U.S. GEOLOGICAL SURVEY CIRCULAR 993 

A Review of Surface-Water Sediment 

Fractions and Their Interactions With 

Persistent Manmade Organic Compounds 

A Review of Surface-Water Sediment 

Fractions and Their Interactions With 

Persistent Manmade Organic Compounds 

By P. J. Witkowski, J. A. Smith, 

T. V. Fusillo, and C. T. Chiou 

U.S. GEOLOGICAL SURVEY CIRCULAR 993 

1987 
DEPARTMENT OF THE INTERIOR 

DONALD PAUL HODEL, Secretary 

U.S. GEOLOGICAL SURVEY 

Dallas L. Peck, Director 

UNITED STATES GOVERNMENT PRINTING OFFICE: 1987 

Free on application to the Books and Open-File Reports Section, 

U.S. Geological Survey, Federal Center, Box 25425, Denver, CO 80225 

Library of Congress Cataloging in Publication Data 

A review of surface-water sediment fractions and their interactions with persistent 

manmade organic compounds. 

(U.S. Geological Sutvey circular; 993) 

Bibliography: p. 

Supt. of Docs. no.: I 19.4/2:993 

1. Organic water pollutants. 2. Sediment, Suspended. 3. Organic compounds. 

4. Water chemistry. I. Witkowski, Patrick J. II. Series. 

1D427.07R48 1987 628.1'68 86-600413 
CONTENTS 

Glossary............................................................................... v 

Abstract . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1 

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1 

Purpose and scope . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1 

Approach........................................................................ 2 

Manmade organic contaminants. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2 

Physical and chemical properties . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3 

Transformation processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . S 

Sorption and bioconcentration...................................................... 8 

Environmental behavior . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10 

Suspended and surficial sediment fractions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ll 

Biological particles. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ll 

Inorganic particles. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . l3 

Humic substances . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14 

Interactions between sediment fractions and manmade organic compounds . . . . . . . . . . . . . . . . . . . . . 18 

Humic substances and nonionic organic compounds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18 

Humic substances and ionic organic compounds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20 

Field studies..................................................................... 22 

:Laboratory studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23 

Transport of sorbed manmade organic compounds. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26 

Hydrodynamics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26 

Flocculation..................................................................... 28 

Estimation of the field distribution of manmade organic compounds . . . . . . . . . . . . . . . . . . . . . . . . . . 29 

Summary and conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 33 

References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 34 

ILLUSTRATIONS 

FiGURE 1. Size spectrum of natural particulates and the hydrodynamic forces affecting their 

settling . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1S 

2. Fractionation of humic substances . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17 

TABLES 

Pace 

TABLE 1. Classification of manmade organic compounds by chemical source................... 3 

2. aassification of manmade organic compounds by chemical structure and chemical 

usage..................................................................... 4 

3. Physical and chemical properties of selected groups of manmade organic compounds. . . S 

4. Physical, chemical, and biological transformation processes that affect manmade organic 

compounds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6 

Ill 
TABlE 5. Physical and chemical characteristics used to assess the distribution and fate of 

manmade organic compounds. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7 

6. Environmental fate processes for selected groups of manmade organic compounds . . . . 12 

7. Estimates of environmental persistence of selected groups of manmade organic 

compounds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14 

8. Colloids, aggregates, and their characteristics..................................... 16 

9. Partition coefficient ranges of polychlorinated biphenyls and various sorbent types . . . . 25 

10. Theoretical settling velocities of particle-size classes............................... 27 

11. Regression equations used for estimating partition coefficients . . . . . . . . . . . . . . . . . . . . . 32 

IV 
GLOSSARY 

Activity. A dimensionless quantity expressing the escaping tendency of a component in a system relative to that in a pure state of 

the same component at the same temperature and total pressure. It is the ratio of the partial pressures (or more precisely, the 

fugacities) of the dissolved component in the system to that of the component at standard state. 

Activity coefficient. A dimensionless quantity that corrects for the deviation of the partial pressure (or fugacity) of a component in 

solution from that value defined by Raoult's law. If the observed partial pressure is greater than that predicted by Raoult's law, 

the activity coefficient is greater than one; conversely, if the observed partial pressure is less than that predicted by Raoult's law, 

the activity coefficient is less than one. Most organic solutes in water show positive deviation from ideality, that is, the activity 

coefficient is greater than one. This term provides an approximation of how much interaction exists between molecules at higher 

concentrations. · 

Adsorption. The interphase accumulation or concentration of substances at a surface or interface. The process can occur at an 

interface between any two phases, such as liquid-liquid, gas-liquid, gas-solid, or liquid-solid interfaces. The material being 

concentrated or adsorbed is the adsorbate, and the adsorbing phase is termed the adsorbent. 

Anthropogenic compounds. Compounds that are produced as a result of the activities of humans as contrasted with compounds 

formed by the actions of natural forces and events. 

Bedload. Sediment that is transported in a stream by rolling, sliding, or skipping along the bed and very close to it, that is, within 

the bedlayer. 

Bedload layer. A thin layer through which the bedload discharges, commonly assumed to be only a few grain diameters thick. 

Bioturbation. Mixing processes in sediment layer caused by the activity of biological organisms. 

Coagulation. A process where charge neutralization of discrete colloids occurs through interactions with available counterions. 

Colloids. Operationally defined as discrete particles with at least one characteristic dimension in the micrometer to nanometer 

range. 

Diagenesis. The gradual and successive chemical and physical changes that take place in sediment previous to or during its 

consolidation. 

Dissolved. Operationally defined as discrete particles less than 0.45 ~m in their maximum dimension. 

Flocculation. Aggregation of neutral discrete colloids into larger masses by the mixing action of water. 

Fulvic acids. The fraction of organic matter that remains dissolved in solution after sequential extraction of the sample with alkali 

and then acid. 

Humic acids. The fraction of organic matter that remains dissolved in solution after extraction with alkali but that precipitates from 

solution upon further extraction with acid. 

Humins. The fraction of organic matter that is not soluble in either alkali or acid. 

Lipophilic. The characteristic of solutes or solvents that are readily miscible in other organic solvents such as lipids. Lipids as a 

heterogeneous group of substances include fatty acids, neutral fats, phosphatides, glycolipids, aliphatic alcohols and waxes, 

terpanes, and steroids. Lipids are categorized by their extractability in nonpolar organic solvents such as chloroform, carbon 

tetrachloride, benzene, ether, carbon disulfide, and petroleum ether. 

Miscible. The ability of two or more substances to mix and to form a single, homogeneous phase. 

Particulate. Operationally defined as discrete particles greater than 0.45 ~min their maximum dimension. 

Partition. The distribution of a compound between two different bulk phases, usually by a solubilization process. 

Polarity. Local charge groupings occurring as a result of geometric asymmetries between atoms of .a given molecule. This 

chemical characteristic can lead to incompatibilities between liquids of different polarities. The relative polarity of large 

macromolecules, such as the humic and fulvic acids, is measured by comparing the oxygen to carbon ratios of the molecules. 

The larger the magnitude of the ratio, the greater the relative polarity of the compound. 

Sediment. Fragmented material that originates from the disintegration of rocks and is transported by, suspended in, or deposited 

by water or air or is accumulated in beds by other natural processes. 

Solute. The compound dissolved in solution. 

Sorption. General expression for a process in which a component moves from one phase to be accumulated in another. The 

material being sorbed is the sorbate, and the sorbing phase is termed the sorbent. Sorption is used when it is not certain if the 

accumulating mechanism is a partitioning, an adsorption, or an absorption process. 

Surface microlayer. Upper, very thin layer of a surface-water body characterized by high surface tension and some physical and 

chemical properties distinguishable from the bulk solution. 

Surficial sediment layer. Upper sediment layer with a thickness that fluctuates depending on the deposition rates of new sediment 

and the degree of resuspension generated by boundary layer turbulence. In this paper, the surficial sediment layer is always 

considered to be submerged. 

v 

A Review of Surface-Water Sediment Fractions 

and Their Interactions With 

Persistent Manmade Organic Compounds 

By P. J. Witkowski, J. A. Smith, T. V. Fusillo, and C. T. Chiou 

ABSTRACT 

This paper reviews the suspended and surficial sediment 

fractions and their interactions with manmade organic compounds. The objective of this review is to isolate and describe 

those contaminant and sediment properties that contribute to 

the persistence of organic compounds in surface-water systems. 

Most persistent, nonionic organic contaminants, such as the 

chlorinated insecticides and polychlorinated biphenyls (PCBs), 

are characterized by low water solubilities and high octanolwater partition coefficients. Consequently, sorptive interactions 

are the primary transformation processes that control their 

environmental behavior. For nonionic organic compounds, sorption is primarily attributed to the partitioning of an organic 

contaminant between a water phase and an organic phase. 

Partitioning processes play a central role in the uptake and 

release of contaminants by sediment organic matter and in the 

bioconcentration of contaminants by aquatic organisms. 

Chemically isolated sediment fractions show that organic 

matter is the primary determinant of the sorptive capacity 

exhibited by sediment. Humic substances, as dissolved organic 

matter, contribute a number of functions to the processes 

cycling organic contaminants. They alter the rate of transformation of contaminants, enhance apparent water solubility, and 

increase the carrying capacity of the water column beyond the 

solubility limits of the contaminant. As a component of sediment 

particles, humic substances, through sorptive interactions, serve 

as vectors for the hydrodynamic transport of organic contaminants. The capabilities of the humic substances stem in part from 

their polyfunctional chemical composition and also from their 

ability to exist in solution as dissolved species, flocculated 

aggregates, surface coatings, and colloidal organomineral and 

organometal complexes. 

The transport properties of manmade organic compounds 

have been investigated by field studies and laboratory experiments that examine the sorption of contaminants by different 

sediment size fractions. Field studies indicate that organic contaminants tend to sorb more to fine-grained sediment, and this 

correlates significantly with sediment organic matter content. 

Laboratory experiments have extended the field studies to a 

wider spectrum of natural particulates and anthropogenic compounds. Quantitation of isotherm results allows the comparison 

of different sediment sorbents as well as the estimation of field 

partition coefficients from laboratory-measured sediment and 

contaminant properties. Detailed analyses made on the basis of 

particle-size classes show that all sediment fractions need to be 

considered in evaluating the fate and distribution of manmade 

organic compounds. This conclusion is based on observations 

from field studies and on the variety of natural organic sorbents 

that demonstrate sorptive capabilities in laboratory isotherm 

experiments. 

INTRODUCTION 

There is a growing need to monitor riverine, 

estuarine, and lake sediment to provide a more complete understanding of contaminant transport and transformations in surface waters. As contaminants move 

through a surface-water body, they interact with natural 

particles suspended in the water column and those 

deposited in the surficial sediment layer. Through 

sorptive processes, sediment particles interact with 

organic contaminants and can become vectors for their 

hydrodynamic transport. Sediment also functions as a 

reservoir for contaminants and has the ability to take up 

and release contaminants in response to concentration 

variations in the overlying water column. Understanding 

the interactions between the sediment and watercolumn compartments helps to reveal how specific 

contaminants are transformed and transported within 

the surface-water body. 

PURPOSE AND SCOPE 

The objective of this paper is to review the 

characteristics of suspended and surficial sediment 

fractions and their interactions with manmade organic 

compounds. The emphasis is on sorptive interactions 

between natural particles and persistent organic contaminants. Sorption is a general term used to describe 

the variety of possible contaminant-sorbent interactions 
when the exact nature of the particle-contaminant 

interaction is not known or when the interaction may 

involve more than one dominant mechanism. Sorptive 

interactions are considered because they are an important mechanism through which persistent manmade 

contaminants interact with sediment particles. The 

influence of sorptive interactions on contaminant distribution and fate is shown to extend to the processes 

controlling the transport and transformation of the 

organic compounds. Consequently, the scope of this 

paper is not limited to interactions within the surficial 

sediment layer; it also encompasses important watercolumn interactions that occur during the deposition 

and resuspension of sediment components. 

This review examines sediment-contaminant 

interactions. The sections on (1) Manmade Organic 

Contaminants and (2) Suspended and Surficial Sediment Fractions examine contaminant and sediment 

properties, respectively. Manmade Organic Contaminants describes the physicochemical properties, transformation processes, and environmental behaviors that 

combine to determine the fate of an anthropogenic 

organic compound. Suspended and Surficial Sediment 

Fractions describes some properties and transformation processes characteristic of biological organisms, 

inorganic particles, and humic substances found suspended in the water column and within the surficial 

sediment layer. Together, these sections provide the 

background information necessary to understand the 

complex interactions between organic contaminants and 

sediment fractions discussed in the section on Interactions Between Sediment Fractions and Manmade 

Organic Compounds. 

The section on interactions looks specifically at 

how the various fractions of the sediment influence the 

fate of organic contaminants. Humic substances, the 

decay products of terrestrial and aquatic organic matter, modify the chemical form, stability, and reactivity of 

contaminants in surface-water systems. Examples show 

how humic substances can change the apparent solubility and affect the sorptive interactions of persistent 

organic contaminants. Field studies are reviewed that 

seek patterns and trends in the sorption of organic 

compounds by sediment sorbents characterized on the 

basis of size and type. Laboratory studies are reviewed 

that seek to quantify more exactly the discernible trends 

and patterns observed in the fiekl data. 

The section on Transport of Sorbed Manmade 

Organic Compounds examines the hydrodynamic 

forces that control the transport of contaminants sorbed 

to different sediment fractions separated on the basis of 

size. Finally, the section on Estimation of the Field 

2 

Distribution of Manmade Organic Compounds deals 

with the methods used in estimating the field sorption 

behavior and bioconcentration of contaminants from 

laboratory-measurable physical properties of the sediment and the anthropogenic organic compounds. 

Overall, the review is designed to provide some 

indication of what happens to an organic contaminant 

when it is introduced into a surface-water system. To 

achieve this objective, it is important to have a basic 

understanding of the nature of the aqueous environment in conjunction with the properties of the anthropogenic compound. 

APPROACH 

In considering the sorptive interaction of 

anthropogenic organic compounds with sediment, interest lies in determining and understanding the physical 

and chemical factors that control the spatial distribution 

and temporal patterns observed in field investigations. 

Field behavior of organic compounds is largely determined by physical and chemical properties of a compound as manifested under a given set of environmental 

conditions. If an organic chemical interacts with a 

natural particulate, the behavior of the compound will 

be modified to an extent that depends on the physical 

and chemical attributes of the compound and the 

sediment particle. Understanding the influence of sediment, in terms of its properties, is achieved by examining different sediment fractions. The different components of the bulk sediment are defined on the basis of 

size and chemical extraction procedure. The way that 

different sediment fractions contribute to the interactions with different organic compounds and the way' 

that this behavior can be accounted for in relation to 

what is observable on the bulk sediment level are the 

subject of this review. 

MANMADE ORGANIC CONTAMINANTS 

This section examines some of the physicochemical properties that are characteristic of different groups 

of anthropogenic organic compounds in aqueous 

systems. In addition, the section considers several 

transformation processes that influence the fate of 

anthropogenic organic compounds. Together, the physicochemical properties and the transformation processes determine the environmental persistence of certain anthropogenic organic compounds. The remainder 

ofthe review will focus on these compound and system 
Table 1.-Classification of manmade organic compounds by chemical source 

Groups Common examples or subgroups 

Petroleum- and coal-derived hydrocarbons 1. Crude oil (alkanes, cycloalkanes, and aromatics). 

2. Refined petroleum products (petrol, kerosene, diesel, heating oils, and 

asphalts). 

3. Combustion and conversion products (P AHs, carbon dioxide, synfuels, and 

byproducts ). 

Synthetic organics 1. Halogenated hydrocarbons (PCBs, DDTs, Kepone, Mirex, dieldrin, 

chlordane, industrial solvents, and other pesticides). 

2. Phthalate esters (plasticizers such as polyvinyl chloride (PVC)). 

3. Others (surface-active agents, organophosphate insecticides, carbamate 

insecticides, and many herbicides). 

Municipal wastes 1. Sewage sludge, secondary effluents, and trihalomethanes (THMs ). 

2. Nutrients (phosphates, nitrogen, carbon, and silica). 

Source: Olsen and others (1982). 

characteristics to examine the interaction of persistent 

organic compounds with sediment fractions. 

The number of existing anthropogenic compounds is large, and their environmental behavior is 

more readily reviewed from the standpoint of compound groups rather than of individual chemicals. 

Table 1 categorizes organic compounds by source, and 

table 2 categorizes them by a combination of chemical 

families and substance usage groups. These tables are 

intended to be representative of the more common 

anthropogenic organic compounds and do not include 

all those that might have an adverse impact on water 

quality. 

PHYSICAL AND CHEMICAL PROPERTIES 

Several common physicochemical properties of 

anthropogenic organic substances contribute to environmental persistence and toxicity. Table 3 lists typical 

ranges of molecular weights, vapor pressures, water 

solubilities, and octanol-water partition coefficients for 

the compound groups and some subgroups from table 

2. The ranges are meant to be representative within 

each group (outliers certainly exist) and also to serve as 

benchmarks for comparisons among different groups. 

From the simple breakdown employed in table 3, 

several distinct physicochemical characteristics are 

apparent. Water solubility is probably the most important property affecting the movement of compounds 

among the various environmental compartments, 

including air, the water column, aquatic biota, and 

sediment (Oschwald, 1972; Pionke and Chesters, 1973). 

Many of the organic compounds are highly water 

insoluble with solubilities in the low milligram per liter 

to microgram per liter range. Within specific chemical 

3 

groups, a decreasing trend in water solubility is discernible as the degree of halogen substitution increases (for 

example, insecticides, aliphatic hydrocarbons, monocyclic aromatics, and phenols). Similarly, within specific 

chemical groups, water solubility decreases as the number of aromatic and cyclic ring substitutions increases 

(for example, polynuclear aromatic hydrocarbons 

(PAHs), ethers, and nitrogen-substituted compounds). 

Much of the following discussion focuses on these 

low-solubility compounds because they are representative of many persistent organic contaminants. 

In general, the low solubility of an organic contaminant in water is commonly reflected by its affinity 

for relatively nonpolar organic phases (solid and liquid). 

This lipophilic tendency is found in compounds possessing high octanol-water partition coefficients, where 

concentrations in the organic phase are 103 to 106 times 

greater than those in the water phase (for example, 

PAHs, polychlorinated biphenyls (PCBs ), and phthalate esters). The octanol-water partition coefficient is a 

dimensionless concentration ratio whose magnitude 

expresses the distribution of a compound between 

equal volumes of two partially miscible solvents, noctanol and water (Leo and others, 1971). The magnitude of the partition coefficient is a useful index of the 

tendency of an organic compound to partition from 

water into sediment organic matter and into aquatic 

organisms (bioconcentration). The greater the magnitude of the partition coefficient, the greater the relative 

solubility of the compound in the organic phase over 

that in the aqueous phase. 

The partition behavior of a compound in a simple 

organic solvent-water system provides a useful estimate 

of the probable behavior of that compound in an 

aqueous environment (Haque and Freed, 197 4; 

U.S. EPA, 1979a). The use of the partition coefficient 

to estimate the environmental distribution of a 
Table 2.-CiassifiCSfion of manmade organic compounds by chemical structure and chemical usage 

Groups Common examples or subgroups 

Halogenated aliphatic hydrocarbons 

Halogenated ethers 

Monocyclic aromatics 

Chlorinated/brominated alkanes and chlorinated/brominated alkenes 

Chloroalkyl ethers and chlorophenylethers 

Chlorinated benzenes, alkyl benzenes, nitrotoluenes, and aminobenzenes 

Chlorinated phenols, nitrophenols, alkylphenols, and substituted cresols 

Dialkyl phthalates 

Phenols and cresols 

Phthalate esters 

Polynuclear aromatic hydrocarbons 

Nitrogen substituted compounds 

Polychlorinated biphenyls 

Organochlorine insecticides 

Organophosphorus insecticides 

Carbamate insecticides and herbicides 

Acid herbicides (phenoxy and benzoic) 

Triazine herbicides 

Naphthalene, anthracene, chrysene, and benzo(a)pr.rene 

Alkylamines, nitrosamines, and halogenated benztdines 

Isomers and Aroclors (1016, 1221, 1232, 1242, 1248, 1254, and 1260) 

Aldrin, DDTs, endosulfan, heptachlor, chlordane, methoxychlor, and toxaphene 

Parathion, malathion, and diazinon 

Carbaryl (Sevin), carbofuran, and aldicarb 

2,4-D, silvex, 2,4,5-T, benzoic acid, and dicamba 

Substituted urea herbicides 

Atrazine, propazine, simazine, and prometryne 

Diuron, flometuron, and Iinuron 

Other herbicides Bipyridylium (diquat and paraquat) 

Dinitroanilines (trifluralin and profluralin) 

Other pesticides 

Others ( alachlor, propachlor, and picloram) 

Nematicides (ethylene dibromide and methyl bromide) 

Miticides (kelthane, chlorobenzilate, and dichlore) 

Fungicides (captan, quinones, and dithiocarbamic acid derivatives) 

compound is based on the recognition that much of the 

molecular driving force controlling distribution among 

water, sediment, and biota is analogous to that found in 

the octanol-water model system (Chiou, 1981 ). The use 

of octanol-water partition coefficients to estimate 

sediment-water partition coefficients and bioconcentration factors (BCF) is discussed in more detail in the 

section on Estimation of the Field Distribution of 

Manmade Organic Compounds. 

The vapor pressure of an organic compound 

indicates the extent to which it will volatilize from 

solution. Although vapor pressure exerts an important 

influence on the volatility of a compound as a pure 

substance, volatilization from an environmental system 

also depends on other physical properties (for example, 

water solubility) and system variables (for example, 

temperature, wind speed, and water turbulence). Lowmolecular-weight compounds ( < 200 amu) such as 

aliphatic hydrocarbons, monocyclic aromatics, and some 

nitrogen-substituted compounds possess relatively high 

vapor pressures and limited water solubilities that in 

conjunction with certain environmental conditions make 

volatilization an important environmental pathway (U.S. 

EPA, 1978, 1981). 

Although water solubility and octanol-water partitioning are of primary importance in controlling the 

environmental distribution of anthropogenic organic 

compounds, other physicochemical properties also 

exert a significant influence on their overall distribution. These properties include polarity, ionization constant, molecular charge, molecular size, and molecu4 

lar configuration (U.S. EPA, 1979a; Verschueren, 

1983). The exact contribution of each of these properties is difficult to isolate because the effects are exerted 

primarily at the molecular level. Howevet; some properties, such as the tendency to ionize in solution, 

determine which compounds respond to changes in pH 

(for example, acid and bipyridylium herbicides) and 

which are not sensitive to its variation (for example, 

chlorinated insecticides). The polarity of the organic 

compound influences its degree of chemical compatibility with polar solvents such as water and other less 

polar solvents such as sediment organic matter. Finally, 

fundamental properties such as molecular size and 

charge exert their influence on the behavior of a 

compound at all levels, from molecular interactions to 

the macroscopic scale of field processes. 

The more environmentally persistent organic 

compounds (for example, PCBs, chlorinated insecticides, PAHs, phthalate esters, highly halogenated 

aliphatics, and aromatics) are generally characterized 

by low water solubility and high octanol-water partition 

coefficients (U.S. EPA, 1982). Most of these chemical 

groups represent nonionic compounds that were synthesized for industrial and (or) agricultural applications 

that require physical stability and specific properties 

(U.S. EPA, 1979b; Verschueren, 1983). On the basis of 

low solubility and high partition coefficients, the distribution and fate of these persistent compounds in the 

environment largely depend on sorptive interactions 

with natural particles. Yet, other physicochemical properties may become important under certain specific 

conditions, and therefore, environmental behavior must 
Table a.-Physical and chemical properties of selected groups of manmade organic compounds 

[ > , greater than; mg!L, milligrams per liter) 

Number Molecular Log vapor Log water Log 

Groups Subgroups of com- weight pressure In solubility In octanol-water References 

pounds range torr* mg/l partition 

coefficient 

Halogenated aliphatic 1-4 Cl!Br 28 50-200 2 to 3 2 to 4 1 to 3 1,2,3,4,5,6 

hydrocarbons >5 Cl/Br 3 200-300 - 2 to -1 0 to 1 3 to 4 1,2,3,4,5,6 

Halogenated ethers Aliphatic 5 100-200 - 1 to 1 3 to 5 1 to 3 1,2,3,4,5,6 

Aromatic 2 200-250 - 3 to -2 0 to 1 4 to 5 1,2,3,4,5,6 

Monocyclic aromatics 19 100-200 - 1 to 2 2 to 3 2 to 4 1,2,3,4,5,6 

Phenols 12 100-200 - 1 to 0 3 to 5 1 to 4 1,2,3,4,5,6 

Monocyclic aromatics >5 0/Br 2 250-300 - 5 to -4 -3 to 1 5 to 6 1,2,3,4,5,6,12 

and phenols 

Phthalate esters 6 200-400 - 3 to -1 0 to 4 2 to 9 1 ,2,3,4,5 ,6, 16 

Polynuclear aromatic 2 benzene rings 4 100-150 - 3 to -2 0 to 2 3 to 5 1,2,3,4,5,6,11 

hydrocarbons 3 benzene rings 3 175-200 - 6 to -4 -2 to 0 4 to 6 1,2,3,4,5,6,11 

4 benzene rings 5 200-250 -11 to -6 -3 to -1 5 to 7 1,2,3,4,5,6,11 

> 4 benzene rings 4 250-275 -11 to -9 -4 to -3 6 to 8 1,2,3,4,5,6,11 

Nitrogen-substituted Aliphatic 10 30-125 1 to 3 >4 -1 to 1 1,2,3,4,5,6 

compounds Aromatic 4 175-250 0 to 3 2 to 3 1,2,3,4,5,6 

Polychlorinated 9 200-375 - 5 to -3 -3 to 1 3 to 7 1,2,3,4,6, 10,12 

biphenyls 

Organochlorine 13 300-400 - 7 to -4 -3 to 0 3 to 6 1,2,3,4,5,6,9,12 

insecticides 

Organophosphorus 

insecticides 

10 275-375 - 7 to -4 1 to 2 8,9 

Carbamate insecticides 8 175-275 - 5 to -1 1 to 5 7,8,9,14,15 

and herbicides 

Acid herbicides 11 175-250 - 7 to -3 2 to 4 7,8,9,12,15 

Triazine herbicides 11 200-250 1 to 3 7,8,9,1$ 

Substituted urea 7 150-275 - 7 to -4 2 to 3 7,9,15 

herbicides 

Bipyridylium herbicides 2 250-350 miscible 7,9,10 

Dinitroaniline 13 200-450 - 8 to -4 -3 to 1 7,9,13 

herbicides 

* 760 torr= 1 atmosphere 

11 U.S. EPA !1979a) ~7 Bailey and White (1965) 2 U.S. EPA 1981) 8 Paris and Lewis (1973) 

3 U.S. EPA 1982) 9 Wauchope (1978) 

4 Verschueren (1983) !10 Garten and Tralbalka (1983) !

13l Helling (1976) 14 Schlagbauer and Schlagbauer (1972) 

15 Bailey and others (1968) 

16 Mathur (1974) 

5 U.S. EPA (1978) 11 Hassett and others (1983) 

6 U.S. EPA (1979b) 12 Mackay and others (1980) 

Note: Base 10 raised to the -1 power = 0.1 

Base 10 raised to the 0 power = 1.0 

Base 10 raised to the + 1 power = 10 

be studied in terms of as many compound properties 

and environmental variables as possible. 

TRANSFORMATION PROCESSES 

Although physicochemical properties are indicative of the potential behavior of a compound in a water 

system, environmental variables are also important. 

This section reviews some of the transformation processes that can play a role in determining the environmental fate of an organic compound. The last part of 

this section (Environmental Behavior) examines how 

the physicochemical properties couple with the environmental variables to determine the primary transformation processes contributing to the persistence of certain 

anthropogenic organic compounds. 

5 

During their movement through the water column and the surficial sediment layer, organic compounds participate in many interactions which determine their ultimate environmental fate. Table 4lists and 

briefly defines some of the more important transformation processes and controlling factors. Table 5 

presents some of the measurable characteristics frequently used to assess the degree to which the behavior 

of a given compound is determined by a particular 

transformation process. The relative importance of the 

competing processes to each compound depends on the 

physicochemical properties of the organic chemical as 

well as the prevailing environmental conditions in the 

water column and in the surficial sediment layer. 

Volatilization, as a transformation process, illustrates some of the information presented in tables 4 and 5. 
Table 4.-Physica/, chemical, and biological transformation processes that affect manmade organic compounds 

Process 

Volatilization 

Sorption 

Photolysis 

Chemical oxidation 

Hydrolysis 

Bioaccumulation 

Biotransformation and 

biodegradation 

Definition and controlling variables 

1. Evaporative loss of a chemical. 

2. De~nds on vapor pressure and water solubility of the chemical and environmental 

vanables like wind, water turbulence, and temperature. 

3. Potentially important for compounds with high vapor pressures, low solubilities, and high 

activity coefficients. 

1. General term encompassing surface attraction (absorption) and partition (solubilization). 

2. Depends on the hydrophilic and lipophilic properties of the chemical and the composition 

of the sorbent. 

3. Indicative parameters are solubility, octanol-water partition coefficient, and sorbent 

organic carbon content. 

1. Nonmetabolic degradation requiring light energy: chemical undergoes either a direct 

transformation reaction from the absorbed energy or an indirect change from a reactionfor example, oxidation, with an excited chemical species or free radical. 

2. Depends on the chemical's absorption spectrum coefficient in the ultraviolet to visible 

range as well as the sunlight intensity distribution for a given time of day, season, latitude, 

depth in water, and ozone thickness. Also depends on the chemical's reaction quantum 

yield. 

1. Breaking down of the chemical bonds in organic compounds through a chemical reaction 

with photochemica]]y derived oxidants (singlet oxygen or free radicals). 

2. Depends on the number and types of possible reactive sites and on the presence of 

oxidants. 

3. Rate constants either measured directly or estimated from structure-activity relations. 

1. Reaction of a chemical with water, hydrogen, or hydroxide ion, commonly resulting in the 

introduction of a hydroxyl group in exchange for the removal of another functional group. 

2. Depends on the presence and number of hydrolyzable functional groups at neutral pH plus 

the catalytic effect of the addition of acids and bases at other pHs. 

1. Uptake and retention of chemicals in the water column by aquatic organisms through 

intake from water or diet. 

2. Depends on the nature of the chemical (lipophilic) and the organism's fat content, 

metabolic, and depuration rates. 

1. Enzyme catalyzed transformation of chemicals as a source of energy, carbon, and nutrients. 

2. Depends on the refractory and toxic nature of the chemical, and the presence of an acclimated microbial population, and a host of environmental factors including pH, 

temperature, dissolved oxygen, available nutrients, and cometabolites. 

It can be an important exchange process between a 

compound dissolved in solution and one vaporized in 

the atmosphere. Organic compounds with high vapor 

pressures, low water solubilities, and high activity coefficients are most susceptible to volatilization (for example, aliphatic hydrocarbons, mooocyclic aromatics, and 

some nitrogen-substituted compounds). The magnitude of Henry's law constant is used as an indicator of 

the volatility of a compound from water because it takes 

into account the physical properties of water solubility 

and vapor pressure. Yet, alone it cannot be used as a 

quantitative measure of the volatility of a compound, 

because volatilization also depends on environmental 

variables such as wind velocity, water turbulence, and 

temperature (U.S. EPA, 1978, 1979a). The balance of 

these environmental variables and physicochemical 

properties determines which groups of organic compounds are susceptible to volatilization and to what 

degree the exchange will alter the water column concentrations. 

Photolysis is another example of a transformation 

process that depends on both the properties of a 

compound and the environmental conditions. The 

degree of photolysis depends on the amount of light a 

particular chemical can absorb (absorption spectrum, 

coefficient) and the efficiency with which it uses that 

energy (reaction quantum yield). Given these properties of the organic compound, the actual degree of 

photolysis depends on climatic and meteorologic conditions such as sunlight intensity, cloud cove:.; time of day, 

season, latitude, and even ozone thickness (U.S. EPA, 

1979a, 1981). In addition, the water turbidity determines the amount of incident sunlight that actually 

penetrates the water column to the depth of the organic 

6 
Table 5.-Physica/ and chemical characteristics used to assess the distribution and fate of manmade organic compounds 

Characteristics 

Water solubility 

Vapor pressure 

Henry's law constant 

Reaeration rate ratio 

Partition coefficient 

Octanol-water partition 

coefficient 

Soil/sediment partition 

coefficient 

Soil/sediment partition 

constant 

Absorp~i?n spectrum 

coefficient 

Reaction quantum yield 

Photolysis rate constant 

Oxidation rate constant 

Hydrolysis rate constant 

Bioconcentration factor 

Biotransformation rate 

constant 

Definitions 

Indicative of a chemical's hydrophobic/hydrophilic nature, the limiting load in water, and its 

potential behavior in solution. Units: mass per volume. 

The partial pressure of vapor that is in equilibrium with a substance in pure state at a s~cified 

temperature. Relative or qualitative measure of the volatility of the chemical in Its pure 

state. Units: torr. 

Relative equilibrium concentration of a compound in air and water at standard temperature 

and pressure. Ratio of the vapor pressure divided by the chemical's water solubility; 

indicative of the compound's volatility. Units: torr per molar. 

Ratio of the first-order rate constant for loss of a chemical from aqueous solution divided by 

the rate constant for oxygen uptake by the same solution. Unitless. 

Estimate of the volatility of compounds with Henry's constants > 3,500 torr per molar for 

which the volatilization rate constant is limited by diffusion through the liquid-phase 

boundary layer. 

Equilibrium distribution of a compound solubilized between two immiscible solvents. 

Unitless. 

E<!uilibrium distribution of a compound between water and n-octanol. Indicative characterIstic of the bioconcentration potential of compounds. Unitless. 

Equilibrium distribution of a compound between water and a sediment/soil substrate. Mass 

of chemical per mass of sediment divided by the dissolved mass of chemical in solution 

per mass of solution volume. Unitless. 

Soil/sediment partition coefficient divided by the sediment/soil percent organic carbon 

content. Normalizes the partition coefficient to a sediment/soil substrate that is 100 percent 

organic carbon, permitting comparison of partition coefficients between substrates of 

differing carbon content. 

Quantity of light absorbed by the chemical at a particular wavelength in the ultraviolet-visible 

range of the electromagnetic spectrum. 

Efficiency of light utilization by a chemical. Ratio of the number of moles of the chemical 

transformed to the quantity (flux) of light absorbed. Unitless. 

First-order rate constant for direct photolysis. Units: reciprocal time. 

Sum of the rate constants for each individual type of reactive site in the compound. Units: 

reciprocal time. · 

First-order rate constant at pH of 7 and the second-order rate constants for acid- and 

base-promoted hydrolysis. Units: reciprocal time. 

Concentration of a chemical in tissue on a dry-weight basis divided by the concentration in 

water. Also the ratio of the uptake to depuration rates for a compound in a given organism. 

Unitless. 

Second-order rate constant dependent on the chemical concentration and the microbial 

biomass. Units: reciprocal time. 

compound. As a result of the large number of physicochemical properties and environmental variables that 

determine photolysis, the rate of photolysis for a given 

compound is commonly expressed by an overall parameter such as the photolysis rate constant. This parameter lumps most of the important factors into one 

measure, and it is useful for relative comparisons of 

photolysis rates between different organic compounds 

measured under similar conditions. 

column and sediment layers. The distribution of different transformation processes reflects preferences dictated by the driving forces for each process (Pionke and 

Chesters, 1973; Haque and Freed, 1974). At the airwater interface, energy supplied by light, wind, or 

surface tension effects is important. Volatilization, photolysis, and surface microlayer concentration of organic 

compounds are the most influential processes near the 

air-water interface. In the bulk solution, factors such as 

pH, ionic strength, dissolved oxygen, and the polarity of 

the chemical relative to water control the behavior of 

compounds. Other processes, such as the biological and 

chemical transformations and sorptive interactions 

A closer examination of the important transformation processes for anthropogenic organic compounds shown in table 4 indicates that different processes are spatially distributed throughout the water 

7 
depend more on the interaction of the bulk solution with 

natural particles and consequently can occur throughout the water column and surficial sediment layers 

(Olsen and others, 1982). At the sediment-water interface, the increase in organic matter, microbial populations, and other sediment components forces diagenetic 

reactions to take precedence (Oschwald, 1972). A brief 

description of these transformation processes follows, 

with the greatest detail given to those processes which 

are most important in controlling the behavior of 

persistent anthropogenic organic compounds (sorption 

and bioconcentration ). 

Atmospheric influence is most pronounced at the 

air-water interface, where volatilization provides an 

exchange mechanism for compounds in solution that 

have relatively high air-water partition coefficients 

(Henry's law constants). Input of light energy into the 

water column is also highest near the air-water interface. Consequently, the importance of photolysis as a 

degradative process for certain organic compounds 

decreases with depth as light radiation decreases. Some 

compounds may be concentrated within the surface 

microlayer (top 1 mm of the water column) to levels 

much greater than those found in the bulk solution 

(Macintyre, 1974; Rice and others, 1982). This surfaceexcess effect is manifested at the air-water interface 

and results from the low affinity of highly nonpolar 

organic compounds for polar water molecules. Despite 

available exchange pathways and low water solubilites, 

many organic compounds still enter into aqueous solution at environmentally significant concentrations. 

Once in the water column, organic compounds 

are subject to chemical and biological transformations 

or interactions with other sorbents. Reactions such as 

solvation, hydrolysis, oxidation, and ionization may affect 

the chemical speciation of certain organic molecules in 

a manner consistent with the oxidation-reduction potential (Eh), pH, and ionic strength of the bulk solution. 

Microbiologically mediated reactions mineralize parent 

organic molecules into more fundamental functional 

groups useful as nutrient and energy sources. Most of 

these reactions either transform the organic contaminant, sometimes to a more toxic form (for example, 

aldrin to dieldrin), or remove it from the water column 

to another environmental compartment such as the 

atmosphere, sediment, or biota. 

Many of the processes that occur in the water 

column continue across the sediment -water interface 

and into the interstitial fluid of the bed sediment 

(Presley and Trefry, 1980). Broadly classified as diagenetic interactions, these physical, chemical, and biological processes include deposition, accumulation, com8 

paction, mineralization, and decomposition of settled 

particulate matter along with its associated contaminant content (Krom and Sholkovitz, 1977; Baker and 

Feely, 1978). In the surficial sediment layer, chemical 

and microbial processes continue the degradation reactions that began in the water column. Biopolymers 

(proteins, lipids, and carbohydrates) are converted to 

monomers (amino acids, fatty acids, and sugars) and 

eventually to simple molecular components for recycling into solution by diffusion (Kuznetsov, 1975). Commonly these reactions are incomplete because new 

sediment is deposited or resuspension results from 

turbulent eddies or biological activity (bioturbation). 

Anthropogenic organic compounds, because of their 

sorptive interactions with natural organic materials, are 

incorporated into this biogeochemical cycling of elements. Consequently, diagenetic interactions provide 

another mechanism for exchange of anthropogenic 

organic compounds between the surficial sediment and 

the water column. 

SORPTION AND BIOCONCENTRATION 

Two other transformation processes-sorption 

and bioconcentration-are of particular importance to 

the fate of anthropogenic organic compounds. Sorptive 

interactions within the water column increase the carrying capacity of the bulk solution beyond the solubility 

limit of the contaminant. Sorptive interactions within 

the surficial sediment layer provide the sediment with a 

mechanism for accumulating and releasing organic 

contaminants. Together, sorptive interactions in the 

water column and in the surficial sediment layer can 

significantly affect the transport and cycling of contaminants (Eisma, 1981; Means and Wijayaratne, 1982; 

Olsen and others, 1982). 

In a similar manner, the bioconcentration of 

organic compounds by biological organisms, in the 

water column and sediment, can affect the distribution 

and fate of contaminants. Biological organisms can 

progressively concentrate organic contaminants, and 

this can directly impact humans because of consumption within the food chain. In addition, because biological organisms do not necessarily move with the average 

flow of surface waters (many species are independently 

motile), contaminants stored within the organisms may 

differ in their transport characteristics from those in 

the bulk solution. Finally, because biological organisms 

play an integral role in the exchange of matter between 

the water column and sediment (for example, bioturbation), they can alter the distribution of contaminants by 
exchanging organic compounds with the water column 

or the sediment (Kenaga, 1975a, b). 

fur sorptive processes, the exact nature of the 

particle-organic-compound interaction is a subject of 

practical importance in understanding the environmental behavior of organic contaminants. Current attention 

has focused on whether the sorption of organic contaminants results from a surface adsorption process or 

from a partitioning process between water and an 

organic sorbent (Chiou and others, 1979, 1983, 1984; 

Pavlou and Dexter, 1979; Mingelgrin and Gerstl, 1983). 

The advocates of a partition model believe that the 

sediment organic matter functions as an immobilized 

amorphous organic phase capable of partitioning 

nonionic organic compounds from the bulk aqueous 

solution in a manner similar to the solubilization taking 

place with a conventional polymer. The surface sorption 

concept treats the same results as adsorption wherein 

the mineral and organic fractions of the sediment have 

different adsorptive capacities. fur metal ions and ionic 

organic solutes it is logical that surface adsorption by 

different colloidal particles can take place as a result of 

electrostatic attraction, covalent bonding, and hydrogen 

bonding (Weber, 1972; Stumm and Morgan, 1981). fur 

nonionic organic solutes, available data indicate that 

sorption by soil/sediment particles is more effectively 

accounted for by the partition mechanism of the solutes 

with the soil/sediment organic phase. 

This review deals primarily with nonionic organics (for example, chlorinated insecticides, PCBs, and 

PAHs ), and therefore, it is pertinent to present evidence supporting the partition model. In addition to the 

recognized dependence of sorption on the soil/sediment 

organic content, Chiou and others (1979, 1983) show 

that the sediment concentration of nonionic organic 

solutes increases linearly with aqueous solute concentration up to the solubility limit of the solute. A nonlinear relation would be expected in a surface adsorptive 

process because the available adsorption sites eventually become saturated as solute concentrations approach 

solubility. In experiments using binary solutes, Chiou 

and others (1983) found that the sorption relation 

showed no competitive effects between different solutes 

and that the magnitudes of the sorptive capacities were 

similar to those for the single-solute cases. In a surfaceadsorption model, solutes compete for a fiXed number 

of adsorption sites, and the overall sorptive capacities 

per contaminant are generally lower than those in 

single-solute cases. Thermodynamically, the heats of 

sorption were found by Chiou and others (1983) to be 

small and less exothermic than the heats of solute 

condensation from water; this is in contrast to the large 

9 

exothermic heats of adsorption characteristic of surface adsorptive processes. Chiou and others (1985) 

suggest that the inorganic fraction of the soil/sediment 

is relatively inert to nonionic organic compounds in the 

presence of water, presumably because water is preferentially adsorbed over the nonionic organic solutes. The 

role of organic matter and other sediment fractions in 

sorptive processes is discussed more thoroughly in the 

section on Suspended and Surficial Fractions. 

Just as nonionic organic compounds partition 

into soil/sediment organic matter, bioconcentration can 

be viewed as a transformation process in which organic 

compounds partition into the lipid fraction of aquatic 

organisms. Bioconcentration is the accumulated concentration of a given compound over that in the surrounding medium (water) by a particular aquatic organism resulting from its diet or from intake of 

contaminated water during respiration (Kenaga, 1975a, 

b). This process generates a reservoir of contaminants 

within the organisms that may enter the food chain. 

Biomagnification is commonly associated with bioconcentration; it is the progressive increase in the concentration of a compound moving from one trophic level to 

the next higher trophic level of the food chain. This 

discussion is limited to bioconcentration because 

biomagnification studies lack a consistent reference 

point for comparison among different trophic levels, 

especially aquatic as opposed to terrestrial organisms. 

Biomagnification has yet to be rigorously defined in a 

manner that accounts for the physiologic differences 

between species and that provides an unequivocal reference concentration for each organic contaminant. 

Once a reference standard is established, it should 

compensate for the different compositions and amounts 

of organic solvents present in the bodily fluids of each 

species of aquatic organism. 

Chiou (1985) has demonstrated that the bioconcentration of nonionic organics by aquatic organisms 

results from the partitioning of compounds into the 

lipid fraction of the organism. Those compounds that 

are strongly bioconcentrated generally have a low water 

solubility and a relatively high affinity for organic phases, 

as is indicated by a high octanol-water partition coefficient. These properties are characteristic of many 

pesticides whose original function was to penetrate the 

cellular structure of the organism and disrupt the 

life-sustaining processes of the plants and insects. Once 

within the organism, these compounds may exhibit 

improved stability against decomposition and transformation. 

From a mass balance standpoint, bioconcentration and biomagnification cycle contaminants as well as 
vital elements such as carbon. From a toxicological 

standpoint, these processes are important because of 

the physiologic effects promoted by elevated concentrations of persistent contaminants. In general, by the time 

most pesticides run off into surface waters, concentration levels are sufficiently low that the acute toxicity of 

the compound is low and frequently unnoticed (Neely 

and others, 1974). Bioconcentration raises the potential 

for chronic effects because it increases the residence 

time of contaminants within the biological reservoir. 

fur these reasons, estimates of the bioconcentration 

potential of synthetic compounds are important in 

order to assess their possible impact prior to their 

widespread use and environmental release. 

fur the persistent manmade organic compounds, 

several physicochemical properties and transformation 

processes play important roles in their environmental 

distribution. Low water solubilities and high octanolwater partition coefficients are common characteristics. 

These properties lead to sorptive interactions and 

bioconcentration as the primary processes controlling 

environmental distribution. They are not solely responsible for the distribution of anthropogenic organics, but 

because their relative contribution is greater than that of 

other transformation processes, sorptive processes are 

the focus of the next discussion of interactions between 

organic compounds and sediment fractions. But first, 

the next part considers the contribution of environmental variables in determining the dominance of certain 

transformation processes under a given set of environmental conditions. 

ENVIRONMENTAL BEHAVIOR 

The specific behavior of organic compounds in 

the environment is primarily a manifestation of the 

interaction between the physical and chemical properties of the compound and the sorbent. The effect of 

these inherent properties in the water column and 

sediment may be subject to changes by environmental 

variables such as pH, Eh, dissolved oxygen, dissolved 

solids, temperature, water hydrodynamics, and atmospheric forcings. Table 6 presents some important 

reaction mechanisms controlling the transformation 

and removal processes for the different organic compound groups. fur some groups, such as halogenated 

aliphatic hydrocarbons, a single property dominates the 

transformation mechanism. For others, apparent 

physicochemical properties can be masked because 

they are more dependent on environmental conditions. 

fur example, the extent of hydrolysis of the carbamates 

and phenoxy herbicides is largely a function of pH. 

10 

The physicochemical properties of a compound 

alone may not adequately indicate its probable field 

behavior without also considering the effect of environmental variables. Physical and chemical characteristics 

are measured under controlled conditions, whereas in 

nature the behavior of organic compounds may be 

influenced by other physicochemical factors not identified in simple laboratory systems. This is especially true 

for periods before a steady-state condition is reached.· 

fur example, insecticides and herbicides may be applied 

in different modes such as water solutions, emulsions, 

gases, wettable powders, and granular pellets. The form 

of introduction of the compound certainly affects its 

residence time and transient behavior in the different 

environmental compartments. fur instance, compounds 

that are sufficiently volatile in water, such as hexachlorobutadiene, are less likely to volatilize if they are 

initially sorbed to some soil or sediment particle. Nevertheless, the physicochemical properties of the compound and the sorbent provide an important criterion 

for predicting and evaluating the initial distribution 

tendency of a compound between different sorbent 

phases. Hence, the physicochemical properties provide 

a useful initial assessment of the ultimate fate of 

organic contaminants. 

The fate of many organic compound groups is 

characterized by competing processes mediated by the 

properties of the compounds and the prevailing environmental conditions. fur example, monocyclic aromatics are susceptible to either mlatilization followed by 

atmospheric photolysis or sorption followed by bioaccumulation (U.S. EPA, 1979a, b). The pair of transformation processes that control the behavior of the compound may vary with environmental factors such as 

wind speed, wave turbulence, temperature, and the 

availability of particulates for sorption in the water 

column. The reactivity of other organic groups, such as 

the substituted phenols, likewise depends on the location of the compounds in the water column. Near the 

air-water interface, photolysis is the primary reaction 

mechanism, whereas closer to the sediment -water interface, photolysis is greatly reduced and biodegradation 

reactions predominate. fur other compounds that 

readily undergo enzymatically catalyzed microbial degradation (for example, phenols, phenoxy herbicides, and 

triazine herbicides), environmental fate strongly 

depends on the presence of a sufficiently large and 

acclimated microbial population. 

Organic compound groups such as the 

organophosphorus insecticides (for example, parathion 

and malathion) undergo rapid transformation to different chemical forms (paraoxon and malaoxon), which 
are more persistent and sometimes more toxic than the 

parent compounds. The dissipation of the parent compounds may appear to be deceptively rapid from an 

environmental standpoint, but their effect is manifested 

primarily through other reaction intermediates. Some 

other compounds (for example, halogenated ethers) 

may persist in an aqueous medium despite dissipation 

by hydrolysis primarily because of their very high 

solubilities in aqueous solution. In other words, consideration of the physical and chemical properties of a 

compound should indicate its tendency for transport 

and partitioning, whereas the environmental fate of the 

chemical may depend on other system conditions. 

The combination of the properties of a compound and available field data give a good indication of 

its potential persistence in the water column and sediment system. Table 7 presents estimates of different 

pesticide half-life ranges based on data collected from 

different ecosystems (rivers, lakes, soils, and estuaries). 

Information on the industrial organic compounds is 

less readily available, and consequently these compounds are not included in the list. fur industrial 

compounds, some estimates can be made on the basis 

of properties and environmental behaviors that are 

similar to those of the pesticides. 

SUSPENDED AND SURFICIAL SEDIMENT 

FRACTIONS 

This section examines the physical and chemical 

characteristics of various suspended and surficial sediment fractions. These sediment fractions are discussed 

to illustrate how their component particles interact with 

organic contaminants. It is recognized that certain 

components of sediment (for example, humic substances) do not exist as particles per se, and consequently, use of the term "particle" is not intended to 

imply a specific physical or geometric form but rather 

to differentiate fractions, some of which behave as 

discrete entities (for example, sand and colloids) when 

transported. Indeed, at the submicron end of the 

particle-size spectrum, the distinction between what is 

and what is not a particle becomes more unclear as the 

scale of observation approaches the limits between 

"particulate" and "dissolved." In general, 0.45 p.m is 

the operationally defined demarcation between particulate and dissolved matter. Nevertheless, colloids are 

treated as discrete particles (Edzwald and others, 1974) 

even when their dimensions fall below the 0.45-p.m 

mark. 

As a result of the wide variety of particle types 

that comprise suspended and surficial sediment, their 

11 

physical and chemical properties are not as readily 

measured as are those of pure anthropogenic organic 

compounds. Sediment varies widely in mineral composition, ultimate analysis, and organic matter content. 

Consequently, much of the description of suspended 

and surficial sediment remains qualitative, and the few 

measures that have been developed only partially characterize the different sediment fractions on the basis of 

physical size and chemical fractions. 

The properties of suspended and surficial sediment are differentiated here in terms of three categories: biological particles, inorganic particles, and humic 

substances. Although these categories do not include 

all sediment components, they include those components currently thought to be most important in influencing the behavior of anthropogenic organic compounds. 

BIOLOGICAL PARTICLES 

Many biological organisms present in the water 

column influence the movement and transformation of 

dissolved and sorbed organic contaminants. They 

include bacteria, protozoans, phytoplankton, zooplankton, and other higher trophic species (fig. 1). 

Although these different biological species are not 

strictly considered suspended sediment, they are 

included here because they are suspended in the water 

column, participate in the transformation and transport 

of organic contaminants, and eventually (after they die, 

start to decay, and settle) become part of what is 

collectively considered surficial sediment. In addition, as 

dead aquatic organisms, they combine with terrestrial 

organic decay products to form detritus, and thus they 

become an integral component of the material cycle of 

elements between the water column and surficial sediment (Sigleo and others, 1980, 1982). 

The ways that biological organisms transform 

organic contaminants are diverse. By exchanging water 

across their cellular membranes, bacteria, protozoans, 

plankton, and lower vertebrates can bioconcentrate 

dissolved contaminants within themselves (Lush and 

Hynes, 1973; Steen and Karickhoff, 1981). Some filter 

feeders, such as zooplankton, ingest lower trophic 

organisms and colloidal size particles and either bioaccumulate the contaminants within their bodies or 

excrete them with fecal pellets (Sheldon and others, 

1973). The fecal pellets are denser and settle faster 

than the ingested colloids or organisms (Meyers and 

others, 1984). Consequently, as has been demonstrated 

for the oceanic distribution of PCBs, the pelletization 

process can provide a significant contribution to the 
Table 6.-EnvironmBntal fate processes for selected groups of manmade organic compounds 

Chemical group Primary 

mechanism 

Secondary 

mechanism Com menta References 

1. Halogenated aliphatic Volatilization Compounds with more than 1,2,4 

hydrocarbons five chlorine atoms are sorptive and have a potential for 

bioaccumulation. 

2. Halo~nated ethers 

aliphatic------ -• Hydrolysis- - -• Volatilization and Some ec?rsistence due to high 1,2,4 

photo-oxidation in solubility. Also persistence 

atmosphere potential due to lipophilicity. 

aromatic • - - - • - -• Sorption - - - - + Bioaccumulation and 

3. Monocyclic aromatics 

biodegradation 

Volatilization---+ Atmospheric Volatilization/sorption are 1,2,4 

oxidation/photolysis competing reactions with 

Sorption and - - -• Slow biodegradation environmental conditions 

bioaccumulation (especially nitrogen determining dominance of 

compounds) one reaction. Some persist4. Phenols Photolysis- - - - - -• ~near air-water surface). ence from hi~ solubility. 

No accumulatiOn/persistence 1,2,4 

Biodegradation- -• near water sediment If except for nitrophenols 

sufficient microbes) which sorb readily to clays. 

5. Monocyclic aromatics and Sorption Chlorine content dominates 1,2,4 

phenols with 5 or more Bioaccumulation compound behavior. 

chlorines 

6. Phthalate esters Sorption Bioaccumulation Biologically reactive in meta- 1,2,4 

Biotransformation bolic processes. Because of 

Biodegradation sorption, transfo:rt depends on hydrogeo ogic conditions. 

7. Porrouclear aromatic Sorption Photolysis (for com- Bioaccumulation short-term 1,2,3,4 

by rocarbons Bioaccumulation pounds with low ring because readily metaboBiodegradation numbers) lized. Sorption increases and 

Volatilization biodegradation decreases 

with increasing number of 

benzene rings. 8. Nitrogen compounds 

aliphatic - - - - - - + Photolysis Mainly hydrospheric photol- 1,2,4 

aromatic - - - - - - + Sorption - - - - -• Biodegradation btis due to high solubility 

9. ~anochlorine insecticides 

ut also atmospheric. 

Sorption Volatilization (if chemi- High degree of chlorination 1,2,3,4,5, 

PCBs Bioaccumulation cal is not immediately fo:in~ low solubility, high Biotransformation sorbed) pop ilicity, and significant 

Acid pH - stable Photo-oxidation enVIronmental persistence. 

Generally short half-lives. 

10. OrganoP!tospllorus Alkaline pH- Biological oxidation Oxidation byproducts (e.g., 3,5,7 ,8,9,10 

insectiCideS rapid chemical- some sorption slows malaoxon and ~araoxon) 

microbial hydro- down hydrolysis may be more persiStent than 

lysis the parent compounds. 

11. Carbamate insecticides and ChemicaVmicrobial Relatively short stability 5,6,11 

herbicides Hydrolysis -- - -• ~sensitive ~weeks) with harmless 

Photo-oxidation me sorption to sedi- ecomposition byproducts. 

ments 

12. Phenoxy acid herbicides Photolytic decomp Slow microbial Low bioaccumulation ~ten- 5,6,7 

under ultraviolet degradation tial due to low lipop ilicity 

light Sorption more pro- and high depuration rates. 

Hy~!Ol)is (pH sen- nounced for esters 

Sittve than for acids 

13. Triazine herbicides Hydrolysis (acid and Microbial degradation 5 

14. Sv.bstituted urea 

base catalyzed) 

Sorption Microbial, chemical, Chemical reactions like oxida- 12 

herbicides Degradation- - -• photo-chemical- tion, hydrolysis, etc., /lay 

slow (months) minor role compare to 

enzymatically mediated degradation reactions. 

12 
Table 6.-Environmtmtal fate procssses for sslsctsd groups of manmade organic compounds-Continued 

Chemical group 

15. Bipyridylium herbicides 

Primary 

mechanism 

Secondary 

mechanism Comments References 

Sorption 

(clays)----~ 

Sorption - - - ~ 

(humic substances) 

Highly sorptive 

Essentially irreversibly sorbed under field conditions. 

Strongly sorbed but unlike clays can undergo some microbial degradation. 

13,14 

16. Dinitroaniline 

herbicides 

Not persistent because of 

photodecomposition, biodegradation, and high 

depuration rates. 

15 

3 U.S. EPA 1978 

Sources: Ill U.S. EPA l1979a) 2 U.S. EPA 1981l 7 Faust and Suffet (1966) !

61 Norris (1981) 

13 Akhavein and Linscott (1968) 

1

111 Schlagbauer and Schlagbauer (1972) 12 Maier-Bode and Hartel (1981) 

4 U.S. EPA 1982 

5 Paris and Lewis (1973) 

8 Van Middelem (1963) 

9 Mulla and Mian (1981) 

(10 Mulla and others (1981) 

14 Hayes and others (1975) 

15 Helling (1976) 

vertical transport of organic compounds through the 

water column (Elder and fuwler, 1977). Finally, some 

biological organisms which are resident in the surficial 

sediment layer (for example, benthic macroinvertebrates) are capable of ingesting and resuspending 

(bioturbation) sediment particles along with their 

sorbed contaminants (Berner, 1980). 

INORGANIC PARTICLES 

The most commonly identified sediment components are the mineral particles of clay, silt, sand, and 

gravel (see fig. 1 ). These particles comprise the bulk of 

the transportable sediment on a mass basis. The clay, 

silt, sand, and gravel fractions are separated on the basis 

of size by sieving, settling, or centrifugation methods. 

Size is used as the distinguishing criterion because of 

its relation to the transport and settling properties of 

the various particle types. Unfortunately, the names 

assigned to the particle-size classes (sand, silt, and clay) 

also imply certain compositional properties. Although 

the implications are partially accurate, many particles 

falling within the sand-sized class may be composed of 

organic matter (shells, seeds, and detritus) rather than, 

say, quartz. Other particles falling within the clay-sized 

class may be metal precipitates (for example, iron and 

manganese oxides and hydroxides) rather than particles of clay mineral composition. 

Mixed within the various size classes of sand, silt, 

and clay are other less readily identifiable mineral 

fragments. They occur in various stages of mineralization and dissolution and include the opaline and calcareous exoskeletons of dead organisms. Although the 

sorptive capacity of these mineral particles for nonionic 

organic compounds in water is small relative to the 

sorptive capacity of organic matter (Chiou and others, 

13 

1985), it is not negligible. The mineral particles are 

commonly coated with organic sheaths that provide a 

microscale organic medium for sorption (partition) of 

nonionic organic solutes. Commonly, this partitioning 

capacity is mistakenly attributed to the mineral grains. 

Organic matter and metal oxides are two groups 

of surface-active substances capable of altering the 

electrochemical properties of suspended mineral particles in surface waters. Acting individually or together, 

these groups are abundant enough to affect the interfacial properties of transported sediment matter 

(Hunter and Liss, 1982). Surface-active substances are 

sufficiently strong to remove the intrinsic differences in 

the laboratory-measured stability of clay minerals 

(Gibbs, 1977). Hunter and Liss (1982) found that the 

charge distribution on suspended particle samples collected from four estuaries was highly uniform despite 

the mixed composition of the suspended matter particles. They attributed the homogeneous electrical properties to both the sorbed organic matter and the metal 

oxides. 

Even though the percentage by weight of dissolved organic matter sorbed on suspended particles in 

lakes, rivers, and marine waters may be small, the 

extent of surface coverage by adsorbed organic matter 

can be large (table 8). Davis (1982) found that within 

the typical pH range of natural waters, almost complete 

surface coverage by dissolved organic matter can be 

expected for hydrous aluminum oxides, hydrous iron 

oxides, and edge sites of aluminosilicates. The natural 

organic matter is characterized by weak, acidic functional groups that form surface complexes with the 

relatively basic surface hydroxyls of the mineral and 

metal oxides (Davis, 1982). Other mineral types (for 

example, silica) have relatively acidic surface hydroxyls 

that form weaker complexes that under natural 
Table 7.-Estimates of environmental persistence of selected 

groups of manmade organic compounds 

Groups Estimated half-lives 

Lead, arsenic, copper, and mercury 

pesticides 

Organochlorine insecticides and 

PCBs 

Bi{'yridylium herbicides 

Tnazine herbicides 

Substituted urea herbicides and 

picloram 

Benzoic acid, dinitroaniline, and 

amide herbicides 

Phenoxy and toluidine herbicides 

Carbamates and aliphatic acid 

herbicides 

Organophosphorus insecticides 

10-30 years 

2- >4 years 

>2 years 

6 months- 2 years 

4 months - 1 year 

3 months - 1 year 

1 month - 6 months 

2 weeks - 3 months 

1 week - 2 months 

Sources: Metcalf and Pitts (1969) 

Wauchope (1978) 

Stojanovic and others (1972) 

Maier-Bode and Hartel (1981) 

Metcalf (1972) 

Paris and Lewis (1973) 

conditions may not be sufficiently energetic to cause 

complete surface coverage. The extent of coverage of 

the oxide-water interface also depends on the pH, the 

availability of surface area sites and sorbable organics, 

the nature of the solid surface, and the inorganic 

electrolyte strength and composition of the bulk solution. Organic coatings may play a major role in the 

surface chemistry of some particulate-sized mattet; 

especially for oonionic organic solutes; they are geochemically important in the interfacial processes of 

coagulation, sedimentation, and sorption. 

Many inorganic particles fall within the colloidal 

range of the particle-size spectrum. They include clay 

mineral colloids as well as oxides and hydroxides of iron 

and manganese. Colloidal size, organic content, and 

composition give the clay particles cation-exchange capabilities for ionic species (for example, metal ions) and a 

micro-organic environment for sorption of nonionic 

organic solutes. Some expansive clays, like montmorillonite, allow interlamellar penetration and entrapment 

of cationic organic species, such as paraquat (Adams, 

1973). Because of the presence of organic matter in the 

form of organomineral colloids, the sorptive capacities 

measured for clay-sized samples are commonly much 

larger than those expected solely on the basis of a pure 

mineral sorbent. 

Depending on water pH and redox potential, 

trace metals such as iron and manganese are present in 

suspension in precipitated forms. The oxides and 

hydroxides of iron and manganese occur as "dissolved" 

colloids, orgaoometal colloids, or aggregated particles 

14 

(Sigleo and Helz, 1981). In conjunction with humic 

materials, the metals coat the mineral particles that 

serve as nucleating surfaces for their precipitation 

reactions (Sholkovitz, 1976). The sorptive and scavenging ability of the metals is attributable to surf~ce 

properties and is more pronounced for freshly precipitated floes than for aged aggregates (Duinker, 1980). 

Sorption of ionic pesticides by iron and aluminum 

oxides has been hard to verify in field situations because 

of the difficulties in sampling the particle aggregates 

without altering or destroying them (Olsen and others, 

1982). 

HUMIC SUBSTANCES 

Based on earlier work indicating the importance 

of organic matter in soil-pesticide interactions 

(Hamaker and Thompson, 1972; Hamaker, 1975; 

Schnitzet; 1978a, b; Chiou and others, 1979, 1983), 

similar procedures have been employed to isolate and 

study the organic matter in sediment. Chemical fractionation, as used here, refers to the various methods of 

extracting the organic matter from the inorganic matrix 

of the bulk sediment sample on the basis of solubility in 

different solvents. As used in this paper, chemical 

fractionation excludes the methodologies employed to' 

separate organic contaminants from bulk sediment 

samples or their fractions even though the extraction 

principles and solvents used for organic contaminant 

and organic matter may be the same. Because most of 

the investigations described in the literature use the 

humic substance extraction method, this review of 

interactions between chemical fractions of the sediment 

and organic contaminants deals almost exclusively with 

the characteristics of the different classes of humic 

substances. 

The most commonly used extraction method for 

organic matter was borrowed from the soil literature. It 

employs sequential isolation in dilute alkali and mineral 

acid to separate the humic substance fractions from the 

sediment. Other solvent extraction methodologies, which 

include inorganic solvents such as aqueous solutions of 

neutral salts (for example, sodium pyrophosphates or 

oxalate salts) and various organic solvents (for example, hexane, ethers, carbon tetrachloride, alcohol, or 

benzene), are less commonly used. These procedures 

are superior to acid and alkali methods because they 

are more selective to specific groups of organics (for 

example, waxes, lipids, fats, or celluloses) and they 

minimize the hydrolysis, decarboxylation, and oxidation 

of the organic matter that may result from the use of a 

stronger alkaline solution. More comprehensive 
... 

Cl'l 

Figure 1.-Size spectrum of natural particulates and the hydrodynamic forces affecting their settling. 

I angstrom nanometer ---- ---micrometer milllmeter-centlmete_r_ meter I 

I 

I -10 -9 -8 -7 -6 -5 -4 -3 -2 -1 o 

iSize, In meters 10 10 10 10 10 10 10 10 10 10 10 10 

I I I I I I I I I I I I I 

I I I I I I I I I I I I I 

I <Molecular>< Colloidai------->I<Cot loldai--><---Coarse suspension---->! I I I 

I < solution>< solution >I< suspension> I I I I I I I 

I I I I I I I I I I I I I 

!Relative number I 21 18i. 181 12 I 12 91 I 9 I I I I I I 

I of particles <10 - 10 ><----10 - 10----->1<10 - 10 I ><------l-10 - 1 1--------> I I I I 

I I I I I I I I I I I I 

I I I I I I I I I I I I l~lnerals IMoleculesi-Col loldsi---------1-><·-Ciayi-><---SIIt----><---1-Sand---- -><---Gravei----><Cobble><Boulder I 

I I I I I I I I I I I I 

I 1---------1---Dissolved matter-----><-------Partlculate matter-------- ---------1---------1---------1---------

1 I I I I I I I I I __ _ I I I I I I I I I 

Metals <Free lons><Metal- ><Dispersed><Partlculate > I I 

I l<humlc >< metal ><metal I > I I 

I l<acld >< collolds><precfpftates> I I 

I <complexes> I I I I ~-------1 I I I I I I , ___ _ I I I I I I I I 

I<Fatty acids> I<VIruses>l <Bacteria> I I 

I I I I I I 

I I<Fulvlc acids> I <Zooplankton----------------- ---------1 

I I I I I I I 

I <Carbohydrates> I I I I I 

I I I I I I I 

organic compounds I I <Humic acids> I <Phytoplankton-------------------------1--------- 1 I I I I I I I t 

I <Am I no acids> I <C I ay-humic-meta I> I I I I I 

I I I< camp I exes > I I I I I 

I <Hydrophilic acids> I I I I I I 

I I I I I I I I I 

!<Hydrocarbons> I I I I I I I 

I I I I I I I I I -------------, I I I I I I I I ~---

1 I I I I I I I I 

1---------1---------1-Browniantmotlon-->I<Fiuld shear forcesland Gravitational forces----- ---------1---------

1 I I I I I I I I I 

I I I<Perlkinetlc> l<ldeal Stokes >I<Nonfdeal Stokes---1--------- ---------1---------

1 I !<coagulation> I< settling >I< settling I I 

I I I I<Orthokinetic> I I I I I 

I I I !<coagulation > I I I I I 

Hydrodynamics 

Sources: Thurman (1985) Stumm (1977) 

0' Me I I a ( 1980 ) 

Sheldon and others (1972) Elsma (1981) 
Table a.-Colloids, aggregates, snd their chsrsctsristics 

Particle group• 

Kaolinite 

Illite 

Chlorite 

Montmorillonite 

Smectites 

Vermiculite 

Allophane 

Iron hydroxides 

Manganese oxide 

Source: Horowitz (1984) 

Cation-exchange capacity 

In meq/fOOg* 

3- 15 

10- 40 

20- 50 

80-120 

80-150 

120-200 

40- 70 

10- 25 

200-300 

Note: • meq = milliequivalents 

extraction techniques available for use on dissolved 

organic matter are capable of isolating hydrophilic 

(acid, base, and neutral) and hydrophobic (acid, base, 

and neutral) organic classes and subclasses (Leenheer, 

1981 ). It is recognized, however, that all extraction 

procedures are incomplete in removing all organic 

carbon because different compounds have different 

solubilities in extractants. A tradeoff is necessary 

between the extent of removal from the sediment and 

the ability to maintain the chemical integrity and characteristics of the organic matter fraction of interest. 

The importance of humic substances is reflected 

by the large quantity of recent research focused on the 

transport and fate of aquatic and terrestrial humic 

materials (Christman and Gjessing, 1983a, b; Aiken 

and others, 1985). As compounds from diverse origins 

and of complex chemical composition, humic substances can function in sediment and water chemistry as 

buffers, ion exchangers, surfactants, sorbents, and 

chelating agents (Kononova, 1966; Jackson, 1975). 

Research also indicates that humic substances interact 

with metal ions and their hydrous oxides, clay minerals, 

and organic pesticides and also serve as precursors to 

halogenated hydrocarbon formation following watersupply chlorination (Josephsen, 1982). A significant 

effect of the sorptive capacity of humic substances is 

the ability to concentrate toxic substances in solution to 

levels above their normal solubilities (Wershaw and 

others, 1969; Poirrier and others, 1972; Schnitzer and 

Khan, 1972; Carter and Suffet, 1982; Chiou and others, 

1986). 

The humic substances have been broadly defined 

as amorphous, brown or black, polydisperse substances 

with polar and acidic functional groups whose molecular weight ranges from several hundred to several 

million (Schnitzer, 1978a). This general definition does 

not describe the diverse physical forms humic substances take on in response to variations in pH and 

aqueous ionic strength. In solution, humic substances 

Surface area In m2/g Dominant surface 

groups 

Surface 

charge 

16 

7- 30 

65-100 

600-800 

600-800 

200-400+ 

100-800 

Oxygen/hydroxyl 

Oxygen 

Oxygen 

Oxygen 

Oxygen/hydroxyl 

Oxygen/hydroxyl 

Oxygen/hydroxyl 

(+,-) 

~~ (+,o,-) 

(+,o,-) 

may be in the form of dissolved species, dispersed 

colloids, precipitated floes or in association with mineral and hydrous metal aggregates. 

Research has shown that the characteristics of a 

humic substance are strongly dependent on its origin. 

Differences have been distinguished between humic 

substances derived from terrigenous plants and planktonic algal sources (Josephsen, 1982). Functional group 

content analyses indicate that terrigenous humic substances have a higher percentage of aromatics (for 

example, phenols) that are characteristic of lignin breakdown products derived from vascular plants (Plechanov 

and others, 1983). Planktonic humic substances are 

more aliphatic and are characterized by protein and 

carbohydrate residues (Ertel and Hedges, 1983). Terrigenous humic substances have higher carbon to nitrogen ratios than do planktonic ones, because of the 

higher content of nitrogen-poor, lignin- and cellulosederived compounds (Meyers and others, 1984). Also, 

chemical fingerprinting has demonstrated that aquatic 

humic substances can be formed in situ, entirely from 

degradation of plankton and without any contribution 

from terrigenous sources (Khan, 1980). 

The categorization and chemical characterization 

of humic substances are difficult because they repre.; 

sent organic matter from diverse origins that react 

degradatively to form an even wider spectrum of breakdown compounds. Consequently, most of the analytical 

research conducted has been aimed at elucidating the 

gross structural features and major functional moieties 

active in the different groups of degradative reactions 

(Christman and Gjessing, 1983a, b). The most commonly used classification of humic substances is operationally defined and founded on the basis of solubility 

in different extraction solvents. The three classes identified on the basis of solubility/insolubility in acid and 

dilute alkali are fulvic acids, humic acids, and humins 

(fig. 2). Further fractionation of the humic acids into 

hymatomelanic acid (alcohol soluble), gray humic acid 
Plants, anfmals, microbes·, and thefr wastes 

BIOPOLYMERS: 1 

Nonhumic substances 

(carbohydrates, ~----------Decay processes 

and pigments) proteins, fats, ! 

GEOPOLYMERS: 

Humic substances 

Insoluble 

Hum ins 

.Jl T~::~r'th 

/\Soluble I Tre:~~~lth 

Precipitated~ Not precipitated 

Humic acids Fulvfc acids 

)\

tract with 

alcohol 

Soluble Insoluble 

Hymatomelanic acid 

Figure 2. Fractionation of humic substances (from Saar and Weber, 1982). 

(precipitated in electrolytic solution), and brown humic 

acid (soluble in electrolytic solution) can be, but is 

infrequently, performed. 

Humic substances originate from the chemical 

and biological degradation of plant and animal residues. 

Attempts to identify the general breakdown mechanisms have not been entirely successful, but it is agreed 

that the more complex, higher-molecular-weight humic 

substances ( > 1,000 amu) are degraded first (oxidized) 

to lower-molecular-weight materials such as the humic 

and fulvic acids (Khan, 1980). Consequently, the structure of the three groups is similar, but they differ in 

ultimate analysis, molecular weight, and functional 

group content (Rashid and King, 1970). In general, the 

lower-molecular-weight humic substances are more 

prevalent in the sediment of rivers, lakes, and seawatet; 

whereas the heavier humic substances predominate in 

soils (Josephsen, 1982). This difference in distribution 

reflects the importance of water column degradative 

processes that are continually active during the transport of the humic materials. 

17 

The fulvic acids are described as water-soluble 

compounds that impart the yellowish to brownish color 

of certain natural waters. In comparison with the 

humins and humic acids, the fulvic acids have progressed farther through the chain of chemical and 

biological degradative processes. Consequently, they 

have lower molecular weights and more hydrophilic 

functional groups than do the heavier humic acids and 

humins (Khan, 1980). 

In aqueous solution, the humic acids are less 

soluble than the fulvic acids and can be precipitated at 

pHs below neutral. Their molecular weight is intermediate between that of the fulvic acids and humins and 

changes with variations in pH aftd ionic streagth 

(Plechanov aiXI others, 1983). As a .result· of tboir 

sensitivity to pH, the molecular stntdu.rc of the lmmic 

acid complex is more aggregated and coiled· in ·acidic 

conditions and more dispersed and extended ·in an 

alkaline environment. The consequences of this reversible change in molecular configuration can potentially 

affect the ability of the humic acid molecule to interact 
with or sorb other chemical species in water. A detailed 

study of this effect and its influence on the behavior of 

humic substances interacting with organic contaminants has yet to be explained fully. 

The last and least studied of the humic substance 

fractions are the humins. In the gradation of bonding 

strengths found when progressing from the fulvic acids 

to the humic acids to the humins, the humins possess 

the most stable configuration. The relative insolubility of 

the humins, even under a wide range of pH conditions, 

seems to be due to their association with inorganic 

constituents of soil and sediment (Schnitzer and Khan, 

1972). Their molecular structure and ability to form 

stable metal and (or) organomineral complexes contribute to their observed resistance to microbial and chemical degradation. 

INTERACTIONS BETWEEN SEDIMENT 

FRACTIONS AND MANMADE ORGANIC 

COMPOUNDS 

The interaction of manmade organic compounds 

with suspended and surficial sediment fractions is examined in the four parts of this section. The first two parts 

consider only the humic substance fraction of the 

sediment. The first focuses on the sorptive interactions 

between sediment-bound humic substances and 

nonionic organic compounds and demonstrates that 

humic substances, as dissolved and colloidal organic 

matter, are capable of enhancing the apparent solubility 

of nonionic organic compounds. The second part 

presents data showing that the sorption of ionic organic 

compounds by sediment humic materials is more 

dependent on bulk solution parameters (for example, 

pH) than is the sorption of nonionic organic compounds. The third part reviews a number of field 

studies that investigate sediment -contaminant interactions on the basis of sediment -size fractions sampled 

from rivers, lakes, and estuaries. The last part reviews 

laboratory studies that attempt to define more accurately the relations observed in the field data. These 

studies lead to a quantitative description of the sorption 

of contaminants by sediment fractions, and this forms 

the basis for the discussion of environmental distribution estimation procedures in the section on Estimation 

of the Field Distribution of Manmade Organic Compounds. 

HUMIC SUBSTANCES AND NONIONIC 

ORGANIC COMPOUNDS 

Humic substances participate in a variety of 

chemical interactions that play important roles in the 

18 

fate and transport of organic contaminants. Humic 

substances may interact with metal ions and hydrous 

oxides through ion-exchange, complexation, chelation, 

coagulation, and peptization (Schnitzer and Khan, 

1972). The resulting organometal complexes can have a 

wide solubility range in water and appear as discrete 

colloids, aggregated masses, or surface coatings of 

mineral particles. As separate particles or in conjunction with other particulates, humic substances also 

influence the behavior of organic contaminants primarily through sorption. 

The process by which humic and fu1vic acids sorb 

as well as enhance the apparent solubility of nonionic 

organic compounds can be described in terms of a 

partitionlike interaction (Chiou and others, 1986). The 

extent of this interaction is determined by the polarity 

and size of the organic matter molecules and the 

magnitude of solute solubility in water. The net effect of 

this partition interaction on the transport of nonionic 

organic compounds is twofold. First, the humic and 

fulvic acids, as dissolved organic matter, can partition 

nonionic organic compounds to give apparent solubilities greater than usual saturation in water and consequently increase the carrying capacity of the water 

column for many organic contaminants that are otherwise only slightly soluble. Second, humic substances 

occurring as components of sediment particles can 

serve, through partitioning interactions, as vehicles for 

the movement of anthropogenic organic compounds. 

Sorptive interactions are the most common process by which humic substances affect the mobility of 

organic contaminants. Sorption to settling particulates 

removes organic compounds from the water column 

and deposits them in sediment reservoirs that can be 

overlain by more recent sediment or resuspended by 

turbulent eddies. Available data indicate that the primary mechanisms for the sorption of cationic species 

are associations by hydrogen bonding and ion exchange, 

whereas the sorption of nonionic and anionic organic 

contaminants is mainly by partitioning into the organic 

matter associated with surficial sediment or suspended 

particulates. In addition, even though fulvic and humic 

acids are more water soluble than the less polar humins, 

they contain enough nonpolar moieties that uptake 

relatively water-insoluble organic contaminants by a 

partitionlike interaction (Chiou and others, 1986). 

The distribution and interaction of the nonionic 

pesticides, especially the chlorinated insecticides, have 

been investigated in much more detail than their ionic 

counterparts. The chlorinated hydrocarbons do not 

dissociate as readily in solution as do the ionic pesticides, and this contributes to their tendency to sorb into 
organic matter by partition (Chiou and others, 1979, 

1983, 1984). Pierce and others (1974) found that much 

of the sorptive activity of dichloro-diphenyltrichloroethane (DDT) could be closely correlated with 

the organic content of the soil or sediment. Based on 

the observed correlation of DDT uptake with the 

organic matter content of the soil and the low heats of 

DDT sorption (ranging from -4 to +2 kcal/mol (kilocalories per mole)), the results can be effectively rationalized in terms of solute partition into the organic 

matter. Carter and Suffet (1982) studied the interaction 

of DDT with dissolved organic matter using a dialysis 

technique and found that 75 percent of the DDT in 

water was bound to a 16 mg!L (milligram per liter) 

solution of dissolved humic acid (that part passing 

through a 0.22-p.m filter), whereas the remainder was 

free in solution. Their laboratory results showed that 

the dissolved organic carbon in natural waters at concentrations greater than 16 mg!L can bind a substantial 

fraction of the total DDT. Thus the DDT that is 

normally analyzed as dissolved and free may in fact be 

bound to a significant extent to dissolved humic materials. In isotherm measurements, Carter and Suffet 

(1982) observed, as did Pierce and others (1974), that 

the partition coefficients for commercial Aklrich humic 

acid were commonly an order of magnitude or more 

higher than comparable values for natural dissolved 

organic matter. 

Landrum and others (1984), in a different experiment using specific resin cartridges instead of dialysis 

to separate bound and free "dissolved" organics, 

obtained similar order of magnitude differences 

between natural dissolved organic matter and humic 

acid. The work of Landrum and others (1984) showed 

that the behavior of several chlorinated biphenyls and 

PAHs was comparable to that of DDT. Landrum and 

others (1984) attributed the disparity in partition coefficients to differences in the chemical composition and 

state of aggregation of the various sorbent materials 

and to the presence of varying amounts of submicron 

particulate matter. In general, these results are in 

agreement with the data of R.L. Malcolm (U.S. Geological Survey, personal commun., 1985), who observed 

distinct differences between the nuclear magnetic resonance spectra of commercial and natural humic and 

fulvic acids. 

Dissolved and colloidal organic matter are categorizations defined on the basis of size but that include 

many of the compounds also classified as humic substances. Colloidal matter falls approximately within the 

1-nm to 1-p.m range (Zsolnay, 1979), whereas dissolved 

matter has been operationally defined in the literature 

19 

as any compound that can pass through a 0.45-p.m 

filter. Humic substances that change state in reponse to 

pH and that flocculate form an important intermediary 

between dissolved and particulate matter. The behavior 

of organic contaminants, which has been attributed to 

the presence of natural organic colloids or dissolved 

organic matte~; is also characteristic of reactions involving humic substances. The similarities among natural 

organic colloids, dissolved organic matte~; and humic 

substances suggest a common compositional linkage 

among these groups and an integral role in bridging the 

hierarchy of chemical structure in going from particulate to truly dissolved organic matter. 

Dissolved humic substances can stabilize certain 

hydrophobic organics and cause their apparent solubility to be greater than their actual solubility. Ogner and 

Schnitzer (1970) and Matsuda and Schnitzer (1971) 

investigated this phenomenon for dialkyl phthalates 

(used as plasticizers and lubricants) and tried to quantify the degree of solubilization in the presence of fulvic 

acid. Both pairs of investigators demonstrated a stoichiometric relation between the amount of fulvic acid 

and the solubility of several dialkyl phthalates. Boehm 

and Quinn (1973), also using extracted fulvic acid (3.7 

mg!L of carbon), found an increase by a factor of 2-5 

in the solubilization of some higher n-alkanes and 

isoprenoids (representative of petroleum hydrocarbons) but no observable effect on some more watersoluble aromatic hydrocarbons. Infrared spectrophotometry of the solubilized compounds revealed no new 

functional groups indicative of a chemical reaction 

between the fulvic acid and the phthalate molecules. 

This led Matsuda and Schnitzer (1971) to hypothesize 

that the enhanced solubility resulted from surface 

adsorption of the substituted phthalates on the fulvic 

acid, forming water-soluble complexes. 

Poirrier and others (1972) investigated the sorption of DDT by color-causing colloids in an acid stream. 

These colloids were filtered to fall within the 5- to 

10-nm range and proved to be iron-organic complexes 

with an organic com position of 68-78 percent fulvic 

acid, 3-9 percent humic acid, and 16-28 percent 

hymatomelanic acid. The colloids were found to concentrate the DDT to levels 15,800 times greater than 

those found in solution. Poirrier and others (1972) 

attributed the affinity of DDT for the colloids to their 

small size, humic composition, and stability in solution. 

Wershaw and others (1969) found that the solubility of DDT in 0.5 percent sodium humate solution 

was at least 200 times the corresponding value in water. 

Hassett and Anderson (1979) reported that the highmolecular-weight fraction of dissolved organic matter 
was responsible for stabilizing and holding cholesterol 

and a tetrachlorobiphenyl isomer in solution. Both 

investigators explain the observations as the combined 

results of the surfactant action of the dissolved organic 

matter and a sorptive uptake of the organic contaminants. Wershaw and others (1969) demonstrated that 

sodium humate lowered the surface tension of water as 

its concentration increased and thereby enhanced solubility. Hassett and Anderson (1979) concluded that the 

increased affinity of organic contaminants for water 

resulted from their association with the hydrophobic 

sites (alkyl chains) on the dissolved organic matter 

molecules. Simultaneously, the hydrophilic sites on the 

dissolved organic matter keep the entire complex in 

solution or colloidal dispersion through their interactions with water. 

Boehm and Quinn (1973) speculate that the 

organic substances in colloids behave as micelles whose 

hydrophobic core attracts the contaminants while its 

hydrophilic moieties ionize to keep them in apparent 

solution. Howevet; direct confirmation is still lacking on 

the micellar formation of organic substances in colloids. 

Chiou and others (1985, 1986) consider the highmolecular-weight organic substances in colloids to 

behave as a microscopic organic phase to which a 

highly water-insoluble organic solute can readily partition. Such a partition effect does not necessarily require 

a micelle formation, as the molecular environment of a 

macromolecular substance would provide an organic 

medium (although reduced compared to a bulk phase) 

for solute partitioning. The significance of this interaction is that part of the highly water-insoluble organic 

contaminants may be held by dissolved organic matter 

in a manner similar to that sorbed to the highmolecular-weight fraction of natural organic matter 

(Hassett and Anderson, 1979; Wijayaratne and Means, 

1984). Undoubtedly, the ability of the humic substances 

to solubilize hydrophobic compounds has a significant 

impact on their transport and fate. 

Recently, Chiou and others (1986) studied the 

solubility enhancement of some nonionic organic contaminants by dissolved humic and fulvic acids. They 

showed that the extent of solubility enhancement is 

linear with humic and fulvic acid concentration and is 

inversely related to solute solubility and proportional to 

the octanol-water partition coefficient. The results also 

showed no competitive effect between solutes, supporting the hypothesis that a partitionlike interaction occurs 

between solutes and dissolved organic matter. This 

observation is in agreement with other data showing the 

sorption of nonionic organics by soil/sediment organic 

matter (Chiou and others, 1979, 1983). 

20 

Chiou and others (1986) found that the magnitude of solubility enhancement caused by dissolved 

organic matter is related to its polarity, size, and 

molecular configuration. The dissolved organic matter 

must be sufficiently large and structurally flexible. 

enough so that localized nonpolar regions of the molecules can partition the organic contaminants. This 

accounts for the greater solubility effect of soil humic 

acid, which is larger and less polat; than that of soil 

fulvic acid. In addition, the role of size and polarity 

support the observation that small organic acids such 

as phenylacetic acid exhibit little effect on enhancing 

solute solubility. Moreover, polyacrylic acid-a polar, 

linear polyelectrolyte similar in composition to fulvic 

acid-also failed to show an enhancement effect owing 

to its polarity and unfavorable molecular configuration. 

The partition hypothesis explains the observation that 

relatively water-soluble organic solutes (lindane and 

1,2,3-trichlorobenzene) show little solubility enhancement at the specified humic and fulvic acid concentration ( < 100 mg!L) where highly water-insoluble solutes 

(DDT and PCBs) exhibit large enhancements of their 

solubilities. This hypothesis is also in keeping with the 

findings of disparate solubility enhancement effects 

found between aliphatic and aromatic organic substances and reported by Boehm and Quinn (1973). It 

thus appears that the dissolved organic matter or the 

organic coating of colloids can effectively function as a 

"microscopic organic environment" for sorbing 

nonionic organic solutes by partition, thereby giving 

concentrations far in excess of those expected in pure 

water. 

HUMIC SUBSTANCES AND IONIC 

ORGANIC COMPOUNDS 

The extent and mechanism of sorption of organic 

pesticides depends on the type of chemical compound 

and the type of sorbent available for sorption. The ionic 

pesticides (for example, phenoxy herbicides and triazine 

herbicides) readily dissociate in solution, and therefore, 

their speciation is more responsive to solution variables 

such as pH and ionic strength, which may affect the 

sorption of the dissociated ions by mechanisms other 

than partition. The nonionic pesticides do not ionize 

significantly, and consequently, their sorptive behavior 

in aqueous systems depends on their ability to partition 

and on the polarity and composition of the organic 

matter in the water and sediment columns. 

Khan (1974a, b) investigated the adsorption of 

two cationic bipyridylium herbicides, diquat and paraquat, by different cation saturated humic acids. The 
choice of cation-saturated humic acid reflects the belief 

that free and dissociated humic acids rarely occur in 

natural waters because of the availability of inorganic 

ions. fur ionic compounds such as paraquat and diquat 

that have high water solubilities, the sorption mechanism is identified as primarily due to ion exchange with 

some contribution from charge transfer processes. The 

extent of the sorption, ranging from 20 to 85 percent of 

the added herbicide, depends strongly on the strength of 

retention of the cations by the humic acids and consequently on the stability of the humic acid-inorganic ion 

complex. The higher valence cations are more firmly 

retained by the humic acid and thus result in smaller 

exchange capacities for the herbicides. Within a group 

of cations of the same valence, differences in adsorption 

were influenced by the size, shape, and nature of the 

bipyridylium cation and the steric hindrances generated 

by the various sizes of cations competing to associate 

with the humic acid complex. 

Basic herbicides tend to undergo significantly 

greater sorption with humic substances at pHs in the 

vicinity of their dissociation constants (p~ values), 

presumably by an ion-exchange process (Khan and 

Schnitze:r; 1972). At pHs much higher than their p~, 

the high water solubility of the herbicides reduces the 

partition effect with humic substances. At pHs lower 

than or around their p~, the basic herbicides (for 

example, organonitrogen triazines) readily become protonated to form positive counterions capable of adsorbing to negative sites of the humic substances through 

complexation or ion exchange. However, at pH much 

lower than the p~, the resulting high hydrophilicity of 

the cations, in combination with the increased competition of hydrogen ions for protonation of the functional 

groups of humic acids, retard the ion-exchange process. This counteracting effect makes the efficiency of 

cation exchange a maximum at pH around the p~ of 

the ionic compound. 

Khan (1974a) examined the sorption of the 

anionic chlorinated phenoxy herbicide, 2,4-D, on a 

fulvic acid-montmorillonite sorbent. The 2,4-D sorption on the organoclay complex was smaller than that 

observed in comparable tests using pure fulvic acid but 

larger than in tests using a pure montmorillonite 

sorbent. The isotherm data gave low values for heats of 

sorption in keeping with the partition hypothesis (Chiou 

and others, 1979, 1983). Miller and Faust (1972), in 

similar studies with 2,4-D and organoclay complexes, 

found that 2,4-D uptake increased significantly when 

the pH was lowered. This agrees with the expectation 

that the anion of an acidic molecule would not be 

strongly sorbed because of its high hydrophilicity at 

21 

high pH (which reduces partition) and because of its 

inability to ion exchange with acidic humic materials. 

At low pH, the increased concentration of the neutral 

form of the acid leads to increased partitioning because 

of its low solubility in water. 

In related studies, Means and Wijayaratne (1982) 

and Wijayaratne and Means (1984) examined the 

behavior of atrazine and linuron in the presence of 

estuarine colloids. They found that sorption of the 

contaminants by the colloids affected the mobility of the 

compounds and contributed directly to their degradation by chemical and microbial processes. The composition of the colloids is strongly dependent on pH and 

salinity (Means and Wijayaratne, 1982) but can be 

generally characterized as a high-molecular-weight 

organic substance that is associated with varying quantities of poorly crystallized clay minerals and trace 

metals (Sigleo and others, 1980, 1982; Sigleo and Helz, 

1981). 

Humic substances have a demonstrable influence 

on the rate of chemical transformation of organic 

contaminants (Perdue, 1983a, b). fur the specific case 

of hydrolysis, transformation rates can increase or 

decrease depending on the type of organic pesticide 

(nonionic, cationic, or anionic) and the pH of the 

aqueous medium. Perdue (1983a, b) observed that for 

the 1-octyl ester of 2,4-D (acidic herbicide), the rate of 

alkaline hydrolysis is greatly decreased in humic acid 

solutions. He states that a portion of the 2,4-D is 

partitioned into the humic acid, and this form is resistant to further hydrolysis. Li and Felbeck (1972) monitored the acid hydrolysis of atrazine (basic herbicide) 

and found that the presence of humic acid strongly 

enhanced the reaction. They hypothesized that the 

increased reaction rate resulted from a combination of 

acid catalysis within the aqueous phase and a surface 

catalysis of atrazine molecules bound to the exterior of 

the humic acid molecules. Khan (1978) demonstrated a 

similar effect for atrazine hydrolysis in a fulvic acid 

solution. 

Although these and other studies show that humic 

substances have a measurable influence on the acidbase hydrolysis of certain organic compounds, the 

exact mechanism (taking into account aqueous solution 

parameters) is still only circumstantially defined. What 

is known is that humic substances, through sorptive 

processes, can significantly alter the reactivity of an 

organic contaminant from that of the unbound compound. The effect of these interactions on the reaction 

kinetics of organic contaminants is significant, and it 

increases directly with the concentration of humic 

substances. 
FIELD STUDIES 

The interactions of nonionic organics with different size sediment fractions affect their transport characteristics to a large extent. If the patterns of association between organic compounds and sediment can be 

better understood, much of the existing information on 

sediment transport can be applied to the movement of 

contaminants. The way that contaminants move provides answers to questions concerning the residence 

time of compounds in a given surface-water system and 

the potential they have for affecting other downstream 

reaches. The preceding part of this section demonstrated how the humic substances affect the water 

chemistry of anthropogenic organic compounds. This 

part reviews how the relations between contaminants 

and different sediment fractions determine their patterns of movement and deposition. 

In the literature reviewed, most of the field 

investigators studying surficial sediment have concentrated their analytical efforts on persistent organic 

compounds such as the chlorinated insecticides and the 

PCBs. It is only since the late 1970s that interest in 

sample analysis has shifted to include the 

organophosphorus insecticides and other herbicides 

(Gilliom and others, 1985). This trend reflects the 

decreasing agricultural use of chlorinated insecticides 

during the last decade and the increasing application of 

more efficient and less environmentally persistent insecticides and herbicides. Another feature of the field 

studies is their emphasis on documenting local concentration levels and then relating (and sometimes extrapolating) those values to the global extent of contamination within a lake basin, riverbed, or estuary. Secondary 

efforts are directed at explaining the observed spatial 

pattern of contamination in terms of usage trends and 

readily measured water-body characteristics. 

In a series of comprehensive studies of the Great 

Lakes, Frank and others (1977, 1979, 1980a, b, 1981) 

collected more than 1,500 surficial (top 3 em) sediment 

samples that revealed an almost ubiquitous distribution 

of PCBs and chlorinated insecticides. The data demonstrate the retentive ability of the sediment, which typically record concentrations (for PCBs and DDTs) in the 

10-300 ppb (parts per billion) range while watercolumn levels are only in the 1-10 ppt (parts per 

trillion) range. Other studies of the Great Lakes (Eisenreich and others, 1979, 1980; Simmons and others, 

1980) and of different surface-water bodies (Pavlou 

and Dexter, 1979; Pavlou and others, 1980; Bopp and 

others, 1981, 1982; Elder and Mattraw, 1984) corroborate these results and show that the concentration 

22 

levels measured by Frank and others (1977, 1979, 

1980a, b, 1981) are not specific to the Midwest. 

In addition to recording the concentration patterns of contaminants, efforts have been made to relate 

these observations to other possible causative processes 

such as sediment transport. Through comparison of 

concentration contours with sediment bathymetry contours, Frank and others (1979, 1980a, b) found that 

contaminant concentrations were 2-3 times higher in 

depositional areas (fine-grained sediments ranging from 

silt to clay) than in nondepositional areas (bedrock, tills, 

and glaciolacustrine clays). Eisenreich and others 

(1979, 1980) in studying Lake Superior and Pavlou and 

Dexter (1979) and Pavlou and others (1980) in studying the Puget Sound found similar trends, implying that 

contaminants are sorbed and transported by the finergrained sediment. 

Pavlou and others (1980), in examining the PCB 

concentrations in a dredged spoil deposition area, found 

that the spatial distribution was fairly heterogeneous 

but closely approximated the fluid flow patterns and 

area bathymetry. Eisenreich and others (1979) discovered the highest DDT concentrations in geologic basins 

where sediment was higher in clay and organic carbon 

content. Bopp and others (1981, 1982) reported that 

PCB movements in the Hudson River are seasonal 

because 80 percent of the sediment is transported 

during spring runoff and PCBs are associated with this 

sediment. Frank and others (1979) and Simmons and 

others (1980) further tested the uptake of contaminants by fine-grained sediment through correlation 

studies between concentration values, percent organic 

carbon content, and percent silt-clay fraction. The statistical results demonstrate a significant relation between 

all combinations of the three variables for two separate 

data sets with approximately 250 samples per set. 

Commonly, sediment cores are taken to complement the surficial sediment samples. Sediment cores' 

encapsulate the history of contaminant accumulation 

within the vertical section of sediment deposition. Core 

samples explicitly demonstrate that contaminants such 

as DDTs and PCBs have persisted for more than 15-25 

years since their widespread industrial and agricultural 

use began (Eisenreich and others, 1979; Frank and 

others, 1979). Cores also contain independent indicators, such as natural and fission-produced radionuclides, that are used to estimate sedimentation rates. 

Knowledge of sedimentation rates along with industrial 

use patterns allows the calculation of basic information 

concerning the response time of the natural system to a 

given contaminant input. Bopp and others (1982) have 

used these techniques to explain the distribution of 
PCBs along the vertical and longitudinal axes of the 

Hudson River. They report that areas of fairly rapid 

sediment accumulation can be used to reconstruct the 

contaminant input and distribution history of persistent 

compounds. 

As a result of the observed association of contaminants with fine-grained sediment, several field studies have attempted to determine which water-column 

or sediment component carries the largest contaminant 

load. Kuntz and Warry (1983) analyzed suspended 

sediment and water-column samples for 19 organochlorine compounds on a weekly basis for 3 years. They 

found that although concentrations in suspended sediment ( > 5 #Lm) were typically higher than water-column 

concentrations, the suspended sediment was responsible 

for only about 40 percent of the total contaminant load. 

Therefore, sediment concentrations arise from contaminant diffusion out of the water column as well as 

through deposition of suspended sediment. Miles 

(1976) compared whole-water samples, suspended sediment, bedload sediment, and bottom sediment. fur the 

five organochlorine compounds compared, he found 

that the extent of uptake by sediment material increased 

in a linear but inverse relation to the solubility of the 

compounds. Additionally, in relating the contaminant 

concentrations to their transport mechanisms, he found 

that the concentration values were highest for the 

suspended sediment and bedload, whereas the bottom 

material had intermediate values and the water column 

had the lowest values. 

To get a more detailed picture of which sediment 

components, if any, are primarily responsible for the 

observed contaminant levels, several field investigators 

have size-fractionated the sediment prior to analysis. 

This multiplies the number of samples needed for 

complete analysis and enlarges the sample sizes of 

whole sediment necessary so that an analyzable quantity 

of the smaller size fractions can be obtained by wet 

sieving, sedimentation, or centrifugation. Choi and Chen 

(1976) examined the distribution of chk>rinated hydrocarbons (DDT, dieldrin, and PCB) and organic carbon 

(total organic carbon, fulvic acid, and humic acid) in 

five size fractions encompassing fine silt and clay. Their 

results show that, for different sediment samples, the 

distribution of the organic carbon among the five 

fractions is roughly equivalent, but the sediment samples containing a higher percentage of fine particles 

also have a higher percentage of organic carbon. In 

addition, Choi and Chen demonstrate a linear relation 

within the five size fractions between the concentration 

of chlorinated hydrocarbons and the organic carbon 

content. These results indicate that the organic content 

23 

of the fine-particle sizes in the sediment is important in 

contributing to contaminant uptake by partition, especially if the organic carbon content is relatively high 

( > 1 percent). 

Goerlitz and Law (1974) examined six sediment 

samples for total chk>rinated hydrocarbons in four size 

fractions (gravel, sand, silt, and clay). In this study the 

linear relation between chlorinated hydrocarbon concentration and sediment particle size was not as clearly 

defined. The trend showing highest concentration in the 

smallest particle fractions was commonly disrupted in 

the six sediment samples. The trend of chlorinated 

hydrocarbon mass with particle-size distribution was 

even more variable. In most cases, deviations were 

generally attributable to contamination of the predominantly mineral fractions in each size range by organic 

sorbents such as clam shells, plant seeds, and other 

organic detritus. This investigation demonstrates that 

all particle-size fractions must be considered if a representative bed material sample is to be collected and 

quantitatively related to stream transport properties. 

LABORATORY STUDIES 

The fractionation of sediment into components 

can be done on the basis of the size or type of sorbent. 

Size fractionation (gravel, sand, silt, and clay) relates 

more to the hydrodynamic properties associated with 

particle-size ranges. Fractionation on the basis of sorbent type (organic, mineral, and organomineral) 

depends on the composition of the particles associated 

with a particular size class. The classifications provide 

order and understanding to the vast continuum of 

particles in the water column and surficial sediment. 

The classifications are not unique or mutually exclusive 

because sorbent types that are grouped purely on a 

compositional basis generally also fall within a corresponding size range (for example, phytoplankton). It is 

important to recognize that the demarcations between 

size and type as well as the finer ones within each class 

are operational and a fair degree of overlap prevails. In 

addition, some particles are not clearly of any size or 

type because their dimensions vary (for example, flocculated masses), and they may consist of various biological and mineral fragments cemented together into 

one mass (for example, detritus). Examples include 

sorbents like organic aggregates and mineral grains 

with surface coatings that are under continuous and 

dynamic modification in response to variations in the 

aqueous environment. 

The laboratory studies seek to better define the 

discernible patterns between contaminants and 
sediment observed in the fiekl. Accordingly, the laboratory studies use model (natural sediment) and sometimes ideal (pure clay) systems to focus on sorption as 

the dominant sediment-contaminant interaction. 

Although column flow-through experiments have been 

thought of as being more representative of the vertical 

movement of contaminants in soils, the majority of 

sediment-related experiments use batch equilibrations. 

Stirred tank reactor experiments are more representative of particle-contaminant mixing and interaction in 

the water column and surficial sediment (Pionke and 

Chesters, 1973; Presley and Trefry, 1980). Sorption 

and desorption isotherms are derived from batch experiments by interpolation between values measured at a 

number of contaminant -sediment and contaminantwater concentrations. These uptake and release isotherms are taken as a quantitative representation of 

actual contaminant -sediment interactions in a wide 

range of surface-water environments. 

In addition to the. determination of partition 

coefficients, batch experiments seek to isolate those 

sediment components and contaminant properties that 

control sorptive behavior. To this end, tests on different 

sediment-size fractions (gravel, sand, silt, and clay) and 

different sediment sorbent types (mineral, organic, 

organomineral, and detrital) seek to ascertain the 

relative contribution of each component compared to 

tests on the bulk sediment. The objective in examining 

the behavior of contaminants with sediment components is to identify the major factors affecting sorption 

from the large set of available contaminant-sediment 

variables. Ideally, the partition coefficient woukl be 

estimated from a two-factor set: one factor descriptive 

of the sediment and the other of the sorptive tendencies 

of the contaminant. 

The isotherm studies based on sorbent type use 

field-identifiable classes of sorbents and subject them to 

various concentrations of contaminants in batchequilibration experiments. The classes of sorbents include bacteria, phytoplankton, zooplankton, suspended 

sediment, natural surficial sediment, natural clay, and 

sand. Commonly, this list is supplemented with other 

"ideal" sorbents such as kaolinite, illite, and montmorillonite or natural sorbents that have been treated to 

remove various components such as organic carbon, 

lipid fractions, or humic fractions. Ideal or treated 

sorbents are compared to natural sorbents for two 

reasons. First, they demonstrate the effect of removal 

of a particular component on the overall sorption 

capacity. Second, they make it possible to interpret the 

observed sorptive behavior in terms of more fundamental characteristics of the sorbent as it relates to the 

sorption mechanism. 

The studies reviewed employ a number of sitespecific sorbent types as well as a large variety of 

organic chemicals. On the basis of these investigations, 

a few generalizations can be made. They should be 

interpreted with the understanding that sorbent groups 

are general classifications that might have significant 

compositional variations from one study site to another. 

The effect of organic matter content on the 

sorption of pesticides by soil has been well documented 

(Hamaker and Thompson, 1972; Hamaker, 1975). A 

' similar dependence wouk.l be expected for particles in 

the water column and sediment because of their variable organic matter content. Table 9 demonstrates for 

PCBs that the magnitude of the partition coefficient 

increases with the organic carbon content of the various 

sorbent types (Nau-Ritter and others, 1982; Nau-Ritter 

and Wurster, 1983). This relation also appears to 

characterize other organochlorine compounds, 

although it has not been documented as extensively as 

in the case of PCB (Hiraizumi and others, 1979; Weber 

and others, 1983). Additional sorbent characteristics 

found to correlate with sorbent uptake concentrations 

include surface area, particle size, mineral content, and 

cation-exchange capacity, but none of these properties 

appears as consistently significant as organic carbon 

content. 

24 

The most common technique for investigating 

how bulk sediment components interact with contaminants is to divide the sediment into fractions on the 

basis of size. Nine such studies were identified from the 

literature, and they encompass laboratory experiments 

conducted with more than a dozen chemical compounds and many different size ranges of the sediment 

fractions. In general, the results indicate that organic 

compounds sorb preferentially to the fine-sediment (silt 

and clay) fraction because of its typically high organic 

carbon content (Morris and Calvert, 1975; Karickhoff 

and Brown, 1978; Karickhoff and others, 1979). This 

observation, which corroborates the field results, is 

consistent with the partitioning of nonionic organic 

solutes from water into the organic matter of the 

soil/sediment components (Chiou and others, 1979, 

1983) and is in general agreement with the trend of 

increasing organic carbon content with decreasing particle size (Wildish and others, 1980; Nkedi-Kizza and 

others, 1983). 

fur those studies that fractionated the fine sediment into size classes, measured partition coefficients 

were essentially homogeneous across groups except for 

the distinct difference of the sand-sized particles 
Table 9.-Partition coefficient ranges of polychlorinated 

biphenyls and various sorbent types 

Substrate type Log partition 

coefficient References 

Live and dead bacteria 5-6 1,5,8 

Live and dead algae 3-6 1,2,8 

Zooplankton 4-5 2 

Susrhended sediment 3-6 1,2,3,8 

Su 1cial sediment 2-5 1 ,2,3,4,6,8, 9 

Organic purged sediment 3-5 1,4,8 

Natural clay 3-4 1,8 

Soil 2-4 2,7 

Montmorillonite, kaolinite, 2-4 1,3,4,6, 7 ,8,9 

illite 

Sand <1-2 2 

Sources: 1 

2 

3 

4 

5 

6 

7 

8 

9 

Voice and others (1983) 

Hiraizumi and others ( 1979) 

Horzempa and Di Toro (1983a, b) 

Weber and others (1983) 

Steen and Karickhoff (1981) 

Di Toro and Horzempa (1982) 

O'Connor and Connolly (1980) 

Voice and Weber (1983) 

Di Toro and others (1982) 

(Karickhoff and Brown, 1978). This distinction in 

behavior between sand and other particle groups was 

still apparent even after the partition coefficients were 

normalized to account for differences in organic carbon content (Nkedi-Kizza and others, 1983; Isaacson 

and Frink, 1984). Choi and Chen (1976) found only 

small differences in the magnitude of partition coefficients when they looked at different subdivisions of the 

clay-sized particles in a number of sediment samples. 

This sorptive uniformity in all but the sand-sized fraction led Frank and others (1979, 1980a, b, 1981) to 

correct all their Great Lakes data for the difference in 

sand content of the different surficial sediment samples. 

The rationale used in this approach is that sand has a 

small retentive capacity for contaminants and therefore 

dilutes the other smaller sediment fractions. 

In contrast to the aforementioned studies, Goerlitz and Law (1974), in assessing sorption, concluded 

that all particle-size fractions are necessary to obtain a 

representative sample. They found that the general 

trend of higher concentration with smaller grain size 

was not always applicable. They concluded that the 

sorptive capabilities of the larger fractions are not 

negligible because of the presence of organic sorbents 

that are sand sized and have significant sorptive capacities. In addition, although not specifically mentioned by 

Goerlitz and Law, sand-sized particles are commonly 

enveloped with metal oxides and (or) organic coatings 

that can alter the sorptive behavior of pure quartz 

grains. Schwarzenbach and Westall (1981) have shown 

that repeated washing of sediment particles (as is 

25 

typical in wet sieving procedures) removes particle 

sheaths and, for more representative isotherm results, 

excessive washing and pretreatment should be avoided. 

The studies on sediment fractions also showed 

that, in many samples, the peaks of contaminant mass 

and concentration commonly fell within different size 

groups (Richardson and Epstein, 1971; Goerlitz and 

Law, 1974; Karickhoff and Brown, 1978). The peak of 

the contaminant-mass distribution depends on the 

particle-size distribution, whereas the peak of concentration reflects the distribution of organic carbon content. Studies comparing samples from different streambeds (Karickhoff and Brown, 1978; Hassett and 

Anderson, 1979; Hassett, 1980), showed that, even 

though the actual concentration in a given size fraction 

varied widely across different sediments, the distribution of contaminant concentration was similar among 

the size fractions of most sediment samples. 

Several investigations have supplemented their 

batch equilibrations with similar tests using organiccarbon-purged sediment. These studies indicate that, 

although organic carbon dominates sorption, it is not 

the sole factor. Schwarzenbach and Westall (1981) 

found that in the absence of organic carbon the type of 

mineral surface has a significant effect on the degree of 

sorption. Richardson and Epstein (1971) and Choi and 

Chen (1976) found that for silt and sand the organic 

carbon controls sorption, whereas for clay, mineral 

composition is also important. Moreover, the greater 

the solubility of the organic contaminant, the smaller 

the influence of organic carbon and consequently the 

more uniform the sorbed concentrations across sediment fractions. In this case, the sorbed concentrations 

would probably be low. 

Isaacson and Frink (1984) found for different 

substituted phenols that sorption is initially affected 

largely by the aqueous solubilities of the contaminants. 

Howevet; as the contaminants move into the bulk of the 

sediment organic matter, the environment becomes 

progressively less aqueous and more strongly affected 

by the nature of the organic matter associated with that 

sediment fraction. From isotherm studies on purged 

sediment fractions, they found that the fine fractions 

had a higher content of labile organic matter-that is, a 

higher percentage of removal by peroxide treatment 

compared to the refractory organic compounds associated with the coarse fractions. Consequently, the 

observed differences in sorption between fractions may 

be attributable to the kind as well as the amount of 

organic matter (Morris and Calvert, 1975). 

Laboratory studies have also been used to examine the release of sorbed contaminants. fur many 
organic contaminants, uptake and release by the sediment is considered an equilibrium and reversible process. However, some laboratory experiments suggest that 

the release of slightly soluble organic compounds from 

the sediment (desorption) is not complete. These studies show that the sorption and desorption isotherms 

generated from the contaminant uptake and release 

data do not always coincide (Di Toro and Horzempa, 

1982; Di Toro and others, 1982). The deviation between 

the uptake and release data is attributed to different 

factors by different investigators. Some researchers 

believe that the difference is a laboratory artifact which 

results from the presence of sediment colloids that are 

not readily separable from the dissolved phase (Horzempa and Di Toro, 1983b; Voice and others, 1983). 

These sediment colloids have an organic matter content 

that is capable of partitioning contaminants and thereby 

preventing a true separation of the dissolved and sorbed 

states of the contaminant in the isotherm measurement 

(Gschwend and Wu, 1985). Other investigators argue 

that a "true" equilibrium is not attained and attribute 

the desorption behavior to a kinetic effect (Karickhoff, 

1980, 1984; Jaffe, 1986). They believe that the desorption kinetics are better described by the combination of 

an initial rapid release of contaminant followed by a 

slowet; time-dependent removal. This process by which 

contaminants desorb from sediment is only partially 

understood and is an area of current research. A more 

detailed understanding of desorption from sediment 

will help provide better estimates of the residence time 

of contaminants in surface-water systems and, consequently, of their availability for interacting by other 

transformation processes. 

TRANSPORT OF SORBED MANMADE 

ORGANIC COMPOUNDS 

The previous section demonstrated that various 

sediment particles, because of their organic matter 

content, can sorb nonionic organic compounds through 

partitioning processes. This sorptive capability makes 

sediment particles into potential carriers for the hydrodynamic transport of organic contaminants. To understand better how contaminants exchange with the surficial sediment layer during their downstream movement, 

this section reviews some of the different physical 

forces controlling sediment particle transport. In addition, a discussion of flocculation is included as a physicochemical process important in the estuarine movement 

of colloids. 

26 

HYDRODYNAMICS 

The distribution of sorbed contaminants depends 

on the physical processes controlling sediment -particle 

transport. Sediment originates from the soil, but transport processes segregate the soil fractions continuously 

throughout their travel history. Consequently, sediment 

may resemble soil in mineral composition, while varying 

dramatically in particle-size distribution (Colby, 1963). 

Contaminants associate with sediment particles, and 

thus the physical processes segregating the soil into 

sediment fractions also determine the organic contaminant deposition patterns. A summary of the different 

mineral and organic matter particle sizes and the 

dynamic forces controlling their transport helps explain 

some of the particle-contaminant associations found in 

surface-water surficial sediment. 

Many physical processes act in conjunction with 

the downstream flow of water to move different sizes of 

particles found in the water column and in the sur~cial 

sediment. O'Melia (1980) and Hunt (1982) have Identified and investigated three major physical processes 

affecting particle transport and removal in the water 

column. First, gravity causes the sedimentation of particles on the basis of their size, shape, and density. The 

influence of gravity relative to other physical processes 

is greater in controlling the transport of silt and largersized particles. Second, and in conjunction with gravity, 

the fluid shearing forces generated by velocity differences between adjacent layers of moving water cause 

transport of silt- and clay-sized particles. Although 

these silt- and clay-sized particles settle vertically by 

gravity, they also are transported laterally and longitudinally by fluid-shearing forces that arise from the 

mixing of water currents of different densities and 

velocities. The third physical process identified by 

O'Melia (1980) and Hunt (1982) is molecular diffusion 

(also called Brownian diffusion) and is only of importance for particles of submicron dimensions. Diffusion 

is the apparently random motion of molecules caused 

by minute thermal gradients in the fluid mediui?. 

Transport by diffusion is independent of both the flUid 

motion and gravitational forces. Its influence is only 

noticeable for very small colloids where gravity and 

fluid-shearing forces exert a smaller relative effect on 

particle transport. The size of particles and the. ran~es 

over which gravity, fluid-shearing forces, and d1ffus10n 

exert the most influence on particle transport are 

depicted graphically in figure 1. 

In contrast to O'Melia's (1980) and Hunt's (1982) 

studies, the description of sediment transport has historically been approached by differentiating between 

the movement of coarse and fine particles. The 
movement of coarse mineral particles results from the 

combination of sedimentation, resuspension, and suspended and bedload transport. The description of the 

vertical settling of gravel, sand, and silt is approximated 

by Stokes' relation for ideal spherical bodies falling 

under quiescent conditions (table 10). The settling 

velocity depends on the balance between the upward 

buoyant forces and the downward gravitational forces 

for a particular particle geometry, particle density, and 

water density (Guy, 1969, 1970). fur larger particles 

(sand and gravel) and for less spherical particle shapes, 

the velocities deviate progressively more from those 

predicted on the basis of an ideal Stokes' settling 

behavior. These deviations increase in magnitude as the 

inertial effects become more significant as a result of 

surface and form drag forces. 

As the coarse particles settle, they are distributed 

longitudinally and laterally as suspended particulates 

caught in the advection of flow. Once settled, the 

particles may resuspend when currents are sufficiently 

strong to cause turbulent eddies capable of scouring the 

surficial sediment layer. Transport of settled particles 

occurs chiefly by rolling or sliding within the bedload 

layer-the uppermost layer of the sediment, commonly 

only a few grain diameters in thickness. Overall, the 

transport of coarse sediment depends primarily on the 

flow characteristics, the particle properties and dimensions, and the streambed geometry. 

Because the supply of coarse particles commonly 

exceeds the carrying capacity of a stream (Guy, 1969, 

1970), the concentration of coarse sediment in suspension is a function of the flow and sediment characteristics. fur fine sediment (clay- and colloid-sized particles) 

the suspended concentration depends less on flow characteristics and more on the available supply of particles 

from erosion and biological processes. fur fine sediment, unlike coarse sediment, the carrying capacity of 

the stream is typically much greater than the supply of 

particles from upstream and adjacent hill slopes (Colby, 

1963). 

The transport of small particles (micron and 

submicron sized) of fine sediment is less dependent on 

the influence of gravity than are larger particles of silt 

and sand dimensions. Oay- and colloid-sized particles 

are more buoyant and more responsive than are largersized particles to fluid shear and Brownian diffusive 

forces within the water column. Mineral colloids settle 

orders of magnitude more slowly than sand grains, and 

organic colloids are even slower as their densities 

approach that of water. These properties give fine 

particles longer residence times in the water column 

and make them more susceptible to the shearing forces 

27 

Table 10.- Theoretical settling velocities of particle-size c/ssses 

Particle 

groups 

Boulders 

Cobbles 

Very coarse 

gravel 

Coarse gravel Medium 

gravel 

Fine gravel 

Very fine 

gravel 

Very coarse 

sand 

Coarse sand 

Medium sand 

Fine sand 

Very fine sand 

Coarse silt 

Medium silt 

Fine silt 

Very fine silt 

Coarse clay 

Medium clay Fine clay 

Very fine clay 

Size In mm 

>256 

256-64 

64-32 

32-16 

16- 8 

8- 4 

4- 2 

2- 1 

Source: Guy (1969, 1970) 

Size In #LI11 

2,000 -1,000 

1,000-

500-

250-

125-

64-

32-

16-

8-

4-

2-

1-

0.5-

500 

250 

125 

64 

32 

16 

8 

4 

2 

1 

0.5 

0.25 

Settling 

velocity 

In stem 

<1- 3 

3- 13 

13- 48 

48- 180 

180- 700 

700- 2,800 

2,800 -11,000 

Note: Settling by Stokes law assumes spherical geometry 

for particles, specific gravity of 2.65, and a temperature of 20 . 

at the benthic boundary layer that resuspend sediment 

components. The combined effect of their transport 

characteristics and their natural abundance makes 

colloid-sized particles significant in terms of their potential impact on contaminant transport (Bilby and Likens, 

1979). 

Of all the particles found in the water column and 

sediment, the transport of organic particles, both dead 

and alive, and of organomineral assemblages has been 

the least accurately described in terms of particle 

dynamics. Discrete particles such as dead organisms 

vary in size, shape, and density, depending on their 

degree of fragmentation, decay, and dissolution 

(Lorenzen and others, 1983). The sizes range from that 

of the parent organisms to those below the 0.45-#Lm 

division between particulate and dissolved matter. Below 

this demarcation, the motion of particles is considered 

indistinguishable from the motion of water molecules 

(Sheldon and others, 1972; Lal, 1977). 

The movement of wing particles is further complicated by the various degrees of motility with which 

different species respond to environmental variables. 

Phytoplankton respond to changes in season, nutrient 

concentration, population age, time of day, and relative 

brightness of the sun. Bums and Rosa (1980) investigated 
the settling velocities of 10 different species of 

phytoplankton (size range of 1~4 #Lm) and found that 

values were as high as 0.8 m/day for flagellates and as 

high as 0.2-0.5 m/day for nonflagellates. These velocities are for ascending and descending movement in the 

water column because light -dependent organisms such 

as phytoplankton follow diurnal cycles of respiration 

and photosynthesis. This type of behavior is important 

within the euphotic zone of the water column because it 

is the primary source of natural particles. These particles also are the main food source for higher trophic 

organisms and form the basis of detrital matter. 

In an analysis of the settling behavior of 

organomineral aggregates in the 5- to 500-#Lm range, 

Chase (1979) found that measured sedimentation velocities were different from values predicted by Stokes' 

law. fur aggregates from lake and marine waters, the 

settling velocities were as much as an order of magnitude faster than the Stokes' law predictions. Edzwald 

and others (1974) found similar order of magnitude 

departures for the difference between discrete and 

flocculated minerals. The deviations from ideal behavior are attributable in part to approximating the actual 

shapes of aggregates by spheroids, but Chase (1979) 

has shown that three other factors are important. He 

found that natural surface coatings, the presence of 

solution electrolytes, and dissolved organic substances 

combine to increase the settling velocity of aggregates 

by reducing the skin friction of their outer layer. 

Although a more exact description of the transport of 

aggregates is hampered by their amorphous structure 

and composition, more evidence is accumulating on 

their integral role in the cycling of organic and mineral 

compounds, natural and anthropogenic, in the water 

column and sediment. 

Because the hydrodynamic behavior of colloidsized particles depends largely on the electrochemical 

balance within the aqueous solution (Stumm, 1977), 

surface coatings can be expected to affect the physical 

and chemical properties of colloid particles (for example, their sorptive capacities). By masking differences 

among mineral colloids, surface coatings affect the 

flocculation behavior of particles and may help to 

explain the lack of field evidence verifying laboratory 

studies that suggest differences in the flocculation 

behavior of different minerals (Gibbs, 1977). From a 

hydrodynamic standpoint, surface coatings reduce the 

frictional drag on settling particles and increase their 

downward velocities (Chase, 1979). All of these results 

indicate the effects surface coatings can have on particle properties and also raise questions on the accuracy 

of laboratory studies that purposefully wash or purge 

28 

particle coatings from suspended and surficial sediment 

(Schwarzenbach and Westall, 1981; Davis, 1982). 

FLOCCULATION 

Flocculation has been identified as an important 

removal mechanism for organic and inorganic compounds in at least 10 estuaries on 16 separate occasions 

(Boyle and others, 1977; fux, 1983a, b). Its primary 

effect is the aggregation of discrete particles (clay 

minerals, metals, and humic substances) into larger 

and denser masses, called floes, that settle 1-2 orders 

of magnitude faster than their component discrete 

particles (Edzwald and others, 1974). This p~ocess 

removes colloidal particles and reduces the residence 

time of other larger particles in the water column. In 

addition, as the flocculated masses settle, they remove 

organic contaminants through surface sorption reactions or by enmeshment within their aggregation 

matrices. 

Estuaries serve as natural coagulation tanks 

because of their salinity variations and the effects of 

tidal motion on mixing processes (Stumm, 1977). Tidal 

excursions provide downstream saline water with abundant electrolytes to interact with and destabilize colloids. 

Natural colloids (organic and inorganic) have a net 

negative surface charge and once in solution become 

surrounded by double layers of ions. The electrochemical balance is such that individual particles repel each 

other and remain stable in suspension until additional 

electrolytes neutralize the surface charges. After the 

repulsive energy barrier is removed, the tidal motion of 

the estuary increases the number and efficiency of 

collisions between discrete particles. Interparticle contact increases as a result of the Brownian motion of 

colloids (perikinetic coagulation), the effects of velocity 

gradients within the water column ( orthokinetic coagulation), and the different settling velocities caused by 

the various shapes and densities of particles (Edzwald 

and others, 1974). The tidally forced mixing provides 

the necessary electrolytes and enhances flocculation so 

that particles can rapidly coalesce into masses that 

settle. 

Flocculation is a complex process also involving 

electrostatic and chemical interactions (Eckert and 

Sholkovitz, 1976; Preston and Riley, 1982). Althou~ 

the exact mechanism is not yet fully understood, salinity has been ascertained to be the primary controlling 

factor, and pH is only of secondary importance (Sholkovitz, 1976; Boyle and others, 1977). The effect ofp~ 

is exerted more on the precipitation behavior of humic 

substances and the role they play in controlling the 
concentration of inorganic trace elements (Sholkovitz, 

1976; Saar and Weber, 1982). Sholkovitz and others 

(1978) have experimentally added various amounts of 

seawater to freshwater samples and demonstrated that 

the reaction kinetics of flocculation are rapid and are on 

the order of tens of minutes. Seemingly, no minimum 

activation salinity is required to initiate flocculation, and 

the process is considered to be continuous throughout 

the entire salinity range. However, maximum removals 

occur when salinities reach the 15-20 ppth (parts per 

thousand) levels (Sholkovitz and others, 1978). Dyer 

(1972) reported that, with adequate particle concentrations, illite and kaolinite are completely flocculated at 

salinities exceeding 4 ppth, whereas montmorillonite 

continues to flocculate until marine salinities are 

reached (35-37 ppth ). The stability of the flocculated 

mineral colloids depends on the type of clay mineral, 

the ionic strength, and the composition (valence of 

counterions) of the aqueous solution. 

As a removal process, the extent of flocculation 

varies for the clay minerals, metal ions, and humic 

materials. As was previously indicated, the three major 

types of clay colloids are rapidly flocculated if salinities 

are sufficient (Dyet; 1972). In estuaries, metal ions 

such as ·iron, manganese, and aluminum are almost 

completely removed and are commonly referred to as 

nonconservative substances (Sholkovitz, 1979). On the 

other hand, because only a small proportion (about 

3-11 percent) of dissolved organic matter, including 

humic substances, is removed, dissolved organic matter 

is considered a conservative substance (Moore and 

others, 1979). aoser examination shows a preferential 

removal of the higher-molecular-weight dissolved 

organic matter constituents such as the humic acids 

(60-80 percent removals), whereas lower-molecularweight dissolved organic matter constituents are 

removed in much smaller percentages (Sholkovitz, 

1976; fux, 1983b). 

Overall, flocculation has a demonstrable effect on 

varying the number of particles in suspension. In 

addition, studies show that particles of colloidal size are 

capable of partitioning organic contaminants to large 

concentrations. Consequently, it seems probable that 

flocculation also exerts a significant effect on the 

movement of organic compounds associated with colloidal particles (O'Melia, 1980; Hunt, 1982). 

ESTIMATION OF THE FIELD DISTRIBUTION OF 

MANMADE ORGANIC COMPOUNDS 

An understanding of the large number of isotherm studies utilizing many chemical compounds and 

29 

many types of sorbents requires a synthesis of the 

collected data within a general, yet simple, physicochemical framework. The first step is reduction of the 

complex contaminant-sorbent interactions of sorption 

to a set of controlling variables that can be related to 

measurable physical properties. Because sorption is a 

solute-sorbent interaction, the initial efforts involved 

identification of a two variable set, with one variable 

characteristic of the sorbent and the other characteristic of the organic contaminant. 

Isotherm studies conducted on soil-pesticide systems and more recently on sediment indicate that 

several sorbent properties affect the sorption capacity 

and mechanism. The soil and sediment investigations 

both point to the organic carbon content as the dominant controlling factor for nonionic and anionic organic 

pesticides. However, cation-exchange capacity, clay mineral composition, and particle-size distribution also 

contribute significantly to the sorption of basic pesticides. Studies correlating the organic carbon content 

and partition coefficients of sorbents, such as aquifer 

constituents, soil, sediment, and activated sludge, show 

a linear relation (Khan, 1972, 1978). This correlation 

holds for natural sorbents with an organic carbon 

content in excess of 0.1 percent and suggests that the 

organic matter composition and partitioning properties 

of the various sorbents are similar (Schwarzenbach and 

Westall, 1981 ). fur sorbents with an organic carbon 

content of less than 0.1 percent, the nature of the 

mineral surface has a greater effect on the degree of 

sorption. 

The strong correlation between organic carbon 

content and partition coefficient can be used to reduce 

the description of sorption by a multicomponent sorbent to that of a single hypothetical reference sorbent 

that is 100 percent organic carbon (Karickhoff, 1984). 

Mathematically, the normalization procedure results in 

a single variable-the organic carbon (matter) partition 

coefficient-that represents the field partition coefficient divided by the organic carbon content of either the 

bulk sediment or any one of its component fractions. 

Strong correlation between the organic carbon content 

and the partition coefficient has been found in several 

soil/sediment studies (Hamaker and Thompson, 1972; 

Chiou and others, 1979, 1983; Means and others, 1980, 

1982), and the normalization approach has been applied 

in sorption quantitation (Karickhoff and Brown, 1978). 

Expressing the sorptive behavior of a specific 

organic compound on the basis of the organic carbon 

content in soil/sediment also reduces the variability in 

partition coefficients across sediment types and within 

sediment fractions. Based on a set of 14 soil/sediment 
samples from a large multistate area, Hassett (1980) 

was able to reduce the twentyfokl to thirtyfokl variation 

in partition coefficient to a factor of 2 by normalizing 

the sorption on the basis of organic carbon content. 

Karickhoff and others (1979) and Nkedi-Kizza and 

others (1983) found a similar reduction in partition 

coefficient variability when looking at sediment fractions separated on the basis of size. After normalizing 

the partition coefficients, they were able to discern 

differences between the sorptive capacity of the sand 

and that of the other fractions. It shoukl be noted that 

these results are specific to the sorptive behavior of 

neutral organic compounds in water, where the mineral 

components are not significantly active in the sorption 

process and where dissociation, precipitation, and other 

chemical interactions are minimal. The relative inertness of the mineral fraction to neutral organic compounds in aqueous systems is attributed to the strong 

dipole interaction between water and the minerals 

(Chiou and others, 1983). In these types of partitioning 

processes, the role of organic matter is similar to that 

of a water-immiscible organic solvent capable of solubilizing organic compounds (Chiou and others, 1979, 

1983). 

Characterizing the sorptive differences between 

organic contaminants has centered on estimating the 

organic-carbon-normalized partition coefficient from 

either the water solubility of the compound or its 

octanol-water partition coefficient. The choice of these 

two chemical properties is based on the assumption 

that partitioning of an organic chemical between an 

organic phase and water is mechanistically similar to 

partitioning an organic chemical between itself and 

water (Chiou and others, 1977). Because organic solutes have better compatibilities with an organic phase 

than with water, the magnitudes of solute partition 

coefficients in a solvent -water system shoukl be inversely 

related to solute-water solubilities. Statistical correlation 

studies have verified that organic-carbon-normalized 

partition coefficients vary linearly with octanol-water 

partition coefficients and inversely with solubility (Chiou 

and others, 1977, 1981a, b; Mackay and others, 1980). 

These studies also show that, in general, the octanolwater partition coefficient is a more convenient basis for 

estimating the organic carbon partition coefficient than 

is water solubility. This occurs because for solid compounds the octanol-water partition coefficient is not 

sensitive to the melting-point effects associated with the 

solid solubility in water (Karickhoff and others, 1979; 

Chiou and others, 1983). Howevet; the melting-point 

effect on solid solubility can be readily corrected if data 

30 

on heat of fusion and melting points are available 

(Mackay and others, 1980; Chiou and others, 1981a, b). 

Many correlations seeking relations between physical properties of chemicals and their organic carbon 

partition coefficients are based on water solubility or the 

octanol-water partition coefficient as the correlating 

variable. Significant correlations have been found 

between octanol-water partition coefficients and aqueous solubilities for more than 50 organic compounds 

with solubilities spanning 8 orders of magnitude (1,000 

ppm (parts per million) to 1 ppb) and octanol-water 

partition coefficients spanning 6 orders of magnitude 

(1 to 1 million) (Chiou and others, 1977, 1979, 1981a, 

b, 1983; Campbell and others, 1983; Yalkowsky and 

others, 1983). More compounds and additional accuracy for the correlations are possible with a meltingpoint correction factor for solid organic compounds 

and by restricting the correlations to chemical families 

rather than to all organic compounds (Banerjee and 

others, 1980). fur highly water-soluble organic compounds, deviations from the linear relation between 

partition coefficient and water solubility may be expected 

because these compounds violate the dilute solution 

approximation assumed for both partition coefficient 

and water solubility (Chiou and others, 1981a, b). In 

general, the correlations and the regression equations 

describing the functional relations are applicable for low 

water solubility, nonionic organics that are not susceptible to speciation changes, or other types of complex 

formation (Brown and Flagg, 1981 ). Other attempts 

have centered on calculating the octanol-water partition 

coefficient by applying the molecular-fragment concept' 

(Leo and others, 1971) and by using activity coefficients of the solute in water (Arbuckle, 1983). 

The correlation studies using solute and sorbent 

properties measured in laboratory studies serve as the 

basis for estimating field partition behavior. With knowledge of the sediment organic carbon content and the 

contaminant octanol-water partition coefficients (or 

water solubility), derived regression equations have been 

tested in many studies of sediment contamination. 

Verification of the regression estimates has been possible to some extent through laboratory experiments 

and some field measurements. The groups and chemical families investigated include pesticides (Chiou and 

others, 1979), methylated and halogenated benzenes 

(Schwarzenbach and Westall, 1981; Chiou and others, 

1983), triazine and dinitroaniline herbicides (Brown 

and Flagg, 1981 ), PAHs and substituted PAHs 

(Karickhoff and others, 1979; Karickhoff, 1980, 1981; 

Hassett, 1980; Hassett and others, 1983), and some 

PCBs (Chiou and others, 1983). The accuracy of the 
estimates for the range of nonionic organic compounds 

and sorbents ( > 0.1 percent organic carbon) tested 

gives results within a factor of 2-3 of the measured 

values (Schwarzenbach and Westall, 1981). If more 

detailed information is available on the individual sediment fractions (organic carbon content and particlesize distribution), the bulk partition coefficient can be 

estimated from a weighted sum of the component 

coefficients (Karickhoff and others, 1979). Table 11 

presents a number of representative regression equations used in estimating field partition coefficients and 

bioconcentration factors from solubility and octanolwater partition coefficients. 

Recent investigations suggest that a third variable-the sorbent concentration-is required to estimate accurately the field partition coefficient of an 

organic contaminant from laboratory measurements 

(Horzempa and Di Toro, 1983b; Weber and others, 

1983). In addition to the sorbent organic carbon content, the concentration of the sorbent has been observed 

to influence the field partition coefficient. These observations indicate that the partition coefficient decreases 

as the concentration of solids increases (O'Connor and 

Connolly, 1980; Connolly and others, 1983). This phenomenon has been called the "solids effect." On a 

log-log scale spanning 4 orders of magnitude in solids 

concentration and partition coefficient, the inverse relation is linear for certain organic compounds and heavy 

metals sorbed on either soil or sediment (O'Connor 

and Connolly, 1980; Voice and others, 1983). The solids 

effect appears most pronounced for organic compounds with large partition coefficients or very low 

water solubilities. 

Observations from a diverse number of solutes 

and solid to water ratios indicate that the solids effect is 

not caused by specific chemical interaction but rather 

involves a nonspecific physical process (Voice and others, 1983). Horzempa and Di Toro (1983a) state that 

direct interactions between solid phase particles at 

different concentrations change the availability of binding sites and potentially the binding energies necessary 

for sorption. Independently, Voice and others (1983) 

and Gschwend and Wu (1985) have demonstrated the 

presence of nonsettling particles with submicron dimensions, whose concentration in the water column 

increases as the solids concentration of the bulk phase 

increases. These particles (which do not settle even 

under centrifugation) are capable of sorbing significant 

quantities of organic compounds and thereby enhancing the apparent solubilities of the organic solutes 

(Chiou and others, 1984; Gschwend and Wu, 1985). 

The extent of such effects increases directly with solids 

31 

concentration, leading to the apparent variation of the 

partition coefficient. Incorporation of the solids concentration as a third variable in the regression equation to 

estimate the field partition coefficient has been proposed as a method to account for the solids effect 

(O'Connor and Connolly, 1980; Voice and others, 1983). 

It should be noted, however, that the solids effect is 

significant only for highly water-insoluble organic solutes at sufficiently high concentrations of suspended 

solids or particulate-bound organic matter. A more 

thorough account of this effect in terms of dissolved 

organic matter has recently been given by Chiou and 

others (1986). 

The ability of sediment organic matter to partition nonionic organic compounds from the water column is paralleled by a similar capability in aquatic 

organisms. The aquatic biota contain lipid pools with 

solvent properties analogous to but more effective than 

the organic matter found in sediment. The effect of 

lipids in bioconcentration and the variables affecting the 

lipid-water partition coefficient have recently been 

accounted for by Chiou (1985). He found that on a unit 

weight basis, the lipid is far more effective in concentrating nonionic organic solutes than soil/sediment 

organic matter. Consequently, biological organisms represent another "dynamic reservoir" capable of accumulating, concentrating, and releasing organic contaminants. 

Neely and others (1974) and Chiou and others 

(1977) first demonstrated a significant correlation 

between the physicochemical property of a variety of 

chemicals ( octanol-water partition coefficient) and their 

ability to bioconcentrate in fish. Kenaga (1975b, 1980), 

Oliver and Nimi (1983), and Mackay (1982) also 

presented BCF values for a wide range of organic 

compounds and developed regression equations relating 

BCFs to water solubility and octanol-water partition 

coefficient (see table 11). Recently, Chiou (1985) specifically dealt with the effect of lipids in bioconcentration and found the lipid-based BCFs to be essentially 

the same as the triolene-water partition coefficients of a 

large number of environmentally significant organic 

compounds. In Chiou's study, triolene (glyceryl trioleate) was used as a model lipid compound because of its 

abundance and structural similarity to other triglycerides commonly found in aquatic organisms. Estimation 

of BCFs on the basis of lipid content appears to help 

converge and normalize the BCFs among fish species. 

Based on this analysis, the amount of a compound 

accumulated by a fish can be readily estimated with 

available data on the solute concentration in water and 
Table 11.-Regression equations used for estimating partition coefficients 

Regression equation 

1. Log Kow = -0.747 log S + 0.730 

Log Kow = -0.670 logS+ 5.00 

2. Log Kom = -0.557 log S + 4.040 

Log Koc = -0.54log S + 0.44 

Log Koc = -0.686log S + 4.273 

Log Koc = -0.594 logS - 0.197 

Log Koc = -0.83 logS - 0.93 - 0.01(MP-25)** 

3. Log Koc = 1.00 log Kow - 0.21 

Log Koc = 1.00 log Kow - 0.317 

Log Koc = 0.72 log Kow + 0.49 

Log Koc = 0.937 log Kow - 0.006 

Log Koc = 0.989 log Kow - 0.346 

Log Koc = 1.029 log Kow - 0.18 

4. Log BCF = -0.802log S- 0.479 

Log BCF = -0.508 log S + 3.41 

Log BCF = -0.873 logS + 0.803 

5. Log BCF = 0.542log Kow + 0.124 

Log BCF = 0.980 log Kow - 0.063 

Log BCF = 0.893 log Kow + 0.607 

Note: NA not applicable 

= hydrocarbons 

Number 

of com- pounds 

156 

33 

15 

10 

22 

5 

47 

10 

22 

13 

9 

5 

13 

7 

6 

6 

8 

5 

18 

= water solubility of the organic solute. 

R2* 

0.874 

0.970 

0.988 

0.940 

0.933 

0.945 

0.93 

1.00 

0.98 

0.95 

0.95 

0.997 

0.91 

0.95 

0.93 

0.93 

0.90 

0.98 

0.904 

= the octanol-water partition coefficient. Unitless. 

Units for Compound 

solubility groups 

Moles per Low-molecular 

liter wei~t-organics 

Micro moles Pesticides and 

per liter PCBs 

Micro moles Pesticides and 

per liter hydrocarbons 

Mole Chlorinated 

fraction hydrocarbons 

Micrograms PAHs 

per liter 

Mole PAHs 

fraction 

Mole Chlorinated 

fraction hydrocarbons 

NA Chlorinated 

NA 

hydrocarbons 

PAHs 

NA Substituted 

benzenes 

NA Herbicides 

NA PAHs 

NA Pesticides 

Moles per Rainbow trout 

liter chlorinated 

HCs 

Micro moles Rainbow trout 

per liter insecticides 

Moles per Guppies chloroliter benzenes 

NA Rainbow trout 

chlorinated 

HCs 

NA Guppies chlorobenzenes 

NA Rainbow trout 

chlorobenzenes 

References 

Hansch and 

others ( 1968) 

Chiou and 

others ( 1977) 

Chiou and 

others (1979) 

Karickhoff and 

others (1979) 

Means and 

others u980) 

Karickho 

(1981) 

Karickhoff 

(1984) 

Karickhoff and 

others (1979) 

Means and 

others (1980) 

Schwarzenbach 

and Westall 

(1981) 

Brown and 

Flagg (1981) 

Karickhoff 

(1981) 

Karickhoff 

(1984) 

Neely and 

others 

(197?:*** 

Chiou 1977) 

Chiou (1981) 

Neely and 

others (1974) 

Chiou (1981) 

Chiou ( 1985) 

HC 

s 

Kow 

Koc = measured partition coefficient/orsanic carbon content of soil or sediment sample. Unitless. Organic matter, 

depending on composition, contains about 58 percent organic carbon (Hamaker and Thompson, 1972) and 

* 

** 

*** 

consequently Koc = 1. 72 Kom. 

Kom = measured partition coefficient/organic matter content of soil or sediment sample. Unitless. 

BCF = bioconcentration factor. Unitless. 

The linear regression coefficient of determination. 

The last term in this equation compensates for melting-point effects which occur in the phase transition of organic 

compounds that are initially solids. 

Original data are from the source cited but the regression equation is found in Chiou (1981). 

32 
the fish lipid content along with the triolene-water (or 

octanol-water) partition coefficient of the compound. 

SUMMARY AND CONCLUSIONS 

The primary determinant of the behavior and 

fate of a contaminant in a surface-water system is its 

physicochemical properties. fur the persistent organic 

contaminants, low water solubility and high octanolwater partition coefficients are the two most common 

characteristics found among the different chemical 

classes. Therefore the primary transformation process 

contributing to the environmental persistence of 

nonionic organic contaminants is sorption. Nonionic 

organic compounds tend to partition from water into 

the organic matter associated with soil/sediment particles and into the lipid deposits of aquatic organisms 

(bioconcentration). A partition model is consistent with 

experimental observations of linear sorption isotherms 

by different sorbents up to solute concentrations 

approaching solubility, the absence of competitive effects 

in binary solute systems, and low exothermic heats of 

sorption. The environmental behavior of ionic organic 

compounds is more complex because sorption depends 

on system variables such as pH and ionic strength. 

Sorption of anionic organic compounds is likely to be 

in~uenced by a partitioning process (although usually 

qmte weak), whereas the sorption of cationic compounds is characterized by a surface adsorption model 

that can involve covalent bonding, electrostatic attraction, and ion-exchange processes. 

Several sediment fraction properties influence the 

transport and transformation of persistent manmade 

organic compounds. Biological organisms, by virtue of 

their lipid contents, bioconcentrate nonionic organic 

compounds. In addition, because of their presence in 

the water column and surficial sediment layer, aquatic 

organisms contribute to the movement and exchange of 

contaminants between these two surface-water compartments. Inorganic particles such as sand, silt, and 

clay possess small sorptive capacities for nonionic 

organic compounds relative to other organic sorbents. 

These inorganic particles are commonly coated with 

organic matter, which significantly enhances their ability to sorb and transport contaminants. Humic substances, in the form of dissolved and colloidal organic 

matter, partition organic contaminants into a microorganic environment and effectively increase the apparent solubility of highly water-insoluble compounds. The 

solubility enhancement is more pronounced as the 

water solubility of the contaminant decreases. This 

33 

phenomenon depends on organic sorbent properties 

such as polarity, molecular size, molecular configuration, and composition. The importance of the humic 

substances lies in their partitioning capability, alone or 

in conjunction with other sediment particles, for 

nonionic organic contaminants. This makes it possible 

for them to increase the carrying capacity of the water 

column beyond the solubility limits of the nonionic 

organic compound. 

The field and laboratory studies reviewed indicate that the organic carbon content of diverse sediment 

samples and fractions controls the sorption of nonionic 

organic compounds. Increased sediment organic carbon content correlates significantly with increased sorption. Similarly, a higher capacity for sorption is present 

in the finer-sized particles of a given sediment sample 

because they tend to have higher organic matter contents. The mineral component of a sediment provides 

only a relatively small contribution to sorption except, 

perhaps, in the case of certain ionic organics because 

water (through dipole interactions) is preferentially 

adsorbed over the nonionic organic compounds by the 

mineral adsorption sites. Normalizing field partition 

coefficients to a hypothetical reference sorbent that is 

100 percent organic matter shows that the variation in 

partition coefficients among different organic contaminants results primarily from differences in the water 

solubility of the contaminants and secondarily from 

differences in the organic matter composition of the 

soil and sediment samples. 

The transport of manmade organic compounds 

is considered in conjunction with the transport of 

sediment. In view of the sorptive interactions between 

sediment particles and contaminants, the hydrodynamic description of the movement of contaminants 

needs to be considered in terms of the transport of a 

dissolved species in addition to the transport of sediment particles having a certain contaminant capacity 

(based on their organic matter contents). In this 

respect, emphasis needs to be placed on the finergrained particles (cohesive sediment) because of their 

typically higher organic matter content. In river-estuary 

systems, additional attention is needed on flocculation 

processes to understand the dynamics of contaminant 

transport because the tidal motion and the salinity 

changes in estuaries combine to destabilize and aggregate colloidal particles into settleable masses that are 

more rapidly removed from the water column. 

One of the valuable results of the basic research 

investigating sediment -contaminant interactions is the 

development of methodologies to estimate the probable 

field behavior of newly synthesized organic compounds. 
Regression equations have been derived that allow the 

estimation of field partition coefficients from laboratorymeasurable contaminant and sorbent properties. fur 

example, sediment partition coefficients can be estimated from information on the organic compound, 

such as water solubility or octanol-water partition coefficient, and the organic matter content of the bulk 

sediment. More detailed estimates are possible that 

take into account the particle-size distribution and the 

organic matter content of each fraction. Similar regression equations are available for estimating the BCFs of 

nonionic organic compounds in aquatic organisms. 

These estimates are based on the triolene-water (or 

octanol-water) partition coefficient of the compound 

and the body content of lipid in the species of interest. 

fur newly synthesized organic compounds, these regression approaches provide a relatively easy way of obtaining an initial glimpse of the environmental behavior of 

compounds that have yet to be monitored under field 

conditions. 

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