the initial water content is above 67% for silts or 75% for clays then it will act as a dense fluid: ambient water will be entrained during the descent phase, and in depths exceeding 20 metres the total volume of the descending jet may be increased several fold before it reaches the seafloor (JBF Scientific Corp., 1975). Dredging of heavily contaminated sediments in British Columbia's Lower Mainland is primarily conducted using clam-shell dredge. Disposal is usually by bottom dumping scows with a capacity of up to 1000 m3. On the basis of the Gordon and JBF studies, this material will pass through a shallow water column relatively intact. In deep water, this dense mass may entrain enough ambient water to create a neutrally bouyant plume. In this case, maximum water column interaction occurs and less bottom impact will occur. Upon impact with the bottom, large blocks of sediment stay in the impact area, but some of the released material spreads rapidly outward from the impact point as a density plume. This bottom plume may be several metres thick, but slows and thins as it travels outward. Initially the plume moves swiftly and, carries sediment particles away from the impact point until its velocity is reduced sufficiently to permit deposition. Over a flat bottom it has been observed to run a few hundred metres, at most, from the impact point (Gordon, 1974). On a sloped bottom, density plumes with a 160 suspension concentration greater than 10 gL- l have been observed to retain their integrity over much greater distances, and move down the slope as a gravitational flow (Meagher and Gorham, 1982). These three steps - descent of the jet, impact on the bottom, and the spread of the bottom density plume - have been observed under a wide range of hydrological conditions and dredged material charac-teristics (Bukoniewicz et al., 1978; Cusstar and Wakeman, 1977). The limiting conditions under which these steps will occur have not been determined. They have been documented in water depths up to 125 metres, but cannot be arbitrarily applied to deeper water sites such as exist in the Strait of Georgia. The Point Grey dumpsite is in 240 metres of water. Literature discussing deeper water disposal mechanics is extremely limited in amount and scope. Yamamoto and Alcauskas (1975) monitored the disposal of 4,000 cubic yards of material dredged from San Francisco Bay into 183 m of water. They determined that much of the material fell in clumps throughout the water column, but that the resultant disposal mound covered only 4000 m2 with an average depth of 0.3 metre. This accounts for only one third of the disposed material. A similar study was conducted by R.M. Towill Corp. (1972) to monitor 250,000 cubic yards of spoil disposed into 1,300 - 1,600 metres of water of Kauai, Hawaii. It was calculated that maximum deposition thickness was approximately 0.02 mm, at a distance of no less than 161 20 km downstream from the disposal site. In this depth of water, it may be reasonably concluded that the disposed material disperses completely during its passage through the water column. In deep water disposal sites in the Strait of Georgia, shear effects on the falling mass may result in considerable erosion of fine material during water column transit. Although the bulk of material disposed will probably impact the bottom, a significant proportion of fines may be lost to the water column. Given that most contaminants are associated with the fine fraction (Boyd et al., 1972), this loss may be of high environmental significance. 6.3 Subsequent to Settling 6.3.1 Physical Factors Effecting Sediment Resuspension Settled sediments may be physically remobilized and dispersed by wave and current turbulence. Studies by Bokuniewicz et al. (1977) of post-depositional movement of dredged materials in Long Island Sound determined that disturbance from tidal current fluctuations is the principal cause: direct, wind-driven flow over the bottom in water depths greater than 20 m is usually weak. This determination would equally apply to Georgia Strait, which is well protected from sea-wave disturbances. 162 The sediment particle size largely determines the speed of current required to erode, transport or deposit naturally settled particles (see Figure 6). For disposed dredged materials, this scheme may vary substantially, dependent on the mode of disposal and the resultant profile of the disposal mound. Numerous computer models have been developed which simulate physical transport dynamics of disposed dredged material. Examples include those of Koh and Chang (1973), Brandsma and Divoky (1976), Krishnappan (1975) and Hunt (1984). However, even if all the important factors pertaining to the dredged material and the water column were known, there appears to be no universally applicable model for quantitatively predicting the time history and ultimate fate of material released in the sea. Resuspension of disposed dredged sediments occurs when the boundary shear stress exceeds the critical stress of the disposal mound. The former is related to the water velocity profile but the latter is a complex function of the geometry of the sediment-water interface and the texture of the sediments. No quantitative testing of dumpsite erosion subsequent to disposal of dredged material has been conducted in British Columbia. The seabed water current regime at our major ocean dumping site at Point Grey has not even been studied (H. Nelson, EPS, pers. comm., 1985). Hoos (1977) determined that the majority of disposed material at the Point Grey dumpsite was located up to 3 km to the northeast of the 163 FIGURE 6 . R e l a t i o n s h i p between C u r r e n t S p e e d , P a r t i c l e D i a m e t e r , and Sediment E r o s i o n , T r a n s p o r t , o r D e p o s i t i o n ( a f t e r K e n n e t t , 1 9 8 2 ) . 100 . •) ErOS'On -90 Water Content (%) Deposiiion CLAY SILT i SAND l 4 20 60 200 2000 PARTICLE DIAMETER (^ m) 164 designated site, but concluded that this was the result of inaccurate dumping rather than bedload movement. This dumpsite has received a diverse and often unrecorded variety of sediment types, so it cannot be determined whether erosion of the fine materials has occurred. Studies conducted elsewhere have shown that clam-shell dredged material deposited in low or moderate energy environments in coastal waters mostly remains where it settles. For clay and silt sediments, the water content and cohesiveness of the disposal mound strongly influences the erosion potential. Sediments dredged by clam-shell and disposed from a bottom dumping scow have low water content and tend to remain cohesive upon deposition. Relatively strong current regimes are therefore reguired to resuspend them once they have settled. Substantial mounds of sediment have accumulated at dumpsites over a number of years. A bathymetric study of a dredged material disposal site in Narrangansett Bay, Rhode Island, showed that the bottom shoaled by about 6 metres after the disposal of 8 x I0*> m3 of sediment (Saila, Pratt and Polgar, 1972). The volume of this mound agreed to within 2.4 percent of the volume that was estimated to have been dumped. Precise bathymetric mapping of the Central Long Island Sound disposal site similarly accounted for 97% of the estimated volume of disposed material (Morton, 1983). 165 Morton (1983) did find that newly deposited dredged materials in Long Island Sound lost approximately 12 percent of their volume when a hurricane passed through the area. However losses due to the hurricane were small at neighbouring older disposal sites, where the disposed sediments had consolidated over time and under the influence of less extreme forces. Bed sediment distributions provide a simple and accurate reflection of sediment transport conditions at a prospective disposal area. This is especially true for clam-shell dredged and scow dumped sediments, where the disposed sediments are structurally similar to their pre-dredged condition. A study by Dexter et al. (1979) confirmed that clam-shell dredged silty-sand deposited on a predominantly sand bottom in Elliott Bay, Puget Sound, did not significantly erode. 6.3.2 Biological Factors Effecting Sediment Resuspension In addition to the above described physical factors effecting trace metal release from disposed sediments, biologically induced remobilization may occur. Feeding, burrowing, excavation and tube construction activities of benthic organisms each result in transport and reworking of the sediments. However, the degree to which these activities alter the stability of a disposed dredge spoil mound has never been definitively studied. The aforementioned studies describing disposal mound stability have been of insufficient duration to determine the effects of recolonization upon material erosion. 166 The biota at dredged material disposal sites is usually substantially reduced as a consequence of disposal activities. However in most cases the biological community at the disposal mound will eventually recover to pre-disposal abundances. Hoos (1977) and Packman (1980) noted large densities of burrowing and deposit feeding fauna at the Point Grey dumpsite. At least two effects of biogenic processes can significantly decrease mound stability. These are: 1. An increase in water content. Figure 6 shows that the current velocity necessary to initiate erosion decreases with increasing water content. Feeding and burrowing activity breaks up sediment aggregates and counteracts compaction, creating an open sediment fabric. 2. A modification of surface topography. Microtopographic features increase turbulence at the sediment-water interface and hence erosion. Rowe (1974) demonstrated that burrows and fecal mounds can act as "catalysts" for erosion by periodic bottom currents. The extent of these effects on any given disposal mound cannot be predicted. Laboratory studies by Young and Southard (1978) showed that after 60 days of reworking by a natural assemblage of benthic organisms, critical erosion velocity decreased by approximately 60% 167 from the initial value. They found that bioturbated muds had lower threshold shear velocities than fine sands, which is contrary to Figure 6. However these researchers were unable to translate these studies to accurate field predictions in the intensively studied Buzzards Bay. This suggests that we are a long way from understanding the biogenic mechanisms influencing the transport of particulates from disposed dredge spoil mounds. Conclusion It is concluded that a significant proportion of clam-shell dredged material that is disposed by bottom opening scow into deep areas of Georgia Strait will be lost to the water column during a dumping operation. The loss will consist predominantly of fine fractions. Because contaminants preferentially sorb to finer fractions, and because fine particles may be transported for substantial distances before resettling, the environmental implications of this loss warrant considerable concern. At shallow water disposal sites, clam-shell dredged sediments will fall to the bottom as a coherent mass. The density plume which emanates from the point of impact should not spread over a flat bottom for more than a few hundred metres. 168 Upon settlement and consolidation on the bottom it is unlikely that sediments disposed in moderate depth, low-energy depositional areas will appreciably erode. Water turbulence below approximately 20 metres depth in Georgia Strait is uneffected by occasional storms. For these areas the energy regime is reasonably reflected by the in situ sedimentary regime. In shoal areas (i.e. < 20 m), low energy depositional areas may be periodically eroded by storm driven wave turbulence that reaches to the bottom. Suitable disposal sites will therefore be sufficiently deep that deposited sediments are protected from both regular and irregular disturbance, yet sufficiently shallow that water entrainment and the resultant disaggregation of the disposed sediments during water column passage will be minimized. On the basis of the available information, suitable disposal sites in Georgia Strait will be within a low energy, depositional environment and in a water depth of not less than 30 m or deeper than 125 m. Dredging and disposal of contaminated sediments should be restricted to clam-shell dredge and bottom dumping scow. Dumping should be restricted to during periods of negligible tidal current and calm sea conditions. Johansen et al. (1976) investigated the potential of placing dredged material in borrow pits or depressions to enhance the post-depositional stability of the dredged material deposit. They offer guidance on how to use a barge or hopper dredge to release dredged material into subaqueous depressions, including suggested effective volumes, navigation techniques, and maneuvering and position main-169 tenance over the pit during disposal. Their study suggests that proper modifications to disposal equipment and careful handling of the material may reduce the potential for erosion and resuspension of the dredged material deposit within the borrow pits. A desirable strategy for disposal that secures confinement of con-taminated sediments can be realized by taking advantage of the lateral or vertical pollution gradients which often exist within a dredging project. Sequential dredging and disposal of polluted and uncontaminated sediments promotes burial of the former below the bio-chemically active zone. Capping of contaminated sediments with clean sands (e.g. from the Fraser River main channel) would achieve the same effect. This process should be considered whenever possible. 170 7.0 DISPOSAL SITE OPPORTUNITIES WITHIN GEORGIA STRAIT 7.1 Recommended Disposal Site Characteristics The important physical oceanographic criteria for dumpsite selection are water depth and circulation/dispersion potential. On the basis of depth alone, the Point Grey dumpsite is an unsuitable location for the disposal of contaminated material. The water depth of 240 m at this site will result in substantial disaggregation of disposed material as it falls through the water column. The disaggregated mass will be frac-tionated by particle size, and the more highly contaminated fine materials will disperse in the water column to contaminate other areas. The extent and distribution of the spread of toxic contami-nates from this site are unpredicatable, but probably substantial. As was concluded in the last section, the preferred locations for disposing of contaminated material should meet the following oceanographic criteria: o water depth should be greater than 20 metres but less than 100-125 meters o bottom topography should be flat, or preferably concave o water circulation should be sufficient to maintain oxygenation of the bottom water, but insufficient to resuspend fine material. Natural sediments at the site should be predominantly mud, or silt and clay (i.e., less than 0.063 mm). 171 In such areas, the disposal of contaminated materials would potentially have substantial adverse effects on resident organisms, but these effects will be restricted to as small an area as possible. There are very few areas within the Strait of Georgia that meet the bathymetric criteria for dumpsite suitability. The Strait is a glaciated marine valley comprising a series of deep basins and troughs separated by shallower ridges and banks. Average depth in the Strait is 157 metres with deeper waters frequently exceeding 250 metres. Few of the shallower banks are expansive, and most are steep-walled. On the other hand, virtually all areas in Georgia Strait are suffi-ciently well circulated to maintain high oxygenation of the bottom waters (Waldichuk, 1983). The Strait lacks a shallow threshold sill, common in many inlets and coastal seas, and replacement of deep water is not substantially impeded. Although there are few data describing bottom-water circulation within the Strait, it can reasonably be assumed for the present purposes that oxygenated water will be maintained at the sediment surface of even the most stagnant shallow water areas. This assumption would need to be confirmed previous to disposing of contaminated sediments at any site. The upper limit to the criterion for water circujation (i.e., where disposed fine sediments may be resuspended) requires more careful monitoring. The large tidal range in the Strait (up to three metres in southern waters and five metres in northern waters) generates 172 strong tidal currents. Waldichuk (1957) divided the Strait into northern, central and southern areas based on the surface current regime, and reported that: o the northern region (north from southern Texada Island) has weak (<0.2 knots) and variable surface currents. Speeds rarely exceed one knot, although they may be faster in some channels. o the central region (to a line joining Point Roberts and the Saanich Peninsula) has strong tidal currents and exhibits complex surface water movement under the influence of Fraser River run-off, wind, tides and Coriolis force. Currents in excess of one knot are common. o the southern region has tidal currents typically greater than one knot. There is virtually no data on bottom current velocities for anywhere within Georgia Strait. Reference is frequently made in the literature to bottom current measurements but these data were not the type of bottom data required. "Bottom" is frequently used as a relative term in measuring water column profiles, and may refer to data collected two or even five metres above the sea bed. Such data are of little or no value for bedload movement calculations, which require data at a maximum of one metre above the sea bed. Accurate information describing sea bed current velocities at potential disposal sites is of fundamental importance to the management of dredged material 173 disposal, and is appropriately a Priority Research Topic for R O D A C funded investigations during 1985-1986 (R. Kussat, EPS, pers. comm., 1985). In the absence of appropriate bottom current data, existing substrate characteristics must be used to reflect maximum resuspension cur-rents. Local geology and sediment sources profoundly effect sediment characteristics. However the predominance of mud or silt and clay particles in the substrate is a reasonable indication of a low current regime. Conversely, where a sea bed is characterized by sand despite an abundant supply of silt and clay into the area, it is safe to speculate that fine grained sediments would be removed (or not deposited) by bottom currents. Identification of Potential Disposal Sites A detailed review of bathymetric charts of Georgia Strait identified only four areas which met the bathymetric criteria for potentially suitable disposal sites. They are described in Table XIII and Figures 7-11. The large size of Area 3 has permitted delineation of two potential dumping sites within it. The area of the circles shown in Figures 8-11 is approximately 2.5 x lO^m?. The Northern Stuart Channel site is located in a shallow, semi-enclosed basin between De Courcy Island and Flewett Point, Vancouver Island (Figure 8). The basin could readily be fully enclosed by strategically dumping clean sediments in the channel extension to 174 TABLE X I I I . P o t e n t i a l D i s p o s a l S i t e s i n the S t r a i t o f G e o r g i a t h a t meet the Recommended B a t h y m e t r i c C r i t e r i a . S i t e S i t e L a t i t u d e L o n g i t u d e Approximate D i s t a n c e Number Depth (m) p t r ™ e y ( n m ) 1. N o r t h e r n S t u a r t Channel 4 9 ° 0 5 ' 2 8 " N 1 2 3 ° 4 5 ' 1 2 " w 64 25 2. McNaughton P o i n t 4 9 ° 3 3 ' 2 7 " N 1 2 4 ° 0 0 ' 3 6 " W 122 45 3 a . Denman I s l a n d 4 9 0 3 5 ' 5 4 " N 1 2 4 ° 4 6 ' 3 0 " W 91 60 3b. Comox 4 9 ° 3 9 ' 5 4 " N 1 2 4 ° 4 9 ' 2 2 " W 77 65 4 . Smelt Bay ( C o r t e s I s . ) 5 0 ° 0 2 ' 3 0 " N 1 2 5 ° 0 0 ' 3 0 " W 100 90 125° © '175 124° 123° 50° 3 ) 49c © , ^ ^ ^ t ? f f X * ' r '41 FIGURE 7. G e o r g i a S t r a i t -'itfedonda Ba/ L o c a t i o n o f Proposed Dumpsites ^ 1a nttiiaskivO 1. S t u a r t Channel 2 . McNaughton P o i n t 3*. Comox - Denman Is 4 . Smelt Bay \ \ . V ,-'^VPowffll Kiver v . --'-''7? ffi *<!••)• '<t-Woe. i f V V v V ' ' ^ ' r t i CoWc [tile Rivet\ \^^\^*=^^\Billingi^NelSo^Y/ ' oysWTK V \ TEXADA ' ^ b - . i a \ l 4 \ V 3 % . H S L A N D is/ Union BayUN IVp W (loen\an Island Gillies E X A D ^ N i y e x a d a \ \ \ N Island v oS,\Madeira Pat* : \ f £ ^ktkpenirtsylai ;| 8 9 l \ V ^ f K Hallmoon\ A a s q u e l i V V S j \ Buccaneer < A . ' . <? r si- i „ v / . . v. <t • Port Mellon y^^- I \ (I {'•', ,\ ,« lambief. ft \ lslan(i,;\ ft * ; n^smui, NANAIMO | I a . D L - U M ! Qualicum Beach ^ /> L 4. i r e ^ W fe^^^^Paxksville * Garth»l<JI-Sjjiiamish M" 50° iORTH I 'ANCOU ^, . ,--V(ESryANC 419 " \ ^ * - < 2 ? ; : . - . ^ - " .riu. VANCOUVER^ illingWn* O \ G a b n o l a ' \ rctsland Q \ \ \ y a l d e s v ^-S ^ i s l a n d ' \ pe Beale*-4y0'(. P a c h e n a P o i n t Carmanah Poiriri^ V--, /' . ».ii!^ ,."tB- ^ry^art ^jua 49° •fender ^ *r ir--A.. Compiled from Canadian E n e r g y , Mines and Resources Map NM-9/10 VIII) 125° 124° 123° 3 5 7 7 48' Adjoining Chart/Carte adjacent e 42' 40' -1X7-124°oc/ J5--LH Francis \ I \<3535/27M \7r\ Peninsulax [too" '-•'•.91 1 • fl(3 : VII. ,0< i H".V U9 ^ ! <4? ' ^ " ^ i ^'4' .5V5 ' : 3i V> i \f v. so ; " \ ,'ii , ~6' 7.'v '•. ; V ,'«r',1 v, • ' ••'.'. . is:-8 0 7 0 ; l-v*t <\ i> i " > &4. ~ -. '-.''V BjerPeShoalV;^ \ ^ (SITE 67 27 . 67 • 13 S E C H V E Li, 'P E N \ I \ N I S \ U A 0 7 H Hires' \ McNajigKcon P.. o o ( *3v 58fc. 65 Turnagain Ii V S • 1 \T/ *6 i V ' 9 > . ^ 6 . . • v t » » _ v ^ v { ;2i. I ^ ^  ' L '.|4, ; Epsorn"Pi'r2i. 54 ! is i. -6- •' 15 I V II 2i 5 5?; [H X M , 16). ^ fi \ ' : ' 9 j * \ I .' . ,t:< 4 8 \ 5 4 F I G i „ ' : „ / /;' •' 6'.'-. Tomnham Gfr2S\ _ jj ; ^ L p T * Chart 3535 535 { 45 VyiefiA ;f 6 ';^:b<LSmuK'V '"A Fl Buccaneeri ••'••.^ .-iy^ /iM> FIGURE 9 . L o c a t i o n o f Proposed McNaughton P o i n t Dumpsite -(Area 2 ) . Area d e s c r i b e d - 2 . 5 x 10 m . Soundings i n f a t h o m s . Bill .360 ° «B . 1 A , Chan 3535 : J s ' .4-41 Kg. Compiled from C a n a d i a n ' ^ - n se: x>x ^sJ" H y d r o g r a p h i c S e r v i c e s , : 4 6 c ^ " N a u t i c a l C h a r t s , 1985 7 S s p ^ 4 ^ 124°otr 50' , 173 9 1 ' 76 92 *° V-.. ••-.U z; 2i • * \ ' ? Ajax 1 k. .-.^ v 15 » i 1 % I s l e t s ^ ^ - * V . . ^ ; . 33 37 " i 7 \ 15 8 14 1. V 16 S. FIGURE 10. Location of Proposed Denman Is land - Comox Dumpsites (Area 3). Area descr ibed • 2.5 x 10 m . Soundings in fathoms. Compiled from Canadian ^ Hydrographic Se rv i ce s , Nautical Charts , 1985 .V,**. V'\ ?7 $•'-4 ,s . • sc a* ' i 50' 125°00l79 FIGURE 11. L o c a t i o n o f Proposed Smelt Bay Dumpsite - (Area 4 ) . Area d e s c r i b e d Soundings i n fathoms 2 . 5 x 10 6 m 2 . Compiled from Canadian ^ H y d r o g r a p h i c S e r v i c e s , ; N a u t i c a l C h a r t s , 1985 ; 8? 125*00' SS' 180 the southwest of the proposed site. The resultant basin would be approximately five metres deep and could accommodate up to 20 x 106 m3 of dredged material. Although Stuart Channel is subject to relatively strong tidal currents in surface waters (1-3 knots during floods; Canadian Hydrographic Service, 1983), the bathymetry indicates that sea bed currents are probably weak. This is supported by the fact that the existing sediments at the proposed disposal site are mud (Lands Directorate, 1981). However given the strength of the surface currents in this area, ocean dumping at this site would need to be restricted to during tidal slacks. The potential site at McNaughton Point is within a submarine trench running between Bjerre Shoal and the Sechelt Peninsula (Figure 9). The narrow entrance to the trench (south of the proposed site) could readily be closed using clean fi l l . The resultant basin would be at least five metres deep and could accommodate at least 30 x 10^ nr)3 of dredged material. There are no data describing water currents inside Bjerre Shoal, but the bathymetry suggests that sea bed currents in this area are also low. The existing sediments at the proposed site are mud (Lands Directorate, 1983). 181 The Denman Island and Comox sites are both located within a trench between the Comox Bar and a shallow bank that extends southwards from Cape Lazo (Figure 10). The trench has a narrow opening to deeper water at its southern end, which could be readily closed with clean sediments. Although Figure 10 describes two specific sites within the resultant basin, the entire basin meets the bathymetric criteria for a potential dumpsite. The basin could be easily designed to accomodate a ten metre mound of spoil throughout. This eguates to approximately 120 x I0& m3 of dredged material. There are no appropriate current data for this area. The existing sediments at the Comox site are silts and clays, probably from the Tsolum River (Lands Directorate, 1981). At the site outside of Denman Island, the sediments are mud (Lands Directorate, 1981). The Smelt Bay site comprises a natural submarine basin within the embayment between Marina and Cortes Islands (Figure II). The basin is approximately 20 metres deeper than the shallow sill at the mouth of the embayment and could accommodate approximately 50 x 10^ of dredged material. The basin has mud sediments (Lands Directorate, 1981), indicative of a low current regime. 182 8.0 RESOURCES IN C O N F L I C T 8.1 Introduction The Strait of Georgia is of crit ical environmental economic and social importance to British Columbia. The advantages of accessibility, moderate climate, rich coastal and fluvial soils, productive estuarine and marine waters, and diverse aesthetic and recreational amenities, have attracted most of British Columbia's population and economic activity to its perimeter. The Strait lies between the two major centres of population in the province, Vancouver and the other urban communities of the Lower Mainland to the east and Victoria, Nanaimo, and other Vancouver Island towns to the west. The Gulf Islands and the Sunshine Coast also contribute to this development base. Approximately 80 percent of the provincial population now live within 80 km of the Strait. The marine resources of Georgia Strait are abundant and diverse. High nutrient inflows from riverine (mainly the Fraser River) and upwelled oceanic sources promote high primary productivity. This provides a rich foundation for organisms higher in the food web. The wide variety of estuarine, protected and open shoreline, and open water habitats in Georgia Strait support more than 500 species of attached marine algae, 200 species of fish, 300 species of invertebrates, 130 species of marine-associated birds and 10 families of marine mammals 183 (Barker, 1974). This abundant animal and plant life contributes a great deal to the commercial and recreational attractiveness of the region. Important commercial and recreational resource uses in Georgia Strait include the salmon, herring, groundfish, shrimp, crab, clam and goeduck fisheries, the culturing of salmon and oysters, and recre-ational activities such as boating, sportfishing, swimming, scuba diving, oyster and clam digging and beachcombing. The exploitation of each of these resources is dependent, to a greater or lesser extent, upon the maintenance of a high level of environmental protection. The potential dredged material disposal sites that were identified in the preceding section each meet the physical oceanographic criteria that will assure minimal short- or long-term release of toxic con-taminants. Although this will restrict the environmental impacts from disposal operations to as small an area as possible, substantial and possibly long-term biological degradation will occur at the disposal site. The present section provides a review of the current and potential resource uses in the vicinity of the proposed ocean dumping sites, to assess their vulnerability to severe but localized biological degradation. 184 TABLE XIV. S p e c i e s Names o f Organisms c i t e d i n S e c t i o n 8 . G r o u n d f i s h P a c i f i c cod E n g l i s h s o l e L i n g c o d W a l l e y e p o l l o c k P a c i f i c hake S p i n y d o g f i s h R o c k f i s h P e l a g i c F i s h Chinook salmon Coho salmon Chum salmon P i n k salmon Sockeye salmon C u t t h r o a t t r o u t P a c i f i c h e r r i n g P a c i f i c anchovy E u l a c h o n S u r f s m e l t Mol1 uses B u t t e r c l a m Japanese l i t t l e n e c k N a t i v e l i t t l e n e c k Geoduck P a c i f i c o y s t e r C r u s t a c e a n s Dungeness c r a b Prawn Shrimp M a r i n e mammals K i l l e r whale Harbour s e a l S t e l l a r sea l i o n C a l i f o r n i a sea l i o n Gadus macrocephal us  Parophys v e t u l i s  Ophiodon e l o n g a t u s  T h e r a g r a Chalcogramma  M e r l u c c i u s p r o d u c t u s  S q u a l u s o c a n t h i a s  S c o r p a e n i d a e Oncorhynchus t s h a w y t s c h a 0 . k i s u t c h 0 . k e t a 0 . gorbuscha 0 . nerka Salmo c l a r k i c l a r k i  Clupea harengus p a l l i s i i  E n g r a u l i s mordax  T h a i e i c h t h y s p a c i f i c u s  Hypomesus p r e t i o s u s Saxidomus g i g a n t e u s  V e n e r u p i s j a p o n i c a  P r o t o t h a c a staminea  Panope generosa  C r a s s o s t r e a g i g a s Cancer m a g i s t e r  Pandalus p l a t y c e r o s  P. j o r d a n i  P. b o r e a l i s  P. h y p s i n o t u s  P. danae P a n d a l o p s i s d i s p a r O r c i n u s o r c a  Phoca v i t u l i n a  Eumetopias j u b a t u s  Zalophus c a l i f o r n T a n u s 185 Groundfish Demersal fish that are commercially exploited by a bottom trawl fishery in the Strait of Georgia include Pacif ic cod, English sole and other flounders, Lingcod, and Walleye pollack. Pacif ic hake and Spiny dogfish have also been increasingly trawled in recent years to sup-plement declines in the traditional species. The Strait of Georgia groundfishery is minor compared with the activity that occurs in offshore areas to the west and north of Vancouver Island, yet it retains a high commercial importance due to its close proximity to markets. Georgia Strait trawl landings of Pacif ic cod alone are between 900 - 1200 tonnes per year (Department of Fisheries and Oceans, 1980). Longlining of Rockfish, Lingcod, Pacif ic cod and Spiny dogfish also provide a minor commercial fishery in the Strait of Georgia. Recreational groundfish harvesting is directed primarily towards Rockfish and Lingcod. Demersal fish that reside within the vicinity of an ocean dumping site are particularly vulnerable to adverse impacts from the disposal of contaminated sediments. Their relatively sedentary and benthic habits result in prolonged and direct exposure to sediment contaminants. The major potential impacts of ocean dumping upon groundfish include: 186 Toxicity. Acute and/or chronic toxic effects of contaminants upon groundfish or their prey may impact the size of the groundfish stock in the area. Groundfish spawning areas are particularly susceptible to toxic effects. McGreer and Munday (1982) found that contaminated sediments from Roberts Bank and False Creek substantially effected survival of fertil ized eggs of the Pacif ic cod. Successful hatching of larvae was virtually eliminated when the eggs were covered to a depth of only I mm with contami-nated sediment, or exposed to suspended sediment concentrations of 7.5 g/L or greater (cf. 30 g/L for "clean" sediments). Although the young of fish species are usually the most sensitive to the toxic effects of contaminants, mortalities of mature groundfish or their prey species at dredged material disposal areas may also occur. Malins et al. (1980) reported high incidences of tissue lesions of varying severity in groundfish of Puget Sound, and related the occurrence of these to the distribution of high levels of sedimentary contaminants. Bioaccumulation. Prolonged and direct exposure to contaminants may result in their accumulation in groundfish tissue to a level which endangers the health of human consumers. Adult groundfish are bottom-feeding, relatively stationary and have a long life span (up to 15 years in many species). In areas of high sediment contamination, contaminant levels in older individuals may be significant. For example, elevated mercury levels that were 187 consistently detected in groundfish and crab species from Howe Sound in the I970's forced the closure of this area as a com-mercial and sport fishery (Garrett, MacLeod and Sneddon, 1980). o Habitat Alteration. Changes to the physical and chemical character of the sediments at a disposal site may induce dramatic changes in the benthic community upon which groundfish feed (Levings, 1982). o Fishery Interference. Ocean dumping of concrete rubble, wood debris or wire cables (which freguently litter areas requiring dredging) will interfere with the operational logistics of a groundfish trawl fishery. The commercial value of the Strait of Georgia groundfish resource, together with the high vulnerability of groundfish species to adverse impact from contaminated sediments, deem important groundfish spawning or fishing areas as unsuitable for the disposal of contami-nated sediments. The Lands Directorate (1981; 1983) has conducted comprehensive reviews of coastal resources in the Strait of Georgia. Their com-pilation of available data describing groundfish stocks in the Strait showed that utilization of the proposed ocean dumping sites is as follows: 188 1. North Stuart Channel. The proposed dumpsite and adjacent deep areas of northern Stuart Channel are productive spawning areas for English sole and other flounders. The reef area to the north of Ruxton Island is a Lingcod spawning area. Walleye pollock are believed to also spawn in Stuart Channel, since mature individuals have been observed there in February, imminent spawners in March-April, and spawned out pollock in May. Commercial trawling for Pacif ic cod, English sole and other flounders, and Walleye pollock is conducted throughout Stuart Channel (i.e. where suitable substrate occurs), including the proposed disposal area. 2. McNaughton Point. There are no groundfish stocks recorded inside Bjerre Shoal. The drop-off zone to the west of Bjerre Shoal is included within the Spiny dogfish longline fishery that occurs along the entire Sunshine Coast. Mid-water trawling for Pacif ic hake occurs further offshore in central Malaspina Strait and south into Georgia Strait. However neither of these fisheries are considered sufficiently proximate to the proposed dumpsite to be vulnerable to impact from contaminants. Bjerre Shoal forms a natural barrier to sediment flow into deeper waters. 3. Denman Island - Comox, Both of the proposed sites in this area are within the principal spawning ground for the northern Georgia Strait stock of Pacif ic cod (the major stock in Georgia Strait) and the Cape Lazo stock of English sole (one of two major stocks in 189 the Strait). The area supports a productive trawl fishery for these species. Pacif ic hake have also been recorded to be very abundant during the summer months in the 50 - 100 metre zone from Cape Lazo to Hornby Island. These fish apparently winter in deeper water of the central Strait, then migrate inshore, especially to the area east of Comox Bar, each summer. 4. Smelt Bay. There are no groundfish stocks recorded between Marina and Cortes Islands. Recreational fishing for Lingcod occurs around the shallow reefs to the north of Marina Island, but this is well distant from the proposed dumpsite. Benthic Crustaceans Five shrimp and one prawn species are fished commercially in British Columbian waters. The major shrimp fishery is by bottom trawl over muddy or sandy bottoms in 95 to 135 metres of water, whilst prawns are predominantly taken by trap. Recreational harvesting of shrimp and prawns is minimal. The Dungeness crab provides a small trap fishery in shallow (<40 m), sand areas of the Strait. This fishery is minor relative to the crab harvest in waters around Cracoft and Gilford Islands, Kincome Inlet and Knight Inlet, which provides 90% of the provincial catch (Ministry of Lands, Parks and Housing, 1984). Recreational fishing for crab is popular among the boating public, but the numbers taken are small (Lands Directorate, 1981). 190 Subsequent to settling from their pelagic larval phase, shrimp, prawns and crabs are sedentary. Populations within a dredged material disposal site will consequently suffer direct and prolonged exposure to contaminant materials. Although data describing the toxicity of sedimentary contaminants to benthic marine crustaceans is often lacking, they are often among the most sensitive organisms to contaminant effects. Shuba, Tatum and Carrel (1978) found the shrimp Paleomonetes pugio to suffer high toxicity from a wide variety of sedimentary contaminants. Chapman (1984) similarly noted the high sensitivity of a benthic amphipod to contaminated sediments and recommended its use in sediment bioassays for toxicity testing. Benthic crustaceans have also been recorded to accumulate high levels of sediment contaminants within their body tissues. The accumulation of sediment bound mercury by crabs in Howe Sound, which forced the closure of the area for fishing, was noted above. Mercury concen-trations above the guideline for human consumption have also be found in crab on Sturgeon Bank (Garrett, MacLeod and Sneddon, 1980). Garrett (1983) similarly reported high concentrations of PCBs in crabs from this area. Nimmo et al. (1971) demonstrated that shrimp and crabs can accumulate high hepatopancreas concentrations of PCBs from ingesting contaminated sediment particles. Shrimp, prawn or crab fisheries occur at three of the four areas proposed for ocean dumping. The area north and west of Denman and Hornby Islands (including proposed sites 3a and 3b) forms one of the three major commercial shrimp trawling grounds in the Strait of 191 Georgia. Shrimp trawling is also conducted in Stuart Channel (proposed site I), which is part of the general grounds comprising the larger channels of the Gulf Islands. The major prawn trapping activity in the Strait occurs at the numerous 50 metre banks along the Sunshine Coast, including the area around Bjerre Shoal (proposed site 2). Shrimp trawling along the Sunshine Coast is conducted well to the south of McNaughton Point, beyond Sechelt. Minor crab fishing activity occurs in shallow water along the northwestern shore of Stuart Channel (west from proposed site I) and north of the proposed Comox dumpsite (site 3b). Molluscs The Strait of Georgia supports rich stocks of molluscs, including oysters, geoducks and other clams. Commercial activity is divided between the less intensive harvest of wild stocks and intensive mariculture operations. Wild stocks are also commonly harvested recreationally, especially near population centres and more accessible locations. The commercial clam fishery has traditionally included Butter clams, Japanese Littleneck clams and the Native Littleneck clam. Each of these species inhabit intertidal areas of suitable substrate throughout Georgia Strait. Landings of geoduck clams, first reported in 1977, 192 now exceed the annual average landings of traditional species. The main populations of geoducks are found in sub-tidal sand or sand-mud, and are fished from the high subtidal to a depth of approximately 15 metres. Pacif ic oysters were initially introduced into B.C. from Japan and now provide a substantial local fishery and culture industry. They normally inhabit intertidal flats of firm mud, sand or gravel, but are often grown in the lowest intertidal area (bottom or near-bottom culture) and subtidally (off-bottom culture). Commercial permit harvesting of wild stocks is limited to scattered populations sustained through local spatfalls. Most wild stocks in the Strait of Georgia are reserved for recreational use. The major oyster harvest in the Strait comes from commercial culture operations. Spat are collected from Pendrell and Hotham Sounds and grown in leased shoreline areas elsewhere on the coast. In the Strait of Georgia the oysters grow to market size in about one to four years, depending on culture technigues, temperature and a variety of other factors. Clams and oysters are filter feeding organisms which select food items on the basis of particle size. Suspended sediment particles that are within the size range of their planktonic feed items will be ingested with the food. Consequently, these animals are prone to dietary intake of contaminants that are sorbed to sediment material suspended in the water column. 193 Reid et al. (1981) found elevated levels of cadmium i n clams (P. staminea) and mussels (Mytilus edulis) from False Creek. E.V.S. Consultants (1984) reported similar body burdens i n the clam Macoma  balthica after six months exposure to sediments from the same area. This species was also shown to bioaccumulate lead from sediments of Vancouver Harbour. Garrett (1983) reported that mussels collected under the Burrard and Granville bridges contained only low PCB concentrations (<20 ug/kg), but mussels from industrial harbour regions in California have been reported to contain up to 1300 ug/kg. The tendency for filter feeding molluscs to bioaccumulate high concen-trations of contaminants has encouraged their use as integrative indicators of water pollution levels i n many coastal areas. The toxicity of Vancouver sediments to the clam M. balthica has been assessed by numerous researchers. McGreer, Reid and Nelson (1981) found no elevated mortalities i n adult clams after 30 days exposure to Vancouver Harbour sediments. Reid et al. (1981) similarly found negligible toxicity to this species from 60 days exposure to False Creek sediments. However E.V.S. Consultants (1984) more recently documented 28% and 24% mortality i n M. balthica after exposing them for six months to the sediments of False Creek and Vancouver Harbour, respectively. This was significantly higher than mortalities i n a cleaner control sediment. From data compiled by the Lands Directorate (1981; 1983) the utilization of areas adjacent to the proposed ocean dumping sites for the cultivation or harvesting of oysters and clams is as follows: 194 1. Northern Stuart Channel. The only oyster leases in the area are in False Narrows and Degnen Bay on the southern shoreline of Gabriola Island, which is remote from the proposed dump site. Recreational digging for clams is popular in sandy intertidal areas near to the proposed site, particularly along the Vancouver Island shoreline. Geoduck clams are found in low abundance in the shallow nearshore region to the west of the proposed site. Geoducks are in higher abundance further south, near Yellow Point. 2. McNaughton Point. There are no harvested mollusc populations in the immediate vicinity of this proposed site. Minor recreational harvesting of clams occurs at Thormanby Island, 12 km to the south. 3. Comox-Denman Island. Comox Bar, eastern Denman Island and the area to the north of Hornby Island comprise the most important commercial clamming region in the Strait of Georgia. Barnes Sound (between Comox Bar - Denman Island and Vancouver Island) and Lambert Channel (between Denman Island and Hornby Island) are among the Strait's most productive areas for oyster cultivation. The shallow subtidal zone throughout the region also supports a rich harvest of geoducks. 4. Smelt Bay. Cortes Island was identified by Valiela (1979) as having major potential for off-bottom culture of oysters. The western shoreline of Cortes Island (north from Marina Island) is 195 currently under extensive use for bottom culture and it is presumed that Valiela (1979) was referring to these more sheltered waters. Closer to the proposed dredged material disposal site, there is minor use of the shoreline adjacent to Mansons Landing and along the eastern side of Marina Island for oyster cultivation. Commercial oyster picking permits have been issued for a section of shoreline i n Smelt Bay. Geoducks are in low abundance i n the immediate vicinity of the proposed disposal site (I-10 clams per square metre; Cox and Charman, 1980), but are in higher abundance (11-50 clarns/m2) further to the north and west. Despite the high sensitivity and bioaccumulation potential of filter feeding molluscs for sedimentary contaminants, the vulnerability of the above described populations to contaminant loadings from dredged sediment disposed at the proposed sites is relatively low. Each of the proposed dump sites are i n more than 60 metres of water and meet criteria that will ensure minimum short or long term loss of disposed material from the immediate dumpsite area. The oysters and beach clams described inhabit intertidal areas near the dumping grounds, so are remote from direct exposure. Geoduck clams are found to 45 metres depth, but commercial harvesting is restricted to more accessible depths between the intertidal and 15 metres. 196 The primary potential source of exposure of these shallow water molluscs to contaminants from disposed dredged material is from contaminated fine materials dispersed in a near-surface plume during the disposal operation. Caution should therefore be exercised to ensure that disposal operations near valuable oyster or clam popula-tions should minimize water column losses. Adequate protection of mollusc resources in the vicinity of proposed areas I, 3 and 4 therefore requires that dumping operations should only be conducted during slack tides and relatively calm sea conditions when the like-lihood of onshore transport of contaminated particles in the water column is minimized. Pelagic Fish Pelagic fish in the Strait of Georgia that are subject to commercial, recreational or native Indian harvesting include five species of salmon, Pacif ic herring, Coastal cutthroat trout, Eulachon, Surf smelt and a variety of other species. The salmon and herring fisheries far exceed the values of other fisheries in the Strait. In contrast to groundfish, pelagic fish are not especially vulnerable to adverse impacts from sediment contaminants. They are surface feeding and migratory, so are not normally exposed to sedimentary contaminants either directly or for prolonged periods. For example, mercury concentrations in salmon, herring, anchovies and eulachons from British Columbia are generally very low even in areas with highly contaminated sediments (Garrett, MacLeod and Sneddon, 1980). 197 Nix and Chapman (1984) regularly sampled herring, anchovies and salmon juveniles in False Creek throughout dredging and disposal operations by B.C. Place and Expo '86, but noted no adverse impacts. The eggs, larvae and young fry of many pelagic species are more prone to extended and direct exposure to sedimentary contaminants than are their adults. Many pelagic fish have demersal eggs, and their fry often remain resident in nursery rearing areas until they approach maturity. For example, Pacif ic herring spawn on kelp in shallow subtidal areas of the protected coast, and their fry spend their first 6-8 months in shallow bays and inlets proximate to their spawning grounds. Thereafter, most Georgia Strait herring migrate out of the Strait to the open ocean and do not return until they are mature fish of 3-4 years. Pacif ic herring, and also anchovies and surf smelt, are virtually exclusively surface feeders throughout their life, so are only indirectly exposed to sediment contaminants even during periods of prolonged residence. However other pelagic fish feed predominantly upon benthic organisms during their early life stages. Benthic feeding habits place the fish in immediate proximity to sediment contaminants and increase the likelihood that prey organisms contain high con-taminant loadings. Coho and chinook salmon in the Strait of Georgia are primarily benthic feeders during their fry residence in estuarine areas, then switch gradually to an exclusively surface feeding dependence when they move into coastal waters. There is a period of 198 uncertain duration as they first move into coastal waters when juvenile salmon retain a high dietary dependence on benthic organisms. The eggs and larvae of many pelagic fish also appear to be more sensitive to contaminants in the sediments than are their parents. Sherk et al. (1974) reviewed the effects of suspended and deposited sediments on a wide range of estuarine species of surface fish, and concluded that eggs and larvae are almost universally the most sensitive life history stages tested. Messiah, Wildish and Peterson (1981) showed that contaminated sediments deposited onto the spawn of Atlantic herring (Clupea harenqus) increased egg mortality. Chapman et al. (1983) exposed eggs of surf smelt in the laboratory to contaminated sediment slurries collected at 22 stations in Puget Sound. Adverse effects on development, which included reduction of hatching success, premature hatching and reduced larval survival, were recorded at 20 of the 22 sites tested. Thus, although it is extremely unlikely that the localized contaminant effects due to ocean dumping at any of the proposed sites in Georgia Strait will impact migratory and surface feeding adult populations of pelagic fish species, caution is required that the location of ocean dumping sites does not conflict with their spawning or nursery grounds. Seasonal dumping restrictions to protect resource species during spawning is a Common practice (Levings, 1982), but these do not address the persistent nature of sedimentary contaminants. Con-199 fidence that ocean dumping will not adversely impact pelagic fish stocks is only achieved by ensuring that disposal sites are geo-graphically remote from important spawning and nursery grounds. Pacif ic herring Pacif ic herring are the dominant fish species (in terms of biomass) in Georgia Strait. They support a lucrative roe fishery during their early spring spawning and a smaller food and bait fishery during the fall months. They also provide a major food source for larger fish, marine mammals and seabirds. Herring spawn between low water and 10-15 metres depth in a limited number of areas scattered throughout the Pacif ic coast. The locations and intensities of spawning vary from year to year, but within the Strait of Georgia, the Comox to Nanaimo region (including areas proximate to proposed ocean dumping area 3) has consistently held the most productive spawning grounds. Spawning ground assessments have indicated that this region produced between 45 -60% of the Georgia Strait herring stock (Lands Directorate, 1981). The coastline near Yellow Point (adjacent to proposed site I) has also been productive, accounting for an estimated 10 - 20% of the Georgia Strait stock (Hourston, 1972). Herring spawning in the vicinity of proposed dump sites 2 (McNaughton Point) and 4 (Smelt Bay) is negligible, but data compiled by the Lands Directorate (1983) indicates 200 that Bjerre Shaol (adjacent to proposed site 2) and southern Cortes Island (to the south of proposed site h) do support rearing areas for herring juveniles. Salmon Salmon stocks in the Strait of Georgia have diverse life histories. Sockeye, pink and most chum stocks from rivers entering the Strait generally migrate to offshore waters soon after leaving the rivers. In contrast, most chinook and coho stocks, as well as some chum stocks, prolong their departure from the Strait with indefinite periods of nearshore residence in protected coastal waters. During this early period, these fish eat benthic as well as pelagic prey, so are prone to dietary uptake of contaminants. Some chinook and coho stocks take up permanent residency in the Strait, but their nomadic lifestyle and exclusively surface feeding habits ensure minimal exposure to localized contamination during their post-juvenile life. Of the four proposed ocean dumping areas, only the northern Stuart Channel area (site I) has been reported to support significant numbers of young salmon. Coho juveniles have been sampled in particularly high numbers in the vicinity of Yellow Point. Further north and closer to the proposed disposal site, chum juveniles were predominant. This area also supports lesser numbers of chinook, pink and sockeye salmon (Lands Directorate, 1981). 201 Juvenile salmon utilization of the other three proposed dumpsite areas has not been reported, and is unlikely. Salmon fry prefer protected waters and have rarely been found in more open water areas of the Strait such as these sites. Each of the proposed sites 2 - 4 are nearby to alternate areas of protected waters where salmon fry from the rivers of the immediate region may be expected to preferentially congregate. Commercial salmon fishing activity in the vicinity of the proposed ocean dumping sites is listed by the Lands Directorate (1981; 1983) as moderate in the Comox-Denman region (area 3) and low at McNaughton Point (area 2) and southern Cortes Island (area 4). Sports fishing activity is described as moderate at proposed areas 2 and 3, and low at proposed areas I and 4. Other Pelagic Fish There are very few data describing the use of areas near to the proposed disposal sites by pelagic fish other than herring and salmon. Rivers utilized by Coastal cutthroat trout are listed in the Lands Directorate reviews (1981; 1983), but none are sufficiently proximate to the proposed sites to warrant concern. No data were found in these reviews or elsewhere indicating utilization of areas proximate to the proposed sites by Eulachon or Surf smelt. The Ministry of Lands, Parks and Housing (1984) also noted the dearth of data on the distribution of these species. However, given the low vulnerability of 202 these surface feeding and nomadic species to impact from sediment contaminants, this information reguirement is considered to be of low priority for the present purposes. Marine Mammals There are ten families of marine mammals that utilize Georgia Strait. The most abundant are Killer whales, Harbour seals and the Stellar and California sea lions. Although none of the marine mammals of Georgia Strait are now harvested by humans, they provide high recreational amenity value. Kil ler whales are nomadic, cruising throughout a range of some 200 nautical miles at approximately 3-4 knots. The only known area where prolonged residence occurs is north of Georgia Strait in Robson Bight. The Kil ler whales migratory routes are commonly 1 - 5 km offshore where they are thought to prey primarily on salmon, herring and smaller marine mammals. The Harbour seal populations in Georgia Strait are permanent residents but are nomadic between haulouts. They frequent estuaries, river deltas, tidal rocks and shallow sublittoral areas, and feed predominantly on littoral fish. Sea lions are resident in Georgia Strait only during the winter months, when the two species intermix on tidal rocks. They feed on a wide range of fish, including numerous groundfish species. 203 Marine mammals warrant particular concern with regard to their potential to accumulate high concentrations of persistent contami-nants. They do not have gills, an effective exchange organ in fish, so contaminant exchange is a function of intake via food and elimination primarily with excretia. The elimination of persistent contaminants by the latter route has been found to be relatively inefficient (Calambokidis et al., 1978). As a result, a portion of the persistent chemicals from contaminated food may be retained, and concentra-tions will tend to increase with age. Although the effects of high contaminant burdens in marine mammals remain unknown, high PCB levels (38 mg/kg wet weight) in the blubber of a Kil ler whale which was washed ashore in Georgia Strait (Herman and Calambodikis, 1978) support high scientific and public concern in this regard. The likelihood of contaminant accumulation in marine mammals from disposed dredged material is dependent upon the mammal's freguency and duration of exposure. Of the four proposed ocean dumping areas, Area I (Stuart Channel) is recognized as a secondary Kil ler whale migration route, but the other proposed sites are remote from the normal migrations of these animals (Lands Directorate, 1981; 1983). The major migration routes of Kil ler whales through the Gulf Islands are well to the south of the proposed site in Stuart Channel. Seal and sea lion populations are open to more direct and prolonged exposure to localized contaminant loadings than are Kil ler whales. In areas proximate to the proposed disposal sites, the rocky foreshore of Link and Round Islands (near Area I in Stuart Channel), and also the 204 Seal Islets (near Area 3), are popular haulouts for seals and sea lions. Their residency in these areas, and their high potential to accumulate substantial contaminant levels through food chain magnification from groundfish prey, warrant concern with the use of the areas for dis-posing contaminated sediments. In contrast, there are no recognized haulouts proximate to proposed sites 2 or 4 (Lands Directorate, 1983). Marine Birds The Strait of Georgia provides a major resting and overwintering environment for migrating birds on the Pacif ic flyway. Most of the birds that reside and feed in the Strait are migrants. Myriad habitat types in the estuaries, inlets, coastal embayments and waters of the Strait provide abundant food and a mild, sheltered winter for the many species that visit. Vermeer et al. (n.d.) reported overwintering shorebird densities, based on aerial surveys, of 63.5 birds per kilometre for the Gulf Islands (including proposed area I), 47.4 birds per kilometre along the Mainland coast (including proposed area 2), 170.2 birds per kilometre along the central Vancouver Island coast (including proposed area 3) and 21.7 birds per kilometre in northern Georgia Strait (including proposed area 4). During the summer months the marine bird population in the Strait of Georgia is comparatively very low (Robertson, 1977). Nesting habitat for the few species that do breed in the Strait differs substantially among species and is usually associated with an immediate source of 205 food (Lands Directorate, 1983). In this sense, intertidal areas, saltmarshes, eelgrass meadows and kelp beds are extremely important to their breeding success. Data describing bird populations in the Strait of Georgia are mostly limited to broad-scale aircraft and boat surveys of migrating and overwintering populations. Although these do not provide the detailed information necessary to assess the potential impact of localized contamination upon seabird populations, they do indicate the relative seabird use of the general regions of the Strait. Vermeer et al. (n.d.) summarized the distribution and densities of marine birds over seven sub-regions in the Strait of Georgia. Further data from individual surveys of bird populations has been summarized by the Lands Directorate (1981, 1983). Util ization of the shoreline and pelagic zone proximate to the proposed disposal areas is indicated by these reports to be as follows: I. Stuart Channel, Gulf Islands. The Gulf Islands provide some of the most popular overwintering and nesting habitats for marine birds in the Strait of Georgia (Rodway and Campbell, 1976). Boat surveys in March-April of 1977 showed 300 shorebirds, I 15 loons, grebes and cormorants, 43 gulls, 3 auks and 138 diving ducks per kilometre of shoreline (Vermeer et al., n.d.). The west coast of DeCourcy Island supports 28 nesting pairs of seabirds (Lands Directorate, 1981) and is rated by Parks Canada (described in Lands Directorate, 1981) as a moderately important nesting area (i.e. it received at rating of II in a four point scheme from I = 206 least significant to IV = most significant). The west coasts of Mudge, Link and Ruxton Islands were also recorded to support nesting seabird colonies (Lands Directorate, 1981). Each of these areas were awarded a rating of I by Parks Canada. Waterfowl surveys by the B.C. Ministry of Environment (n.d.) and Ducks Unlimited (n.d.) also identified moderate numbers of seabirds along the Vancouver Island coast of Stuart Channel during the fall and higher numbers during spring. 2. McNaughton Point, Sechelt Peninsula. The coastal region proximate to McNaughton Point appears to be of relatively low importance to nesting or overwintering seabird populations. Boat observations of overwintering populations along the entire Sunshine Coast (i.e. from Lund to Gibsons) noted 87 shorebirds, 2 loons, grebes and cormorants, 10 gulls, 4 auks and 71 diving ducks per kilometre of shoreline (Vermeer et al., n.d.). It is likely that the preponderance of these sitings were made to the south of North Thormanby Island, which is 5 km south of McNaughton Point. This southern stretch of the Sechelt Coast was rated III by Parks Canada, whereas the region north of North Thormanby Island did not earn a rating (Lands Directorate, 1983). An aerial survey from Francis Point to Smuggler Cove by Ducks Unlimited (n.d.) noted only 102 overwintering birds along this 10 km tronsect. 3. Denman Island-Comox. The Sandy Island - Seal Islets area is a very important staging area for shorebirds (Lands Directorate, 1981). A linear survey by the B.C. Provincial Museum during the 207 winter of 1976 reported 5600 seabirds, including especially high numbers of Surf scoters and Greater scaups, along a six kilometre transect between Comox and Cape Lazo (Lands Directorate, 1981). Kye Bay, immediately to the north of Cape Lazo, has been reported to provide overwintering support for 300 seaducks per kilometre (Lands Directorate, 1981). Vermeer et al. (n.d.) report overwintering birds densities along the central Vancouver Island coastline of 124 shorebirds, 18 loons, grebes and cormorants, 45 gulls, 5 auks and 52 diving ducks per kilometre. 4. Smelt Bay, Cortes Island. Boat surveys of the northern Georgia Strait coastline in March-Apri l, 1977, documented only 21 shore-birds, 2 loons, grebes and cormorants, 3 gulls, 5 auks and I I diving ducks per kilometre of shoreline (Vermeer et al., n.d.). Surveys by the B.C. Ministry of Environment and Ducks Unlimited during October, 1979 and January, 1980 recorded similarly low numbers along the mainland coast to the east of Cortes Island (Lands Directorate, 1983). No information specific to the southern Cortes Island coastline was uncovered by the literature review for this study, but it appears likely that utilization of the area by marine birds is low. Marine birds, like marine mammals, warrant particular concern with regard to their potential to accumulate high body burdens of per-sistent contaminants. Seabirds prey on a diverse array of marine organisms from several trophic levels. Because their elimination (primarily with ecretia) of dietary accumulated contaminants is 208 relatively inefficient, a portion of their dietary intake of these substances is retained and concentrations tend to increase with age. Very high tissue concentrations of persistent pesticides and PCBs have been recorded in adult seabirds from diverse geographic regions (Kreitzer and Heinz, 1974). Although the toxicity of persistent contaminants such as PCBs to most avian species is usually guite low, experimental work with PCBs has shown reproductive impairment in birds through reduced hatch-ability (Cecil et al, 1974), behavioural modifications (Kreitzer and Heinz, 1974), decreased egg production (Tumasonis, Bush and Baker, 1973) and embryo abnormalities (Cecil et al., 1974). Studies by Lincer and Peakall (1970) suggest that PCBs may effect breeding cycles by stimulating the production of liver hydroxylases which reduce circulating estrogen levels. Other effects of PCB exposure in birds include decreased avoidance responses (Kreitzer and Heinz, 1974) and decreased hatching of eggs due to inadeguate parental care (Peakall and Peakall, 1973). Very little information is available on the effects of other persistent organic contaminants of concern, or on the effects of environmental pollutants in general, upon wild populations of marine birds. However the potential for impact is possibly substantial. Many marine birds v survive by adapting one or all of the functions (e.g reproduction) to the behaviour of a prey species (e.g. herring spawn). As the popula-tions of prey species are often subject to sharp fluctuations and shifts in location, the populations of seabirds can also fluctuate substan-209 tidily. The .additional consequences of unnatural stresses, such as reproductive failures or high adult mortality, may be catastrophic but may not be revealed by adult enumerations until during periods of high natural population decline. Until this occurs the inherent variability of seabird populations may effectively mask the compli-cating unnatural perturbations. By the time the effect is unmasked, the population may have entered an irrecoverable decline. The high concern that is warranted by the above is moderated in the present case by regard to the very low potential for seabird exposure to sedimentary contaminants. The principal feeding reliance of marine birds is upon zooplankton, pelagic fish and intertidal benthos. As has been outlined above, the disposal of contaminated dredged material at the locations proposed by this study should have minimal impact upon these prey items. Each of the proposed disposal sites is within at least 60 metres of water and has been selected on the basis of oceanographic criteria which will minimize release and dispersion of contaminants to effect nearby areas. It is therefore considered unlikely that the controlled disposal of dredged material at these sites (i.e. during periods of slack tides and moderately calm weather) will significantly effect nearby marine birds. Human Recreation Recreational activities other than sports fishing that may be impacted by dredged material disposal in the Strait of Georgia are swimming, scuba diving and boating. The potential impact of ocean dumping 210 upon the health of swimmers and boaters is probably more imaginary than real: a very large volume of water containing suspended contaminated sediments would need to be swallowed before effects may occur. However the high priority accorded to human health concerns, real or perceived, requires that any potential for conflict between water contact recreational activities and ocean dumping should be minimized. Most of the coast of the Strait of Georgia is rated by the Canada Land Inventory as having moderate to very high capability for outdoor recreation (Department of Environment, 1978). Nearby population base and accessibility appear to be the primary determinants of actual recreational use. Accordingly, of the four proposed ocean dumping areas, Areas I and 2 have high recreational use, Area 3 has moderate use and Area 4 has low recreational use. This applies equally to shoreline and waterborn recreational activities. Popular scuba diving sites in the Strait of Georgia perhaps deserve heightened environmental protection as sites where recreational and biological amenity values coincide. Of the areas proximate to the proposed dump sites, Pratt-Johnson (1977) lists Round Island in northern Stuart Channel (3 km northwest from proposed site I) and Woods Bay on the Sechelt Peninsula (2 km southwest from proposed site 2) among the best 141 dives in the Georgia Strait - Puget Sound region. 211 Parks and Reserves The Strait of Georgia has 30 Provincial Parks with water frontage, including several Provincial Marine Parks. There are also nine Ecological Reserves preserved by the provincial government as part of the International Biological Program, and a number of conservation areas established by regional and municipal authorities. These out-standing scenic or ecological areas are preserved for public recreation and/or the protection of wildlife: other uses are strictly controlled. None of the proposed dredged material disposal sites impinge directy upon established parks, reserves or conservation areas. However there are parks or reserves sufficiently proximate to each of the proposed sites to warrant concern. These are: 1. North Stuart Channel. Mudge, Link and De Courcy Islands, to the east of proposed site I, are conservation areas under the Gulf Islands Regional Plan. Pirates Cove on the southeastern end of De Courcy Island, is a Provincial Marine Park. 2. McNaughton Point. Smuggler Cove Provincial Marine Park is within 10 km to the south of the proposed dump site. Thomanby Island, some 12 km to the south of the site,, has no declared conservation status but is generally recognized as an important ecological area. 212 3. Comox-Denman Island. The Sandy Island - Seal Islets area (approximately 4 km from each of the two proposed dump sites) is a Provincial Park (upland only). The intertidal flats surrounding this park are recognized as an important staging area for shorebirds, but have no reserve status. 4. Smelt Bay, Cortes Island. Smelt Bay Provincial Park is an upland park abutting a small section of the southern shoreline of Smelt Bay. The Mansons Landing Provincial Marine Park is approxi-mately 2 km north of the proposed dump site. The boundary to this park corresponds closely with the 5 metre depth contour. Although the potential impacts of the water disposal of contaminated dredge material upon upland parks and reserves is negligible, an implicit requirement to leave a suitable buffer between the park boundary and waste disposal of any kind is recognized by the author. Of concern therefore, is the proposed ocean dumping in the Comox-Denman Island area, within 4 km of the important intertidal sandflats surrounding the Sandy Island -Seal Islets provincial park. Of slightly less concern would be ocean dumping proximate to the Smelt Bay Provincial Park and the conserved islands of Northern Stuart Channel. The status of these reserves is not based upon important marine associations. The marine parks are substantially more vulnerable to impact from dredged material disposal. Although the Mansons Landing marine park (site 4) extends only to the shallow subtidal so should not be effected 213 by confined disposal at a site some 2 km distant and in 100 metres water depth, concern for unforseen potential impacts should be recognized. Disposal during slack water or offshore tides will minimize the likelihood that a surface plume of contaminated particulates may be transported into shallow water areas. Similar precautions may be necessary to protect the less proximate marine parks near proposed sites I and 2. 8 . 1 0 Conclusion None of the proposed dredged material disposal sites that were identified on the basis of bathymetric criteria in Section 7 are free for use as disposal sites without conflict with other resource uses of Georgia Strait. Although other areas may exist in Georgia Strait where ocean dumping may better avoid conflict with other resource uses at the immediate site, the unpredictable and perhaps substantial impacts of non-confined disposal upon non-proximate resources prohibits the use of these areas for disposing contaminated material. Therefore, if marine disposal of contaminated material is to be permitted within the Strait, the selection of a suitable location must be made from the nominated sites. Table XV provides a quantitative summary of the potentially con-flicting resource uses near to the proposed disposal sites, and permits comparative selection to be made between the sites. Relative activity scores have been awarded for each of the nominated resource uses at each of the proposed sites. These scores were derived from 214 TABLE XV. Q u a n t i f i c a t i o n o f Resource C o n f l i c t s a t t h e Proposed Ocean Dumping S i t e s . Resource V u l n e r a b i l i t y L e v e l o f A c t i v i t y / A b u n d a n c e ( P o t e n t i a l Impact) Use F a c t o r S i t e 1 S i t e 2 S i t e 3 S i t e 4 G r o u n d f i s h : - spawning a r e a 8 2 (16) - ( -) 3 (24) - (-) - commercial f i s h e r y 8 2 (16) - (-) 2 (16) - (-) - s p o r t f i s h e r y 8 2 (16) - (-) 1 (8) - ( - ) C r u s t a c e a n f i s h e r y : 8 2 (16) 3 (24) 3 (24) - (-) M o l l u s c s : - commercial 4 1 (4) - (-) 3 (12) 2 (8) - r e c r e a t i o n a l 4 1 (4) 2 (8) 2 (8) 2 (8) H e r r i n g : - spawning a r e a 4 3 (12) - ( - ) 3 (12) - (-) - r e a r i n g a r e a 2 3 (6) 1 (2) 1 (2) 2 (4) - f i s h e r y 1 2 (2) - ( - ) 2 (2) - (-) Salmon: - r e a r i n g a r e a 4 2 (8) - (-) - ( - ) - ( -) - commercial f i s h e r y 1 - (") 1 (1) 2 (2) 1 (1) - s p o r t f i s h e r y 1 1 (1) 2 (2) 2 (2) 1 (1) M a r i n e mammals: 4 2 (8) - (-) 2 (8) - <-> M a r i n e B i r d s : 4 3 (12) 1 (4) 3 (12) 1 (4) R e c r e a t i o n : 1 3 (3) 3 (3) 2 (2) 1 (1) Parks and R e s e r v e s : - marine 8 1 (8) 1 (8) - ( - ) 2 (16) - u p l a n d 1 1 (1) - (-) 2 (2) 1 (1) TOTAL (133) (52) (136) (44) LEGEND SITE LOCATIONS 1. NORTH STUART CHANNEL, GULF ISLANDS 2. MCNAUGHTON POINT, SECHELT PENINSULAR 3. COMOX - DENMAN ISLAND 4. SMELT BAY. CORTES ISLAND VULNERABILITY FACTOR - A measure o f the degree o f exp o s u r e and s e n s a t i v l t y o f the r e s o u r c e use t o d i s p o s e d s e d i m e n t a r y c o n t a m i n a n t s . 1. NEGLIGIBLE 2. LOW 4. MODERATE 8. HIGH LEVEL OF ACTIVITY / ABUNDANCE - The a c t i v i t y s c o r e s t h a t a r e awarded a r e d e r i v e d from c o n s i d e r a t i o n o f t h e r e l a t i v e i n t e n s i t y o f t h e r e s o u r c e use a c t i v i t y and/or t h e abundance o f t h e resource,, and t h e g e o g r a p h i c p r o x i m i t y o f t h e r e s o u r c e t o t h e proposed ocean dumping s i t e . 1. LOW 2. MODERATE 3. HIGH LEVEL OF POTENTIAL IMPACT - The p r o d u c t o f t h e V u l n e r a b i l i t y F a c t o r and the L e v e l o f A c t i v i t y / A b u n d a n c e . 215 consideration of the relative intensity of the activity and/or the abundance of the resource, and the geographic proximity of the resource to the proposed disposal site. For example, the proposed Denmand Island-Comox disposal area is directly within the principal spawning ground for two major stocks of groundfish. The high abundance and immediate proximity of this resource warrants a high rating. In comparison, the Northern Stuart Channel site is within productive spawning areas for minor stocks of groundfish, and is close to (but not within) spawning areas for Lingcod. The moderate abundance and proximity of this resource justifies a moderate rating. Both of these areas support a commercial groundfishery, but in each case the fishery extends throughout the general region (i.e. the disposal site comprises only a small area of the fishery) and is a minor fishery relative to groundfisheries elsewhere in the province. The level of activity of this fishery is therefore considered to be moderate. Similarly for the commercial harvesting of molluscs. The proposed Denman Island-Comox disposal area is in close proximity to the most important clamming and oyster production region in the Strait of Georgia, so warrants a high rating. Cortes Island has a high potential for oyster cultivation, as well as supporting existing harvests of clams, oysters and geoducks. However the productive areas for these activities are relatively distant from the proposed disposal site. Commercial mollusc activity at Site k is therefore rated as moderate. In Northern Stuart Channel the harvesting of oysters and geoduck 216 clams is a low level activity and is primarily conducted at sites that are geographically distant from the proposed disposal site. The level of activity/abundance at this site is therefore rated as low. Activity/abundance scores awarded range from low or distant act i -vity/abundance (score = I) to high and proximate activity/abundance (score = 3). Reference by the reader to the narrative in Sections 8.2 to 8.8 is recommended to clarify the awarded scores. Although the scores that are awarded in each case are not beyond argument, they do provide a general indication of the relative potential for resource conflicts at the proposed dump sites. Minor alteration of the activity scores awarded has little effect on the overall outcome of Table XV. Superimposed upon the relative activity of each conflicting resource use is a "vulnerability factor" which is dependent upon the degree of exposure and sensitivity of the resource use to disposed sedimentary contaminants. As has been described above, resource uses in Georgia Strait vary substantially in their sensitivity and exposure to the disposal of contaminated material at confined, relatively deep water disposal sites. Consistent with the scientific literature describing dose-response and exposure time-response relationships between nonessential contaminants and marine organisms (see Connell and Miller (1984) for a review of this topic), a logarithmjc scoring system has been used. Thus, resident and benthic feeding groundfish at the disposal site have arbitrarily been determined to be twice as vul-nerable as nomadic but benthic feeding juvenile salmon, who are in 217 turn twice as vulnerable as temporarily resident but surface feeding juvenile herring, and so on. Vulnerability scores awarded range from a low of I to a high of 8. The relative potential impact of dredged material disposal at each of the proposed sites is the product of the relative activity of resource utilization (or abundance of the resource) and the relative vulnerability of the resource. These values are shown in parentheses in Table XV. The total of these potential impacts provides an estimate of the conflict that exists at each of the proposed sites between dredged material disposal and other resource uses. Comparison of the potential impact scores for each of the four proposed areas indicates that dredged material disposal at Smelt Bay (Area 4) will conflict less with competing resource uses in Georgia Strait than disposal at any of the other proposed sites. Disposal at McNaughton Point (Area 2), which is only half the distance from Vancouver, may cause marginally greater conflict. There are sub-stantial conflicts with existing resource uses at Areas I (Stuart Channel) and 3 (Comox-Denman Island) and these areas are recom-mended as inappropriate for dredged material disposal. The costs of towing dredged material the considerable distances to the two recommended disposal sites are estimated in Table XVI. As can be seen from the Table, the costs vary substantially depending upon 218 TABLE XVI. Comparative Towing C o s t s f o r Recommended D i s p o s a l S i t e s . 3 Component D i s p o s a l S i t e s McNaught on P o i n t Smelt Bay Low volume High volume Low volume High volume o p e r a t i o n o p e r a t i o n o p e r a t i o n o p e r a t i o n Tug r e q u i r e m e n t s 500 HP 3600 HP 500 HP 3600 HP Tug h i r e $3000/day $14000/day $3000/day $14000/day Barge r e q u i r e m e n t s 1000 m3 7000 m3 1000 m3 7000 m3 Barge h i r e $300/day $1500/day $300/day $1500/day T r i p D u r a t i o n ^ 1 . 3 days 1 . 3 days 2 . 5 days 2 . 5 days T r i p c o s t $4290 $20150 $8250 $38750 3 Cost per metre $ 4 . 2 9 $ 2 . 8 8 $ 8 . 2 5 $ 5 . 5 4 a . I n f o r m a t i o n d e r i v e d from e x t r a p o l a t i o n o f i n f o r m a t i o n p r o v i d e d i n T a b l e I . b. Assuming a t o w i n g s p e e d , averaged o v e r t h e r e t u r n j o u r n e y , o f 3 k n o t s . T h i s f i g u r e i s based on i n f o r m a t i o n from T. H i l l i e r , R i v t o w , p e r s . comm., 1985. 219 the scale of operation. In either case however, disposal at McNaughton Point offers considerable cost savings over the use of Smelt Bay. Despite the value of Table XV in selecting between possible dredged material disposal sites on the basis of potential impacts to conflicting resources, the Table does not permit final recommendation as to the suitability of the proposed sites for ocean dumping. An additional important variable that requires consideration but cannot be readily guantified in the above manner is the relative value of each of the different resource uses. The value of a resource is determined by mutifarious and often intangible economic and social factors, some of which can only be resolved by the political process. For example, although the economic returns of the commercial salmon harvest can be readily quantified, salmon also offer an abstract social value based on romantic perceptions of life on the west coast and similar sub-jective and intangible values. The recent controversy over harvesting non-endangered species of whales further exemplifies the intangible resource values that accrue to some aspects of nature. Economists have derived various methods for estimating these values (e.g. willingness to pay or to be compensated) but without significant success. Until the relative value of each of the conflicting resource uses listed in Table XV can be accurately reckoned, the final deter-mination of disposal site suitability cannot be made. In the present case, a determination is reguired as to whether the impacts of ocean dumping at McNaughton Point upon the prawn fishery, the recre-ational salmon fishery and swimming and boating activities are of 220 greater environmental, economic and social cost than the impacts of dumping at Smelt Bay on the local mollusc stocks and the provincial parks. This issue involves subjective judgements that can only be addressed by the political process. A related but separate question that requires consideration pertains to the limits of sediment quality which might be disposed at any selected disposal site. Dredged sediments from the Vancouver region comprise a continuum from pristine to highly polluted. Relatively unconta-minated sediments could appropriately be disposed at either the McNaughton Point or the Smelt Bay sites with minimal risk to conflicting resources. However uncertainties prevent our prediction of the ultimate impacts of highly contaminated sediments at these sites. Caution in response to these uncertainties should reasonably prohibit disposal of highly contaminated sediments here or elsewhere within our coastal waters. The uncertainty of ultimate impact also prevents authoritative determination of the degree of contamination of sediments at which such caution should be expressed. The means of addressing the uncertainties of environmental impact, and of measuring sediment quality in a manner reflecting the potential for causing environmental degradation, is considered in the concluding section to this thesis. 221 9 . 0 CONCLUSIONS 9 . 1 Introduction The pressures to dispose of dredged material in Georgia Strait can be expected to continue, and they will not be limited to clean sedi-mentary material. Areas within Vancouver's extensive waterways have been reported to contain substantial sediment concentrations of a range of toxic chemicals. Numerous other toxic compounds have been identified in this thesis as likely additional pollutants of the sedi-ments, but for which analyses have not been conducted. Cadmium, mercury, PCBs, hexachlorobutadiene, polychlorinated dibenzofurans, pentachlorophenol, polychlorinated benzenes, various halogenated and non-halogenated polyaromatic hydrocarbons and phthalate esters are identified as each reguiring particular concern in the management of the disposal of dredged material from the Vancouver region. Concerns for the environmental impacts from the disposal of con-taminated dredged materials relate primarily to the release of the contaminants from the disposed materials. Biological degradation of the disposal site is inevitable, but appropriate site selection and confinement of the contaminants within as small an area as possible will minimize the ecological and economic conseguences of this degradation. Without contaminant release, the question is essentially: where do we relocate our pollution? But if contaminant release and 222 dispersion occur, the act of relocation accentuates the potential harm, adding unpredictable and perhaps substantial environmental costs to the dredging program. Dredging and dredged material disposal involves the mechanical aggitation of the sediments, their exposure to the water column, and their relocation into a physical, chemical and biological environment that may differ substantially from the dredged site. Each of these processes may induce release of contaminants from the disposed material. This study has reviewed available information describing the processes, transformations and pathways of contaminant release from sediments in the marine environment. Desorptive release of contaminants to the water column was found to be negligible, except for perhaps cadmium. Biological transformation of contaminants and release of the more soluble metabolites to the water column is potentially significant for mercury and for some of the less chlorinated and lower molecular weight organic compounds of concern (e.g. PCP, low molecular weight PAHs, PAEs). Biological uptake and dispersion from the disposal site via the food chain is also a potentially significant pathway for contaminant release, particularly for the highly lipophilic organic contaminants and methylmercury. However the primary route of potential release and dispersion of both trace metals and organic contaminants from disposed dredged materials is via the erosion and transport of contaminated particulates. Minimizing the resuspension 223 of material during and subseguent to disposal is consequently adopted as the primary objective for the selection of disposal sites where substantial contaminant release will be avoided. A review of the processes that promote erosion and resuspension of particulate material from disposed dredged material, and the appli-cation of this information to local oceanographic conditions, identified only four areas within the Strait of Georgia where particulate release from disposed dredged material may not be substantial. Two of these areas are adjacent to valuable biological resources and were con-sequently deemed unsuitable for dredged material disposal. Ocean dumping at the other two sites, one near Smelt Bay, Cortes Island and the other off McNaughton Point on the Sechelt Peninsula, may also result in localized degradation of adjacent resources. However the relatively low biological and economic value of these resources, and the compliance of these areas with oceanographic criteria which ensure minimal release of contaminants from disposed material, deem these two sites as the most environmentally suitable locations for contaminated dredged material disposal within the Strait. The principal requirements for further information to ensure the validity of the arguments presented in this thesis are considered to be the following: I. Oceanographic information is required to characterize the near sea-bed environment at the recommended disposal sites. The principal data requirements are the dissolved oxygen regime at the 224 sediment-water interface (to ensure the maintenance of oxygenated conditions that will minimize dissolution of trace metals); sea-bed current profiles (to ensure that sediment resuspension will be minimal); and sediment deposition rates (to determine the need for capping the disposal sediment mound with clean sediments). 2. An ecologically appropriate and practicable testing procedure is required to measure the quality of dredged material proposed to be disposed at the recommended sites. Specifically, we need to distinguish between dredged materials that can be disposed at the sites with negligible ecological impact and those where the environmental risks remain unacceptable. 3. Associated with this second information reguirement, further research is required to define the potential long-term distribution and fate of toxic contaminants from dredged material disposed into our coastal waters. Substantial uncertainties of knowledge have been identified by this review with regards to the long term potential for chemical release of toxic trace metals and the biologically induced release of trace metal and organic contaminants. The required oceanographic information is comparatively straight-forward and well within the realms of practicable environmental investigation. An adequate data base could readily be obtained within a twelve month research program, and should be a prerequisite to the use of any area proposed for ocean dumping. 225 The latter two requirements for information present considerably greater demands upon our research committment. The last decade has witnessed a vigorous research effort into the impacts of ocean dumping, but the results have been freguently inconsistent or otherwise inconclusive, and have failed to calm our environmental concerns. This is primarily a reflection of the complexity and inherent variability in the characteristics of both dredged materials and the receiving environment. There are no clear case studies from which one can infer that contaminated dredged materials have impacted the food chain (Kester et al., 1983). But it is clear that the physically, chemically or biologically induced release of contaminants from disposed sediments may be possible in time and under certain conditions. Credible and practicable procedures are reguired to assess dredged materials for their potential to deleteriously impact the environment proximate to the disposal site. The problem is that the science of measuring, much less predicting, the fate and effects of pollutants in the marine environment is in its infancy. Dredged Sediment Cr i ter ia Several attempts have been made to establish criteria to distinguish natural and contaminated sediments. In Canada, the Ocean Dumping Control Ac t (1975) specifies bulk sediment concentrations for mercury (acid digestible fraction <Q.75 mg/kg), cadmium (acid digestible fraction <0.60 mg/kg), persistent plastics (4% by volume) and oils and 226 grease (10.0 mg/kg of n-hexane extractables), beyond which ocean disposal is prohibited. A poorly defined criterion for organohalogens prohibits ocean dumping when the concentration of these compounds exceeds 0.1 parts of a concentration shown to be toxic to sensitive marine organisms: there is no definition of test organisms or bioassay protocols. The Act also "restricts" ocean dumping of materials containing arsenic, lead, copper, zinc, organosilicon, cyanides, flourides, pesticides not prohibited as organohalogens, beryllium, chromium, nickel and vanadium. Although no criteria have been specified for restricted metals under the Act, permitting currently relies on bulk concentration data (R. Kussat, EPS, pers. comm., 1984). It is now well recognized that bulk chemical analyses provide a poor basis for assessing the biological availability or mobilization potential of contaminants, particularly trace metals, in sediments. For example, Hirsch, Di Salvo and Peddicord (1978) concluded that there was little or no correlation between bulk sediment trace metal content and environmental impact, and recommended development of site-specific toxicity evaluation procedures using sediment bioassays. In an extensive review of the impacts of ocean dumping on the Great Lakes, the Board of the International Joint Commission (1978) expressed similar concerns, stating that research results "have shown no relationship between bulk chemical content for .sediment and the adverse effect that may result from open water disposal of that sediment". And ..."in other words, dredged material may be classified as polluted on the basis of bulk analyses, while in fact few of the contaminants may be available to the environment". The corollary of 227 this, that bulk chemical analyses may allow disposal of material where a low bulk concentration belies the environmental significance of its high availability to the biological system, has also been found to occur. Chapman and Barlow (1984) analysed sediments from 12 harbour sites in British Columbia, and recorded some of the highest toxicities in sediments where the concentrations of regulated metals were comparatively low. The continued use of bulk concentrations to characterize the potential environmental effects of dredged material disposal merely confuses an already complex question. Studies which have attempted to devise other reliable chemical methods to predict biological and ecological effects of trace metals in marine sediments have had limited success (e.g. Hall and Bindra, 1979; Luoma and Bryan, 1978) due to the extreme complexity of the predictive relationship involved (Burton, 1979). Predictive relationships become even more difficult to compute when more than one con-taminant is present, as is invariably the case in the "real world". It is generally agreed that no adeguate chemical procedures exist for assessing the ecological effects of contaminated sediment disposal (U.S. E P A / C O E , 1977). The criteria adopted in the U.S.A. (Federal Register 142, No. 7, 11 January, 1977) are much more comprehensive than those used in Canada, and emphasize biological effects of the contaminants rather than simple chemical presence. The cornerstone to these criteria is the Elutriate Test, where the dredged material is experimentally mixed with disposal site water under defined operational conditions, 228 then the liquid, suspended particulate and solid phases are each separated for evaluation. The independent evaluation of the liquid and suspended particulate phases is considered to correspond with potential contaminant releases via desorption and erosion/resuspension respec-tively, with the solid phase containing residual contaminants. The evaluation requirements for these three phases are notable for their requirement for extensive use of bioassay toxicity testing and testing for bioaccumulation potential. Bioassay toxicity responses indicate overall sediment contamination including both inorganic and organic compounds, as well as synergistic/antagonistic reactions between chemicals. Bioaccumulation testing indicates the potential for contaminants to be released from disposed dredged material via entry into the food chain. As such, these tests provide data of more direct relevance to environmental effects of ocean dumping than do chemical analyses alone. The principal objections to the adoption of the U.S. criteria in Canada are that they oblige proponents of ocean dumping to conduct very prolonged and expensive sediment analyses. Concern has also been expressed in the U.S. that the strict application of these criteria unreasonably restricts opportunities for environmentally acceptable ocean dumping (Brannon, 1978). Implicit recognition of these d i f f i -culties by the U.S. EPA has been indicated by their frequent resort to their authority to waive requirements of strict compliance with the evaluation criteria, even in cases where substantial contamination has been suspected (Kamlet, 1983). 229 The need for credible and practicable procedures to assess dredged materials for their potential environmental effects remains. To be credible, these assessment procedures must be sensitive, ecologically relevant and analytically reliable. To be practicable, their routine inclusion within the ocean dumping permitting process must be economically and logistically feasible, and the data produced must facilitate management decisions. Obviously, these objectives are more easily stated that achieved. However a beginning would be made by discarding the requirement for bulk chemical analyses of sediments and replacing it with potentially more relevant evaluative criteria. Chapman (1984) has suggested sediment bioassay testing incorporating acute lethal, sublethal and genotoxic endpoints. Using suitably sensitive indicator species, the sediment bioassay provides an integrated, rapid and inexpensive method of pollution assessment whose validity is not dependent on correlation with sediment chemistry. Death as an end-point in acute lethal tests provides compelling and convincing evidence for decision makers, while sublethal effects determined in more sensitive tests provide more subtle indications of possible ecological effects. The lack of toxicity in sediments tested by a range of sensitive bioassay technigues provides similarly convincing evidence that the material is not prone to causing gross ecological degradation in the marine environment. Sediment bioassay testing using the sensitive benthic amphipod Rhepoxynius abronius, described by Swartz et al. (1982), is widely used by the U.S. E P A as a screening tool in environmental monitoring, and 230 consistent results have been determined in interlaboratory calibration testing (A. Mearns, U.S. NOAA, pers. comm., 1985). It, and an oyster larvae bioassay test (Chapman and Morgan, 1983), have recently been adopted by the U.S. EPA as part of their interim decision criteria for disposal of contaminated dredged material at Four Mile Rock in Puget Sound (U.S. EPA, 1984). Conseguently, substantial data is accumu-lating which will permit detailed evaluation of the relevance and appropriateness of criteria based on these tests. The principal failure of bioassay procedures is that they assume that effects that cannot be readily measured in short-term laboratory tests either are not occurring or cannot be of environmental significance. They frequently underestimate the resourcefullness of the ecosystem; for example, the ability of microorganisms to alkylate inorganic forms of heavy metals with resulting enhancement of their toxicity and biological availability. Also, the use of bioassay toxicity tests as determinants of the suitability of dredged materials for ocean dumping emphasizes concern for environmental impacts at the disposal site. As was argued in Section 2.2 of this study, areally confined degradation at a disposal site is of less concern to ocean dumping management than is release and dispersion of contaminants. The Need for Field Assessment The bottom line is that the long-term environmental impacts of dredged material disposal is only truly assessible in the field. Long-term degradation will first manifest itself in the form of sub-231 lethal effects occurring within the biological community that inhabits the seabed at or near the disposal site. To assess these effects reguires carefully designed studies of the physiology and chemical composition of organisms within this community. These organisms have successfully overcome the short-term effects of turbidity and burial and represent integrators of the multifarious processes of contaminant release. Whatever sediment analysis procedures and criteria are established under the permitting process, comprehensive field monitoring should be conducted to coordinate and correlate with the laboratory tests. Dependent on whether adverse effects are found in the field, then the laboratory tests can be adjusted with suitable correction factors or safety margins so they better serve their predictive function. The recommended disposal sites at Smelt Bay and McNaughton Point are each in sufficiently shallow water that intensive biological sampling remains within technical and financial feasibility. Given the many uncertainties that preclude definitive prediction of the environmental effects of ocean dumping, it is recommended that the initial use of the selected site should be for the purpose of large scale field experimentation. Chemical and biological monitoring at the perimeter of the disposal site should emphasize guantification of the nature and extent of contaminant release. The results could then be applied to test the site-specific relevance and appropriateness of a range of procedures for determining sediment guality. In the unlikely case that contaminant release from the disposed sediments exceeds environmental acceptability, remedial actions (e.g. capping, redredging) 232 would have to be implemented. Redredging disposed sediments in 100-125 m of water would be expensive, but is technically and logistically feasible. 233 10.0 LITERATURE CITED Aller, R.C. 1978. Experimental studies of changes produced by deposit feeders on pore water, sediment, and overlying water chemistry. Am. J . Sci., 278, 193-237. Aller, R.C. and J.Y. Yingst. 1978. Biogeochemistry of tube-dwellings: A study of the sedentary polychaete Amphitrite ornata (Feidy). Journal of Marine Research, 35, 201-254. American Institute of Biological Sciences. 1978. Aquatic hazards of pesticides task group. In: J . Cairns, K.L. Dickson and A.W. Macki (eds.) Estimating the hazard of chemical substances to aquatic life. ASTM STP 657. Andreae, M.O. 1978. Distribution and speciation of arsenic in natural waters and some marine algae. Deep-sea Res. 25: 391-402. Arimoto, R. and S.Y. Feng. 1983. Changes in the levels of PCBs in Mytilus  edulis associated with dredged material disposal. In: D.R. Kester et al. (eds.). Wastes in the ocean. Vol. 2. Dredged material disposal in the ocean. Wiley Interscience. pp. 199-213. Bailey, G.W. and J.L. White. 1965. Herbicides. Residue Rev., 10: 97-122. Baker, M.D., P.T.S. Wong, Y.K. Chau, C.I. Mayfield and W.E. Inniss. 1981. Methylation of lead, mercury, arsenic and selenium in the acidic aquatic environment. In: Proc. Int. Conf. Heavy Metals in the Environment, Amsterdam, Sept. 1981, C E P Consultants Ltd., pp. 645-648. Barker, M.L. 1974. Water resources and related land uses, Strait of Georgia -Puget Sound Basin. Environment Canada, Lands Directorate, Geographical Paper No. 56. Berner, R.A. 1971. Principles of chemical sedimentology. McGraw Hi l l , New York. 240 pp. Biddinger, G.R. and S.P. Gloss. 1984. The importance of trophic transfer in the bioaccumulation of chemical contaminants in aquatic ecosystems. Residue Reviews, 91: 103-145. Bindra, K.S. 1983. Mobilization of selected trace metals/in the aquatic environment: sediment to water column and benthic invertebrates. Ph.D. Thesis, Civi l Engineering, U.B.C., Vancouver. Bisogni, Jr., J . J . and A.W. Lawrence. 1975. Kinetics of mercury methylation in aerobic and anaerobic aquatic environments. J . Water Pollut. Contr. Fed. 47: 135-152. 234 Black, C P . 1965. Soil Chemical Analysis. Madison Wl. Am. Soc. Agronomy. Blazevich, J.N., A.R. Gahler, G.J. Vasconcelos, R.H. Rieck, and S.V.W. Pope. 1977. Monitoring of trace constituents during PCB recovery dredging operations, Duwamish Waterway. Report No. EPA-910/9-77-039, U.S. Environmental Protection Agency, Region X, Seattle, Washington. 67 pp and 6 appendices. Boehm, P.D. and D.L. Fiest. 1983. Ocean dumping of dredged material in the New York Bight: organic chemistry studies. In: D.R. Kester et al. Wastes in the ocean. Vol. 2. Dredged material disposal in the ocean. Wiley and Sons. pp. 151-172. Bokuniewicz, H.J., J.A. Gerbert, R.B. Gordon, J .L. Higgins, P. Kaminsky, C.C. Pilbeam, M.W. Reed and C B . Tuttle. 1978. Field study of the mechanics of the placement of dredged material in open water disposal sites. Report D-78-F. Vol. I. U.S. C O E Waterways Expt. Station, Vicksburg. 94 pp. Bokuniewicz, H.J., J.A. Gerbert, R.B. Gordon, J .L. Higgins and C B . Tuttle. 1977. Field study of the effects of storms on the stability and fate of dredged material in subaqueous disposal areas. Tech. Rep. D-77-22. U.S. Army Engineers Waterways Expt. Station, Vicksburg, Miss. Borneff, J . and H. Kunte. 1965. Carcinogenic substances in water and soil: about the origin and evaluation of P A H in water. Arch. Hyg. Bakt. 149: 226. Boyd, B., R.T. Saucier, J.W. Keeley, R.L. Montgomery, R.D. Brown, D.B. Mathis and C .J . Guice. 1972. Disposal of dredge spoil, problem identification and research program development. Tech. Rep. D-72-8. U.S. Army Eng. Waterways Expt. Station, Vicksburg. Braman, R.S. 1976. Molecular forms of arsenic in the environment, jn: R.W. Andrew, P.V. Hodson and D.E. Konasewich (eds.), Proceedings of a Workshop on Toxicity to Biota of Metal Forms in Natural Waters. International Joint Commission. Windsor, Ontario. Brandsma, M.G. and D.J. Divoky. 1976. Development of models for prediction of short-term fate of dredged material discharged in the estuarine environment, Technical Report D-76-5, U.S. Army Engineer Waterways Experiment Station, C E , Vicksburg. Brannon, J.M. 1978. Evaluation of dredged material pollution potential. Technical Report D-78-6, U.S. Army Engineers Waterways Experiment Station, C.E., Vicksburg. 235 Brannon, J.M., R.M. Engler, J.R. Rose, P.G. Hunt and I. Smith. 1976. Distribution of toxic heavy metals in marine and freshwater sediments, jn: Dredging and Its Environmental Effects, P.A. Krenkel, J . Harrison and J .C . Burdick III (Eds.). American Society of Civi l Engineers, New York, pp. 455-495. Brannon, J.M., R.H., Plumb and I. Smith. 1978. Long-term release of contaminants from dredged material. Tech. Rep. D-78-49. U.S. C O E Waterways Experiment Station, Vicksburg. Branson, D.R., G.E. Blau, H.C. Alexander and W.B. Neely. 1975. Bioconcentra-tion of 2,2',4,4'-tetrachlorobiphenyl in Rainbow Trout as measured by an accelerated test. Trans. Am. Fish Soc. 4: 784-792. British Columbia Ministry of Environment. No date. Flight line data sheets. Surveys and Resource Mapping Branch, Victoria. B.C. Place, B.C. Ministry of Environment, Fisheries and Oceans Canada, Environment Canada. 1983. B.C. Place False Creek redevelopment marine sediment disposal problem. Unpublished briefing summary, March 1983. Brothers, D.E. and D.L. Sullivan. 1984. Flase Creek benthic sediment survey 1982/83. Dept. of Environment, EPS, Pacif ic and Yukon Regional Program 84-08. Vane , B.C. Brown, B.E. 1976. Observations on the tolerance of the isopod Asellus  meridianus Rac. to copper and lead. Water Res. 10: 555-559. Brown, D.A., C.A. Bawden, K.W. Chatel and T.R. Parsons. 1977. The wildlife community of lona Island jetty, Vancouver, B.C. and heavy metal pollution effects. Environm. Conserv. 4: 213-216. Brown, D.S., and E.W. Flagg. 1981. Emperical prediction of organic sorption in natural sediments, J . Environ. Qual., 10(4): 382. Brown, R.F., V.D. Chakley and D.G. Demontier. 1977. Preliminary catalogue of salmon streams and spawning escapements of Statistical Area 14 (Comox-Parksville). Canada Fisheries and Marine Service, Data Report PAC/D-77-12. Bryan, G.W. 1979. Bioaccumulation of marine pollutants. Phil. Trans. R. Soc. Lond. Ser. B286, 483. Bryan, G.W. 1976. Metal Contamination in the sea. ]n: R. Johnston, Ed., Marine Pollution, New York: Academic Press, Inc., 185-302. 236 Bryan, G.W. and L.G. Hummerstone. 1973. Adaptation of the polychaete Nereis  diversicolor to estuarine sediments containing high concentrations of zinc and cadmium. J . Mar. Biol. Assoc. of the U.K. 53: 839-857. Buikema, A.L., M.J. McGinniss and J . Cairns. 1979. Phenolics in aquatic ecosystems: a selected review of recent literature. Mar. Environ. Res., 2: 87-181. Burks, S.L. and R.M. Engler. 1978. Water quality impacts of aquatic dredged disposal (laboratory investigations). Synthesis report DS-78-4. U.S. C O E Waterways Expt. Station, Vicksburg. 35 pp. Burns, K and J . Teal. 1979. The West Falmouth oil spill: hydrocarbons in the salt marsh ecosystem. Estuar. Coastal Mar. Sci. 8: 349-360. Burton, J.D. 1979. Physical - chemical limitations in experimental investi-gations. Phil. Trans. R. Soc. Land. B 286: 443-456. Cain, R.T., M.J.R. Clark and N.R. Zorkin. 1979. Fraser River Estuary Study, Water Quality, The trace organic constituents in discharges. Min. of Environment, Victoria, B.C. Calabrese, A., F.P. Thurberg, M.A. Dawson and D.R. Wenzloff. 1975. Sublethal physiological stress induced by cadmium and mercury in winter flounder. ]n: J .H. Koeman and J.J.T.A.W. Stik (eds.), Sublethal Effects of Toxic Chemical in Aguatic Animals. Elsevier Publications, New York. pp. 15-21. Calambokidis, J . and Other Students. 1978. Chlorinated hydrocarbon con-centrations and the ecology and behaviour of harbor seals in Washington State waters, a student-oriented study. The Evergreen State College, Olympia, Washington. 121 pp. Calambokidis, J . , J . Mowrer, M. Beug and S. Herman. 1979. Selective retention of polychlorinated biphenyl components in the mussel, Mytilus edulis. Arch. Environ. Contam. Toxicol. 8: 299-308. Callahan, M.A. and M.W. Slimak. 1979. Water-related environmental fate of 129 priority pollutants. EPA 440/4-79-029a. Canadian Hydrographic Service. 1983. Current Atlas. Juan de Fuca Strait to Strait of Georgia. Dept. of Fisheries and Oceans, Ottawa, Ont. Carey, A.E. and G.F. Harvey. 1978. Metabolism of polychlorinated biphenyls by marine bacteria. Bull. Environ. Contam. Toxicol. 20: 527-534. 237 Carpenter, R. and A.W. Fairhall. 1979. Hydrocarbon studies in Puget Sound and off the Washington coast. Progress Report. March 78 - Feb. 79. U.S. Dept. of Energy. Contract No. E476-S-06-2225-TA40. Dept. of Oceanography, Univ. of Washington, Seattle. 28 pp and appendix. Carson, R. 1962. Silent Spring. Houghton Mifflin Co., Boston, Mass. Cec i l , H.C., J . Betman, R.J. L i I lie and J . Verrett. 1974. Embryotoxic and teratogenic effects in unhatched fertile eggs from hens fed PCBs. Bull. Environ. Contam. Toxicol. 11(6): 489-495. Chapman, P.M. 1984. Sediment bioassay tests provide toxicity data necessary for assessment and regulation. Paper presented at the Nth Annual Aquatic Toxicity Workshop, November 13-15, 1984, Vancouver, B.C. Chapman, P.M. and C.T. Barlow. 1984. Sediment bioassays in various B.C. coastal areas. Unpubl. report prepared by E.V.S. Consultants for the Environmental Protection Service. Chapman, P.M. and J.D. Morgan. 1983. Sediment bioassays with oyster larvae. Bull. Environm. Contam. Toxicol. 3U 438-444. Chapman, P.M., D.R. Munday, J . Morgan, R. Fink, R.M. Kocan, M.L. Landolt and R.N. Dexter. 1983. Survey of biological effects of toxicants upon Puget Sound biota. II. Tests of reproductive impairment. N O A A Tech. Rep. NOS 102 OMS-I. 58 pp. Chapman, P.M., D. Munday and G.A. Vigers. 1982. Monitoring program for heavy metals and elemental sulphur at Vancouver Wharves Ltd. Unpublished report to Vancouver Wharves Ltd. by E.V.S. Consultants Ltd., North Vancouver. Chapman, W.H., H.L. Fisher and M.W. Pratt. 1968. Concentration factors of chemical elements in edible aguatic organisms. UCRL-50564. Un i -versity of California, Livermore. Lawrence Radiation Lab. 50 pp. Chau, Y.K., P.T.S. Wong, B.A. Silverberg, P.L. Luxon and C.A. Bengert. 1976. Methylation of selenium in the aguatic environment. Science 192: 1130-1131-Chen, K.Y., S.K. Gupta, A.Z. Sycip, J.C.S. Lu, M. Knezevic and W.W. Choi. 1976. Research study on the effect of dispersion, settling and resed-imentation on migration of chemical constituents during open water disposal of dredged materials. U.S. Army Corps of Engineers, D.R.M.P. Report D-76-1. Waterways Expt. Station, Vicksburg, Miss. 238 Chen K.Y., J.C.S. Lu and A.Z. Sycip. 1976. Mobility of trace metals during open water disposal of dredged material and following resedimentation. Proc. Spec. Conf. Dredging and its Environmental Effects, Mobile, A la. New York: A S C E pp. 435-454. Chiou, C.T., L.J. Peters and V.H. Freed. 1979. A physical concept of soil-water equilibria for nonionic organic compounds. Science. 206: 831-832. Clayton, J.R., Jr., S.P. Pavlou and N.F. Breitner. 1977. Polychlorinated biphenyls in coastal marine zooplankton: Bioaccumulation by equil-ibrium partitioning. Env i ron . Sc i .Techno l . i l : 676-682. Connell, D.W. and G.J. Miller. 1984. Chemistry and ecotoxicology of pollution. John Wiley and Sons. 444 pp. Cox, R.K. and E.M. Charman. 1980. A survey of the geoduck clam (Panope  qenerosa) in Queen Charlotte, Johnstone and Georgia Straits, B.C. Fisheries Dept. Report No. 16. B.C. Min of Environment, Mar. Res. Branch, Victoria. Custar, C . and T. Wakeman. 1977. Dredge disposal study, San Francisco Bay and estuary. U.S. C O E San Francisco. 83 pp. DeMora, S.J. 1981. Manganese chemistry in the Fraser River estuary. Ph.D. Thesis, University of British Columbia, Vancouver, B.C. 177 p. Department of Environment. 1978. Canada Land Inventory: land capability for recreation. Map. Lands Directorate, Ottawa. Department of Environment and Department of National Health and Welfare. 1979. Canadian environmental contaminants act priority chemicals. Canada Gazette, Part I, I December, 1979. p. 7365. Department of Fisheries and Oceans. 1980. Pacif ic cod. Fact sheet. Information Branch, Vancouver. Dexter, R.N., D.E. Anderson, E.A. Quinlan, W. Horn, and S.P. Pavlou. 1979. Long-term impacts induced by disposal of contaminated river sediments in Ell iott Bay, Seattle, Washington: results from studies conducted in 1979. U.S. Army Corps of Engineers Waterways Experiment Station, Vicksburg. 126 pp. DiToro, D.M. and L.M. Horzempa. 1982. Reversible and resistant components of P C B adsorption-desorption: isotherms, Environ. Sci. .Technol., 16(9): 594. 239 Ducks Unlimited. No date. Flight line data sheets, Delta, B.C. Dunn, B.P. and H.F. Stich. 1976. Monitoring procedures for chemical carcino-gens in coastal waters. J . Fish. Res. Bd. Can., 33: 2040-2046. Eder, G. 1976. Polychlorinated biphenyls and compounds of the DDT group in sediments of the central North Sea and the Norwegian Depression. Chemosphere 2: 101-106. Eganhouse, R.P. and J.A. Calder. 1976. The solubility of medium molecular weight aromatic hydrocarbons and the effects of hydrocarbon co-solutes and salinity. Geochim. Cosmochim. Acta. 40: 555-561. Engel, D.W. and B.A. Fowler. 1979. Factors influencing cadmium accumulation and its toxicity to marine organisms. Environ. Health Persp. 28: 81-88. Engler, R.M. 1981. Impacts asociated with the discharge of dredged material: Management approaches. In: Use of the Ocean for Man's Wastes: Engineering and Scientific Aspects. Marine Board, National Research Council, National Academy Press, Washington, D . C , pp. 129-185. Esser, H.O. and P. Moser. 1982. An appraisal of problems related to the measurement and evaluation of bioaccumulation. Ecotoxicol. Environ. Safety. 6, 131. Evans, D.W. and N.H. Cutshall. 1973. Effects of ocean water on the soluble-suspended distribution of Columbia River radionuclides. ]n: Radio-active contamination of the marine environment. I.A.E.A. pp. 125-138. E.V.S. Consultants. 1984. Study of the rate of uptake and bioaccumulation of contaminants from marine sediments for ocean dumping. Report prepared for Environmental Protection Service, West Vancouver. Fabricand, B.P., R.R. Sawyer, S.G. Ungar and S. Adler. 1962. Trace metals in the ocean by atomic absorption spectroscopy. Geochim. Cosmochim. Acta., 26, 1023-1027. Ferguson, K.D. and K.J. Hall. 1979. Fraser River Estuary Study, Water Quality, Stormwater Discharges. Environmental Protection Service, Environ-ment Canada; Westwater Research Centre, U.B.C. Fisheries and Environment Canada. 1977. Crab, shrimp and prawn: British Columbia fishery resources. Fisheries Information Bulletin 77-I0E. Forrester, C.R., K.S. Ketchen and C.C. Wong. 1972. Mercury content of spriny dogfish in the Strait of Georgia, B.C. J . Fish. Res. Bd. Can., 29: 1487-1490. 240 Forstner, U. and S.R. Patchineelam. 1976. Chemical associations of heavy metals in polluted sediment from the Lower Rhine River, jn: Part i -culates in water, Advances in chemistry. Am. Chem. Soc. 189, 177-193. Forstner, U. and G.T.W. Wittman. 1981. Metal Pollution in the Aquatic Environment. Second Revised Edition. Springer-Verlag, Berlin, 486 pp. Frazier, J.M. 1979. Bioaccumulation of cadmium in marine organisms. Environ. Health Persp. 28: 75-79. Fulk, R., D. Gruber and E. Wullschleger. 1975. Laboratory study of the release of pesticide and PCB materials to the water column during dredging and disposal operations; Dredged Materials Research Program, U.S. Army Engineer Waterways Expt. Station, NTIS Report No. AB-A026-685. Furukawa, K. and F. Matsumura. 1976. Microbial metabolism of polychlorinated biphenyls. Studies on the relative degradability of PCB components by Alkaliqenes sp. J . Agric. Food Chem. 24: 251-256. Gambrell, R.P., R.A. Khalid, V.R. Col lard, C.N. Reddy and W.H. Patrick, Jr . 1976. The effect of pH and redox potential on heavy metal chemistry in sediment-water systems affecting toxic metal bioavailabiltiy. jn: Dredging: Environmental Effects and Technology. Proc. World Dredging Conference, San Pedro/Calif. Gambrell, R.P., R.A. Khalid and W.H. Patrick, Jr. 1976. Physiochemical parameters that regulate mobilization and immobilization of toxic heavy metals. jn: Dredging and Its Environmental Effects, P.A. Krenkel, J . Harrison, and J .C . Burdick III (Eds.). American Society of Civi l Engineers, New York, pp. 418-434. Gambrell, R.P., R.A. Khalid, M.G. Verloo and W.H. Patrick. 1977. Trans-formations of heavy metals and plant nutrients in dredged sediments as affected by oxidation reduction potential and pH. II. Materials and methods/results and discussion. U.S. Army Corps of Engineers, Dredged Material Research Program. Vicksburg, Miss., Rept. D-77-4, 309 pp. Gardner, W.S., R.L. Lee, K.R. Tenore and C.W. Smith. 1979. Degradation of selected polycyclic aromatic hydrocarbons in coastal sediments: im-portance of microbes and polychaete worms. Water, Air, Soil Pollut. II: 339-347. Garrett, C L . 1983. An overview of PCBs and their current status in British Columbia. Dept. of Environment, EPS Pacif ic and Yukon Regional Program. Report 83-16, Vancouver, B.C. 241 Garrett, C.L. 1982. Pacif ic and Yukon region toxic chemicals profile. Environment Canada, Pacif ic and Yukon Region. 259 p. Garrett, C.L. 1980. Fraser River Estuary Study, Water Quality; Toxic Organic Contaminants. Environmental Protection Service, Environment Canada, Vancouver, B.C. Garrett, C.L., L.A. MacLeod and H.J. Sneddon. 1980. Mercury in the British Columbia and Yukon Environments - Summary of data to January, 1979. Dept. of Environment, EPS Pacif ic Regional Program Report 80-4. GESAMP. 1977. Impact of oil on the marine environment. Rep. Study No. 6. Food and Agriculture Organization. Rome. Giam, C.S., H.S. Chan, G.S. Neff, and E.L. Atlas. 1978. Phthalate ester plasticizers: a new class of marine pollutant. Science 199: 419-420. Gibson, D.T. 1976. Microbial degradation of carcinogenic hydrocarbons and related compounds. In: Symp. Proc. Sources, effects and sinks of hydrocarbons in the aquatic environment. Washington D.C. August 9-I I, 1976, American Inst. Biol. Sci., Washington D.C. pp. 224-238. Gilbert, T.R., A.M. Clay and D.A. Leighty. 1976. Influence of the sed-iment/water interface on the aquatic chemistry of heavy metals. Environmental Chemistry Research Div., Tyndall A i r Force Base, Fla., Rep. AFCEC -TR -6 -22 , 89 pp. Gledhill, W.E., R.G. Kaley, W.J. Adams, 0 . Hicks, P.R. Michael, V.W. Saeger and G.A. LeBlanc. 1980. An environmental safety assessment of butyl benzyl phthalate. Environ. Sci. Technol. 14: 301-305. Goldbach, R.W., H. van Genderen, and P. Leeuwangh. 1976. Hexachloro-butadiene residues in aquatic fauna from surface water fed by the river Rhine. Sci. Total Environ. 6: 31-30. Gordon, R.B. 1974. Dispersion of dredge spoil dumped in near-shore waters. Estuar. Coast. Mar. Sci. 2: 349-358. Gordon, D., J . Dale and P. Keizer. 1978. Importance of sediment reworking by the deposit-feeding polychaete Arenicola marina on the weathering rate of sediment-bound oil, J . Fish. Res. Bd. Can. 35: 591-603. Grant, N. 1971. Mercury in man. Environment. 13:2. 242 Green, F.A., Jr. and J.M. Neff. 1977. Toxicity, accumulation, and release of three polychlorinated naphthalenes (Halowax 1000, 1013, 1099) in post-larvael and adult grass shrimp, Palaemonetes pugio. Bull. Environ. Contam. Toxicol. 17: 399-407. Greig, R.A. and D.R. Wenzloff. 1978. Metal accumulation and depuration by the American oyster, Crassostrea virginica. Bull. Environ. Contam. Toxicol. 20: 499-504: Hall, K.J. 1976. The quality of water in the lower Fraser River and sources of pollution. In: A.H.J. Dorcey (ed.) The Uncertain Future of the Lower Fraser, Westwater Research Centre, U.B.C., Vancouver, B.C. pp. 21-48. Hall, K.J. and K.S. Bindra. 1979. Geo-chemistry of selected metals in sediments and factors affecting organism concentration, pp. 337-340. jn: Inter-national Conference, Management and Control of Heavy Metals in the Environment. London, Sept. 1979. C E P Consultants Ltd., Edinburgh. Hall, K.J., P. Parkinson and T. Ma. 1983. Determination of selected trace organic contaminants in marine sediments. Unpublished report by Westwater Reserach Centre and Dept. of Civi l Engineering, U.B.C. to Environmental Protection Service. 3 pp. Hall, K.J. and J .H. Weins. 1976. The guality of water in tributaries of the lower Fraser and sources of pollution. In: A.H.J. Dorcey (ed.). The Uncertain Future of the Lower Fraser. Westwater Research Centre, U .B .C, Vancouver, Canada, pp. 49-84. Hall, K.J., I. Yesaki and J . Chan. 1976. Trace metals and chlorinated hydrocarbons in the sediments of a metropolitan watershed. Westwater Research Centre, Tech. Rep. 10., U.B.C, Vancouver. Hallberg, R.O. 1974. Metal distribution along a profile in an intertidal area. Estuar. Coast. Mar. Sci. 2. 153-170. Haque, R., J . Falco, S. Cohen and C. Riordan. 1980. Role of transport and fate studies in the exposure, assessment and screening of toxic chemicals. In: R. Haque (ed.). Dynamics, exposure and hazard assessment of toxic chemicals. Ann Arbor Science, Ann Arbor, Michigan, pp. 47-67. Hartung, R. 1976. Pharmokinetic approaches to the evaluation of methyl-mercury to fish, jn: R.W. Andrew, P.V. Hodson, D.E. Konasewich (eds.), Toxicity to Biota of Metal Forms in Natural Water (workshop proceedings). 1975. Great Lakes Reg. Off. International Joint Commi-ssion, Windsor, Ontario, pp. 233-248. 243 Hartung, R. and G.W. Klingler. 1970. Concentration of DDT in sedimented polluting oils. Envir. Sci. & Technol. 4:5, pp. 407-410. Hassett, J .J . , J .C . Means, W.L. Banwart, S.G. Wood, S. Al i and A. Khan. 1980. Sorption of dibenzothiophene by soils and sediments, J . Environ. Qual., 9(2), 184. Herman, S.G. and J . Calambokidis. 1978. Personal communications to Ed Long, MESA Project Off ice, Seattle. Cited by E. Long during address to the Regional Ocean Dumping Advisory Council Annual Workshop, I.O.S., Saanich, B.C., Apri l , 1985. Hiraizumi, Y., M. Takahashi and H. Nishimura. 1979. Adsorption of poly-chlorinated biphenyl onto seabed sediment, marine plankton, and other adsorbing agents. Environ. Sci. Technol. 13: 580-584. Hirsch, N.D., L.H. Di Salvo and R. Peddicord. 1978. Effects of dredging and disposal on aquatic organisms. Technical Report DS-78-5 by Naval Biosciences Laboratory, University of California under contract to U.S. Army Engineer Waterways Experiment Station, C E , Vicksburg, Miss. 41 P-Hites, R.A., R.E. LaFlame and J.W. Farrington. 1977. Sedimentary polycyclic aromatic hydrocarbons. The historical record. Science 198: 829-831. Hollifield, H.C. 1979. Rapid nephelometric estimate of water solubility of highly insoluble organic compounds of environmental interest. Bull. Environ. Contam. Toxicol. 23: 579-586. Horn, W. 1978. Polychlorinated biphenyls in northern Puget Sound. M.Sc. Thesis, Dept. of Oceanography, Univ. of Washington, Seattle. 91 pp and appendix. Horn, W., R.W. Riseborough, A. Soutar and D.R. Young. 1974. Deposition of DDE and PCB in dated sediments of the Santa Barbara Basin. Science 184(4142), I 197-1 199. Hoos, R.A.W. 1977. Environmental assessment of an ocean dumpsite in the Strait of Georgia, British Columbia. Environmental Protection Service, Pacif ic Region, Report No. EPS-5-PR-77-2. Home, R.A. 1969. Marine chemistry: the structure of water and the chemistry of the hydrosphere. Wiley Interscience. 244 Hourston, A.S. 1982. Publications and reports on Pacif ic herring arising from investigations conducted at or in cooperation with the Pacific Bio-logical Station - Vol. 2, 1977-1981. Department of Fisheries and Oceans, Resource Services Branch, Nanaimo. Huang, J . and C. Liao. 1970. Adsorption of Pesticides by Clay Minerals. J . of San. Engineering 96: 1057-1075. Hunt, J.R. 1984. Incorporating particle coagulation into a transport model for predicting the fate of particle wastes in the ocean. Presented at the 5th International Ocean Disposal Symposium, Corvallis, Oregon, 10-14 September, 1984. Hunt, C D . and D.L. Smith. 1981. Remobilization of metals from polluted marine sediments, jn: Ocean Pollution 1981, Halifax, Nova Scotia. Hutzinger, O., S. Safe and V. Zitko. 1974. The Chemistry of PCBs. Chemical Rubber Publishing Co., Cleveland, Ohio. International Joint Commission. 1978. Annual Report of the Great Lakes Advisory Board, July, 1978, Windsor, Ontario. International Joint Commission. 1977. The waters of Lake Huron and Lake Superior. Volume II (Part B). Windsor, Ontario. Jamieson, J.W.S. 1977(a). Polybrominated biphenyls in the environment. Environmental Protection Service, Economic and Technical Review Report EPS-3-EC-77-I8. Jamieson, J.W.S. 1977(b). Polychlorinated terphenyls in the environment. Environmental Protection Service, Economic and Technical Review Report EPS-3-EC-77-22. JBF Scientific Corp. 1975. Dredging technology study: San Francisco Bay and estuary. Report to U.S. C O E , San Francisco District. Contract DACW 07-75-C-0045. 240 pp. Jenne, E.A. and S.N. Luoma. 1977. Forms of trace metals in soils, sediments and associated waters, jn: Biological implications of metals in the environment, R.E. Wildung and H. Druker (eds.), Springfield, Virginia, pp. 110-143. Jphansen E. 1976. State of the art survey and evaluation of open water dredged material placement methodology. Contract Rep D-76-3. U.S. Army Engineers Waterways Expt. Station, Vicksburg, Miss. 245 Johnson, B.T., D.L. Stalling, J.W. Hogan and R.A. Schoettger. 1977. Dynamics of phthalate acid esters in aquatic organisms. Adv. Environ. Sci. Technol. 8: 283-300. Johnson, N.T., L.J. Albright, T.G. Northcote, P.C. Oloffs and K. Tsumura. 1975. Chlorinated hydrocarbon residue in fishes from the lower Fraser River. Westwater Research Centre, Tech. Report 9, U.B.C., Vancouver, B.C. Kamlet, K.S. 1983. Dredged material ocean dumping: perspectives on legal and environmental impacts. In: Kester D.R. et al. (eds), Wastes in the ocean, Vol. 2, Dredged material disposal in the ocean. Wiley and Sons, pp. 29-70. Karickoff, S.W. 1985. Pollutant sorption in environmental systems. In: W.B. Neely and G.E. Blau (eds.). Environmental exposure from chemicals. Vol. I. C R C Press, Boca Raton. Karickhoff, S.W. 1980. Sorption kinetics of hydrophobic pollutants in natural sediments, Contaminants and Sediments, 2, Baker, R.A., ed., 193. Karickhoff, S.W., D.S. Brown and T.A. Scott. 1979. Sorption of hydrophobic pollutants on natural sediments, Water Res., 13, 241. Kaufman, D.S. 1978. Degradation of PCP in soil, and by soil microorganisms. In: K.R. Rao (ed.). Pentachlorophenol. Plenum Press, New York. pp. 27-40. Kay, S.H. 1984. Potential for biomagnification of contaminants within marine and freshwater food webs. U.S. Army Engineers Waterways Expt. Station, Tech. Rept. D-84-7, Vicksburg, Miss. Keckes, S. and J.K. Miettinen. 1972. Mercury as a marine pollutant. In: M. Ruivo (ed.), Marine Pollution and Sea Li fe, F A O , Fishing News (Books) Ltd., London. Kenaga, E.D. and C.A.I. Goring. 1978. Relationship between water solubility, soil sorption, octanol-water partitioning and bioconcentration of chemi-cals in biota. In: Proceedings of the American Society for Testing and Materials, 3rd Aquatic Toxicology Symposium, STP 707, 78. Kennett, J.P. 1982. Marine Geology. Prentice-Hall, Englewood Cl iffs, New Jersey. 813 pp. Kester, D.R., B.H. Ketchum, I.W. Duedall, P.K. Park. 1983. .The problem of dredged material disposal, \n: Wastes in the ocean. Vol. 2. Dredged material disposal in the ocean. Kester et al. (eds). Wiley Interscience Publ. 3-28. 246 Khalid, R.A., W.H. Patrick, Jr. and R.P. Gombrell. 1978. Effect of dissolved oxygen on chemical transformations of heavy metals phosphorus and nitrogen in an estuarine sediment. Estuar. Coastal Mar. Sci. 6: 21-35. Kharkar, D.P., K.K. Turekian and K.K. Bertine. 1968. Stream supply of dissolved silver, molybdenum, antimony, selenium, chromium, cobalt, rubidium and cesium to the oceans. Geochim. Cosmochim. Acta . 32, 285-298. Klaassen, C D . 1980. Absorption, distribution and excretion of toxicants. In J . Doull, C D . Klaassen and M.O. Amdur (eds.). Toxicology - the basic science of poisons. 2nd ed. MacMillan, New York. Kobayashi, K. 1978. Metabolism of pentachlorophenol in fish. In: K.R. Rao (ed.). Pentachlorophenol. Plenum Press, New York, pp. 89-106. Koh, R . C Y . and Y . C Chang. 1973. Mathematical model for barged ocean disposal of wastes, EPA-660/2-73-029, U.S. Environmental Protection Agency, Washington, D . C Konemann, W.H. 1979. Quantitative structure-activity relationships for kinetics and toxicity of aquatic pollutants and their mixtures in fish. Univ. Utrecht, Netherlands. Kopperman, H.L., D.W. Juehl and G.E. Glass. 1976. Chlorinated carbons found in waste treatment effluents and their capacity to bioaccumulate. N.T.I.S. PB2653I0. Koroki, H. and Y. Masuda. 1978. Determination of polychlorinated dibenzofuran isomers retained in patients with Yusho. Chemosphere 10: 771-777. Kraueter, J . 1976. Biodeposition by salt-marsh invertebrates. Mar. Biol. 35. 215-223. Kreitzer, J .H. and G.H. Heinz. 1974. The effect of sublethal dosages of five pesticides and a PCB on the avoidance response of Coturnix guail chicks. Environ. Pollut., 6; 21-29. Krishnappan, B.G. 1975. Dispersion of granular material dumped in deep water, Scientific Series No. 55, Canada Centre for Inland Waters, Burlington, Ontario. Ladner, L. 1971. Biochemical model for the biological methylation of mercury suggested from methylation studies jn vitro with Neurospora crassa. Nature 230: 252-253. Lake, J.L., P.F. Rogerson and C B . Norwood. 1981. A polychlorinated dibenzofuran and related compounds in an estuarine ecosystem. Environ. Sci. Technol. 15: 549-553. 247 Lands Directorate. 1983. Coastal Resources Folio. South mainland coast. Environment Canada, Vancouver. Lands Directorate. 1981. Coastal Resources Folio. East coast of Vancouver Island. Environment Canada, Vancouver. Larson, R.A., D.W. Blankenship and L.L. Hunt. 1976. Toxic hyproperoxides: photochemical formation from petroleum constituents. In: Symp. Proc. Sources, effects and sinks of hydrocarbons in the aquatic environ-ment. Washington D . C , August 9-11, 1976, American Inst. Biol. Sc., Washington D . C pp. 198-208. Laseter, J.L., C K . Bartell, A.L. Laska, D . C Holmquist, D.B. Condie, J.-W. Brown and R.L. Evans. 1976. An ecological study of hexachloro-butadiene (HCBD). EPA 560/6-76-010. Lee, G.F. 1977. Significance of chemical contaminants in dredged sediment on estuarine water quality. Proc. Estuarine Pollution Control and Assess-ment. U.S. EPA Off ice of Water Planning and Standards, Washington, D.C. Vol. I: 211-216. Lee, G.F. 1976. Environmental impact of dredging, dredged material disposal and dredged material research in the U.S. Occasional Paper No. 10, Center for Environmental Studies, Univ. of Texas. Lee, G.F. 1975. Potential environmental problems of dredging and dredged material disposal. Proc. 6th Pacif ic Coast Conf. on Dredging. 62-68. Lee, G.F. 1973. Role of hydrous metal oxides in the transport of heavy metals in the environment. Proceedings of Conference on heavy metals in the environment, Vanderbilt Univ., December, 1973. Lee, G.F., C . McDonald, F. Salek, P. Bandyopadhyay, J . Butler, D. Homer, A. Jones, J . Lopez, C Mariani, M. Piwoni and M. Nicar. 1978. The development of criteria for dredged material disposal. Tech. Rep. D-78-4 U.S. A C O E Waterways Expt. Station, Vicksburg. Lee, G.F., M. Piwoni, J . Lopez, G. Mariani, J . Richardson, D. Homer and F. Saleh. 1975. Research study for the development of dredged material disposal criteria. U.S. Army C.O.E. Dredged Material Research Program, Vicksburg, Miss. Contract Report D-75-4. Lee, H. and R . C Swartz. 1980. Biological processes affecting the distribution of pollutants in marine sediments. Part II, Biodeposition and biotur-bation. ]n: R.A. Baker (ed.). Contaminants and Sediments V2. Ann Arbor Science, Ann Arbor, Mich. pp. 555-606. 248 Lee, R.F. 1977. Fate of petroleum components in estuarine waters of the Southeastern United States. ]n: Proceedings of the 1977 Oil Spill Conference. American Petroleum Institute, Washington, D.C. pp. 611-616. Lee, R.F., W.F. Gardner, J.W. Anderson, J.W. Blaylock and J . Barwell-Clarke. 1978. Fate of polycyclic aromatic hydrocarbons in controlled eco-system enclosures. Environ. Sci. Technol. 123: 832-838. Leeuwangh, P., H.J. Bult and L. Schneiders. 1975. Toxicity of hexachloro-butadiene in aquatic organisms. ]n: Sublethal Effects of Toxic Chemicals on Aquatic Animals. Proceedings Swedish-Netherlands Symposium, September 2-5. Elsevier Scientific Publishing Co., New York. Leo, A., C . Hansen and D. Elkins. 1971. Partition coefficients and their uses. Chemical Reviews 71: 525-616. Levings, C D . 1982. The Ecological Consequences of Dredging and Dredge Spoil Disposal in Canadian Waters. National Research Council of Canada, Associate Committee on Scientific Criteria for Environmental Quality, N R C C Pub. No. 18130. 142 p. Lincer, J .L. and D.B. Peakall. 1970. Metabolic effects of PCBs in the American kestrel. Nature, 228(5273); 783-784. Lu, J . C S . and K.Y. Chen. 1977. Migration of trace metals in interfaces of seawater and polluted surficial sediments. Environ. Sci. and Technol. II. 174-182. Lu, P.Y. and R.L. Metcalf. 1975. Environmental fate and biodegradability of benzene derivatives as studied in a model aquatic ecosystem. Environ. Health Perspect. 10: 269-284. Lu, P.Y., R.L. Metcalf and E.M. Carlson. 1978. Environmental fate of five radiolabelled organohalogens and conversion byproducts evaluated in a laboratory model ecosystem. Environ. Health Perspect. 24: 201-208. Lu, P.Y., R.L. Metcalf and L.K. Cole. 1978. The environmental fate of ' ^ C -Pentachlorophenol in laboratory model ecosystems. ]n: K.R. Rao (ed.). Pentachlorophenol. Plenum Press, New York. pp. 53-66. Lu, P.Y., R.L. Metcalf, N. Plummer and D. Mandel. 1977. The environmental fate of three carcinogens: benzo(a)pyrene, benzidine, apd vinyl chloride evaluated in laboratory model ecosystems. Arch. Environ. Contam. Toxicol. 6: 129-142. 249 Luckey, T.D., B. Venngopal and D. Hutcheson. 1975. Heavy Metal Toxicity, Safety and Hormonology. Academic Press, New York. Luoma, S.N. 1977. The dynamics of biologically available mercury in a small estuary. Estuar. Cst l . Mar. Sci., 5: 643-652. Luoma, S.N. 1976. The uptake and interorgan distribution of mercury in a carnivorous crab. Bull. Environ. Contam. Toxicol. 16: 719-723. Luoma, S.N. and G.W. Bryan. 1978. Factors controlling the availability of sediment-bound lead to the estuarine bivalve Scrobicularia plana. J . Mar. Biol. Ass. U.K. 58: 793-802. McCaffrey, R.J., A .C . Myers, E. Davey, G. Morrison, M. Bender, N. Luedtke, D. Cullen, P. Froelich and G. Klinkhammer. 1980. The relation between pore water chemistry and benthic fluxes of nutrients and manganese in Narragansett Bay. Limnology and Oceanography, 25, 31-44. McDermott, D.J., D.R. Young and T .C . Heesen. 1976. PCB contamination of Southern California marine organisms. ]n: Proceedings of the National Conference on Polychlorinated Biphenyls, November 19-21, 1975. Chicago. EPA-560/6-75-004. pp. 209-217. Macek, K.J., S.R. Petrocelli and B.H. Sleight. 1977. Considerations in assessing the potential for, and significance of biomagnification of chemical residues in aquatic food chains. Presented at: ASTM 2nd Symposium on Aguatic Toxicology, Oct. 31 - Nov. I, Cleveland, Ohio. McFarland, V.A., A.B. Gibson and L.E. Meade. 1984. Application of physico-chemical estimation methods to bioaccumulation from contaminated sediments. ]n: Proceedings - Applications in water guality control. U.S. Army Hydrologic Engineering Centre, Dains, Calif. 250 pp. McGreer, E.R. 1981. Factors affecting the distribution of the bivalve (Macoma  balthica) (L.) on a mudflat receiving sewage effluent Fraser River estuary, British Columbia. Marine Environment Research. McGreer, E. 1979. Sublethal effects of heavy metal contaminated sediments on the bivalve Macoma balthica (L). Marine Pollution Bull., 10, 259-262. McGreer, E.R. and D.E. Konasewich. 1981. A review of criteria for evaluating ocean dumping permit applications with recommendations for further monitoring requirements. Unpublished report prepared for Environ-mental Protection Service, West Vancouver, B.C. McGreer, E.R. and D.R. Munday. 1982. Effects of silt and contaminated sediment on eggs and larvae of Pacif ic cod (Gadus macrocephaIus). Unpublished report to Department of Fisheries and Oceans, Vancouver, B.C. 250 McGreer, E.R., B.J. Reid and H. Nelson. 1981. Mobilization, bioaccumulation and sublethal effects of contaminants from marine sediments. Can. Tech. Fish. Aquatic. Sci. 990: 130-151. MacKay, D. 1982. Correlation of bioconcentration factors. Environ. Sci. Technol. 16, 274. MacKay, D. and P.J. Leinonen. 1975. Rate of evaporation of low-solubility contaminants from water bodies to atmosphere. Environ. Sci. Technol. 9: 1178-1 180. MacKay, D. and W.Y. Shui. 1977. Aqueous solubility of polynuclear aromatic hydrocarbons. Chem. Eng. Data 22: 399-402. McKim, J.M., G.F. Olson, G.W. Holcombe and E.P. Hunt. 1975. Long term effects of methyl-mercuric chloride on three generations of brook trout (Salvelinus fontinalis): toxicity, accumulation, distribution, and el imi-nation. (Manuscript). Environmental Protection Agency, Environ-mental Research Laboratory, Duluth, Minn. MacLeod, J .C . and E. Pessah. 1973. Temperature effects on mercury accumu-lation, toxicity and metabolic rate in rainbow trout (Salmo qairdneri). J . Fish. Res. Board Can. 30: 485-492. Mai ins, D . C , B.B. McCain, D.W. Brown, A.K. Sparks and H.O. Hodgins. 1980. Chemical contaminants and biological abnormalities in central and southern Puget Sound. N O A A Tech. Mem. OMPA-2, National Oceanic and Atmospheric Administration, Boulder, Colorado, pp. 186. Manheim, F.T. 1976. Interstitial waters of marine sediments. Jru Riley and Chester (eds.). Chemical Oceanography. Vol. 6., Academic Press, New York. pp. I 15-186. Martens, C S . 1976. Control of methane sediment-water transport by macro-infaunal irrigation in Cape Lookout Bight, North Carolina. Science, 192, 998-1000. Maugh, T.D. 1978. Chemical carcinogens: how dangerous are low doses. Science 202: 37-41. Mayer, F.L., Jr. and H.O. Sanders. 1973. Toxicity of phthalate esters in aguatic organisms. Environ. Health Perspectives 3: 153. Maynard, A. and G. Vigers. 1979. Monitoring environmental contamination from chlorophenol contaminated wastes generatged in the wood preservation industry. Environmental Protection Service, Pacific and Yukon Region, Regional Program Report 79-24. 251 Meagher, T.D. and R.A. Gorham. 1982. Results of monitoring program for sub-sea pipeline dredging. Unpublished report prepared for Woodside Petroleum Ltd., Perth, Australia by T.D. Meagher and Associates. Means, J.C., S.G. Wood, J . J . Hassett and W.L. Banwart. 1980. Sorption of polynuclear aromatic hydrocarbons by sediments and soils, Environ. Sci. Technol., 14, 1524. Messieh, S.N., D.J. Wildish and R.H. Peterson. 1981. Possible impact of sediment from dredging and spoil disposal on the Miramichi Bay herring fishery. Can. Tech. Rep. Fish. Aquat. Sci. No. 1008. 33 p. Metcalf, R.L., J.B. Sanborn, P.Y. Lu and D. Nye. 1975. Laboratory model ecosystem studies of the degradation and fate of radiolabeled tr i - , tetra-, and pentachlorobiphenyl compared with DDE. Arch. Environm. Contam. Toxicol. 3(2), 151-165. Mingelgrin, U., and Z. Gerstl. 1982. Reevaluation of partitioning as a mechanism of nonionic chemicals, submitted to J . Agric. Food Chem., 1982. Ministry of Lands, Parks and Housing. 1984. Eastern Vancouver Island coastal and marine data atlas. Provincial Govt, of B.C. Morton, R.W. 1983. Precision bathymetric study of dredged material capping experiment in Long Island Sound, jn: Kester, D.R. et al. (eds.). Wastes in the ocean. Vol. 2. Dredged material disposal in the ocean. Wiley Interscience Publ. 99-122. Murray, J.W. and G. Gi l l . 1978. The geochemistry of iron in Puget Sound. Geochim. Cosmochim. Acta . 42.9-19. Nimmo, D.R., L.H. Bahner, R.A. Rigby, J.M. Sheppard and A . J . Wilson, Jr. 1977. Mysidopsis bahia: An estuarine species suitable for l ife-cycle toxicity tests to determine the effects of a pollutant, jn: F.L. Mayer and J.L. Hamelink (eds.), Aguatic Toxicol. Hazard Eval. ASTM STP 634. pp. 109-116. Nimmo, D.R., R.R. Blackman, A . J . Wilson and J . Forester. 1971. Toxicity and distribution of Arochlor 1254 in the pink shrimp. Mar. Biol. I 1(3): 191-197. Nissenbaum, A. and I.R. Kaplan. 1972. Chemical and isotropic evidence for the in situ origin of marine humic substances. Limnol. and' Oceanog. 17: 570-582. Nissenbaum, A. and D.J. Swaine. 1976. Organic matter - metal interactions in recent sediments. Geochim. Cosmochim. Acta. 40: 809-816. 252 Nittrouer, C.A. and R.W. Sternberg. 1975. The fate of fine grained dredge spoils deposited in a tidal channel of Puget Sound, Washington. J . Sed. Pet. 45:1, 160-170. Nix, P.G. and P.M. Chapman. 1984. Monitoring program for Expo '86 dredging and dumping activities in False Creek, B.C. Unpubl. report prepared by E.V.S. Consultants for Expo '86, Vancouver. Norstrom, R.J., R.W. Risebrough and D.J. Cartwright. 1976. Elimination of chlorinated dibenzofurans associated with polychlorinated biphenyls fed to mallards (Anas platyrhynchos). Toxicol. Appl. Pharmacol. 37: 217-228. Northcote, T.G., N.T. Johnston and K. Tsumura. 1975. Trace metal con-centrations in lower Fraser River fishes. Westwater Research Centre Tech. Rept. No. 7: 41 pp. O'Connor, J.T. and D.R. Kester. 1975. Adsorption of copper and cobalt from fresh water and marine system. Geochim. Cosmochim. Acta . 39: 1531-1543. Packman, G.A. 1980. An environmental assessment of the Point Grey Ocean Disposal Area in the Strait of Georgia, British Columbia. Department of Environment, Environmental Protection Service, Pacif ic Region, Regional Program Report: 80-3, 128 p. Paris, D.F., D.L. Lewis, J.T. Barnett and G.L. Baughman. 1975. Microbial degradation and accumulation of pesticides in aquatic systems. U.S. EPA Report No. EPA-660/3-75-007. Parizek, J . , J . Kalouskova, A. Babicky, J . Benes and L. Pavlick. 1974. Interaction of selenium with mercury, cadmium and other toxic metals. Jn: W.G. Hockstra (ed.), Trace Element Metabolism in Animals. Park Press, New York. Parks, G.A. 1975. Adsorption in the marine environment. In: Chemical oceanography, Vol. I. Riley, J.P. and G. Skirrow (edsT)^ Academic Press, New York. Parsons, T.R., C.A. Bawder and W.A. Heath. 1973. Preliminary survey of mercury and other metals contained in animals from the Fraser River mudflats. J . Fish. Res. Bd. Can., 30(7), 1014-1016. Patchineelam, S.R. and C. Calmano. 1981. Unpublished study cited in Forstner U. and G.T.W. Wittman. 1981. Metal pollution in the aguatic environment. Second edition. Springer-Verlag, Berlin, p. 248. 253 Patrick, W.H. and D.S. Mikkelsen. 1971. Plant nutrition behaviour in flooded soil, jn: Fertil izer technology and use, 2nd edition, Soil Science Soc. of Amer., Madison, Wisconsin, 1971. pp. 187-215. Peakall, D.B. and M.L. Peakall. 1973. Effects of a PCB on the reproduction of artificially and naturally incubated dove eggs. J . Appl. Ecol., 10(3): 363-868. I 2 Pearson, C R . and G. McConnell. 1975. Chlorinated C and C hydrocarbons in the marine environment. Proc. Royal Soc. London B 189: 305-332. Peddicord, R.K. 1980. Direct effects of suspended sediments on aguatic organisms, jn: R.A. Baker (ed.). Contaminants and sediments Ann Arbor Science, Ann Arbor, Michigan, Vol. I, pp. 501-536. Pemberton, C , M. Risk and D. Buckley. 1976. Supershrimp: Deep Bioturbation in the Strait of Canso, Nova Scotia. Science, 192, 790-791. Petr, T. 1977. Bioturbation and exchange of chemicals in the mud-water interface. jn: Interactions between sediments and fresh water. Golterman, H.L. (ed.). The Hague: Junk Publ., 1977. 216-226. Phillips, D.J.H. 1980. Quantitative aguatic biological indicators. Applied Science Publishers, London. Pierce, R.H. and D.M. Victor. 1978. The fate of PCP in an aquatic ecosystem. jn: K.R. Rao (ed.). Pentachlorophenol. Plenum Press, New York. pp. 51-52. Poirrier, M.A., B.R. Bordelon and J.L. Laseter. 1972. Adsorption and con-centration of dissolved carbon-14 DDT by coloring colloids in surface waters. Environ. Sci. Technol. 6(12): 1033-1035. Pratt-Johnson, B. 1977. 141 dives in the proctected waters of Washington and British Columbia. Gordon Soules Book Publishers, Vancouver, B.C. Prosi, F. 1979. Heavy metals in aquatic organisms, jn: U. Forstner and G.W. Wittman (eds.). Metal pollution in the aquatic environment. Springer-Verlag, Berlin, pp. 271-323. Pryor, W. 1975. Biogenic sedimentation and alteration of argillaceous sediments in shallow marine environments. Geol. Soc. Am. Bull. 86. 1244-1254. Quinlivan, S.C., M. Ghassemi and T.V. Teshendok. 1977. Sources, character-istics, treatment and disposal of industrial wastes containing hexa-chlorobenzene. J . Hazardous Material, I: 349-359. 254 Radding, S.B., T. Mil l, C.W. Gould, C.H. Liu, H.L. Johnson, D.C. Bomberger and C.V. Fogo. 1976. The Environmental Fate of Selected Polynuclear Aromatic Hydrocarbons. EPA 560/5-75-009. 122 pp. Rao, P.S.C., J.M. Davidson and D.P. Kilerease. 1982. Examination of non-singularity of pesticides adsorption-desorption isotherms for soil pest-icide systems, Agronomy Abstracts, 34. Rappe, C , A. Gara and H.R. Buser. 1978. Identification of polychlorinated dibenzofurans (PCDFs) in commercial chlorophenol formulations. Chemosphere. 7: 981-991. Ratkowsky, D.A., T.G. Dix and K.C. Wilson. 1975. Mercury in fish in the Derwent estuary, Tasmania and its relation to the position of the fish in the food chain. Aust. J . Mar. Freshwater Res. 26: 223-231. Reddy, K.R. and W.H. Patrick. 1975. Effect of alternate aerobic and anaerobic conditions on redox potential, organic matter decomposition and nitro-gen loss in a flooded soil. Soil Biol, and Biochem. 7.87-94. Reid, B.J., R.W. Deverall, P.M. Chapman and A.W. Maynard. 1981. Exper-imental investigation into the accumulation of cadmium by the poly-chaete worm, Capitella capitata and the bivalve Macoma balthica. Report prepared for Environment Canada, Dept. of Fisheries and Oceans, Sidney, B.C. by E.V.S. Consultants Ltd., North Vancouver. 65 Renfro, W. 1973. Transfer of Zn from sediments by marine polychaete worms, Mar. Biol. 21: 305-316. Rhoads, D., P. McCal l and J . Yingst. 1978. Disturbance and Production on the estuarine seaf loor. Am. Sci. 66. 577-586. Riley, R.G., E.A. Crecelius, D.C. Mann, K.H. Abel, B.L. Thomas and R.M. Bean. 1980. Quantitation of Pollutants in Suspended Matter from Puget Sound. N O A A Technical Memorandum E R L MESA-49, Marine Eco-systems Analysis Program, Environmental Research Laboratories, Boulder, Colo. Risk, M.J. and H.D. Craig. 1976. Flatfish feeding traces in the Minas Basin. Sed. Petrology 46: 411-413. Robertson, I. 1977. Low seabird densities in the pelagic environment of the Strait of Georgia. Pacif ic Science 31(3). Rodway, M.S. and R.W. Campbell. 1976. Natural history theme study of marine bird and mammal habitats in the Gulf Islands, British Columbia. Department of Indian and Northern Affairs, Parks Canada. Ottawa, Ont. 255 Roesijadi, G., J.W. Anderson and J.W. Baylock. 1978. Uptake of hydrocarbons from marine sediments contaminated with Prudhoe Bay oil: Influence of feeding type of test species and availability of polycyclic aromatic hydrocarbons. J . Fish Res. Board Can. 35: 608-614. Rogers, R.D., J .C. McFarlane and A . J . Cross. 1980. Adsorption and desorption of benzene in two soils and montmorillonite clay, Environ. Sci. Technol., 14(4), 457. Rossi, S.S. and J.M. Neff. 1978. Toxicity of polynuclear aromatic hydrocarbons to the polychaete Neanthes arenoceodentata. Mar. Pollut. Bull. 9: 220-223. Rowe, G. 1974. The effects of the benthic fauna on the physical properties of deep sea sediments. jn: Deep-sea Sediments: Physical and Mechanical Properties. A. Inderbitzen (ed.), New York: Plenum Publishing Co. Safe, S.., C. Wyndham, A. Crawford and J . Kohli. 1979. Metabolism: detoxification or toxification. jn: Proceedings of Second International Symposium on Aquatic Pollutants, September 26-28, 1977. Amsterdam, pp. 299-307. Saila, S.B., S.D. Pratt and T.T. Polgar. 1972. Dredge spoil disposal in Rhode Island Sound. Tech. Rep. No. 2, Univ. of Rhode Island. Sanders, H.O., F.L. Mayer, Jr. and D.F. Walsh. 1973. Toxicity, residue dynamics, and reproductive effects of phthalate esters in aquatic invertebrates. Environ. Res. 6: 84-90. Schimmel, S.C., J.M. Patrick and L.F. Faas. 1978. Effects of Na-PCP on Several Estuarine Animals, jn: K.R. Rao (ed.), Pentachlorophenol. Plenum Press, New York. pp. 147-156. Schneider, R. 1982. Polychlorinated biphenyls (PCBs) in Cod tissues from the western Baltic: significance of eguilibrium partitioning and lipid composition in the bioaccumulation of lipophilic pollutants in gi l l -breathing animals. Meeresforschung-Reports on Marine Research. Sonderdruck aus Bd. 29:H.2, S. 69-79. Sherk, J.A., J.M. O'Connor, D.A. Newmann, R.D. Prince and K.V. Wood. 1974. Effects of suspended and deposited sediments on estuarine organisms. Department of Environmental Research, Chesapeake Biological Labor-atory, MA. Ref. No. 74-20. Shokes, R.F. 1976. Rate dependent distributions of lead "210" and interstitial sulfate in sediments of the Mississippi Delta. Tech. Rep. 76-1-T. Dept. of Oceanogr., Texas A & M Univ. 256 Sholkovich, E.R. 1976. Flocculation of dissolved organic and inorganic matter during the mixing of river water and seawater. Geochim. Cosmochim. Acta. 40,831-845. Shuba, P.J., H.E. Tatum and J.H. Carrol. 1978. Biological assessment methods to predict the impact of open water disposal of dredged material. U.S. Army Corps of Engineers, Dredged Material Research Program, Tech. Rept. D-78-50, Vicksburg, Miss. Shubel, J.R. 1978. Field Investigation of Nature, Degree and Extent of Turbidity Generated by Open Water Pipeline Disposal Operations, Technical report D-78-30, U.S. Army Engineer Waterways Experiment Station, C E , Vicksburg. Shutz, D.F. and K.K. Turekian. 1965. The investigation of the geographical and vertical distribution of several trace elements in seawater using neut-ron activation analysis. Geochim. Cosmochim. Acta., 29, 259-313. Sloan, J.P., J .A .J . Thompson and P.A. Larkin. 1974. The biological half-l ife of inorganic mercury in the Dungeness crab (Cancer magister). J . Fish. Res. Board Canada 31: 1571 -1576. Southworth, G.R., J . J . Beauchamp and P.K. Schmieder. 1978. Bioaccumulation potential and acute toxicity of synthetic fuels effluents in freshwater biota: Azaarenes. Environ. Sci. Technol. 12(9): 1062-1066. Spangler, W.J., J .L . Spigarelli, J.M. Rose and H.M. Miller. 1973. Methylmercury bacterial degradation in lake sediments. Science 180: 192-193. Spehar, R.L., R.L. Anderson and J.T. Fiandt. 1978. Toxicity and bioac-cumulation of cadmium and lead in aquatic invertebrates. Environ. Pollut. 15: 195-208. Stainken, D. and J . Rollwagen. 1979. PCB residues in bivalves and sediments of Raritan Bay. Bull. Environ. Contam. Toxicol. 231: 690-697. Stancil, D.E. 1980. Fraser River Estuary Study Water Quality; Aquatic Biota and Sediments. Environmental Protection Service, Environment Canada, Vancouver, B.C. Steen, W.C., D.F. Paris and G.L. Baughman. 1980. Effects of sediment sorption on microbial degradation of toxic substances. ]n: R.A. Baker (ed.). Contaminants and Sediments, Vol. I. Ann Arbor Science, Ann Arbor, Michigan, pp. 477-482. Stendall, R.C. 1976. Summary of recent information regarding effects of PCBs on birds and mammals. ]n: Proceedings of a National Conference on Polychlorinated Biphenyls, November 19-21, 1975. Chicago. E P A 560/6-75-009. pp. 262-267. 257 Stout, V.F. 1980. Organochlorine residues in fishes from the northwest Atlantic Ocean and Gulf of Mexico. Fish. Bull. 78(1): 51-58. Sundstrom, G., O. Hutzinger and S. Safe. 1976. The metabolism of chloro-biphenyls. Chemosphere 5: 267-298. Swaine, L.G. 1980. Fraser River Estuary Study, Water Quality, Industrial Effluents. Environmental Protection Service, Environment Canada, Vane. B.C. Swartz, R.C., W.A. DeBen, K.A. Sercu and J.O. Lamberson. 1982. Sediment toxicity and the distribution of amphipods in Commencement E(ay, Washington, U.S.A. Mar. Poll. Bull. \3: 359-364. Swartz, R.C. and H. Lee. 1980. Biological processes affecting the distribution of pollutants in marine sediments. Part I. Accumulation, trophic transfer, biodegradation and migration. Jn: Contaminants and Sediments, Vol. 2. Analysis, Chemistry, Biology, R.A. Baker (ed.). Ann Arbor Science, Michigan, pp. 533-553. Tagatz, M.E., J.M. Ivey, J .C . Moore and M. Tobia. 1977. Effects of pentachlorophenol on the development of estuarine communities. J . Toxicol. Environ. Health 3: 501-506. Thibodeaux, L .J . 1979. Chemodynamics. John Wiley and Sons, New York. Thorn, N.S. and A.R. Agg. 1975. The breakdown of synthetic organic compounds in biological processes. Proc. Roy. Soc. Lona*. B. 189: 347-357. Thomann, R.V. 1978. Size Dependent Model of Hazardous Substances in Aguatic Food Chains. Report No. EPA-600/3-78-036. Duluth, Minnesota: U.S. Environmental Protection Agency, 39 pp. Thomas, D.J. and E.V. Gri l l . 1977. The effect of exchange reactions between Fraser River sediment and seawater on dissolved Cu and Zn con-centrations in the Strait of Georiga. Estuarine Coastal Mar. Sci. 5: 421-427. Thompson, J .A .J . 1981. Production of lead alkyls in marine sediments. Jri: Conf. Proceedings, Heavy metals in the environment. C E P Consultants, Amsterdam, Sept.'81. pp. 653-656. Thompson, J .A .J , and J.A. Crerar. 1980. Methylation of lead in marine sediments. M a r . P o l l . B u l l . i l : 251-253. Thomson, J . , K.K. Turekian and R.J. McCaffrey. 1975. The accumulation of metals in and release from the sediments of Long Island Sound, jn: Estuarine Research, Cronin, L.E. (ed.), New York, Academic Press, pp. 28-44. 258 Tinsley, I.J. 1979. Chemical concepts in pollutant behaviour. John Wiley and Sons, New York. Towill Corporation. 1972. Environmental assessment of maintenance dredging operations. Reprot to U.S. C O E , Pacif ic Division, Honolulu, Hawaii. Trefrey, J .H. and B.J. Presley. 1976. Heavy metals in sediments from San Antonio Bay and the northwest Gulf of Mexico. Environ. Geol. I. 283-294. Tulp, M.Th.M. and 0. Hutzinger. 1978. Some thoughts on aqueous solubilities and partition coefficients of PCB, and the mathematical correlations between bioaccumulation and physico-chemical properties. Chemosphere 7: 849-860. Tumasonis, C.F., B. Bush and F.D. Baker. 1973. PCB levels in egg yolks associated with embryonic mortality and deformity of hatched chicks. Arch. Environ. Contam. Toxicol. 1(4): 312-324. Tunnicliffe, V. and M.J. Risk. 1977. Relationships between the bivalve (Macoma  balthica) and bacteria in intertidal sediments: Minos Basin, Bay of Funday. J . Mar. Res. 35: 499-507. Turekian, K.K. 1977. The fate of metals in the oceans. Geochim. Cosmochim. Acta . 41, 1139-1 144. Tute, M.S. 1971. Principles and practice of Hansch analysis: A guide to structure-activity correlation for the medicinal chemist. Adv. Drug Res. 6: 1-77. United States Environmental Protection Agency. 1984. Interim decision criteria for disposal of dredged material at the Four Mile Rock open water disposal site. Unpublished memorandum of the Environmental Protection Authority, Seattle, WA. June 13, 1984. U.S. Environmental Protection Agency. 1980. Treatability Manual. 5 Volumes. EPA-600/8-80-042. U.S. Environmental Protection Agency. 1978. Ambient Water Quality Criteria: Cadmium. NTIS (PB 292 423). U.S. Environmental Protection Agency. 1978. Ambient Water Quality Criteria: Chlorinated Naphthalenes. NTIS (PB 292 426). U.S. Environmental Protection Agency. 1978. Ambient Water Quality Criteria: Naphthalene. NTIS PB 296 786. 259 U.S. Environmental Protection Agency. 1977. Sampling and analysis procedures for screening of industrial effluents for priority pollutants. EPA Monitoring and Support Laboratory, Cincinnati, Ohio. 45 268. U.S. Environmental Protection Agency. 1976a. Quality Criteria for Water, U.S. EPA 440/9-76-023, Washington, D.C. U.S. Environmental Protection Agency. 1976b. Semi-annual report, Apr i l -September, 1976. Environmental Research Lab., Gulf Breeze, Fla. U.S. Environmental Protection Agency. 1975. Survey of industrial processing data. Task I - Hexachlorobenzene and hexachlorobutadiene pollution from chlorocarbon processing. EPA 560/6-76-010. U.S. Environmental Protection Agency/Corps of Engineers. 1977. Ecological evaluation of proposed discharge of dredged material into ocean waters. Implementation Manual for Section 103 of Public Law 92-532 (Marine Protection, Research and Sanctuaries Act of 1972). Environmental Effects Laboratory, U.S. Army Engineer Waterways Experiment Station, Vicksburg, Miss., July, 1977; Second Printing Apri l , 1978. 84 p. Uthe, J.F., F.M. Atton and L.M. Royer. 1973. Uptake of mercury by caged rainbow trout (Salmo gairdneri) in the South Saskatchewan River. J . Fish. Res. Board Can. 30: 643-650. Valiela, D. 1979. The B.C. oyster industry: policy analysis for coastal resource management. Vol. I, oyster ecology and culture. Westwater Research Centre, Tech. Publ. No. 19, U.B.C. Van der Weiden, C.H., M.J.H.L. Arnoldus and C.T. Meurs. 1977. Desorption of metals from suspended material in the Rhue Estuary. Neth. J . Sea. Res., I 1, 130-145. Varanasi, V. and D.C. Malins. 1977. Metabolism of petroleum hydrocarbons: accumulation and biotransformation in marine organisms, jn: D.C. Malins (ed.), Effects of petroleum on Arct ic and subarctic marine environments and organisms. Vol. 2. Academic Press, New York., p. 175. Veith, G.D., K.J. Macek, S.R. Petrocelli and J . Carroll. 1980. An evaluation of using partition coefficients and water solubility to estimate biocon-centration factors for organic chemicals in fish. In: J .G. Eaton, P.R. Parrish, and A.C. Hendricks (eds.), Aquatic Toxicology. Proceedings of the Third Annual Symposium on Aquatic Toxicology. ASTM, Philadelphia. 260 Vermeer, K., I. Robertson, R.W. Campbell, G. Kaiser and M. Lemon. No date. Distributions and densities of marine birds on the Canadian west coast. Draft report prepared by Canadian Wildlife Service, Delta, B.C. and the British Columbian Provincial Museum, Victoria, B.C. under funding from Chevron Standard Limited. Vernberg, W.B., P.J. DeCoursey, M. Kelly and D.M. Johns. 1977. Effects of sublethal concentrations of cadmium on adult Palaemonetes puqio under static and flow through conditions. Bull. Environ. Contamin. Toxicol. 17: 16-24. Verschueren, K. 1983. Handbook of environmental data on organic chemicals. 2nd edition. Van Nostrand Reinhold, New York. 1310 pp. Wakeman, T. 1976. The biological ramifications of dredging and disposal activities, pp. 53-68 \n Dredging: Environmental effects and tech-nology. Proceedings of WODCON VII, San Francisco, July 10-12, 1976. Waldichuk, M. 1983. Pollution of the Strait of Georgia: a review. Can. J . Fish. Aquat. Sci.o 40: 1042-1167. Waldichuk, M. 1957. Oceanography of the Strait of Georgia. J . Fish. Res. Bd. Can. 14(3): 321-486. Ward, A.B. and D.L. Sullivan. 1980. A review of existing and historical ocean dumpsites in the Pacif ic region. Environment Canada, Pacif ic Region, Environmental Protection Service. Regional Program Rep. 80-5: 121 p. Wildish, D.J., C D . Metcalfe, H.M. Akagi and D.W. McLeese. 1980. Flux of Aroclor 1254 between estuarine sediments and water. Bull. Environ. Contam. Toxicol. 24: 20-26. Williams, G.A. 1976. Dredging in Canada. In: Proc. WODCON VII: Dredging, Environmental Effects and Technology. San Francisco, Calif. July 10-12, 1976. pp. 19-27. Williams, P.M. and H.V. Weiss. 1973. Mercury in the marine environment: concentration in seawater and in a pelagic food chain. J . Fish. Res. Board Can. 30: 293-295. Windom, H.L. 1976. Environmental aspects of dredging in the coastal zone. C R C Crit . Rev. Environ. Control 5, 91-109. Windom, H.L., W.S. Gardner, W.M. Dunstan and G.A. Paffenhofer. 1976. Cadmium and mercury transfer in a coastal marine ecosystem. ]n: H.L. Windom and R.A. Duce (eds.), Mar. Pollut. Transfer. D . C Heath and Co., Lexington, Mass. pp. 135-157. 261 Wolfe, N.L.I, D.F. Paris, W.C. Steen and G.L. Baughman. 1980. Correlation of microbial degradation rates with chemical structure. Amer. Chem. Soc. 14: 9, I 143-1 144. Wood, J.M. 1974. Biological cycles for toxic elements in the environment. Science 183: 1049-1052. Wood, J.M., F.S. Kennedy and C.G. Rosen. 1968. Synthesis of methylmercury compounds by extracts of a methanogenic bacteria. Nature 220: 173-174. Wright, T.D. 1978. Aquatic dredged material disposal impacts. Dredged Material Research Program. Technical Report DS-78-1, U.S. Army Engineer Waterways Experiment Station, Vicksburg, Mississippi, 57 pp. Yamamoto, S. and J.B. Alcauskas. 1975. Ocean disposal of dredged material. Appendix L in: Dredge disposal study, San Francisco Bay and Estuary. U.S. C O E , San Fransciso, 43 pp. Young, D.R., A . J . Mearns, J . Tsu-Kai, T .C. Heisen, M.D. Moore, R.P. Eganhouse, G.P. Hershelman and R.W. Gassett. 1980. Trophic struc-ture and pollutant concentrations in marine ecosystems of southern California. Cal C O F I Rep. Vol. XXI, 197-206. Young, R. and J . Southard. 1978. Erosion of fine-grained marine sediments: seafloor and laboratory experiment. Geol. Soc. Am. Bull. 89, 663-672. Zaroogian, G.E. 1979. Studies on the depuration of cadmium and copper by the American oyster Crassostrea virqinica. Bull. Environ. Contam. Toxicol. 23: 117-122. Zitko, V. 1977. Uptake and excretion of chlorinated and brominated hydro-carbons by fish. Fisheries and Marine Service Tech. Rep. No. 737. Fisheries and Environmental Sciences Resource Branch, St. Andrews, New Brunswick. 14 pp. Zitko, V. 1974. Uptake of chlorinated paraffins and PCB from suspended solids and food by juvenile Atlantic salmon. Bull. Environ. Contam. Toxicol. 12: 406-412. 


Citation Scheme:


Citations by CSL (citeproc-js)

Usage Statistics



Customize your widget with the following options, then copy and paste the code below into the HTML of your page to embed this item in your website.
                            <div id="ubcOpenCollectionsWidgetDisplay">
                            <script id="ubcOpenCollectionsWidget"
                            async >
IIIF logo Our image viewer uses the IIIF 2.0 standard. To load this item in other compatible viewers, use this url:


Related Items