U.S. GEOLOGICAL SURVEY CIRCULAR 1007 

Manmade Organic Compounds 

in the Surface Waters of the 

United States-A Review of 

Current Understanding 

Manmade Organic Compounds 

in the Surface Waters of the 

United States-A Review of 

Current Understanding 

By JAMES A. SMITH, PATRICK j. WITKOWSKI and 

THOMAS V. FUSILLO 

U.S. GEOLOGICAL SURVEY CIRCULAR 1007 
U.S. DEPARTMENT OF THE INTERIOR 

BRUCE BABBITT, Secretary 

U.S. GEOLOGICAL SURVEY 

GORDON P. EATON, Director 

Any use of trade, product, or firm names 

in this publication is for descriptive purposes only 

and does not imply endorsement by the U.S. Government 

First printing 1988 

Second printing 1993 

Third printing 1996 

UNITED STATES GOVERNMENT PRINTING OFFICE:1988 

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

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

Ubrary of Congress Cataloging In Publication Data 

Smith, James A. (James Albert) 

Manmade organic compounds in the surface waters of the 

United States. 

(U.S. Geological Survey circular ; 1007) 

Bibliography: p; 

Supt. of Docs. no.: I 19.4/2:1007. 

1. Organic water pollutants-United States. 2. Organic 

compounds. I. Witkowski, P.J. (Patrick J.) II. Fusillo, 

Thomas V. Ill. Title. IV. Series. 

T0427.07S55 1988 363.7'394 87-600258 
CONTENTS 

Abstract 1 

Introduction 1 

Background 1 

Purpose and scope 2 

Processes affecting the fate and distribution of organic compounds 2 

Sorption 2 

Definitions 2 

Adsorption and partitioning 3 

Sorption of nonionic organic compounds 3 

Importance of particulate organic matter 6 

Importance of dissolved organic matter 8 

Sorption of ionic organic compounds 8 

Organic acids 9 

Organic bases 9 

Bioaccumulation 11 

Importance of biological lipid reservoirs 11 

Biomagnification 14 

Transformation processes 1S 

Photolysis 16 

Hydrolysis 18 

Biodegradation 19 

Volatilization 21 

Polychlorinated biphenyls and chlorinated insecticides · 25 

Uses, production, properties, and regulatory limits 25 

Environmental fate 27 

Environmental distribution 32 

Carbamate and organophosphorus insecticides 36 

Uses, production, properties, and regulatory limits 36 

Environmental fate 37 

Environmental distribution 39 

Herbicides 39 

Uses, production, properties, and regulatory limits 39 

Environmental fate 40 

Environmental distribution 43 

Phenols 44 

Uses, production, properties, and regulatory limits 44 

Environmental fate 45 

Environmental distribution 49 

Halogenated aliphatic and monocyclic aromatic hydrocarbons 50 

Uses, production, and properties 50 

Environmental fate 52 

Environmental distribution 55 

Phthalate esters 56 

Uses, production, and properties 56 

Environmental fate 57 

Environmental distribution 59 

Contents Ill 
Polychlorinated dibenzo-p-dioxins 60 

Uses, production, and properties 60 

Environmental fate 62 

Environmental distribution 63 

Polycyclic aromatic hydrocarbons 64 

Uses, production, and properties 64 

Environmental fate 66 

Environmental distribution 70 

Summary 73 

Selected references 7 4 

FIGURES 

1-10. Plots showing: 

1. Correlation of octanol-water partition coefficient with water 

solubility for selected aromatic liquids and solids at 25 oc 4 

2. Soil-water equilibrium isotherms of some neutral organic 

compounds at 20 oc 4 

3. Typical soil-water equilibrium isotherms for 1,3-dichlorobenzene 

and 1 ,2,4-trichlorobenzene as single solutes and as binary solutes 

on a Woodburn silt loam soil at 20 oc 5 

4. Soil organics-water sorption coefficients plotted as a function of 

the aqueous solubilities of selected neutral organic 

compounds 5 

5. Sorption coefficient normalized for organic-carbon content as a 

function of compound water solubility and octanol-water 

partition coefficient 7 

6. Adsorption isotherms of paraquat in soils 10. 

7. Correlation of bioconcentration factor versus octanol-water 

partition coefficient showing the correlations of Veith and others 

(1979), Neely and others (1974), and Mackay (1982a) 12 

8. Relation between log water solubility of pesticides and log 

bioconcentration factor for Triaenodes tardus egg masses 13 

9. Correlation of log bioconcentration factor on the basis of fish 

lipid content with log triolein-water partition coefficient for a 

variety of nonionic organic compounds 14 

10. Relation between microbial specific growth rate and substrate 

concentration 20 

11. Structural diagrams of selected chlorinated insecticides and a 

polychlorinated biphenyl 25 

12. Plot showing relation of summer mean polychlorinated biphenyl (PCB) 

concentrations in water and lipid-bas~ PCB concentrations in fiiiets of 

whole yearling pumpkinseed collected September 1977-83 at Stillwater, 

N.Y 35 

13. Frequency of detection of organochlorine insecticides in water and 

bed-material samples from stations in the U.S. Geological Survey-U.S. 

Environmental Protection Agency Pesticide Monitoring Network, 

1975-80 35 

14. Structural diagrams of selected carbamate and organophosphorus 

insecticides 36 

IV Contents 
15, 16. Plots showing: 

15. Cumulative 14C02 (percentage of added ring-labeled 14C-insecticide) emlved in a soil planted to rice under flooded 

and nonflooded conditions 38 

16. Trends in national use of herbicides and insecticides on major 

crops, 1964-82 40 

17, 18. Structural diagrams of: 

17. Selected herbicides 40 

18. Selected phenolic compounds 44 

19. Structural diagrams showing degradation pathway of pentachlorophenol in 

soils 48 

20. Structural diagrams of selected halogenated aliphatic and monocyclic 

aromatic hydrocarbons 51 

21. Graph of predicted volatilization rates of methylene chloride and 

dichlorobenzene in the Rio Grande River, N.Mex. 55 

22. Structural diagrams of some common phthalate esters 57 

23, 24. Structural diagrams showing: 

23. Degradation of di-n-butyl phthalate, mono-n-butyl phthalate, and 

phthalic acid by different microorganisms: organisms isolated 

with di-n-butyl phthalate as carbon source for enrichment, stock 

cultures, and bacteria isolated from phthalic acid enrichment 

medium 59 

24. Formation of 2,3,7,8-tetrachlorodibenzo-p-dioxin as a byproduct 

in the synthesis of 2,4,5-trichlorophenoxyacetic acid 60 

25. Structural diagrams of selected polycyclic aromatic hydrocarbons 66 

26. Structural diagrams showing proposed degradation pathway of 

phenanthrene by soil pseudomonads 70 

TABLES 

1. Contrasting characteristics of adsorption and partition processes 3 

2. Selected anionic and cationic organic compounds of environmental 

significance 9 

3. Approximate molecular composition of some Aroclor mixtures 26 

4. Some uses of polychlorinated biphenyls suggested by the manufacturer 27 

5. Detection limits, water-quality criteria, and use data for selected organochlorine insecticides 28 

6. Experimentally determined octanol-water partition coefficients, solubilities, and 

vapor pressures of selected polychlorinated biphenyls and chlorinated insecticides at room temperature 29 

7. Experimentally determined sorption coefficients for some polychlorinated 

biphenyls and organochlorine insecticides 30 

8. Experimentally determined bioconcentration factors for some polychlorinated 

biphenyls and organochlorine insecticides 31 

9. Means and ranges of detectable sediment residues of polychlorinated biphenyls 

and organochlorine insecticides 33 

10. Detectable residues of polychlorinated biphenyls and organochlorine insecticides in biota 34 

11. Experimentally determined octanol-water partition coefficients, solubilities, and 

vapor pressures of selected organophosphorus and carbamate insecticides at 

room temperature 36 

Contents V 
12. Detection limits, water-quality criteria, and use data for selected 

organophosphorus insecticides 37 

13. Characteristics of selected carbamate insecticides 38 

14. Approximate persistence of pesticides in soils 38 

15. Detection limits, water-quality criteria, and use data for some common 

herbicides 40 

16. List of and magnitude of application of herbicides in 11 agricultural watersheds in Ontario, 1975 41 

17. Experimentally determined aqueous solubilities and vapor pressures of 

selected herbicides at room temperature 41 

18. Structural groupings of selected herbicides 41 

19. Herbicide residues in North American surface waters 43 

20. World production of phenol in 1978 45 

21. Experimentally determined physicochemical properties of selected phenolic 

compounds at room temperature 46 

22. Phenol bacteria present during the winter and (or) summer season in St. 

Lawrence River water and maximum phenol levels that could be tolerated 47 

23. Time required for bacteria to use 95 percent of parent substrate 48 

24. Photolysis of phenols and their degradation products 48 

25. Experimentally determined sorption coefficients of nonionized chlorinated 

phenols for different natural sorbents 49 

26. Concentration of phenol and some of its chlorinated derivatives in municipal 

and industrial wastewater 50 

27. U.S. production of selected halogenated aliphatic and monocyclic aromatic 

hydrocarbons 51 

28. Uses of selected halogenated aliphatic and monocyclic aromatic hydrocarbons 52 

29. Experimentally determined octanol-water partition coefficients, solubilities, and 

vapor pressures of selected halogenated aliphatic and monocyclic aromatic 

hydrocarbons at room temperature 53 

30. Evaporation times for 50 and 90 percent removal of chlorinated aliphatic 

hydrocarbons (1 milligram per liter) at 25 oc, 200 revolutions per minute of 

stirring, and a depth of 6.5 centimeters 54 

31. Ratios of overall transfer coefficients for selected aliphatic and monocyclic 

aromatic hydrocarbons to overall transfer coefficients for oxygen 54 

32. Uses of phthalate esters in the United States 57 

33. Solubilities and vapor pressures of selected phthalate esters 58 

34. Selected bioconcentration factors of phthalate esters in aquatic systems 58 

35. Chlorinated dibenzo-p-dioxins found in commercial pesticides 61 

36. Single oral doses of dioxins necessary to kill 50 percent of exposed 

animals 62 

37. Bioconcentration factors of 2,3,7,8-tetrachlorodibenzo-p-dioxin for a variety of 

aquatic organisms 62 

38. Estimated annual inputs of polychlorinated dibenzo-p-dioxins to the Canadian 

environment from major sources 64 

39. Residues of chlorodibenzo-p-dioxins in soil samples from various locations in 

the United States 65 

40. Mean and maximum concentrations of detected residues of 2,3,7 ,8-

tetrachlorodibenzo-p-dioxin in a variety of freshwater and marine fish 65 

VI Contents 
41. Residues of polycyclic aromatic hydrocarbons in oils, gasoline, bitumen, and 

sludge 67 

42. Estimated annual benzo(a)pyrene emissions to the atmosphere 68 

43. Experimentally determined octanol-water partition coefficients, solubilities, and 

vapor pressures of selected polycyclic aromatic hydrocarbons 68 

44. Calculated and experimental sediment-water sorption coefficients normalized 

for organic carbon for a variety of polycyclic aromatic hydrocarbons 69 

45. Bioconcentration factors for selected polycyclic aromatic hydrocarbons 69 

46. Sediment residues of polycyclic aromatic hydrocarbons 71 

47. Residues of polycyclic aromatic hydrocarbons in biota 72 

Contents VII 
ABBREVIATIONS 

A Angstroms mg!L Milligrams per liter 

atm-m3/mol Atmosphere-cubic meter Mglyr Megagrams per year 

per mole mi Miles 

oc Degrees Celsius min Minutes 

em Centimeters mm Millimeters 

cm/h Centimeters per hour mo Months 

d Days mol!L Moles per liter 

gld Grams per day mol/m3 Moles per cubic meter 

giL Grams per liter m/s Meters per second 

h Hours ng/L Nanograms per liter 

kcal/mol Kilocalories per mole nm Nanometers 

kg Kilograms pg/L Picograms per liter 

kgld Kilograms per day ppb Parts per billion 

kg/ha Kilograms per hectare ppm Parts per million 

kglyr Kilograms per year ppt. Parts per trillion 

km Kilometers rpm Revolutions per minute 

lb Pounds t Metric tons 

lb/yr Pounds per year torr-L/mol Torr-liters per mole 

L-torr/K-mol Liter-torr per degree Kelvin- torr/molar Torr-liters per mole 

mole yr Years 

m Meter p.g Micrograms 

m2/h Square meters per hour p.g!g Micrograms per gram 

m3 Cubic meters p.glkg Micrograms per kilogram 

mg!kg Milligrams per kilogram p.g!L Micrograms per liter 

VIII Abbrev .. tlona 
Manmade Organic Compounds in the 

Surface Waters of the United StatesA Review of Current Understanding 

By James A. Smith, Patrick J. Witkowski, and Thomas V. Fusillo 

Abstract 

This report reviews the occurrence and distribution of 

manmade organic compounds in the surface waters of the 

United States. On the basis of their aqueous solubilities, 

nonionic organic compounds partition themselves among 

water, dissolved organic matter, particulate organic matter, 

and the lipid reservoirs of aquatic organisms. Ionized organic 

compounds can be adsorbed to sediments, thereby reducing their aqueous concentrations. Transformation processes 

of photolysis, hydrolysis, biodegradation, and volatilization 

can attenuate organic compounds, and attenuation rates 

commonly follow a first-order kinetic process. 

Eight groups of manmade organic compounds are 

discussed: 

1. Polychlorinated biphenyls and organochlorine insecticides, 

2. Carbamate and organophosphorus insecticides, 

3. Herbicides, 

4. Phenols, 

5. Halogenated aliphatic and monocyclic aromatic hydrocarbons, 

6. Phthalate esters, 

7. Polychlorinated dibenzo-p-dioxins, and 

8. Polycyclic aromatic hydrocarbons. 

For each compound group, data pertaining to use, production, and properties are presented and discussed. Processes 

that influence the environmental fate of each group, as 

determined primarily through laboratory studies, are 

reviewed, and important fate processes are identified. Environmental concentrations of compounds from each group in 

water, biota, and sediment are given to demonstrate representative values for comparison with concentrations determined during ongoing research. Finally, where data are 

sufficient, regional and temporal contamination trends in the 

United States are discussed. 

INTRODUCTION 

Background 

Because of the dramatic increase in the production 

and use of manmade (anthropogenic) organic compounds in the United States over the past 50 yr, the quality 

of our Nation's surface-water systems has gradually deteriorated. Organic contaminants have been detected in the 

water, sediments, and biota of rivers, estuaries, and lakes 

throughout the United States. Shackelford and Keith 

(1976) documented the occurrence of 1,258 organic 

compounds in 5,720 observations in 33 water types from 

1970 to 1976. Gilliam (1985) reported the detection of 

one or more pesticides in almost 20 percent of sediment 

samples collected from more than 150 river sites in the 

United States .. Residues of persistent organochlorine 

compounds have been detected around the world, including in antarctic snow (Peterle, 1969) and fish from depths 

of 2,500 m in the Atlantic Ocean (Barber and Warlen, 

1979). 

Manmade organic compounds enter surface-water 

systems in point -source discharges, nonpoint -source runoff, atmospheric deposition, and ground-water discharges. 

The discharge of polychlorinated biphenyls (PCB's) into 

the Hudson River (N.Y) by two capacitor-manufacturing 

plants is a notorious example of point-source contamination. Over a 30-yr period, these plants combined to 

discharge PCB's to the Hudson at an average rate of 14 

kgld (Brown and others, 1985). Although the discharge of 

PCB's to the Hudson by these plants was stopped by 1977, 

significant residues of the persistent organochlorine contaminants remain in both the sediment and the biota. 

Nonpoint-source discharges also can contribute to 

surface-water contamination (Novotny, 1980), particularly with regard to pesticides applied to farmland and 

forests. In many areas, sediment resulting from eroded 

soil contains elevated concentrations of sorbed pesticides 

and is regarded as the primary source of contamination of 

Introduction 1 
the surface-water system. Based on land-use data, greater 

than 97 percent of land area in the United States is a 

potential source of nonpoint contaminant discharges 

(McElroy and others, 1975). Atmospheric deposition of 

organic contaminants into surface waters also has been 

documented. Heit and others (1984) determined that 

PCB contamination of four remote lakes in Rocky Mountain National Park (Colo.) was attributable to the atmospheric transport and subsequent deposition of the organic 

contaminants. They found that surficial sediments in the 

lakes had total PCB concentrations as high as 540 p.glkg 

dry weight. Finally, ground water also can supply a 

surface-water system with organic contaminants (Frank 

and others, 1981 ), but the persistence of soluble groundwater contaminants in surface waters is typically not great 

because of their volatility (Connor, 1984b ). 

Purpose and Scope 

The increasing production and use of manmade 

organic compounds coupled with their ease of transport 

into surface-water systems makes knowledge of their 

environmental behavior essential to the management of 

the Nation's freshwater resources. The purpose of this 

report is to provide an interpretative, critical review of the 

scientific literature with regard to the occurrence, distribution, and fate of manmade organic compounds in the 

surface-water systems of the United States. More specifically, the review examines organic-contaminant interactions with bottom and suspended sediment, the water 

column, and the biota of rivers, estuaries, and lakes. Both 

laboratory and field studies are reviewed, and, where 

applicable, their results are compared and contrasted. 

Examples of organic-contaminant concentrations reported 

in field studies are given to serve as a means of comparison with ongoing research. Special emphasis is given to 

the environmental variables and the physicochemical properties of the organic compounds that are intrinsic to the 

distribution and fate of the contaminant. Finally, where 

data are sufficient, the regional and temporal distribution 

patterns of manmade organic compounds in the surfacewater systems of the United States are discussed. 

PROCESSES AFFECTING THE FATE AND 

DISTRIBUTION OF ORGANIC COMPOUNDS 

This section is divided into three main parts. The 

first part reviews sorption interactions between sediments 

and organic contaminants, the second discusses bioaccumulation of organic contaminants, and the third reviews 

transport and transformation processes affecting organic 

contaminants in surface-water systems. 

Sorption 

Water quality has traditionally been assessed in 

terms of concentrations of dissolved constituents. However, certain organic contaminants can be sorbed to 

suspended and bottom sediment and may be present in 

sediment at concentrations that are orders of magnitude 

higher than their dissolved concentrations. Sediments can 

be a reservoir for many persistent organic compounds, 

causing the compounds to remain in the surface-water 

system years after input of the contaminant has stopped. 

Additionally, the sorbed organic contaminant may represent a large fraction of the total mass of contaminant 

transported downstream. The ability to predict the extent 

of sorption for a contaminant in a surface-water system 

may, therefore, be extremely important in assessing water 

quality. 

Definitions 

Unlike aqueous concentrations, concentrations of 

anthropogenic organic compounds sorbed to sediment or 

soil are expressed as a ratio of the mass of contaminant to 

the mass of sediment or soil (for example, p,g/g, p,g/kg, 

mglkg). The extent to which an organic contaminant is 

sorbed to a given sediment at equilibrium is quantified by 

its distribution (sorption) coefficient,~ , which is defined 

as 

v =Cs 

.l.'d c ' e 

where Cs is the concentration of a contaminant sorbed to 

a specific weight of sediment and ce is the concentration 

of the same contaminant dissolved in an equal weight of 

water (Olsen and others, 1982). 

For a given contaminant-sediment combination, C9 

is a function of ce and commonly is related to ce by the 

empirical Freundlich isotherm, given by 

where K and n are empirical constants that are a function 

·of the sorbent and sorb ate. If n has a value of unity, a plot 

of Cs versus Ce will be linear and K will equal the 

previously defined sorption coefficient. 

The sorption coefficient commonly is expressed on 

an organic-matter basis and is defined as 

Kd ~m=-f -, 

om, 

where fom is the fractional mass of organic matter in the 

sediment. Similarly, the sorption coefficient expressed on 

an organic-carbon basis is defined as 

2 Manmade Organic Compounds In the Surface Waters of the United States 
where foe is the fractional mass of organic carbon in the 

sediment. The values of fom and foe can be related by the 

Van Bemmelen factor of 1.724 {Allison, 1965), which 

assumes that soil or sediment organic matter is composed 

of 58 percent carbon. This conversion factor is, however, 

only an approximation, and determinations of foe and fom 

typically are obtained by laboratory analyses. 

An organic compound's octanol-water partition 

coefficient, I<aw, is defined as the ratio of the compound's 

concentration in a known volume of n-octanol to its 

concentration in a known volume of water after the 

octanol and water have reached equilibrium. 

Adsorption and Partitioning 

Sorption of organic compounds by sediments can be 

accomplished by either an adsorption or a partition process. Adsorption is a surface phenomenon that involves 

formation of either a physical or a chemical bond between 

the organic compound and a site on the sediment particle's 

surface. Physical adsorption is caused by van der Waal's 

forces, whereas chemical adsorption (chemisorption) is 

the result of formation of a chemical bond between the 

organic compound and the sediment (Khan, 1978). Ion 

exchange is an example of chemical adsorption. For a 

partition process, the organic compound distributes itself 

between water and the sediment organic matter by molecular forces common to solution, similar to the distribution 

between water and an organic solvent such as octanol 

{Chiou and others, i979; Hassett and others, 1980; 

Chiou, 1981; Chiou, Schmedding, and Manes, 1982; 

Chiou, Porter, and Schmedding, 1983). 

Table 1 contrasts the characteristics of partition and 

adsorption processes (Chiou and others, 1979; Chiou, 

Porter, and Schmedding, 1983). Adsorption is a surfacecondensation process that occurs with high exothermic 

heat in compensation for losses in entropy. The enthalpy 

change for adsorption of a solute from water must be 

greater than the solute's heat of condensation from water 

(except when adsorption is very weak). For a partition 

process, the heat of sorption equ.als the difference between 

the molar enthalpies of solution of the solute in the 

organic and water phases and typically is less exothermic 

than the solute's heat of condensation. Adsorption processes are characterized by nonlinear isotherms, whereas 

partition processes yield linear isotherms over a wide 

range of solute solubility. An adsorption process is also 

characterized by a competitive effect between solutes for a 

fixed number of surface sites, while a partition process 

involves independent sorption between individual solutes. 

Finally, the magnitude of the partition coefficient of an 

organic solute between an organic phase and water is 

Table 1. Contrasting characteristics of adsorption and partition processes 

Characteristics 

Adsorntion Partition 

High heats of sorption Low heats of sorption 

Nonlinear isotherms Linear isotherms 

Competitive sorption Noncompetitive sorption 

Solubility dependence 

determined primarily by the solute's solubility in water. 

The dependence of the octanol-water partition coefficient 

on solute-water solubility has repeatedly been demonstrated for various organic solutes (Chiou and others, 

1977; Chiou and Freed, 1977; Mackay, 1977; Chiou, 

Schmedding, and Manes, 1982; Miller and others, 1985); 

water solubility was found to be the major factor affecting 

the partition coefficient. In analogy to the solvent-water 

system, Chiou, Porter, and Schmedding {1983) demonstrated that solute solubility also was the primary· factor 

influencing the partitioning of 12 nonionic organic compounds between water and soil organic matter. The 

relation between aqueous solubility and Kow developed by 

Chiou, Schmedding, and Manes {1982) is given in figure 

1. 

Because of significant differences between the 

adsorption and partition processes, it is important to 

determine the mechanism responsible for the sorption of 

manmade organic compounds to sediments in surfacewater systems. Although the issue was considered unsettled by some investigators (Kyle, 1981; Mingelgrin and 

Gerstl, 1983; Curl and Keoleian, 1984; Macintyre and 

Smith, 1984 ), there is strong evidence that sorption of 

nonionic and acidic organic compounds from water by soil 

or sediment is caused primarily by the partitioning of the 

solute into the soil/sediment organic matter (Chiou and 

others, 1979, 1985; Chiou, Porter, and Schmedding, 1983; 

Witkowski and others, 1987). Sorption of organic bases, 

on the other hand, appears to occur primarily by adsorpt~on. 

Sorption of Nonionic Organic Compounds 

The sorptive behavior of nonionic organic compounds characteristic of a partitioning process has been 

repeatedly observed. Chiou and others (1979) calculated 

low heats of sorption for 1,2-dichlorobenzene and 1,1,1-

trichloroethane sorbed to a Willamette silt loam. Pierce 

and others {1974) calculated low isosteric heats of sorption for dichlorodiphenyltrichloroethane (DDT) sorbed 

Processes Affecting the Fate and Distribution of Organic Compounds 3 
~ 7 ~ 

g 

..J 

..: 6 z 

w 

§ 

It 5 w 

0 

0 

z 

Q 4 .... 

i= 

a: 

c 

~ 3 w 

.... 

c 

~ 

~ 2 0 

z 

< 

.... 

0 

0 

Aromatic liQuid• and aollda: 

log K0w • -0.862 log Sw + 0.710 

(r • 0.994; n • 36) 

0 

SOLUBILITY IN WATER, LOG Sw,(MOLES PER LITER) 

Figure 1. Correlation of octanol-water partition coefficient 

with water solubility for selected aromatic liquids and solids at 

25 oc. The solubilities of solid compounds in the plot are 

those of their supercooled liquids at 25 oc. [Reprinted with 

permission from Environmental Science and Technology, v. 

16, 9 (1982). Copyright 1981 by the American Chemical 

Society.) 

z::e 

-:< ::EO: 

..... (!) 

xa: 

·w ::!a. 

0(/) 

(/):::!: 

>-< ala: 

WC!J 

~0 

<a: 

1--0 

~:::!: 

to several types of particulate matter ranging from -2 to 

+ 4 kcaVmol. Because the heat of condensation for DDT 

is approximately 10 kcal/mol, an adsorption process would 

require at least 10 kcal/mol (exothermic). The low heats 

calculated by Pierce and others (1974) are in keeping with 

a partition process. Khan (1974b) also witnessed low 

isosteric heats of sorption of the weakly acidic herbicide 

2,4-D when sorbed to a fulvic acid-clay complex, again 

indicating a partition process. 

Figure 2 presents sorption isotherms for a variety of 

nonionic organic compounds partitioned between water 

and a Willamette silt loam (Chiou and others, 1979). All 

the isotherms are linear, and, for 1 ,2-dichloropropane and 

1,2-dichloroethane, the isotherms are linear up to equilibrium aqueous concentrations that are 75 and 95 percent of the solute's solubility. This observation, along with 

the low exothermic heat, cannot be reconciled by an 

adsorption model. Lotse and others (1968) observed 

approximately linear sorption isotherms for lindane 

sorbed to lake sediments. 'Karickhoff and others (1979) 

observed isotherm linearity for the sorption of pyrene and 

methoxychlor to natural sediments. Pierce and others 

(1974) and Picer and others (1977) witnessed linear 

Freundlich isotherms for DDT sorption to natural sediments. Hassett and others (1980) reported linear isotherms for dibenzothiophene sorption to a variety of soils 

and sediments. Finally, Herbes (1977) also observed 

200 400 600 800 1000 1200 1400 1600 1800 2000 2200 2400 

EQUILIBRIUM CONCENTRATION, Ce, IN PARTS PER MILLION 

Figure 2. Soil-water equilibrium isotherms of some neutral organic compounds at 20 °C. The soil used is a Willamette silt 

loam with 1.6 percent organic matter, 26 percent clay, 3.3 percent sand, and 69 percent silt. [Reprinted with permission from 

Science, v. 206, p. 831 (1979). Copyright 1979 by the American Association for the Advancement of Science.] 

4 Manmade Organic Compounds In the Surface Waters of the United States 
linear isotherms for anthracene sorption to particulate 

organic matter. All these examples of linear isotherms are 

indicative of a partition process rather than an adsorption 

process. 

The noncompetitive effect of binary nonionic solutes on sorption isotherms was demonstrated by Chiou, 

Porter, and Schmedding (1983), Chiou and others (1985), 

and Chiou and Shoup (1985). Figure 3 presents soil-water 

isotherms for single and binary solutes (Chiou, Porter, 

and Schmedding, 1983). No competitive effects are 

observed. Instead, the data indicate independent sorption 

that is characteristic of a partition process. Additiona11y, 

both isotherms are linear up to 60 to 90 percent of solute 

saturation. 

The solubility dependence of the sorption process 

also has been observed. Senesi and Testini (1980) noted 

increased sorption of some nitrogenated herbicides with 

decreasing solubility. Sharom and others (1980a) noted 

that the sorption of 12 insecticides by three types of soil 

was inversely correlated with the insecticides' solubilities. 

Figure 4 presents data from Chiou and others (1979) 

showing the linear log-log relation between solubility and 

the soil organic matter-water distribution coefficient (~m) 

for a variety of nonionic organic compounds. The data 

demonstrate that the distribution coefficient decreases 

with increasing solubility. The observed solubility dependz 

0 

.... 

::l 

~ 

a: 

.... 

~ 

cw 

a:..J 

w< .,_o 

<UJ 

3:~ 

10 C/)..J 

0 .. -t- Zz 

<w 

(!)-

a:O 

oU: 

u. -lw 

0) 

~ 200 

lr 

CJ 

0 

lr 160 0 

i 

_; 

0 

0) 

> 

m 

w 

:IC 

< 

IG. 

~ 

Single solute Isotherms: 

01,3-Dichlorobenzene 

ll. 1,2 ,4-Trichlorobenzene 

Binary solute isotherms: 

81,3-Dichlorobenzene 

•1,2,4-Trlchlorobenzene 

20 40 80 80 

EQUILIBRIUM CONCENTRATION, C8 , 

IN MILLIGRAMS PER LITER 

Figure 3. Typical soil-water equilibrium isotherms for 1 ,3-

dichlorobenzene and 1,2,4-trichlorobenzene as single solutes and as binary solutes on a Woodburn silt loam soil at 20 

°C. [Reprinted with permission from Environmental Science 

and Technology, v. 17, p. 229 (1983). Copyright 1983 by the 

American Chemical Society.] 

1,2-Dichlorobenzene .1,1,1-Trichloroethane 

1,2-Bromo-3 chloropropane 1,1,2,2-Tetra1,2-0i bromoethone chloroethane 

1,2-Dichloropropone oo 

cno 10 I 2-Dichloroethone 

10-3 10- 2 10 1 I 10 102 103 104 105 106 

SOLUBILITY IN WATER, IN MICRO MOLES PER LITER. LOG SCALE 

Figure 4. Soil organics-water sorption coefficients plotted as a function of the aqueous solubilities of selected neutral 

organic compounds. [Reprinted with permission from Science, v. 206, p. 831 (1979). Copyright 1979 by the American 

Association for the Advancement of Science.] 

Processes Affecting the Fate and Distribution of Organic Compounds S 
ence of the sorption process again lends support to the 

hypothesis that partitioning between water and sediment 

organic matter is primarily responsible for the sorption of 

nonionic organic compounds by sediments. 

Adsorption of nonionic organic compounds appears 

not to be a significant mechanism in sediment -water 

systems because of the ability of polar water molecules to 

compete successfully with nonionic organic solutes for 

mineral adsorption sites (Yaron, 1978; Chiou, Peters, and 

Freed, 1981; Chiou, Schmedding, and Block, 1981; Chiou 

and Shoup, 1985; Chiou and others, 1985). On the other 

hand, sorption of nonionic organic compounds in the 

absence of water may result primarily from an adsorption 

mechanism. Yaron and Saltzman (1972) found that 

parathion sorption from a hexane solution decreased with 

increasing soil-moisture content. Chiou and others (1985), 

when observing parathion and lindane sorption from 

hexane onto dry and partially hydrated Woodburn soil, 

observed nonlinear isotherms, strongly exothermic heats 

of sorption, competitive sorption effects, and dependence 

on clay content. Additionally, increased soil-moisture content decreased sorption. In the absence of the competitive 

effects of water for surface adsorption sites, the two 

nonionic insecticides were apparently adsorbed to the 

mineral surfaces. In hexane solution, little solute would 

partition into the soil organic matter because of its high 

solubility in the organic solvent. A slight increase in the 

soil-moisture content, however, resulted in a sharp decline 

in sorption owing to adsorptive competition by the water 

molecules. Similar moisture-content effects have also 

been observed by Yaron (1978) for the sorption of 

organophosphorus insecticides to soils. Strong sorption of 

organic vapors by dry soils has also been attributed to 

adsorption by minerals since water is not present to 

compete with the nonionic org~nic molecules for adsorption sites (Chiou and Shoup, 1985). 

The above hypothesis explains the results of Walker 

and others (1975), who observed greater extraction efficiencies for petroleum hydrocarbons from dry sediment 

when a more polar solvent was used. Polar solvents, such 

as methylene chloride, can replace nonionic organic compounds that are weakly adsorbed to the sediment mineral 

surfaces more effectively than can nonpolar solvents, such 

as hexane (Chiou and Shoup, 1985). Likewise, the sorption of a nonionic organic compound from a polar solvent, 

such as methanol or methylene chloride, onto dry soil or 

sediment would likely be insignificant (Chiou and others, 

1985). The compound's high solubility in the solvent 

would greatly limit partitioning into the soil or sediment 

organic matter, and the polar solvent would successfully 

compete with the solute for surface-adsorption sites. This 

observation is in agreement with the data of Nkedi-Kizza 

and others (1985), who observed decreased sorption of 

anthracene, atrazine, and diuron from water onto soil 

when methanol or aceton~ was added as a cosolvent. The 

cosolvent enhanced solute solubility (thus reducing the 

partition of solute into organic matter), while significant 

adsorption by minerals was prevented by the polar mixed 

solvents. In surface-water systems, however, sorption of 

nonionic organic compounds will always occur from water. 

Therefore, adsorption processes are likely to be insignificant compared with the partition of the organic compound into the sediment organic matter. 

Importance of particulate organic matter 

Because sorption of nonionic organic compounds 

appears to be caused by partition between water and 

sediment organic matter, sediments having large organicmatter contents would be expected to sorb greater 

amounts of solute than sediments having small organicmatter contents. A correlation between organic-matter 

content and sorption has been observed by numerous 

investigators (Pierce and others, 1974; Spencer and others, 1974; Choi and Chen, 1976; Herbes, 1977; Rao and 

Davidson, 1979; Hassett and others, 1980; Sharom and 

others, 1980a; Abdullah and Ringstad, 1982; Banwart 

and others, 1982; Hassett and Anderson, 1982; NkediKizza and others, 1983; Glotfelty and others, 1984; Oliver 

and Charlton, 1984; Brown and others, 1985; Stamer and 

others, 1985) and is an important consideration when 

comparing sorption data obtained from different sediments. Two sediment samples collected from the same 

river cross section may have significantly different contaminant concentrations if their organic-matter contents 

are dissimilar. 

To allow comparisons of sorption coefficients 

obtained from different sediments, Karickhoff and others 

(1979) normalized experimentally determined sedimentwater sorption coefficients by dividing by the organiccarbon content of the sediment. The resulting organiccarbon sorption coefficient, I<ac, differed little from 

sediment to sediment. Rao and Davidson (1979) measured soil-water sorption coefficients for several pesticides using three soil types. The sorption coefficients 

varied greatly with soil type, but when the coefficients 

were adjusted to reflect the organic-carbon content of the 

soils used, the resulting I<ac values showed much greater 

independence from soil type. Similarly, Nkedi-Kizza and 

others (1983) observed that soil-water sorption coefficients obtained using six particle-size fractions of a Webster soil differed by a factor of 7, while calculated ~c 

values differed by a factor of only 1.5. In a field study of 

atrazine and simazine sorption, Glotfelty and others 

(1984) also observed much less variability in I<ac values 

than in ~rption coefficients not normalized for the 

organic-carbon effect. The results of these studies indicate that sorption· coefficients normalized for sediment 

organic-carbon content will best reflect spatial contamination trends by eliminating much of the variability caused 

6 Manmade Organic Compounds In the Surface Waters of the United States 
by different concentrations of organic matter in sediments. 

Because solute solubility in water (S) is the primary 

factor affecting both the soil organic-matter sorption 

coefficient (Chiou, Porter, and Schmedding, 1983) and 

the octanol-water partition coefficient (Chiou, Schmedding, and Manes, 1982), a relation can be established 

between S or ~w and the sediment organic-carbon sorption coefficient. Such a relation is depicted in figure 5 for 

10 nonionic organic compounds partitioned between 

water and the coarse-silt fractions of two sediments 

(Karickhoff and others, 1979). Sorption differences 

caused by different solute types are accounted for by S or 

~w' and differences caused by different sediment types 

are accounted for by organic-carbon content. Similar 

relationships have been developed by Banwart and others 

(1982) and Oliver. and Charlton (1984). Chiou and others 

(1979), Chiou, Porter, and Schmedding (1983) also have 

developed such relations, but they have normalized the 

sediment-water sorption coefficient by the organic-matter 

content rather than the organic-carbon content. 

Relations such as the one depicted in figure 5 

potentially are very useful because they allow prediction 

of the sediment-water sorption coefficient for a multitude 

of nonionic organic contaminants with a knowledge of 

only the sediment organic-matter content and the solute's 

solubility or octanol-water partition coefficient. The ~m 

(or ~c) value for a given compound will likely show some 

variation from sediment to sediment because of variations 

in the composition of the sediment organic matter (Means 

and others, 1982; Chiou, Porter, and Schmedding, 1983; 

Chiou and others, 1984). However, this variation is small 

compared with variations in the sediment-water sorption 

coefficient prior to normalization for organic-matter content. 

In general, sediment organic-matter content has 

been shown to increase with decreasing particle size ( Choi 

and Chen, 1976; Frank and others, 1979; Scott and others, 

1982; Simmons and Kotz, 1982; Ricci and others, 1983; 

Stamer and others, 1985). Therefore, the highest sorbed 

contaminant concentrations are commonly found in the 

silt and clay fractions of surface-water sediments. Bottom 

sediments containing large percentages of silts and clays 

typically accumulate in areas of low water velocity, such as 

along the banks of rivers or lakes. Hunter and others 

(1980), in a study of an Oklahoma stream, observed the 

highest PCB concentrations in areas of sediment deposition. Routh (1972) likewise detected elevated DDT concentrations in Salinas River (Calif.) sediments in areas of 

low velocity. 

It should be noted that although the highest contaminant concentrations are usually found in the silt- and 

clay-size fractions of a sediment (because of their usually 

high organic-carbon content), numerous exceptions have 

been documented. Stainken and Rollwagen (1979) found 

LOG WATER SOLUBILITY (MOLE FRACTION x 109 ) 

10.0 -1.0 0 1.0 2.0 3.0 4.0 5D 

9.0 

.& Octanoi/Water 

8.0 <> Water solubility 

7.0 

6.0 0 

0 

~ 

(!) 5.0 

0 

..J 4.0 

2 3 4 6 7 

LOG K0 w 

Figure 5. Sorption coefficient normalized for organiccarbon content (K0 J as a function of compound water 

solubility and octanol-water partition coefficient (Kow)· 

[Reprinted with permission from Water Research, v. 13, p. 

247 (1979). Copyright 1979 by Pergamon Press Ltd.] 

no correlation between percent silt -clay and PCB concentrations in sediment samples from Raritan Bay (N.J.). Van 

Luik (1984) similarly found no correlation between percent silt-clay and either PCB or polycyclic aromatic 

hydrocarbon (PAH) concentrations in sediment samples 

from the Calumet Sag Channel (Ill.). In a laboratory study 

of dibenzothiophene sorption to soils and sediments, 

Hassett and others (1980) observed no relation between 

the partition coefficient and either the cation-exchange 

capacity, the clay mineralogy, the percentage of clay, or the 

surface area of the soil-sediment. However, a positive 

correlation was observed between the partition coefficient 

and the organic-matter content of the soil-sediment. 

Organic-matter content, therefore, seems to be the most 

important indicator of a sediment's sorptive capacity for 

nonionic organic solutes. 

Frank and others (1979, 1981), Frank, Thomas, 

Braun, and others (1980), and Frank, Thomas, Holdrinet, 

and Damiani (1980), after observing the relatively low 

sorptive capacity of the sand-size fraction of Great Lakes' 

sediment compared with the silt- and clay-size fractions, 

corrected their sorption data for the variable content of 

sand in the different samples. Although this correction 

was an improvement, the approach depended on the 

Processes Affecting the Fate and Distribution of Organic Compounds 7 
relation between organic-carbon content and particle size 

and did not consider differences in the organic-carbon 

contents of the different sediment samples. The studies by 

Frank and others dealt with nonionic sorbates, such as 

PCB's and organochlorine insecticides, that will likely 

sorb by partition into the sediment organic matter. Therefore, a better normalization approach might have been to 

divide the sorbed contaminant concentration by the 

organic-matter content of the sediment sample, as illustrated by Smith and others (1987). This approach does 

not depend on the organic-matter content to increase with 

decreasing particle size. It also accounts for differences in 

the organic-matter contents of the silt-day fractions of the 

different sediment samples. 

Importance of dissolved organic matter 

Although the sorption coefficient is strongly influenced by sediment organic-matter content, the presence 

of dissolved organic matter (DOM) also can influence the 

sorption coefficient by serving as an additional organic 

phase into which a solute can partition. In this manner, 

DOM increases the apparent solubility of the solute in 

water while the amount of solute partitioned into the 

sediment organic matter may remain relatively unchanged 

because of low concentrations of DOM. The resulting net 

effect will be a decrease in the sediment -water sorption 

coefficient. Chiou and others (1986) have demonstrated 

that the apparent aqueous solubilities of DDT and two 

PCB isomers increase linearly with DOM content, with 

no indication of competitive effects. They attribute the 

solubility enhancement of highly water insoluble organic 

solutes by DOM to a partitionlike interaction of the 

compounds with the microscopic organic environment of 

the high-molecular-weight DOM. The solubility enhancement caused by DOM has also been observed by other 

researchers for DDT (Wershaw and others, 1969; Carter 

and Suffet, 1982; Caron and others, 1985), phthalate 

esters (Matsuda and Schnitzer, 1971), and the alkanes, 

hexadecane and eicosane (Boehm and Quinn, 1973). For 

more soluble compounds such as lindane, however, DOM 

concentrations up to 100 mg!L have been shown to have 

no significant enhancement effect on solute solubility 

(Caron and others, 1985; Chiou and others, 1986), which 

is in agreement with the proposed partition interaction 

between solute and DOM. Therefore, only relatively 

insoluble organic compounds, such as DDT or PCB's, are 

likely to exhibit enhanced solubility and reduced sorption 

coefficients in the presence of sufficient concentrations of 

DOM. 

The ability of DOM to affect the sorption coefficient helps to explain the phenomenon known as the 

"solids effect," or the observed decrease in sorption 

coefficients with increasing sediment concentration. This 

effect has been studied by numerous researchers (Lotse 

and others, 1968; Grover and Hance, 1970; Herbes, 1977; 

O'Connor and Connolly, 1980; Horzempa and Di Toro, 

1983b; Koskinen and Cheng, 1983; Voice and others, 

1983; Curl and Keoleian, 1984; Bowman and Sans, 1985a; 

Gschwend and Wu, 1985; Voice and Weber, 1985), and in 

some cases an order of magnitude increase in the sediment concentration resulted in a decrease of comparable 

magnitude in the sorption coefficient. Voice and others 

(1983) and Gschwend and Wu (1985) have found that 

increased sediment· concentrations result in increased 

DOM concentrations. Because increased DOM can 

reduce the apparent sorption coefficients of relatively 

insoluble organic compounds, the relation between the 

sediment concentration and the sorption coefficient can be 

at least partially attributed to the solubility-enhancing 

effect of DOM (Voice and others, 1983; Chiou and 

others, 1984; Gschwend and Wu, 1985; Voice and Weber, 

1985). 

DOM, therefore, may be very important in determining the environmental fate of extremely hydrophobic 

organic pollutants. Sorption coefficients may differ 

greatly-between bottom sediments and suspended sediments because of the large variation in sediment concentration and, consequently, DOM concentration (Voice and 

Weber, 1985); organic compounds that partition strongly 

to DOM may not exhibit the same bioavailability as they 

would if they existed in a free, solubilized state. Similarly, 

other fate-influencing processes such as volatilization, 

hydrolysis, or photolysis also may be substantially 

affected by the partitioning of organic compounds into 

DOM. However, available data suggest that under typical 

environmental conditions the solubilities of only the least 

soluble organic compounds will be enhanced by DOM, 

and that the extent to which these compounds partition 

into DOM will depend on the polarity and size of the 

dissolved organic molecules along with the solute's solubility (Chiou and others, 1986). 

Sorption of Ionic Organic Compounds 

Herbicides and phenolic compounds represent the 

majority of environmentally significant ionic organic compounds. Compared with nonionic organic compounds, 

such as PCB's and chlorinated insecticides, they generally 

have much higher aqueous solubilities, and, as a result, 

solute partitioning into sediment organic matter is greatly 

reduced. Their solubilities are functions of system pH, so 

sorption can be expected also to vary with pH. Ionic 

organic compounds dissociate to either cations or anions, 

and, based on the surface charge of sediment particles, 

the ionized organic compound may be able to compete 

successfully with water molecules for surface adsorption 

sites. Table 2lists some of the environmentally important 

anionic and cationic organic compounds. 

8 Manmade Organic Compounds In the Surface Waters of the United States 
Table 2. Selected anionic and cationic organic compounds 

of environmental significance 

Anionic 

Dalapon 

Dicamba 

2,4-D 

2,4,5-T 

Glyphosate 

Pic lor am 

Phenol 

Chlorophenols 

Nitrophenols 

Cresols 

Organic acids 

Cationic 

Ametryn 

Atrazine 

Cyperquat 

Difenzoquat 

Diquat 

Paraquat 

Benzidine 

Prometon 

Prometryn 

The solubility of organic acids increases with pH as 

a result of increased deprotonation and the resulting 

anion formation. The organic anion presumably does not 

displace polar water molecules (or other inorganic cations) bound to surface adsorption sites inasmuch as 

particle surfaces, like the organic anion, generally are 

negatively charged (Khan, 1974c, 1980; Hayes and others, 

1975; Jackson, 1975; Wang, 1975; Mulholland, 1981; 

Davis, 1982; Olsen and others, 1982). Therefore, because 

increased dissociation of the organic add will not cause 

increased adsorption and because increased solute solubility will reduce solute partitioning into the sediment 

organic matter, the sorption of organic acids will be 

reduced at elevated pH. 

Schellenberg and others (1984) noted decreased 

sorption coefficients for chlorinated phenols when pH was 

increased from 6.5 to 8.5. The sorption isotherms were 

linear even at the higher pH values, indicating that solute 

adsorption was small compared with solute partitioning 

into the sediment organic matter despite the presence of 

the phenolate anion. This is consistent with the hypothesis 

that organic anions are repelled from negatively charged 

particle surfaces. Linear isotherms for phenolic compounds also have been determined by Scott and others 

(1982). Khan (1974b) observed nearly linear isotherms 

and low isosteric heats of sorption for 2,4-D sorption by a 

fulvic acid-clay complex, despite the fact that more than 

50 percent of the 2,4-D was present in its anionic form. 

Nkedi-Kizza and others (1983} observed approximately 

linear isotherms for 2,4-D sorption to different soil 

particle-size fractions and noted the dependence of the 

sorption coefficient on the soil's organic-matter content. 

Linear isotherms also have been observed for 2,4,5-T 

sorption onto soil from water (Koskinen and Cheng, 

1983). These results indicate that the organic anions 

present in solution generally do not exhibit appreciable 

adsorption owing to their negative charge, and that any 

sorption that does occur is attributable primarily to 

partitioning of the undissociated organic acid into the 

sediment organic matter (Chiou, 1981). Sorption of 

organic acids as influenced by partitioning will be reduced 

at higher pH because of increased solute solubility and 

decreased undissociated solute concentration in accordance with the results of Schellenberg and others (1984). 

Organic bases 

Sorption of organic bases to sediment also will be 

largely controlled by pH, but, in contrast to organic acids, 

adsorption of the protonated organic cation to particle 

surfaces (presumably by ion exchange or ion pairing) can 

be a significant sorption mechanism. Khan (1972a) has 

noted that sorption of basic herbicides is greatest at pH 

values close to the herbicides' p~ (-log base dissociation constant) values. At pH > > p~, the basic herbicides are almost completely in their neutral form and will 

be unable to compete with water molecules for mineral 

adsorption sites of the sediment becaus~ of their neutral 

charge. Partitioning into the sediment organic matter may 

also be low owing to the generally high solubility of the 

basic herbicides. At pH < < p~, the organic base will be 

almost completely in its cationic form, but adsorption will 

be retarded because of the high hydrophilicity of the 

cations in combination with the increased protonation of 

many of the acidic functional groups on the sediment 

organic matter. This counteracting effect makes the efficiency of the adsorption process a maximum at pH 

around p~. 

Zierath and others (1980) found that system pH 

controlled the sorption of benzidine to soils and sediments 

by varying the amount of benzidine present in the cationic 

form~ Benzidine sorption increased as pH was lowered to 

values approaching the compound's p~. Sorption of the 

ionized species also was correlated with the surface area 

of the soil-sediment used. Narine and Guy (1982} 

observed diquat sorption to humic acid to increase as pH 

was raised from 2.0 to 9.0. Because diquat is essentially 

100 percent in its cationic form within this experimental 

pH range, the increased diquat sorption at high pH can be 

attributed to a charge interaction (ion exchange or ion 

pairing) of diquat with the negatively charged acidic 

phenolic and carboxylic functional groups (having p:K.t 

(-log acid dissociation constant) values of about 10 and 

4, respectively) of the humic acid. The increased protonation of the functional groups at low pH decreases the 

availability of adsorption sites for the organic cations. The 

importance of the adsorption process to the sorption of 

organic bases is demonstrated by the data in figure 6 (Juo 

and Oginni, 1978), which presents sorption isotherms for 

paraquat using a variety of soils. The nonlinear isotherms 

are indicative of an adsorption process. Distinctly nonlinear isotherms also have been reported by Karickhoff and 

Processes Affecting the Fate and Distribution of Organic Compounds 9 
(/) 

:E 7 

< 

a: 

(!) 

~ 6 

...J 

-...J 

:eo 

~(/)5 

•LL 

co 

w 

Ol::i4 

O:c( 

Oa: 

(/)(!) 

ca: 

<w 

D 

1-a. ---------------------v 

~ 2 

a 

< 

a: 

< 1 

a. 

----v- v 

0 

50 100 150 200 250 300 350 

C. MILLIGRAMS PER LITER 

Figure 6. Adsorption isotherms of paraquat in soils. The X-axis denotes the paraquat concentration, C, in 

water. [Reprinted with permission from Journal of Environmental Quality, v. 7, p. 11 (1978). Copyright 1978 by 

the American Society of Agronomy, Crop Science Society of America, and Soil Science Society of America.] 

Brown (1978) for the sorption of paraquat by sediment 

and by Khan (1974a) for the sorption of paraquat and 

diquat. 

Organic cations appear to be adsorbed primarily by 

an ion:-exchange mechanism, as is evidenced by the 

infrared spectroscopic analyses of Khan (1974a) and 

Senesi and Testini (1980). The adsorption process is, 

therefore, sensitive to both the charge and atomic radius 

of the saturating cation associated with the sediment. 

Khan (1974c) stated that the sediment-associated cations 

that enhanced adsorption of both diquat and paraquat to 

a humic acid followed the sequence Al+3 < Fe +3 < Cu +2 < 

Ni+2 <Zn+2 < Co+2 <Mn+2 <H+ < Ca+2 <Mg+2 • This 

relation is inversely related to the stability of the element 

when complexed with humic acid (Khan, 1974c). The 

stability of an element typically increases with both 

valence and ionic radius. Strongly bound cations, such as 

Al+3 or Fe+3 , are more resistant to displacement by an 

organic cation than are weakly bound cations, such as 

Ca +2 or Mg +2 • The variation in organic-cation sorption as 

a function of the type of inorganic cation saturating the 

soil or sediment also has been observed by El-Dib and 

Aly (1976b) and Hayes and others (1975). Hayes and 

others (1975) also noted that while different saturating 

inorganic cations may influence the sorption of bipyridylium herbicides, this influence will also depend on the type 

of clay mineral present. Karickhoff and Brown (1978) 

observed sorption coefficients for paraquat to be correlated with the cation-exchange capacity of different 

particle-size fractions of a natural sediment. Zierath and 

others (1980) similarly noted that benzidine sorption 

increased directly with particle surface area. 

Because sorption of organic bases, such as the 

bipyridylium herbicides, is due primarily to an adsorption 

process, solute solubility and sediment organic-matter 

content may not be sufficient parameters to predict 

sorption coefficients. As described above, the solute's p~ 

and sediment factors, such as cation exchange capacity, 

surface area, or the charge or atomic radius of the 

saturating inorganic cation, also may be important in 

determining a sorption coefficient. System parameters, 

such as pH and ionic strength (Narine and Guy, 1982), 

may also influence sorption. For these reasons, predictive 

sorption modeling is more complex for organic bases than 

for organic acids or nonionic organic compounds. However, the environmentally significant compounds that 

10 Manmade Organic Compounds In the Surface Waters of the United States 
exhibit the greatest tendency to sorb to surface-water 

sediments are the hydrophobic, nonionic organic compounds, such as the organochlorine insecticides and 

PCB's. 

Bioaccumulation 

Bioaccumulation (or bioconcentration) is the uptake 

of organic compounds by biota from either water or food. 

Many toxic organic chemicals attain concentrations in 

biota several orders of magnitude greater than their 

aqueous concentrations, and, therefore, bioaccumulation 

poses a serious threat to both the biota of surface waters 

and the humans that feed on these surface-water species. 

The extent to which an organic compound tends to 

accumulate in an aquatic organism is quantified by the 

bioconcentration factor (BCF). Chiou (1981) defines 

BCF as the ratio of the concentration of a contaminant in 

an organism or the tissue of an organism to the concentration in water. BCF also can be defined as the ratio of an 

organism's accumulation rate to its depuration rate (Neely 

and others, 1974). BCF's have been applied to quantify 

organic-contaminant concentrations in a variety of 

surface-water organisms. Agarwal and others (1982) documented the accumulation of DDT isomers by the freshwater ciliate Tetrahymena pyrifonnis. Barker (1984) noted 

the bioconcentration of organochlorine compounds at 

four different trophic levels in the Schuylkill River (Pa.) 

Belluck and Felsot ( 1981) observed pesticide bioaccumulation by caddisfly egg masses. Birmingham and Colman 

(1983) documented diquat accumulation by freshwater 

plants. Microorganisms also have been shown to accumulate organic compounds (Grimes and Morrison, 1975; 

Paris and others, 1977; Blair and others, 1982). Additionally, numerous examples of organic-compound bioaccumulation have been documented for fish and shellfish 

(Biddinger and Gloss, 1984). 

Importance of Biological Lipid Reservoirs 

Hamelink and others (1971) first proposed that the 

uptake of hydrophobic compounds by fish is accomplished primarily by exchange equilibria between the 

water and blood followed by storage in the fish's lipid 

reservoirs. Since then, researchers have investigated the 

importance of lipid-water partitioning with regard to 

bioaccumulation. 

Hansch and others (1968) implied that the bioaccumulation of organic contaminants from water is comparable to the partitioning of the contaminant between 

water and an organic solvent such as octanol. In surfacewater systems, hydrophobic organic contaminants will 

partition themselves between water and the lipid reservoirs of biological organisms, with solute solubility in 

water being a primary factor affecting the magnitude of 

the BCF for any given organism. However, unlike a 

solvent -water system, biological lipid-water systems do not 

attain rapid equilibrium. Slow solute diffusion to the lipid 

reservoir, combined with metabolic breakdown or aqueous phase transformation processes, can delay or even 

prevent the attainment of an equilibrium condition (Chiou, 

1981). Generally, stable organic compounds having low 

aqueous solubility exhibit the longest equilibration times. 

Sanders and Chandler (1972) did not observe an equilibrium condition for PCB uptake from water by crayfish 

even after 21 d of exposure. Similarly, DDT uptake by 

goldfish did not result in steady-state concentrations even 

after 50 d (Rhead and Perkins, 1984). Oliver and Niimi 

(1983) monitored the bioconcentration of a variety of 

chlorinated benzenes by rainbow trout and noted that, as 

compound solubility decreased, the time to reach equilibration generally increased. Dichlorobenzene isomers 

attained equilibrium after 10 d, but pentachlorobenzene 

did not reach a steady-state concentration after 119 d. The 

tri- and tetrachlorobenzenes reached equilibrium after 

about 15 and 50 d, respectively. For the more soluble 

organic compounds, such as the herbicides alachlor and 

dinoseb, bioaccumulation by fathead minnows has been 

observed to reach equilibrium after only 24 h (Call and 

others, 1984). 

The time required to reach equilibrium has also 

been shown to be a function of fish size. Murphy (1971 ), 

in a laboratory bioaccumulation study using female mos- . 

quito fish, observed that the larger fish accumulated DDT 

from the water at a slower rate than smaller fish exposed 

to the same aqueous concentration. Similar results have 

been observed for the bioaccumulation of dieldrin (Reinert, 1972). 

The hypothesis that bioaccumulation is caused primarily by partition into the lipid reservoirs of aquatic 

organisms is supported by the repeated observation of 

increased BCF's with increased lipid content. Canton and 

others (1977) noted that the greatest accumulation of 

hexachlorocyclohexane occurred in Chlamydomonas cells 

having the greatest lipid contents. Helwig and Hora 

(1983) confirmed the lipophilicity ofPCB's by quantifying 

the compound's elevated concentration in fat tissue relative to meat tissue in Minnesota snapping turtles. Niimi 

(1979) measured the concentrations of hexachlorobenzene (HCB) in lake trout, rainbow trout, and coho 

salmon from Lake Ontario. The mean fat contents of each 

species of fish were proportional to their whole-body 

HCB concentrations. Additionally, HCB concentrations in 

fish samples within each species increased with fish size. 

Similar lipid-content effects were observed by Rudling 

(1970) with pentachlorophenol accumulation, by Ryan 

and others (1984) for dioxin and PCB accumulation, and 

by Seguchi and Asaka (1981) for diazinon accumulation. 

Skea and others (1979) placed fish in submerged cages in 

the Hudson River (N.Y.) to quantify the bioaccumulation 

Processes Affecting the Fate and Distribution of Organic Compounds 11 
u. 

(.) 

m 

Figure 7. Correlation of bioconcentration factor (BCF) versus octanol-water partition coefficient (Kow) 

showing the correlations of Veith and others (1979), Neely and others (197 4), and Mackay (1982a) (labeled 

•this work•). Only solid points are included in the regression. [Reprinted with permission from Environmental 

Science and Technology, v. 16, p. 277 (1982). Copyright 1982 by the American Chemical Society.] 

of PCB's. Again, the highest PCB concentrations were 

found in the largest fish (which possessed the highest lipid 

contents). The results of these studies indicate the importance of lipid content to the bioaccumulation of organic 

compounds. For a given exposure level, organisms having 

the highest lipid contents can be expected to accumulate 

the greatest amount of organic solute on a concentration 

basis. Similarly, the accumulated contaminant will likely 

concentrate itself in an organism's lipid reservoirs. 

The extent to which an organic compound is bioaccumulated by a given aquatic organism depends largely on 

its aqueous solubility. As discussed earlier, decreasing 

solubility correlates with increasing ~w· Because bioaccumulation appears to be a partitioning process, the BCF 

can be expected to vary inversely with solubility. Also, the 

BCF can be expected to vary with ~w such that increasing ~w yields increasing BCF. Examination of the scientific literature reveals that these relations have been 

demonstrated in the laboratory. Mackay (1982a) compiled 

data from a variety of sources to develop figure 7, in 

which the BCF's based on whole-fish mass are used. The 

majority of the data are from the work of Veith and others 

(1979); their plot of log ~w versus log BCF is labeled in 

the figure along with the plot developed by Neely and 

others (1974). By selectively disregarding some of the 

data in figure 7, Mackay (1982a) developed his own 

relation (labeled "this work" in the figure), having the 

equation 

log BCF= log ~w -1.32 

with a correlation coefficient of 0.95. The data of Neely 

and others (1974), not included in the above correlation, 

resulted in the following relation: 

log BCF=0.52 log ~w +0.124 

with a correlation coefficient of 0.948. Casserly and others 

(1983) also correlated Kow with BCF for the green alga 

Selenastrum capricomutum and established the equation 

log BCF=0.46 log ~w +2.36 

with a correlation coefficient of 0.83. Similar correlations 

have been reported by Oliver and Niimi (1983, 1985), 

KOnemann and van Leeuwen (1980), and Southworth and 

others (1978a, 1978b ). Figure 8 presents a plot of BCF as 

a function of aqueous solubility for a variety of pesticides 

12 Manmade Organic Compounds in the Surface Waters of the United States 
10 ------------------~--------~------~~------~------------------

a: •ooo 0 103 t-

(.) 

< 

u. •HCB z 

0 

102 •coT 

1-

< 

a: 

tz 

w Diflubenzuron • (.) 

101 z 

0 

(.) 

Q 

m 

~ 1 0 

.J 

10- ~------~--------~--------~------~--------~--------~------~ 10-3 1 

LOG SOLUBILITY, IN MICROMOLES PER LITER 

Figure 8. Relation between log water solubility of pesticides and log bioconcentration factor (BCF) for Triaenodes 

tardus egg masses. HCB is hexachlorobenzene. [Reprinted with permission from Bulletin of Environmental Contamination and Toxicology, v. 26, p. 304 (1981). Copyright 1981 by Springer-Verlag New York, Inc.] 

accumulated by caddisfly egg masses (Belluck and Felsot, 

1981 ). Again, a linear log-log relationship is assumed and 

defined by the equation 

log BCF=-0.52log S+1.67. 

Other solubility-BCF relations have been demonstrated by 

Chiou and others (1977) and Metcalf and others (1975). 

These relations indicate the dependence of the BCF on 

organic compound solubility and on Kow· The BCF tends 

to increase with decreasing solubility and (or) increasing 

octanol-water partition coefficient. 

Although numerous researchers have successfully 

performed regression analyses for BCF's as a function of 

either solubility or octanol-water partition coefficient, the 

BCF's for a given compound predicted by many of the 

correlation equations are significant I y different, sometimes by orders of magnitude. Chiou (1981, 1985) has 

suggested that these differences are accounted for largely 

by differences in the lipid contents of the aquatic organisms used to develop the different correlations. To account 

for this effect, Chiou (1985) normalized BCF values 

obtained from the work of KOnemann and van Leeuwen 

(1980) and Oliver and Niimi (1983) by dividing them by 

the fraction of lipid present in the organism. Figure 9 

presents a log-log plot of the lipid-normalized BCF's for 

rainbow trout and guppies versus the triolein-water partition coefficients C~w) for a variety of organic compounds. The parameter K1w was used instead of l<ow 

because triolein more closely resembles the fish lipid than 

does octanol. The data are well correlated, and the 

regression equation takes the form 

log BCF=0.957 log ~w +0.245 

with n= 18 and a correlation coefficient=0.915. Unlike 

previous regressions, no systematic differences between 

the two fish species are evident. Just as sorption coefficients can be estimated from sediment organic-matter 

content and a compound's solubility, or ~w' the above 

correlation allows the prediction of the BCF from only a 

knowledge of the organism's lipid content and the 

compound's ~w· Chiou (1985) also related Krw to Kow• so 

the BCF can also be determined from the compound's l<ow 

Processes Affecting the Fate and Distribution of Organic Compounds 13 
&L • Rainbow trout 

0 

m 

CJ ~ Guppies 

0 

.J 

r£ 5 

0 

1-

0 

< IL 

z 

0 

i= 

< a: 

~ 4 w 

0 

z 

0 

0 

0 

m 

TRIOLEIN-WATER PARTITION COEFFICIENT, LOG Ktw 

Figure 9. Correlation of log bioconcentration factor (BCF) 

on the basis of fish lipid content with log triolein-water 

partition coefficient (~) for a variety of nonionic organic 

compounds. Vertical limits of bars for rainbow trout are the 

minimum and maximum BCF's reported. (Reprinted with 

permission from Environmental Science and Technology, v. 

19, p. 61 (1985). Copyright 1984 by the American Chemical 

Society.] 

using the following equation: 

log BCF=0.893 log ~w+0.607 

with n=18 and a correlation coefficient=0.904. 

Differences between predicted and actual BCF's 

can arise because of a variety of factors (Chiou, 1985). 

The BCFs determined in figure 9 are assumed to be at an 

equilibrium condition. If a compound is unstable in water 

or is metabolized at a rate greater than the rate of 

equilibration with the organism's lipid reservoir, actual 

BCF's will be lower than predicted values. Also, partitioning of an organic compound into the nonlipid phases of an 

organism may be significant for compounds of high 

aqueous solubility. Third, dissolved organic matter may 

enhance the solubility of organic compounds, which in 

turn would result in lower apparent BCF's than predicted 

by a compound's ~w· Finally, Niimi and McFadden 

(1982) and Oliver and Niimi (1983) have observed that 

BCFs increase slightly with increasing solute concentration (although Isensee and Jones (1975) and Canton and 

others (1977) did not observe variance of BCF's with the 

solute's aqueous concentration). Despite these possible 

sources of variation in the prediction of BCF's, consideration of the lipid contents of different aquatic organisms 

will likely result in better comparability of BCFs. 

Biomag nification 

There has been considerable debate as to whether 

bioaccumulation of organic compounds in aqueous systems is based on the physicochemical properties of the 

compound acting in combination with the physiology of 

the aquatic organism or is attributable to the trophic 

transfer of these compounds through the food chain 

(Biddinger and Gloss, 1984 ). The latter concept of 

aqueous-system biomagnification, or the accumulation of 

organic compounds through trophic transfer, is not well 

substantiated by the scientific literature. 

Previous assumptions of biomagnification have been 

based on data that neglected to account for physiological 

differences between organisms. In most cases, reports of 

aqueous system biomagnification were confused with the 

concept of bioaccumulation resulting from the uptake of 

organic chemicals directly from water. To determine if 

biomagnification is an important process in aqueous 

systems, recent studies have been conducted to isolate the 

relative importance of food and water to the uptake of 

organic compounds by aquatic organisms. The results of 

these studies have indicated that the uptake of contaminants through food has only a small effect on the accumulation of organic contaminants by aquatic organisms 

relative to uptake resulting from exchange equilibria with 

water. 

Canton and others (1975) studied the bioaccumulation of hexachlorocyclohexane (HCH) using the following 

food chain combinations: algae-daphnia, daphnia-guppy, 

and guppy-trout. When compared with contamination by 

water alone, the transfer of HCH through the food chain 

was minimal, and the authors concluded that uptake of 

HCH from water was much more important than uptake 

through food. In a study of atrazine uptake by freshwater 

mollusks and fish, Gunkel and Streit (1980) illustrated 

that the feeding of contaminated food to the organisms 

did not increase their equilibrium whole-body atrazine 

concentration above the level obtained by exposure to 

. contaminated water only. Additionally, by noting that the 

highest BCF's were occurring in the organs of the fish 

that had high blood circulation, they postulated that 

atrazine is taken in by the gills and transported in the 

blood to the organs. They went on to use tritiated-waterbalance calculations to determine that the swallowing of 

water and the exchange of water through the skin are of 

little importance to either the fish's water balance or its 

accumulation of atrazine. Niimi and Cho (1983) analyzed 

fish from Lake Ontario representing two trophic levels 

and noted that pentachlorophenol (PCP) levels in the fish 

from the different trophic levels were not significantly 

different. In an earlier study, Niimi and McFadden (1982) 

14 Manmade Organic Compounds in the Surface Waters of the United States 
illustrated that rainbow trout exposed to aqueous PCP 

concentrations of 35 ng!L and 660 ng!L accumulated 9 p.g 

and 115 p.g of PCP, respectively, after an exposure period 

of 115 d. Although the fish's diet had a PCP concentration 

of 3 p.glkg, the authors estimated that no more than 2 p.g 

of PCP were accumulated through diet. Southworth and 

others (1979) also confirmed the importance of the 

uptake of organic contaminants from water. They showed 

that the uptake rates of acridine by freshwater fish via 

ingestion of contaminated invertebrates (0.02 p.glglh) or 

by ingestion of contaminated sediment (0.01 p.glglh) were 

insignificant relative to the rate of direct uptake from 

water (1.40 p.glglh). 

In further support of the hypothesis that aqueous 

solute concentrations control the extent of bioaccumulation are the data of Canton and others (1977) and Paris 

and others (1977). The former study determined that the 

accumulation of HCH by the marine algae Chlamydomonas was the same for both living and dead cells. The 

study of Paris and others (1977) reported the same results 

for the bioaccumulation of toxaphene by a variety of 

aquatic microorganisms. The metabolic processes of the 

aquatic organisms, therefore, had no measurable effect 

on the equilibria established between the water and the 

lipid reservoirs of the microbes. 

Some laboratory studies have noted that the uptake 

of organic contaminants from food has significantly contributed to the bioaccumulation of the compound in the 

aquatic organism. However, many of these studies did not 

allow an equilibrium condition (essentially zero change of 

contaminant concentration in the aquatic organism with 

time) to be established. Rhead and Perkins (1984) found 

that either dietary exposure or solution exposure of DDT 

to goldfish could be the dominant source of accumulation, depending on their relative concentrations. However, 

equilibrium conditions were apparently not obtained in 

any of their experiments. Therefore, their experiments 

demonstrated only that dietary exposure to DDT can 

significantly enhance the rate of DDT bioaccumulation. It 

did not demonstrate that dietary exposure could increase 

the magnitude of an equilibrium BCF determined from 

water exposure only. In a similar study, Shannon (1977) 

observed the relative importance of food versus water 

uptake to the bioaccumulation of dieldrin by channel 

catfish. Catfish exposed to 2 mg!kg of dieldrin through 

their diets combined with exposure to 75 ng!L of aqueous 

dieldrin accumulated significantly more of the compound 

than catfish exposed only to aqueous dieldrin. Here again, 

however, an equilibrium condition was not attained in 

either of the exposure experiments. Skea and others 

(1981) fed brook trout a mirex-contaminated diet for 100 

days and noted a substantial accumulation of the compound in the fish. Equilibrium conditions were not 

attained, and no comparison was made with mirex exposure through water only. These studies indicate that prior 

to equilibrium, uptake of organic contaminants by aquatic 

organisms can occur through both dietary and aqueous 

exposures, but they do not support the concept of 

biomagnification. 

In some cases, increases in BCF's are mistakenly 

attributed to biomagnification when, in fact, they are 

caused by differences in lipid content. In a study of four 

different trophic levels in the Schuylkill River (Pa. ), 

Barker (1984) reported PCB levels of 56 p.g!kg in the 

primary producers, 125 p.glkg in the primary consumers, 

and 650 to 700 p.g!kg in the secondary and tertiary 

consumers. Although data such as these commonly are 

attributed to biomagnification, Barker attributes the variations primarily to differences in the lipid contents of the 

trophic levels along with other physiological factors such 

as age, sex, sex cycle, and diet. Of the aquatic organisms 

examined by Barker, the highest organochlorine residues 

were detected in the American eels, which typically have 

very high lipid contents relative to other aquatic organisms. 

The studies discussed above indicate that the popular concept of biomagnification in aquatic systems is not 

well supported by the scientific literature. Future studies 

need to take into account physiological differences 

between aquatic species (and even different specimens), 

and BCF's should be compared only if the values are 

obtained at an equilibrium condition. The available data 

suggest that accumulation of organic contaminants by 

aquatic organisms is caused primarily by solute partitioning between water and the lipid reservoirs of the organIsm. 

Transformation Processes 

This section reviews the following four transformation processes that remove organic contaminants from 

aqueous solution: 

1. Photolysis, 

2. Hydrolysis, 

3. Biodegradation, and 

4. Volatilization. 

Each process attenuates the concentration of an organic 

compound in solution by a different chemical, physical, or 

biological mechanism. The magnitude of the attenuation 

depends on the combination of physicochemical properties of the compound and the environmental conditions of 

the water body in which the compound is present. The 

removal mechanism, as well as the environmental and 

compound-specific properties influencing the rate of 

removal, are reviewed for each of the four processes. 

In addition to detailing the environmental and 

physicochemical factors influencing the removal of contaminants by these four processes, this section also examines the kinetics of the removal processes. Each transformation process can be represented, mathematically, by a 

Processes Affecting the Fate and Distribution of Organic Compounds 15 
first-order rate law. Other representations, such as higher 

order rate laws, also are reviewed where appropriate. The 

first-order rate law is emphasized because it is the simplest kinetic description and is applicable to all four 

processes. The rate law for each process is described by a 

constant whose magnitude reflects the influence of both 

environmental factors and physicochemical properties. 

Consequently, the rate constant is useful as a relative 

measure for determining which of the four competing 

processes dominates the organic-contaminant removal 

under a specific set of environmental conditions. 

Photolysis 

The study of photochemistry in natural waters is a 

relatively new and emerging field. Only in the last decade 

has the knowledge obtained about atmospheric photochemistry during the last century been extended to the 

photochemical processes in marine and fresh waters 

(Miller, 1983; Zafiriou and others, 1984). The environmental photolysis of organic compounds is a complicated 

series of processes that depends on 

1. Properties of the particular organic molecule, 

2. Characteristics of the water body, and 

3. Input of sunlight. 

This subsection reviews the component factors from these 

three categories that appear to have the greatest influence 

on the process of photolysis in surface-water systems. 

Photolysis is the degradation of compounds by the 

action of light. The aqueous photochemical reactions 

collectively referred to as photolysis can be further categorized as direct and indirect (or "sensitized") photolysis 

(Zepp and Oine, 1977). Direct photolysis is defined as 

the absorption of light by a compound causing a chemical 

reaction involving the energized molecule. The absorbed 

radiation causes the molecule to become excited, and it 

has a tendency to react in order to return to a more stable 

ground-state configuration. The possible resulting reaction mechanisms can include, but are not limited to 

molecular rearrangement, dissociation, and oxidation. ' 

Indirect photolysis refers to those photochemical 

reactions in which an organic compound interacts with 

other previously energized molecules rather than directly 

with solar radiation. Natural waters provide an abundance 

of potential reactive species that can absorb radiation and 

then react with organic contaminants by indirect photolytic mechanisms. These light-absorbing species (called 

chromo ph ores) include humic substances, transitionmetal ions, organic-free radicals, and oxidizing species, 

such as peroxides, singlet oxygen, and ozone (Miller, 

1983). Because it is difficult to separate and isolate the 

dominant photochemical mechanisms in natural waters 

most studies do not attempt to differentiate between th~ different types of photolytic reactions (Callahan and 

others, 1979). Much of the available laboratory data and 

the remainder of the current discussion focuses on the 

physicochemical properties and environmental variables 

that influence the direct photo,ysis of organic contaminants. 

The importance of photolysis to the degradation of 

organic compounds in surface-water systems is measured 

by the photolysis-rate constant, which is a measure of an 

organic compound's photodecomposition over time. The 

magnitude of the photolysis-rate constant depends on the 

amount and wavelength of the solar radiation that reaches 

a particular compound and on the ability of the compound 

to absorb that energy source. Environmental variables 

determine the intensity and spectral distribution of wavelengths of light available to a compound. Physicochemical 

properties, such as the light-absorption coefficient, determine the efficiency of the conversion of the energy into the 

photochemical transformation of the compound. 

The intensity and spectral distribution of solar radiation is attenuated in the atmosphere through absorption 

by gases-for example, ozone-and scattering by molecules and aerosols. The degree of attenuation varies with 

time of day, season, geographic location, and thickness of 

the ozone layer. At the air-water interface, the two sources 

of radiation originating from the atmosphere are direct 

radiation and sky radiation. The latter type of radiation is 

derived ·from the scattering of direct radiation into the 

water body. The radiation reaching the air-water interface 

either is reflected back into the atmosphere (generally 

less than 10 percent) or is refracted into the water body at 

an angle different from the angle of incidence. Once in the 

bulk solution, light is further attenuated by absorption and 

scattering. Light is absorbed by water molecules, dissolved solids, and suspended sediment. Absorption of light 

also occurs as a result of the biological activity of photosynthetic organisms. Attenuation by scattering in the 

water column has a smaller impact than absorption on the 

penetration of light and is generally ignored in obtaining 

first-order approximations of light intensity (Zepp and 

Cline, 1977). 

In addition to the environmental variables, the 

photolysis rate also depends on certain characteristics of 

the organic contaminant. Of the light that reaches the 

compound, only light absorbed by the chemical can cause 

photolytic degradation. The percentage of available light 

absorbed by the contaminant depends on its molar extinction coefficient and its concentration. In other words, the 

presence of the solute changes the value of the absorption 

coefficient of the solution from the value attributable 

solely to pure water. Based on the fundamental law of 

photochemistry, the average photolysis rate, at a specific 

wavelength in a completely mixed water body, is directly 

proportional to the rate of light absorption by the solute 

per unit volume (Zepp and Cline, 1977). To obtain the 

actual, average photolysis rate, this relation is summed 

over all the wavelengths in the ultraviolet- and visible-light 

16 Manmade Organic Compounds in the Surface Waters of the United States 
range and multiplied by a constant. The constant, called 

the quantum yield of the reaction, represents the efficiency of usage of the absorbed radiation by the solute for 

the particular photochemical reaction. From the intensity 

of radiation at a particular depth in a water body, the 

average photolysis rate is estimated from the properties of 

the solute-namely, the molar extinction coefficient, the 

concentration, and the quantum yield for the photochemical reaction. 

The most commonly used representation for a 

direct photolytic process is a first -order kinetic expression, 

d[C]/dt=- ~[C). 

In this expression, the rate of change (d/dt) of a 

compound's concentration (C) (resulting from photolysis) 

is directly proportional to the concentration of the compound multiplied by a first-order rate constant, kP, with 

units of reciprocal time. 

The first-order rate constant is the product of the 

rate constant for the absorption oflight by the chemical, 

~' and the quantum yield (0) (the efficiency of energy 

conversion for the photochemical reaction), or 

~=(~) (0). 

The determination of the absorption-rate constant takes 

into account the environmental variables for a specific 

system, such as time of day, season, latitude, ozone 

thickness, and depth in the water body, as well as the 

absorption characteristics, such as the molar extinction 

coefficient, of the particular compound. 

The influence of the compound's absorption properties on the rate of direct photolysis is illustrated by the 

following example. Zepp and Cline (1977) calculated the 

direct photolysis rates for carbaryl (a carbamate insecticide) and trifluralin (a dinitroaniline herbicide). The 

compounds have similar quantum yields for reactions in 

air-saturated water, but their peak absorption of radiation 

occurs in different parts of the ultraviolet- and visiblelight ranges. Consequently, for the same set of environmental conditions (midday, midsummer, latitude 40° N.), 

the absorption rate for trifluralin is 400 times larger than 

that for carbaryl, and the direct photolysis of trifluralin is 

two orders of magnitude more rapid than that of carbaryl 

(Zepp and Cline, 1977). 

The intensity and spectral distribution of light are as 

important as the absorption properties of the compound 

in determining the photolysis rate. The data of Zepp and 

Cline (1977) illustrate the difference that chemical properties can have on the photolysis rate. This difference is 

not constant, but varies with environmental conditions. 

Different combinations of time of day, season, latitude, 

and ozone thickness will selective] y filter out different 

wavelengths of light. Because the two compounds absorb 

radiation of different wavelengths, this filtering process 

may alter the intensity of usable radiation available to the 

compounds. Thus, for the same set of environmental 

conditions, two compounds can photolyze at different 

rates because of differences in the energy content of the 

range of solar radiation (the specific set of wavelengths) 

they absorb. 

Using the first-order rate equation for direct photolysis, an expression can be derived for a compound's 

half life. The half life (originally devised to measure the 

decay of radioactive material) refers to the time required 

for one-half of the initial concentration of a compound to 

be removed solely by direct photolysis. The half-life 

expression for a first-order, direct photolysis process is 

tv~ =0.693~, 

and it is dependent on only the quantum yield and the 

absorption coefficient for a particular compound. Compounds that have high absorption coefficients and high 

quantum yields will have low half-life values, indicating 

rapid removal of the compound by photochemical processes. Because the half-life concept provides an indication of the rate of reaction in units of time (as opposed to 

rate constants in units of reciprocal time), it is a tangible 

measure for comparing photolysis with other first-order 

rate processes, such as hydrolysis, volatilization, and biodegradation. 

Although estimates of the photolytic half-lives of 

certain organic chemicals account for many environmental and compound-specific properties, additional factors 

should be considered to improve the approximation of the 

photolytic degradation of an organic contaminant. For 

example, cloud cover can reduce the intensity of ultraviolet radiation by as much as 50 percent and thereby alter 

the degradation kinetics of an organic compound. In 

contrast, variations in the elevation of a stream within a 

geographical area can increase light intensity by 15-20 

percent for every kilometer increase in elevation. Also, the 

thickness of the ozone layer can vary daily as much as 30 

percent, causing changes in atmospheric scattering and 

absorption. 

The influence of other factors on the photolysis 

process has not been well quantified. For example, it has 

yet to be determined how varying concentrations of 

chromophores, such as dissolved and particulate organic 

matter, affect the rate and extent of indirect photolysis 

reactions. In many studies, the effect of light scattering in 

the water column is ignored, but it may be significant in 

turbid lakes and rivers. Finally, the mixing of water 

continually changes the temporal and spatial distribution 

of available sunlight and chemical concentrations. Current photolysis models still assume a spatially homogeProcesses Affecting the Fate and Distribution of Organic Compounds 17 
neous distribution of the contaminant throughout the 

depth of the water column. 

Although laboratory protocols and kinetic representations for direct photolysis have been fairly well 

established, similar procedures for indirect photolysis are 

only beginning to be understood. Indirect photolysis is a 

much more complicated phenomenon, and it is mediated 

by any one of a variety of natural compounds. The 

chemical kinetics are represented by a second-order process where the rate of removal depends on both the 

compound's concentration and the concentration of the 

reactive intermediate (U.S. Environmental Protection 

Agency, 1979b ). The difficulty encountered in applying 

this kinetic approach is in experimentally measuring the 

concentrations of the reactive intermediates, which frequently have half-lives on the order of microseconds or 

less (Zafiriou and others, 1984). Nevertheless, Miller 

(1983) states that the contribution of indirect photolytic 

processes to the overall photolysis rate may be significant. 

The increasing ability of researchers to isolate, monitor, 

and measure the chromophores present in natural waters 

(for example, humic substances, transition metal ions, and 

free radical and other oxidizing species) will likely result 

in an increased understanding of both indirect and direct 

photolysis of anthropogenic organic compounds in 

surface-water systems. 

Hydrolysis 

Hydrolysis is a transformation process that changes 

the chemical speciation of an organic contaminant. Like 

photolysis, biodegradation, and volatilization, the extent of 

contaminant removal by hydrolysis depends on both the 

chemical properties of the organic compound and the 

aqueous medium. Also, like other transformation processes, the hydrolysis of organic contaminants is mathematically represented by a first-order kinetic process. 

This subsection summarizes some of the equations and 

physicochemical factors influencing the rate and extent of 

the hydrolysis reaction for different classes of organic 

compounds. 

Hydrolysis is the reaction of a compound with water 

resulting in an exchange of some functional group, x-, 

from the organic molecule with a hydroxyl group, OH-, 

from the aqueous medium. The overall mechanism can be 

represented by the following reaction sequence 

RX+H2 0-+ROH+HX, 

where RX is the organic molecule and X is a particular 

functional group. The individual reaction steps involve the 

formation of one or more intermediates which combine to 

transform the compound by the simplified mechanism 

given above. On the basis of this mechanism, the hydrolysis rate is expressed by a first-order rate law such that 

where 

-d[RX]/dt=Kh[RX], 

-d[RX]/dt=rate of hydrolysis of compound RX 

(mass volume - 1 time - 1 ), 

Kh=hydrolysis rate (time- 1), and 

[RX]=concentration of compound RX (mass 

volume- 1 ). 

In a manner similar to other transformation processes, the rate of contaminant removal by hydrolysis 

depends on the chemical structure of the compound, the 

influence of pH, temperature, and ionic strength, and the 

presence of other compounds, such as humic substances. 

The contribution of each factor to the overall rate of 

hydrolysis is accommodated by modification of the firstorder rate law, given above, to explicitly account for the 

additional variable. For example, the rate of hydrolysis 

depends on the pH of the aqueous solution. Depending on 

the particular organic compound, the hydrolysis rate is 

increased by the addition of acid or base to the system. In 

this case, the rate expression is rewritten to reflect the 

explicit dependence of the reaction rate on the concentration of acid, [H+], and base, [OH-]: 

-d[RX]/dt=Ka[H+][RX]+ l<o[OH-][RX]+ ~[RX] 

=Kh[RX], 

where Ka and~ are second-order rate constants (mass 

volume- 1 time- 1 ). This relation shows that the rate of 

hydrolysis depends on the product of the acid or base 

concentration and the contaminant concentration. Kn is 

the first-order rate constant for the pH-independent reaction. 

Incorporation of the effects of acid- or basepromoted catalysis increases the order of the overall 

hydrolysis expression from first to second order. The rate 

now depends, explicitly, on the concentration of two 

components, [RX] and [OH-] or [H+]. Rearranging the 

rate equation to express it solely in terms of [H+] is done 

using the expression for the dissociation of water, 

~<w= [OH-][H+], and by canceling the [RX] in each term. 

From the above equation, this gives 

From this expression, it is apparent that at elevated pH, 

the overall rate constant is controlled by the first term on 

the right side of the equation (assuming ~ 1<w is sufficiently large). Under acidic conditions, the elevated H+ 

concentration makes the second term on the right side 

dominate in the determination of the overall rate constant 

(assuming Ka is sufficiently large). At neutral pH, the last 

18 Manmade Organic Compounds In the Surface Waters of the United States 
term on the right side controls the rate of reaction. Thus, 

for a particular reaction, the rate depends on the pH of the 

aqueous medium as well as on the relative magnitude of 

the three rate constants, ~' ~' and ~<u. Because the 

order of the reaction varies with pH (second order for 

acid or base conditions and first order at neutral pH), the 

overall hydrolysis rate constant, Kh, is sometimes termed 

a pseudo-first-order constant (Mabey and Mill, 1978). It 

is the observed or estimated rate constant for hydrolysis at 

a constant pH. 

In addition to pH, other factors influence the rate of 

hydrolysis of organic contaminants; these include temperature, ionic strength of the solution, and presence of other 

natural organic compounds, such as humic substances. In 

general, the rate of hydrolysis increases with temperature, 

and this observation follows the commonly cited Arrhenius equation. This equation relates the rate constant of a 

chemical reaction to the absolute temperature of the 

system and, for a specific reaction, takes the form 

k= A exp [-E/RT], 

where k is the rate constant, A is a constant, E is the 

activation energy, R is the universal gas constant, and T is 

the absolute temperature. 

The presence of dissolved ions in solution also can 

increase or decrease the rate of hydrolysis. The extent of 

this effect depends on the organic contaminant and specific ions and their concentrations. In some natural waters, 

the total concentration of dissolved anions and cations is 

low (less than 0.01 molar) and the influence of ionic 

strength on the rate of reaction is small (Mabey and Mill, 

1978). In other types of waters, such as brackish, estuarine, and marine, the significance of the effect increases 

with dissolved-ion concentration. 

Many of the recent investigations of dissolved 

organic matter, including humic substances, show that 

these substances collectively play an important role in the 

kinetics of contaminant hydrolysis. Humic substances can 

increase or decrease the rate of hydrolysis depending on 

the concentration of organic matter and the pH of the 

solution (Purdue, 1983). In general, humic substances will 

inhibit base-catalyzed hydrolysis reactions (for example, 

2,4-D ester; Purdue, 1983) and accelerate acid-catalyzed 

hydrolysis reactions (for example, atrazine; Li and 

Felbeck, 1972). The ability of humic substances to partition certain organic contaminants affects the rate of 

hydrolysis, because the reactivity of the humiccontaminant mixture differs from that of the free dissolved contaminant in solution (Purdue, 1983). Consequently, whether the contaminant is free or partitioned 

influences the overall rate of hydrolysis as well as the 

extent of catalysis of the reaction based on changes in pH. 

Biodegradation 

Microorganisms decompose organic compounds 

through enzymatically mediated reactions. The driving 

force behind these reactions is the microorganisms' need 

for energy, carbon, and other essential nutrients. Some 

manmade organic compounds can serve as organic substrates which are degraded by enzymes and used for 

cellular growth. Consequently, the degradation of an 

organic contaminant can be monitored by examining the 

growth of a microbial population that uses the compound 

as a substrate. In addition to the chemical structure of the 

organic compound and the type of microbial population, 

certain environmental factors influence the biodegradation process. Among the most important are temperature, 

pH, and amount of oxygen present in the system. These 

factors are discussed in this subsection, along with the 

kinetics of the biodegradation process. 

The rate of biodegradation is a function of the size 

of the microbial population and the concentration of the 

organic contaminant. To achieve a noticeable amount of 

contaminant degradation, a sufficiently large microbial 

population must be present. In addition, for a single 

substrate system, sufficient contaminant must be present 

to result in enough cell growth to sustain the microbial 

population. Under ideal conditions (a large resident population and ample substrate for growth), the rate of 

growth is limited by the cellular metabolic and reproductive systems of the microbial population. 

The rate of biodegradation of an organic contaminant is examined in terms of the rate of microbial growth, 

assuming all of the consumed contaminant is converted 

into mass for biological growth and no other competing 

substrates are available. The relation betWeen growth rate 

and substrate concentration is represented by the Monod 

equation for microbial growth: 

where 

[C] 

J.L=J.Lmax~+[C]' 

J.L =specific growth rate, which represents the 

rate of growth per unit amount of biomass 

(time- 1 ), 

#lmax =maximum specific growth rate (time- 1), 

[C] =contaminant or substrate concentration 

(mass volume - 1 ), and 

~=concentration of contaminant corresponding 

to a growth rate, J.L, equal to J.l.max/2 (mass 

volume- 1 ). 

Figure 10 graphically depicts the relation between substrate concentration and specific growth rate (Benefield 

and Randall, 1980). At low substrate concentrations, the 

specific growth rate increases linearly with an increase in 

substrate concentration. At high substrate concentrations, 

the growth rate approaches and maintains a limiting 

growth rate that is independent of substrate concentration. 

Processes Affecting the Fate and Distribution of Organic Compounds 19 
:1 

w 

1-

<( 

a: 

::I: I-I max 

1-

~ 

0 

a: 

CJ IJmax 

0 2 

u. 

0 

UJ 

0.. 

Cl) 

Ks 

SUBSTRATE CONCENTRATION, S 

Figure 10. Relation between microbial specific growth rate 

p,, and substrate c'oncentration. K5 is the contaminant concentration corresponding to a growth rate equal to one-half 

the maximum specific growth rate. 

Because the rate of biodegradation and consequently the 

rate of microbial growth is the result of many enzymecatalyzed processes, the Monod equation is sometimes 

replaced by the analagous Michaelis-Menten equation for 

enzymatic reactions (Banerjee, Howard, and others, 

1984). The mathematical forms of the Monod and 

Michaelis-Menten equations are the same. The difference 

lies in the scale of observation used to interpret the 

equation parameters and variables. The Monod equation 

applies specifically to microbial growth, whereas the more 

general Michaelis-Menten equation applies to any biochemical process whose rate of substrate reaction is 

catalyzed by enzymes. 

Using the Monod equation for microbial growth, an 

expression for the rate of contaminant removal is formulated in terms of the contaminant concentration and the 

size of the microbial population. This expression is written 

as 

X [C] X [C]X 

-d[C]/dt=J.L y=#Lmax~+[C] y=~ ~+[C], 

where 

X =biomass per unit volume (mass volume- 1 ), 

Y =the growth yield, equal to the biomass produced per unit amount of chemical consumed, and 

~ =biodegradation rate constant (time -l). 

The constants J.Lmax' l<s, and Y depend on the characteristics of the microbes, pH, temperature, and the presence 

of other nutrients. For example, the amount of biomass 

produced for a unit amount of contaminant consumed, Y, 

depends on the ease of assimilation and conversion of that 

substrate into cellular components. Values of Y vary with 

the species of microbes, the chemical structure of the 

contaminant, and the environmental conditions of the 

system. 

The above kinetic equation for the rate of contaminant removal simplifies under a number of specific 

environmental conditions. For example, at high contaminant concentration, [C] > > l<s, the equation reduces to 

the following form: 

-d[C]/dt=~ X. 

In this case, the rate of degradation is first order and 

depends only on biomass concentration. Another possible 

and more likely case is that the contaminant concentration 

is very low, [ C] < < Ks . In this case the equation becomes 

-d[C]/dt=Ko [C] X/~=l<u2 [C] X, 

where 2 is the second-order rate constant. The rate of 

degradation depends on both the contaminant concentration and the biomass population size. A third possibility is 

commun to many surface waters and occurs when the 

biomass population is relatively large and constant. In 

addition, the contaminant concentration is relatively low, 

so that the consumption of contaminant will not radically 

alter the population concentration. These conditions represent the specific case in which microbes are acclimated 

to the chemical and actively use the compound. In this 

case, the degradation rate is termed pseudo-first order 

and is represented mathematically as 

-d[C]/dt=~' [C], 

where 

~I =Ko2 X. 

The above equations show that the two primary 

factors influencing the rate of biodegradation of an 

organic contaminant are microbial-population density 

and concentration of the contaminant (substrate). Different combinations of substrate concentration and population density can significantly alter the shapes of substratedisappearance curves that are functions of time (Simkins 

and Alexander, 1984; Alexander, 1985). Additional factors are embedded within the biodegradation rate constant. These include differences in assimilability between 

contaminants as well as differences between various species of microbes responsible for converting the contaminants into biomass. Microbial populations are sensitive to 

variations in environmental conditions, and different combinations of temperature, pH, and oxygen content of the 

water may provide an advantage to one species competing 

with another species for the same substrate. The contributions of these environmental variables to the rate of 

biodegradation commonly are significant. 

20 Manmade Organic Compounds In the Surface Waters of the United States 
A change in temperature can alter the balance of 

competition between microbial species using a contaminant as substrate. Each class of bacteria has an optimum 

temperature range within which a maximum growth rate 

is attainable assuming a sufficient supply of all other 

essential nutrients. Bacteria are commonly classified on 

the basis of the temperature range within which their 

optimum growth occurs. Psychrophilic bacteria grow best 

at temperatures ranging from -5 to 30 °C. Mesophiles 

grow best at temperatures between 15 and 45 oc. Thermophilic (heat-loving) bacteria grow best at temperatures 

ranging from 45 to 75 °C (Benefield and Randall, 1980). 

Within the temperature range of each group of bacteria, 

the rate of biochemical reaction approximately doubles 

for every increase of 10 °C. However, this relation, known 

as the van't Hoff rule, is valid over only a limited range for 

bacteria, because further increases in temperature eventually break down the cellular growth and reproductive 

mechanisms (Benefield and Randall, 1980). Other, more 

complicated relations describing the influence of temperature on the rate of reaction are based on the Arrhenius 

equation. These relations include correction factors that 

show how temperature affects the microbial-growth and 

decay rate, the dominant organism groups, and the availability of ·nutrients-factors that contribute to alteration 

of the rate of contaminant removal (Novak, 1974). 

Like temperature, the oxygen content of a water 

system contributes to the determination of which class of 

microbes can best degrade a particular organic contaminant. Depending on the mechanism by which microorganisms obtain their energy, different classes of organisms 

can exist under aerobic, anaerobic, or anoxic conditions. 

Thus, the oxygen content of the water-sediment system 

determines which group of microorganisms competes for 

the contaminant as a substrate and, ultimately, the rate of 

the contaminant's degradation. 

In summary, the biodegradation of an organic compound is primarily determined by two components of the 

system: the microorganisms and their substrate (the 

organic contaminant). The microorganisms are sensitive 

and responsive to the condition of their immediate environment, including temperature and oxygen content, as 

well as to the chemical nature of the organic substrate. 

For rapid removal and conversion into biomass, the 

substrate must be readily decomposable and more assimilable than other available organic compounds. For this 

system, the rate of biodegradation depends largely on the 

density of the microbial population and the concentration 

of the contaminant. The reaction dynamics are approximated by a kinetic relation that is limited by the maximum 

growth rate of the microbial population. The order of the 

reaction rate for the degradation of the contaminant 

depends on the abundance of the microorganisms and the 

organic substrate. For many environmental cases, characterized by a stable microbial population and a small 

contaminant concentration, the equation simplifies to a 

pseudo-first-order expression: Only one parameter-the 

biodegradation rate constant-is needed to parameterize 

this kinetic representation. In this approach, the rate 

constant is specific for a single organic compound, a 

particular population of microorganisms, and the temperature and oxygen content characterizing the aqueous 

system. 

Volatilization 

Volatilization and.condensation are two complementary processes contributing to the exchange of organic 

compounds across the air-water interface. Both processes 

involve phase changes of the organic compound, either 

from gas to liquid or from liquid to gas, and are driven by 

concentration gradients between the respective bulk 

phases. Volatilization, in the present context, refers to the 

evaporation and vapor-phase transport of an organic 

compound from aqueous solution into ~e atmosphere. 

This subsection examines only volatilization, even though 

condensation and volatilization of organic compounds are 

controlled by many of the same chemical properties and 

environmental variables. 

A number of compound properties and system 

variables combine to determine the extent of volatilization 

of an organic chemical. The most important chemical 

properties are vapor pressure and water solubility. Commonly, the ratio of these two properties is expressed as the 

Henry's law constant, where the constant, H, is equal to 

vapor pressure divided by water solubility. The Henry's 

law constant represents the equilibrium distribution of 

concentrations of the organic chemical between the air 

and water phases at' equilibrium. Air and water temperatures are important environmental variables controlling 

volatilization. In addition, the degree of air turbulence 

caused by the wind and the amount of subsurface mixing 

generated by velocity gradients in the water have a significant influence on the rate of volatilization. This subsection discusses the relative contributions of these chemical 

and environmental factors and examines some models 

currently used to represent the volatilization process. 

From a physical standpoint, the potential volatility of 

an organic compound in a pure solution is related to its 

vapor pressure. In a natural water body, the actual rate of 

volatilization depends on the extent to which the environmental conditions attenuate the potential volatility of the 

compound. Competing processes, such as sorption with 

water-column particulates, act to suppress the vapor 

pressure attainable if all of the compound remained in a 

dissolved and unassociated state. In general, however, the 

greater the vapor pressure of a compound, the greater its 

mass loss rate (flux). Other factors also influence the 

volatilization rate, but most do so through their effect on 

vapor pressure or on the movement of. the compound 

Processes Affecting the Fate and Distribution of Organic Compounds 21 
away from the air-water interface. Vapor pressure, for 

example, increases with temperature and thereby increases 

the rate of volatilization. In addition, increasing the ternperature also increases the rate of diffusion, which serves 

to enhance the movement of organic molecules away from 

the interfacial boundaries. 

Volatilization is a transformation process that acts 

to establish an equilibrium between the concentrations of 

a compound in the bulk phases of air and water. The 

driving force behind this process is the difference in 

concentrations (more precisely, the chemical potentials) 

of the compound in the two bulk phases. The retarding 

forces that prevent the instantaneous attainment of an 

equilibrium state are those that reduce the chemical 

potential difference of the compound between the two 

bulk phases. The overall resistance can be considered to 

be the sum of the resistance experienced in a different 

part df the entire physical system. The greatest of the 

component resistances serves to control the rate for the 

overall movement of a compound from the water to the 

air. 

If the physical system is assumed to consist of two 

bulk phases--air and water-separated by a stagnant 

near-surface water layer and a stagnant near-surface air 

layer, the component resistances in the system can be 

examined. These include the resistance to diffusion in the 

(1) water column, (2) near-surface water layer, (3) 

surface-film layer (if any), (4) near-surface air layer, and 

(5) air column. Ignoring the presence of surface films 

(such as oil spills) for now, the majority of the resistance 

occurs in the near-surface gas and liquid layers (Smith 

and others, 1980). The magnitude of the resistance in 

these interfacial layers varies with the thickness of the two 

interfacial layers; layer thickness, in turn, depends on 

such factors as temperature, wind and water turbulence, 

and the specific properties of the compound of interestfor example, diffusivity and (or) molecular weight. 

Many investigators use a first -order process as a 

mathematical representation of volatilization (Mackay and 

Leinonen, 1975; Dilling, 1977; Mackay and others, 1979, 

1983; Southworth, 1979; Chiou and others, 1980; Smith 

and others, 1980, 1981, 1983; Rathbun and Tai, 1982, 

1984; Chiou, Kohnert, and others, 1983). For this representation, the rate of removal of a compound from the 

water phase by volatilization depends on the concentration 

. of the compound in solution and on the volatilization rate 

constant. An expression for this relation is 

d[C]/dt= l<v [C], 

where [ C] is the dissolved concentration of the compound 

in solution and Kv is the volatilization rate constant. The 

usefulness of this approach depends on the ability to 

obtain an accurate estimate of the volatilization rate 

constant for a particular compound and a given set of 

environmental conditions. 

Initial attempts to estimate volatilization rates were 

based on the belief that organic compounds codistill with 

water into the atmosphere (Acree and others, 1963; 

Mackay and Wolkoff, 1973). The codistillation hypothesis 

implies that the simultaneous evaporation of water and 

organic compounds somehow enhances the volatility of 

the organic compound to rates not normally expected 

solely on the basis of the chemical's physical properties. 

From a theoretical standpoint, it has been shown that the 

volatilization of organic compounds and water are mechanistically independent processes, and no evidence for a 

thermodynamic coupling has yet been presented (Spencer 

and others, 1973; Chiou and Manes, 1980). In addition, 

Chiou and Manes (1980) have reported experimental 

results demonstrating that, under certain conditions, 

organic solutes evaporate while water vapor condenses. 

The use of the term "codistillation" should be avoided 

because it incorrectly implies that water loss enhances 

volatilization of organics by codistillation and that water 

loss is necessary for volatilization of significant quantities 

of organic compounds (Spencer and others, 1973). 

Estimation of the volatilization-rate constant is 

based on a two-film-layer model because, for many compounds, the resistance to volatilization occurs in either the 

gas or the liquid near-surface layer. This approach is 

based on the two-film, mass-transfer model origina11y 

proposed by Whitman (1923). The model assumes uniformly mixed air and water phases separated by thin films 

of air and water in which transport is by molecular 

diffusion. The model's volatilization-rate constant is 

expressed in terms of the mass-transfer rates of the 

organic compound across the liquid-phase and gas-phase 

boundary layers. The general expression for the volatilization rate constant is 

where Kv is the volatilization-rate constant (h- 1), Lis the 

depth (em), which equals the interfacial area divided by 

the liquid volume, k1 is the liquid-film mass-transfer coefficient (em/h), R is the gas constant (L-torr/K-mol), Tis 

the absolute temperature (K), H is the Henry's law 

constant (torr-L/mol), and kg is the gas-film mass-transfer 

coefficient ( cm/h ). The first term within parentheses 

represents the liquid-film resistance, and the second term 

represents the gas-film resistance. An important assumption behind the two-film-layer model is that the organic 

solute is at equilibrium between the water and air phases 

at the interface (Smith and others, 1983). This equilibrium is described by the Henry's law constant, which is the 

organic-solute concentration in the air (mol/m3

) divided 

by the organic-solute concentration in water. Because 

22 Manmade Organic Compounds in the Surface Waters of the United States 
atmospheric concentrations are frequently expressed in 

partial pressures, P, they can be converted to mol/m3 by 

using the ideal-gas law (nN=P/RT). 

From the above expression, it is seen that the 

volatilization-rate constant depends on the Henry's law 

constant for the organic ~. om pound as well as on the liquid 

and gas mass-transfer coefficients. The Henry's law constant is a function of the compound properties, whereas 

the mass-transfer coefficients are a function of both the 

compound properties and environmental variables. The 

liquid-film, mass-transfer coefficient is expressed as the 

diffusion coefficient (D, in m2

/h) for the given organic 

solute divided by the thickness of the stagnant liquid film. 

The diffusion coefficient in a particular phase is related to 

the molecular weight of the organic solute. Transport in 

the boundary layer is by molecular diffusion, because the 

air and liquid films. are considered stagnant. The thickness 

of the liquid film depends on the level of water turbulence 

in the system. In other words, the thickness of the stagnant 

water layer decreases with increasing water velocities 

(turbulence). The same relations hold true for the gasfilm, mass-transfer coefficient except that they apply to 

the gas phase and air turbulence imparts the windshearing force. Consequently, the rate of volatilization 

depends on the diffusion properties of the organic compound in the respective stagnant film phases. Diffusion, in 

turn, is related to the molecular weight of the organic 

molecule, the concentration or vapor density of the 

organic compound, and temperature, which influences 

both the vapor density and the diffusion process. 

The difficulty in applying the two-film model is 

trying to measure or estimate the liquid and gas masstransfer coefficients accurately in actual water bodies. 

The overall mass-transport rate of a substance may 

depend on liquid-phase and (or) gas-phase resistance, 

based on the relative magnitude of k1 and Hkg. A simplifying approach is to determine, from the two-film equation, which film-phase resistance dominates for particular 

values of the Henry's law constant (Dilling, 1977; Mackay, 

1979; Mackay and others, 1979; Smith and others, 1980, 

1981, 1983). Smith and others (1980) found that, for 

compounds having Henry's law constants greater than 

3,500 torr/molar (4.4X10- 3 atm-m3

/mol), the liquidphase resistance controlled approximately 95 percent of 

the mass transfer for alkanes and benzene. The compounds having constants above the 3,500-torr/molar 

demarcation are considered highly volatile. For compounds having constants below 10 torr/molar (1.2X 10-5 

atm-m3 /mol), the gas-phase resistance controls mass 

transfer, and these compounds are considered to be of low 

volatility. For compounds having Henry's law constants 

between 10 and 3,500 torr/molar, the liquid- and gasphase resistances are significant, and both terms in the 

two-film equation must be included in the calculation of 

the volatilization-rate constant. These compounds are 

considered to have intermediate volatility-for example, 

DDT (Mackay, 1979). These three groupings, based on 

Henry's law constants, provide a convenient way to categorize the potential volatility of organic substances so that 

experimental methods can be devised to measure masstransfer coefficients for the phase (air, water, or both) 

imparting the dominant resistance to volatilization. 

Smith and others (1980) have proposed a method of 

predicting the volatilization rate of highly volatile organic 

compounds in streams on the basis of laboratory measurements. Highly volatile compounds experience resistance to mass transport, primarily in the liquid-film 

phase, whereas resistance in the gas filn1 is considered 

negligible. Consequently, estimation of the volatilizationrate constant simplifies to estimation of only the liquidphase, mass-transfer coefficient. Using oxygen as a highvolatility reference compound, Smith and others (1980) 

used results of previous studies to show that the ratio of 

the liquid-film, mass-transfer coefficient of an organic 

compound to the reaeration coefficient for oxygen is 

constant over a wide range of turbulence levels. If this 

ratio is measured in the laboratory, and if the oxygen 

reaeration constant is measured in an actual water body, 

the corresponding field liquid-phase, mass-transfer coefficient can then be calculated for the organic compound. 

Theoretical justification for the constancy of the masstransfer coefficient ratio for the organic compound and 

oxygen as well as a test of the method of predicting 

volatilization of benzene and six other chlorinated hydrocarbons is reported in Smith and others (1980). 

The rate of mass transfer of low-volatility compounds is controlled by resistance in the gas film, whereas 

resistance in the liquid film is relatively small. Smith and 

others (1981, 1983) suggest the use of water as a reference compound, because its rate of evaporation is controlled entirely by gas-phase, mass-transfer resistance. 

The ratio of the mass-transfer coefficients for water and 

the organic compound is constant, and any effect of 

gas-phase turbulence is equal for the two compounds 

(and therefore is canceled). Based on the ratio, the 

gas-film, mass-transfer coefficient is estimated from field 

measurements of the water evaporation flux, relative 

humidity, and air temperature. Smith and others (1981, 

1983) demonstrated this approach for two organic compounds of low volatility and two organic compounds of 

intermediate volatility. For compounds of intermediate 

volatility, estimates are required for both the gas-film and 

liquid-film, mass-transfer coefficients. These cannot be 

estimated by a single experiment; independent measurements of each coefficient are required. Methods of 

obtaining these estimates are described by Southworth 

(1979) and Smith and others (1981, 1983). 

Despite the fact that the two-film theory is commonly used to model volatilization, it has several important limitations. One of the basic assumptions of the 

Processes Affecting the Fate and Distribution of Organic Compounds 23 
two-film theory is that the surface layer of the solvent 

(water) is stationary and is not evaporating, so that a 

solute has to move to the liquid surface by diffusion to 

evaporate. This assumption applies only to solutes that 

are evaporating much faster than the solvent; it breaks 

down when the loss rate of the solvent is comparable to or 

greater than the loss rate of the solute. For the latter case, 

the solute is not required to diffuse to the liquid surface to 

evaporate, inasmuch as the solvent at the old surface is 

gone. Therefore, even if no diffusion occurred, the solute 

would then be at the newly generated surface and would 

evaporate. The failure of the two-film theory to account 

for solvent evaporation may make the model somewhat 

inaccurate for describing the evaporative behavior of 

solutes that are relatively nonvolatile. Another possible 

weakness of the two-film theory is its emphasis on the 

diffusion of the solute and lack of consideration that the 

solvent also is diffusing. Because the solvent is the dominant component of the mixture, its diffusive motion may 

cause additional movement of the solute to the surface. 

Therefore, the size of the solute molecule may not have a 

strong influence on solute diffusion in water; in other 

words, the solute may be "pushed" by the solvent's 

diffusive motion. Finally, the two-film theory treats the 

concentration gradient of the solute in the liquid film as 

the driving force for evaporation. In fact, the concentration gradient is merely the result of uneven evaporative 

loss rates between solute and solvent and should not be 

treated as the driving force for evaporation of the solute. 

Chiou and others (1980; Chiou, Kohnert, and others, 1983) have presented an alternative first -order model 

for volatilization that explicitly accounts for the effects of 

solute concentration and system turbulence. The model 

differs from the previously described two-film approach 

in that it explicitly treats the influence of air turbulence 

and subsurface mixing on the evaporative behavior of 

organic solutes of different volatility. In addition, although 

the two-film model gives only relative resistances, the 

model proposed by Chiou and others (1980; Chiou, 

Kohnert, and others, 1983) gives relative as well as 

absolute resistances. The absolute liquid-phase and airphase resistances are based on pure liquid and vacuum as 

the reference states. 

Chiou and others (1980; Chiou, Kohnert, and others, 1983) report that for compounds that have small 

Henry's law constants (low volatility and gas-phasecontrolled diffusion), the rate of volatilization is enhanced 

by air turbulence but not by subsurface mixing. On the 

other hand, the volatilization rate of volatile solutes 

increases with air turbulence and subsurface mixing. The 

degree of enhancement by liquid mixing is a function of 

the magnitude of the Henry's law constant. In addition, 

Chiou, Kohnert, and others (1983) examined the influence of system conditions on the ratio of mass-transfer 

coefficients for different organic-solute combinations in 

water. They found that the ratios of mass-transfer coefficients for organic solutes are constant with respect to 

changes in air turbulence and water mixing if the organic 

compounds have similar or low Henry's law constants. 

The ratios are not constant for compounds that have 

vastly different or high Henry's law constants. These 

results show that careful examination of compound properties and system conditions is necessary before reference 

compounds, such as oxygen, can be used to estimate 

volatilization rates for organic compounds in natural 

waters. 

A number of statistical approaches have been proposed to estimate mass-transfer coefficients (either liquid 

or gas, or both) of organic compounds from field properties that characterize the degree of air and water 

turbulence. For example, Rathbun and Tai (1982) measured liquid-film, mass-transfer coefficients for ethylene 

and propane and correlated their magnitudes with a 

number of hydraulic and geometric properties of streams 

(stream energy, depth of flow, average velocity, and stream 

slope). The regression equations derived for ethylene and 

propane are then used to estimate mass-transfer coefficients for other highly volatile organic compounds,,after 

the proper adjustment is made on the basis of molecular 

diffusivity, molecular weight, or molecular diameter. 

Mackay and Yeun (1983) used a similar statistical 

approach but chose for their correlation studies stream 

and compound properties different from those used by 

Rathbun and Tai. Mackay and Yeun (1983) developed 

separate regression equations for the gas- and liquid-film, 

mass-transfer coefficients. They used friction velocity as 

the hydrodynamic parameter and the Schmidt number to 

characterize differences between organic solutes. The 

Schmidt number (viscosity divided by density multiplied 

by diffusivity) incorporates the solute-diffusion properties 

as wen as a temperature dependence, whereas the friction 

velocity is related to the windspeed. The correlations were 

found to yield reasonable results for 11 organic compounds under a number of windspeed conditions. 

Although these correlations do allow estimates of masstransfer coefficients necessary to apply the two-film 

model to the process of volatilization, they also show that 

more research needs to be conducted to increase understanding of factors that control the mass-transfer coefficients in the gas and liquid films. 

In summary, the current approach to representing 

the volatilization of organic compounds from surfacewater bodies is through the use of a first-order-rate 

expression. In this form, the rate of volatilization depends 

on the concentration of the compound in solution and on 

the volatilization-rate constant. The most common way to 

represent the volatilization-rate constant is by the two-film 

resistance model. This model is parameterized by the 

Henry's law constant and a gas-film and liquid-film, 

mass-transfer coefficient. In some cases, depending on the 

24 Manmade Organic Compounds In the Surface Waters of the United States 
magnitude of the Henry's law constant of the compound 

of interest, the resistance to volatilization can be approximated by the gas-film resistance and (or) the liquid-film 

resistance. Generally, the Henry's law constant is obtainable from independent measurem€nts of the organic 

compound's water solubility and vapor pressure. Vapor 

pressures can be estimated from correlation equations 

based on boiling or melting points (Mackay and others, 

1982) or by experimental procedures. Burkhard and 

others (1985) review and assess the predictive ability of 11 

methods as they are applied to 15 chlorinated biphenyl 

isomers. Because of the large number of existing 

approaches to estimating the Henry's law constant, the 

usefulness of the two-film model depends on obtaining 

values for the gas- and liquid-film, mass-transfer coefficients. 

The difficulty in estimating the gas- and liquid-film, 

mass-transfer coefficients stems from their dependence 

on system conditions (such as temperature, wind turbulence, and water mixing) and compound properties 

(such as diffusivity and molecular weight). Statistical 

approaches have tried to correlate mass-transfer coefficients with system variables for organic compounds for 

which some experimental data are available. These correlation equations are extrapolated to other organic solutes of similar volatility by making corrections on the basis 

of diffusivity, molecular weight, or molecular diameter. 

Experimental approaches derive estimates by the use of 

reference compounds-for example, oxygen, carbon dioxide, radon, or other organic solutes-whose variation in 

parameter values is constant relative to the organic compound of interest over a particular range of conditions. 

These studies indicate that no entirely satisfactory 

method has been developed to estimate the different 

mass-transfer resistances (gas or liquid) over the full 

range pf volatilities exhibited by the organic solutes. 

Although the two-film model is a promising 

approach to quantifying the volatilization rate of organic 

compounds, a number of other factors still need to be 

incorporated in the model. The presence of natural or 

manmade organic films (or surface-active agents) creates 

another layer that acts to further retard volatilization. This 

surface film possesses a characteristic thickness and a 

resistance to diffusion that has to be considered in addition to the air- and water-boundary layers already treated 

in the two-film model (Smith and others, 1980; Mackay, 

1982b ). The two-film model gives the theoretical maximum volatilization rate for a given set of environmental 

conditions. Other water and atmospheric phenomena 

(such as stratification and aerosol formation) as well as 

competing processes (such as sorption, photolysis, hydrolysis, and biodegradation) reduce the amount of organic 

solute available for volatilization. For some of these 

processes, relative comparisons of first-order rate conDDT ODE DOD 

Cl 

~ Cl 

Cl 

~ Cl 

O Cl ~CI Cl Cl 

Cl 

Aldrin 

~~~~ Clr~~CI 

Cl Cl 

Lindane 

Mirex 

Cl 

Dieldrin 

CH~CI CH2CI 

Cl Cl 

Cl Cl 

CHCJ 2 Cl 

Toxaphene 

Cl 

Z

CI 

Cl 

Cl 

Cl 

Cl PCB 

Cl Cl 

Cl 

Chlordane 

CI-6-CI 

ca¥c• 

Cl 

HCB 

Figure 11. Structural diagrams of selected chlorinated 

insecticides and a polychlorinated biphenyl (PCB). HCB is 

hexachlorobenzene. 

slants can be used to determine the transformations that 

control the aquatic fate of an organic compound. 

POLYCHLORINATED BIPHENYLS AND 

CHLORINATED INSECTICIDES 

Uses, Production, Properties, and 

Regulatory Limits 

Polychlorinated biphenyls (PCB's) and chlorinated 

insecticides have been grouped together in this section 

because of their similar behavior in surface-water systems. As a group, they are the most persistent anthropogenic organic compounds introduced into the environment since the lead and arsenic pesticides of the early 

1900's. Although the use of many of these contaminants 

has been banned in the United States for more than a 

decade, significant residues contaminate the sediment, 

water, and biota of the country's surface waters. Figure 11 

shows the structural diagrams of 2,2',4,4',5,5'-PCB and 

some of the more common chlorinated insecticides. 

PCB's are synthetically produced by the catalyzed 

chlorination of biphenyl. Because of their chemical and 

thermal stability, inertness, and dielectric nature, they 

Polychlorinated Biphenyls and Chlorinated Insecticides 25 
Table 3. Approximate molecular composition of some Aroclor mixtures 

[nd, not detected. From Hutzinger and others, 1974] 

Percentage in each Aroclor mixture 

Empirical Molecular Percent 

foonula weight chlorine 1016 

c1f1o 154.21 0 <0.1 

C1-JI9Cl 188.65 18.79 1 

crJis0 2 223.10 31.77 20 

C1:f7°3 257.54 41.30 57 

C1f6°4 291.99 48.56 21 

clZI5°s 326.43 54.30 1 

C1.Jll:l.6 360.88 58.93 < .1 

Clt13Cl7 395.32 62.77 nd 

------ Average molecular weight 257.9 

have been used as plasticizers, as hydraulic lubricants in 

gas turbines and vacuum pumps, in heat-transfer systems, 

and as dielectric fluids in electrical capacitors and transformers. Aroclor1 is the trade name for one manufacturer's PCB's. Aroclors are mixtures of a variety of 

polychlorinated bi- and terphenyls that are identified by a 

four-digit number. The first two numbers refer to the type 

of molecule primarily used in the mixture (12- refers to 

biphenyl, 54- refers to terphenyl), and the second two 

numbers specify the percent chlorine (by weight) in the 

mixture. Aroclor 1016 was marketed more recently in 

response to the environmental problems associated with 

the persistence of the other Aroclor mixtures. This product contains predominantly mono-, di-, and trichlorobiphenyl isomers. The approximate molecular compositions of 

Aroclor mixtures are given in table 3. Additionally, some 

uses of PCB's suggested by the manufacturer are detailed 

in table 4. 

PCB's were first synthesized in 1929, with production escalating in the United States during the 1960's. 

Production reached a peak in 1970, when 38,600 t were 

produced (Moore and Ramamoorthy, 1984). The manufacturer voluntarily decreased and eventually ceased pro1 Use of trade names in this report is for identification purposes 

only and does not constitute endorsement by the U.S. Geological 

Sutvey. 

1221 1232 1242 1248 1254 1260 

11 <0.1 <0.1 nd <0.1 nd 

51 31 1 nd < .1 nd 

32 24 16 2 .5 nd 

4 28 49 18 1 nd 

2 12 25 40 21 1 

< .5 4 8 36 48 12 

nd < .1 1 4 23 38 

nd nd < .1 nd 6 41 

200.7 232.2 266.5 299.5 328.4 375.7 

duction of PCB's during the 1970's. Because of the 

environmental persistence and toxicity of PCB's, their 

production, sale, distribution, and disposal were first 

regulated under the provisions of the Toxic Substances 

Control Act of 1976. In the same year, the U.S. Environmental Protection Agency (USEPA) set 0.001 p.g!L as the 

water-quality criterion for PCB's to protect freshwater 

and marine aquatic life (U.S. Environmental Protection 

Agency, 1976). In 1977, PCB use was restricted by the 

USEPA, and in 1979, all manufacture, sale, and distribution of PCB's in the United States was prohibited (Haley, 

1984). Despite the cessation of production, PCB's are 

estimated to be present in 2 million transformers and 2.8 

million capacitors currently in use in the United States 

(Moore and Ramamoorthy, 1984). 

The use of organochlorine insecticides began with 

·the discovery of the insecticidal properties of DDT by 

Paul Muller in 1939. In that year, DDT was chosen as the 

most promising new synthetic organic insecticide. Muller 

was awarded the Nobel Prize for his discovery 9 yr later 

(Reutergardh, 1980). Technical-grade DDT contains 

approximately 80 percent p,p'-DDT and 15-20 percent 

o,p'-DDT (Verschueren, 1983 ). The metabolicdegradation products of p,p'-DDT are p,p'-DDE and 

p,p'-DDD, and their residues are commonly found in the 

environment. 

Following the discovery of DDT, numerous other 

organochlorine insecticides were developed and used. 

26 Manmade Organic Compounds in the Surface Waters of the United States 
Table 4. Some uses of polychlorinated biphenyls suggested by the manufacturer 

[From Moore and Ramamoorthy, 1984) 

Base material 

Type of Aroelor 

used and its 

percent content 

in final product 

Purpose of Aroclor use 

and beneficial result 

Polyvinyl chloride 1248, 1254, 1260 

Polyvinyl acetate 1221, 1232, 1242 

Polyester resins 1260 (10-20%) 

Polystyrene 1221 (2%) 

Epoxy resins 1221, 1248 (20%) 

Styrene-butadiene 1254 (8%) 

eo-polymer 

Neoprene 1268 (40%) 

(1. 5%) 

Crepe rubber 1262 (5-50%) 

Nitrocellulose 1262 (7%) 

lacquers 

Ethylene vinyl 1254 (41%) 

acetates 

Chlorinated rubber 1254 (5-10%) 

Varnish 1260 (25% of oil) 

Yax 1262 (5%) 

Lindane, the gamma isomer of hexachlorocyclohexane 

(HCH), was found to be an effective insecticide in 1942, 

followed by chlordane in 1945, and heptachlor, aldrin, 

dieldrin, and toxaphene in 1948. Toxaphene is produced 

by the chlorination of terpene derivatives and is a mixture 

of more than 170 components. Endrin and endosulfan 

were first introduced in the 1950's. Mirex was first 

prepared in 1946 but was not used extensively until1969, 

when it was used to control fire ants in the Southern 

United States. The degradation product of mirex is 

kepone. Methoxychlor also became widely used in 1969 in 

response to the ban of DDT in the United States. In the 

past 10 yr, it has been estimated that the quantities of 

insecticides applied has decreased owing to the use of 

more potent insecticides, such as the pyrethroids permethrin and fenvalerate, which require much lower application rates than older organochlorine insecticides, such as 

toxaphene (Gilliom, 1985). 

The production of organochlorine insecticides in 

the United States has been greatly curtailed over the last 

15 yr owing to their environmental persistence, toxic 

(7-8%) 

(11%) 

Secondary plasticizers to 

increase flame retardance 

and chemical resistance 

Improved quick-track and 

fiber-tear properties 

Stronger fiberglass; 

reinforced resins and 

economical fire retardants 

Plasticizer 

Increased resistance to 

oxidation and chemical 

attack; better adhesive 

properties 

Better chemical resistance 

Fire retardant 

injection moldings 

Plasticizer in paints 

Co-plasticizer 

Pressure-sensitive adhesives 

Enhances resistance, flame 

retardance, electrical 

insulation properties 

Improved water.and alkali 

resistance 

Improved moisture and flame 

resistance 

effects, and lipophilic nature. DDT and aldrin/dieldrin 

use peaked in the mid-1960's at annual values of 70,000 

and 9,000 t, respectively. By 1970, these values had 

decreased to 36,000 and 5,000 t, respectively. Currently 

(1987), production of these compounds in the United 

States is negligible. The use of other organochlorine 

insecticides currently applied to crops (such as toxaphene) is being restricted (Moore and Ramamoorthy, 1984). 

Additional use data for these and other chlorinated 

insecticides, along with detection limits and water-quality 

criteria, are presented in table 5. 

Environmental Fate 

Organochlorine insecticides and PCB's are characterized by relatively low aqueous solubilities and high 

octanol-water partition coefficients (table 6). Because low 

solute solubility correlates with high sediment -water sorption coefficients for nonionic organic compounds, it is 

expected that PCB's and organochlorine insecticides 

would strongly partition into sediment organic matter and 

Polychlorinated Biphenyls and Chlorinated Insecticides 27 
Table 5. Detection limits, water-quality criteria, and use data for selected organochlorine insecticides 

[p.g/L, micrograms per liter; lb/yr, pounds per year. From Gilliom, 1985] 

Water-Quality Total use, 

Criteria Principal National use on farms 1981 

Detection ~i!Slll uses and ~milli2n lb,!:r:rl (million limit H~n Aquat1c sources lb/yr) Chemical (~-&91L> 1 health2 life2 1966 1971 1976 19823 

Aldrin 0.01 0.0007 0.002 Corn 15 7.9 0.9 nr 0.8 

Chlordane .15 .005 .004 Corn, termites, 0.5 1.9 nr nr 9.6 

general purpose (Most farm uses cancelled in 1974) 

DOD .05 .0002 .001 Fruits and vegetables, 2.9 0.2 nr nr 0 

degradation product (Cancelled in 1972) of DDT • 

DOE • 03 .0002 .001 Degradation product nr nr nr nr 0 

of DDT and DOD. 

DDT .05 .0002 .001 Cotton, fruits, 27 0.1 nr nr 0 

vegetables, general (Cancelled in 1972) 

purpose. 

Dieldrin .03 .0007 .002 Termite control, de~· 0.7 0.3 nr nr 0 

radation product o (Most farm uses cancelled in 1974) aldrin • 

Endrfn .05 • 1* .002 Cotton, wheat 0.6 1.4 0.8 nr 0.3 

Heptachlor .01 .003 .004 De~radation product 1.5 1.2 0.6 nr 2.0 

epoxide o heptachlor' used on corn and term1tes • 

Lindane .01 • 4* 0.08 Livestock, seed 0.7 0.7 0.2 nr 0.8 

100* .o3* 

treatment, general purpose. 

Methoxychlor .10 Livestock, alfalfa, 2.6 3.0 3.8 0.6 5.0 

general purpose. 

Toxaphene .25 .007 .013 Cotton, livestock 35 37 33 5.9 16 

nr = none report&:l 1. Detection limits are for water samples. Bed-sediment detection limits are 10 times greater and in units of micrograms 

per kilogram (lucas and others, 1980). 2. Except for values marked with an asterisk, which are from u.s. Environmental Protection Agency (1977), all criteria are from U.S. Environmental Protection Agency (1980b). The human health criteria from the 1980 publication represent 

the average concentrations associated with an incremental increase in cancer risk of 10·5 (one additional cancer per 

1001 000 people) over a lff~time of exposure. The aquatic life criteria are for freshwater environments and represent 24·nour average concentrat1ons. 

3. Data for 1966, Eichers and others (1970); for 1971~ Andrilenas (1974); for 1976, Efchers and others (1978); for 1982, U.S. DeP.Brtment of Agriculture (1983). Data for 19o2 do not include use on livestock or use in California, Colorado, 

Connect1cut, Ma!ne, ~assachusett~, Nevada, New Hampshire, New Jersey, New Mexico, Oregon, Rhode Island, Utah, Vermont, West V1rg1n1a, and Wyom1ng. 

dissolved organic matter. Laboratory and field studies 

have corroborated this hypothesis. 

Table 7 presents some partition coefficients that 

have been determined for PCB's and chlorinated insecticides in laboratory and field studies. The values for the 

distribution coefficient, ~, and the distribution coefficients normalized for organic carbon, ~c' or organic 

matter, ~m' are very high, especially for DDT, DDE, and 

the PCB's. These high sorption coefficients indicate the 

strong tendency of these compounds to partition into the 

sediment organic matter. Even lindane, one of the more 

soluble of the organochlorine insecticides, has been documented as having ~m values greater than 800 (Sharom 

and others, 1980a). 

As a result of their high sorption coefficients, 

insecticides and PCB's having low dissolved concentrations in water (in the parts-per-trillion range) can still 

have very high concentrations in sediment (up to the 

parts-per-million range). In an Oklahoma stream, Hunter 

and others (1980) determined that PCB concentrations in 

sediment ranged from 0.23 to 7.2 ppm, yet the concentration of PCB's in the water was undetectable. Likewise, at 

one location in the Hudson River (N.Y), Nadeau and 

Davis (1976) detected PCB concentrations in sediment of 

almost 3,000 ppm, even though the aqueous concentration 

in the overlying water was only 3.0 ppb. These studies 

indicate the importance of sorption to the transport, fate, 

and distribution of organochlorine insecticides and PCB's 

in surface-water systems. 

In addition to their affinity for sediment organic 

matter, these chlorinated organic compounds also tend to 

partition strongly into the lipid reservoirs of aquatic 

organisms. The resulting concentration of the compound 

in the biota can be as much as 1 million times greater than 

the compound's aqueous concentration. Table 8 presents 

bioconcentration factors (BCF's) for a variety of PCB's 

28 Manmade Organic Compounds In the Surface Waters of the United States 
Table 6. Experimentally determined octanol-water partition coefficients (i<ow), solubilities, 

and vapor pressures of selected polychlorinated biphenyls (PCB's) and chlorinated 

insecticides at room temperature 

[mg!L, milligrams per liter; mm, millimeters. Dash indicates no data available] 

Compound Log K Solubility Vapor pressure Reference 

ow (mg/L) (mm of Hg) 

Biphenyl 4.09 7.5 1.8 1,2,3 

2,4'-PCB 5.10 0.28 4,3 

4,4'-PCB 5.58 _s 

0.062 8.99 X 10 5,2,3 

2,5,2',5'-PCB 5.81 _3 

0.016 4.27 X 10 5,2,3 

2,4,5,2'5'-PCB 6.11 0.010 5,2 

2,4,5,2',4',5'-PCB 6.72 -· 

p,p'-DDT 6.36 

o,p' -DDT 

0.00095 4.57 X 10 5,2,3 

_7 

0.0031 1.9 X 10 1,2,6 

p,p' -DDD 

o,p' ·DDD 

p,p' -DDE 

o,p' -DOE 

Lindane 

Aldrin 

Dieldrin 

Chlordane 

Toxaphene 

Endrin 

Heptachlor 

a-Endosulfan 

5.99 

6.08 

5.69 

5.78 

3.70 

5.48 * 

0.026 

0.020 

0.040 

7.87 

0.017 

0.195 

0.056 

0.74 

0.26 

0.056 

0.53 

.s 5.5 X 10 6 

_7 

10.2 X 10 6 

_7 

18.9 X 10 6 

_e 

6.5 X 10 6 

_e 

6.2 X 10 6 _ .. 1.6 X 10 7,7,6 

_8 

6.0 X 10 6 

.7 

1. 8 X 10 6,2 

_5 

1.0 X 10 8,6,6 

0.2 - 0.4 6 

_7 

2 X 10 6 

3 X 10 -· 6 

_s 

1 X 10 6 

* - Value calculated from structural parameters according to Mabey and 

others (1982) 

1. Chiou, Schmedding, and Manes (1982) 5. Chiou and others (1985) 

2. Verschueren (1983) 

3. Burkhard and others (1985) 

4. Chiou, Porter, and Schmedding 

and chlorinated insecticides and for a variety of aquatic 

organisms ranging from microorganisms to fish. Differences in the BCF values among different compounds 

result largely from differences in each compound's solubility in water. For a given compound, the BCF can be 

expected to vary with the lipid contents of the different 

organisms. BCF's greater than 10,000 have been observed, 

and even the more soluble insecticides such as lindane 

have been bioconcentrated to values hundreds of times 

greater than the compound's aqueous concentration. 

Therefore, even low aqueous concentrations of organochlorine insecticides and PCB's can produce environmentally significant concentrations of. these compounds in 

aquatic organisms. 

6. Callahan and others (1979) 

7. Chiou and others (1986) 

(1983) 8. Mabey and others (1982) 

In addition to their typically low solubilities and 

high octanol-water partition coefficients, PCB's and chlorinated insecticides also are characterized by relatively 

low vapor pressures (table 6). Compounds that have low 

vapor pressures typically have low mass-loss rates (fluxes) 

from solution. However, PCB's and some of the chlorinated insecticides in water are relatively volatile, because 

their low water solubilities cause relatively high rates of 

loss from water. The rate of volatilization for a given 

compound depends on such factors as suspendedsediment concentration and the compound's organicmatter content, the concentration of dissolved organic 

matter, temperature, and water turbulence. Increased 

concentrations of organic matter result in increased sorpPolychlorinated Biphenyls and Chlorinated Insecticides 29 
Table 7. Experimentally determined sorption coefficients for some polychlorinated biphenyls (PCB's) and organochlorine insecticides 

[OC, organic carbon; OM, organic matter; Kci. distribution coefficient; Koc and Kom•· distribution 

coefficients normalized for organic carbon and organic matter, respectively. Dash indicates no data available] 

Compound Soil/sediment 

2,4,4'-PCB Silt loam 

2,5,2'-PCB Suspended river 

sediment 

Hexachloro- Lake Michigan 

biphenyl sediment 

Hexachloro- Suspended river 

biphenyl sediment 

Aroclor 1254 Lake Michigan 

sediment 

DDT 

DDT 

p,p' -DDE 

Lindane 

Lindane 

a-chlordane 

{J-chlordane 

Endrin 

Kepone 

Marine sediment 

Soil samples 

Suspended river 

sediment 

Silt loam 

Creek sediment 

Suspended river 

sediment 

Suspended river 

sediment 

Sand 

Estuarine sediment 

Methoxychlor Clay 

1. Chiou, Porter, and Schmedding 

2. Oliver and Charlton (1984) 

3. Voice and others (1983) 

4. Voice and Weber (1985) 

5. Pierce and others (1974) 

OC or OM (%) Kd Koc or Kom Reference 

1. 9 (om) 

4.1 (oc) 

2.9 (oc) 

4.1 (oc) 

1. 7 (oc) 

2.7 (oc) 

-(om) 

4.1 (oc) 

1. 9 (om) 

2.8 (om) 

4.1 (oc) 

4.1 (oc) 

0.7 (om) 

1. 2 (oc) 

(1983) 6. 

7. 

8. 

9. 

10. 

460 

10,000 

9,000 

13,000 

7,000 

24,000 

250,000 

310,000 

300,000 

410,000 

48,000 1,800,000 

140,000 

41,000 1,000,000 

14 

24 

13,000 

10,000 

58 

1,700 

1,100 

740 

860 

300,000 

250,000 

8,300 

90,000 

Chiou and others (1979) 

Chiou and others (1985) 

Sharom and others (1980a) 

Huggett and others (1980) 

Karickhoff and others (1979) 

1 

2 

3 

2 

4 

5 

6 

2 

7 

8 

2 

2 

8 

9 

10 

tion and, consequently, decreased volatilization (Haque 

and others, 1974; Callahan and others, 1979). Increases in 

temperature or turbulence, however, will increase the rate 

of volatilization. 

in the lakes were due to atmospheric deposition of the 

contaminants. Similarly, atmospheric deposition has been 

considered the primary source of PCB contamination to 

the Great Lakes (Murphy and Rzeszutko, 1977; Eisenreich and others, 1979). Haines (1983) found residues of a 

variety of organochlorine insecticides in brook trout taken 

from six remote lakes in New England. Because the lakes 

were surrounded by forested land that had not been 

treated with pesticides for at least 80 yr, it was concluded 

that the contaminant residues were atmospherically 

deposited. Volatilization, atmospheric transport, and subsequent deposition, therefore, partly explain the numerous detections of PCB's and organochlorine insecticides 

around the globe. 

Volatilization from solution and other atmospheric 

discharges (such as incineration) combine with atmospheric transport to contribute to the widespread distribution of PCB's and some of the chlorinated pesticides. 

Munson (1976) detected both toxaphene and PCB's in 

rainwater collected at Sollers Point, Md. Heit and others 

(1984) have determined PCB concentrations up to 194 

p.g!kg in the sediments of remote Rocky Mountain lakes 

that had not been directly affected by point source 

discharges of PCB's. They concluded that PCB's present 

30 Manmade Organic Compounds in the Surface Waters of the United States 
Table 8. Experimentally determined bioconcentration factors for some polychlorinated 

biphenyls (PCB's) and organochlorine insecticides 

Aquatic Bioconcentration 

Compound organism factor Reference 

PCB Fish 12,000 Metcalf and others (1975) 

PCB Largemouth bass 2,800 Martell and others (1975) 

Aroclor 1254 Guppies 200 Gooch and Hamdy (1983) 

4-PCB Brown trout 4,025 Sugiura and others (1979) 

DDE Fish 12,000 Metcalf and others (1975) 

trans-chlordane B. Subti1is 2,300 Grimes and Morrison (1975) 

cis-chlordane B. Subtilis 2,000 Grimes and Morrison (1975) 

a-HCH ~blam:~Z:domonas 310 Canton and others (1977) 

{J-HCH Brown trout 658 Sugiura and others (1979) 

Lindane B. Subti1is 100 Grimes and Morrison (1975) 

Lindane Brown Trout 442 Sugiura and others (1979) 

Toxaphene ~. Subt111s 3,400 Paris and others (1977) 

Endrin Algae 140-222 Grant (1976) 

Dieldrin B. Subtilis 

Heptachlor B. Subtilis 

Epoxide 

In addition to their ability to partition into organic 

matter and biological lipid reservoirs and their tendency to 

volatilize and be transported atmospherically, the chlorinated insecticides and PCB's are, in general, highly 

resistant to chemical or biological transformation. As a 

result, they are extremely persistent environmental contaminants whose residues commonly remain years after 

use at the source of contamination has ceased. 

The biodegradability of PCB's decreases with 

increased chlorination of the molecule, with the more 

highly chlorinated biphenyls being extremely resistant to 

biodegradation (Sayler and others, 1977). The biotransformation of PCB's is enhanced by an increase in the 

number of C-H bonds available for hydroxylation by 

microbial enzymes (Haley, 1984). Increased chlorination 

decreases the number of C-H bonds and, therefore, 

inhibits biodegradation. PC~'s also are extremely resistant to oxidation and to both acidic and basic hydrolysis 

(Callahan and others, 1979). Neither oxidation nor hydrolysis are likely to transform PCB's significantly in the 

1,200 Grimes and Morrison (1975) 

800 Grimes and Morrison (1975) 

environment. There is some evidence that PCB's can be 

partially dechlorinated by photolysis, particularly when 

exposed to short-wave ultraviolet light. Bunce and Kumar 

(1978) calculated that highly chlorinated biphenyl molecules would lose one chlorine molecule per year in 

shallow waters. In the same 1-yr period, however, they 

predicted that only 5 percent of the lesser chlorinated 

biphenyls would lose a chlorine molecule photolytically. 

Although both biodegradation and photolysis transform 

PCB's at a very slow rate, they may be the most significant 

long-term transformation processes because of the negligible degradative affects of hydrolysis and oxidation. 

The chlorinated insecticides also are relatively resistant to biodegradation. Aldrin has been shown to be 

biotransformed to dieldrin, but this product is strongly 

resistant to further degradation (Callahan and others, 

1979). Likewise, DDT has been biotransformed to both 

DDD and DDE in natural systems, with the latter compound being highly resistant to continued biodegradation. 

Microbial decomposition of DDT occurs primarily 

Polychlorinated Biphenyls and Chlorinated Insecticides 31 
through DDD formation by reductive dechlorination under 

anaerobic conditions (Johnsen, 1976). Other chlorinated 

insecticides such as chlordane, endosulfan, heptachlor, 

toxaphene, and mirex do not appear to biodegrade at 

significant rates in natural surface-water systems (Callahan and others, 1979; Aslanzadeh and Hedrick, 1985). 

Lindane is one of the few chlorinated insecticides that has 

been shown to biodegrade at significant rates in natural 

systems, with documented degradation products being 

penta- and tetrachloro-1-cyclohexanes and penta- and 

tetrachlorobenzenes (Moore and Ramamoorthy, 1984). 

Oxidation of the organochlorine insecticides is not 

an important factor affecting their fate, with the possible 

exception of the oxidation of aldrin to dieldrin in the 

presence of oxygen or ozone (Callahan and others, 1979). 

Hydrolytic half-lives for the chlorinated insecticides typically are greater than 1 yr, with the exception of heptachlor, which is r~idly hydrolyzed to 1-hydroxychlordene. 

Photolytic reactions degrade the chlorinated insecticides 

in laboratory experiments, and this process may be important to the long-term fate of these compounds. Moore and 

Ramamoorthy (1984) state that sunlight reaching the 

Earth may have enough energy to cleave aromatic C-CI 

bonds or an aliphatic alcohol's C-H bonds but not aromatic C-C and H-OH bonds. 

Environmental Distribution 

Because of the tendency of PCB's and chlorinated 

insecticides to partition into sediment organic matter and 

biological lipid resermirs, elevated concentrations of these 

contaminants have been reported in soils, sediments, and 

aquatic organisms throughout the United States and other 

parts of the world. Tables 9 and 10 present some of the 

documented residues of PCB's and organochlorine insecticides in sediments and biota, respectively. The highest 

sediment PCB residues have been observed in the Hudson 

River (N.Y.) near the former discharges of two capacitormanufacturing plants. It has been estimated that more 

than 150,000 kg of PCB's were discharged to the Hudson 

River from these two plants over a 30-yr period ending in 

1976 (Brown and others, 1985). Sediment residues have 

been reported to be greater than 2,000 mg!kg (Nadeau 

and Davis, 1976). The sediments of Los Angeles Harbor 

have also been grossly contaminated with high concentrations of organochlorine compounds, including dieldrin 

(1,800 ~g!kg), total DDT (770 ~g!kg), and PCB's (140 

~g!kg) ( Choi and Chen, 197 6). 

Biota from diverse locations across the United 

States also have been contaminated with organochlorine 

residues (table 10). Again, the Hudson River serves as an 

example of the ''worst-case" PCB contamination, with 

concentrations approaching 100 mg!kg found in the livers 

of Atlantic tomcods (Klauda and others, 1981). Brown 

and others (1985) have documented lipid-based PCB 

concentrations greater than 1,000 mg!kg in the fillets of 

whole yearling pumpkinseeds in the Hudson River. The 

8-yr discharge of kepone into the James River estuary in 

Virginia also caused extreme contamination of biota, with 

concentrations as high as 10 mg!kg in the muscle of white 

perch (Huggett and others, 1980). Concentrations of 

organochlorine insecticides and PCB's commonly are 

highest in American eels, inasmuch as this species has a 

relatively high lipid content (Zitko, 1971; Barker, 1984). 

In contrast to sediment and biological residues, 

aqueous concentrations of PCB's and chlorinated insecticides are typically low, with values typically in the partsper-trillion range. In a study of a Tennessee river, Jaffe 

and others (1982) detected the insecticides hexachlorobenzene, heptachlor, aldrin, chlordane, dieldrin, and endrin 

in water samples at mean concentrations of 5.3, 3.2, 5.1, 

16.7, 0.8, and 1.2 ppt, respectively. Rihan and others 

(1978) similarly detected aqueous concentrations of aldrin 

(0.01 to 0.49 ppt), lindane (0.02 to 0.16 ppt), p,p' -DDT 

(2.73 to 12.31 ppt), and heptachlor (undetectable to 0.18 

ppt) in the surface waters of northern Mississippi. These 

concentrations are typical of contaminated surface waters, 

and their low magnitudes are consistent with the low 

solubilities of the contaminants. 

PCB's and chlorinated insecticides have been shown 

in field studies to be extremely persistent and ubiquitous 

environmental contaminants. Nevertheless, increased 

restrictions on the use and production of these compounds generally have resulted in a slow but steady 

decline of their environmental residues over the last 10 to 

15 yr. 

Several researchers have noted decreases in biological residues of PCB's and chlorinated insecticides over 

time. Wszolek and others (1979) determined that DDE 

concentrations in 12-yr-old fish from Cayuga Lake in 

central New York had decreased from 11.7 to 3 to 4 ppm 

over an 8-yr period. Over the same timespan, PCB 

residues exhibited less of a decline, with concentrations 

decreasing only from 15.7 to 13 ppm. As a result of an 

accidental discharge of PCB's to Escambia Bay, Fla., in 

1969, Wilson and Forester (1978) monitored PCB con-

·centrations in the bay's oysters from 1969 to 1976. Despite 

the fact that no known point-source discharges of PCB's 

contaminated the bay after 1969, residues in the oysters 

showed only a slight decline over the study period, with an 

apparent equilibrium PCB concentration of about 0.5 

ppm recorded in 1976. Sullivan and others (1983) 

reported that PCB concentrations in fish of the lower Fox 

River, Wis., had declined steadily from 1976 to 1981. 

Figure 12 presents some of the temporal trends in PCB 

contamination in water and fish of the Hudson River at a 

site approximately 50 km below two capacitormanufacturing plants that had formerly discharged PCB's 

32 Manmade Organic Compounds In the Surface Waters of the United States 
Table 9. Means and ranges of detectable sediment residues of polychlorinated biphenyls 

(PCB's) and organochlorine insecticides 

[p.glkg, micrograms per kilogram) 

Compound 

PCB 

PCB 

PCB 

PCB 

PCB 

PCB 

PCB 

PCB 

PCB 

Aroclor 1016 

** t-DDT 

t-DDT 

DDT 

DDE 

DDD 

Chlordane 

Chlordane 

Mirex 

Mirex 

Dieldrin 

Dieldrin 

Endrin 

Heptachlor 

Heptachlor 

Epoxide 

Lindane 

a- and .8-HCH 

Locality 

Delaware River 

Los Angeles Harbor 

Lake Ontario 

shoreline marsh 

Rocky Mountain 

National Park 

Southern Puget 

Sound. WA 

Niagara River. 

Lake Ontario 

Rock Reservoir. IA 

Bay of Quinta, 

Lake Ontario 

Raritan Bay, NY 

Hudson River, NY 

Los Angeles Harbor 

Salinas River, CA 

Trinity River, TX 

Rock Reservoir, IA 

Rock Reservoir, IA 

Belmont Lake, NY 

Trinity River, TX 

Lake Ontario 

shoreline marsh 

Niagara River, 

Lake Ontario 

Los Angeles Harbor 

Rock Reservoir, IA 

Trinity River, TX 

Trinity River, TX 

Rockaway River, NJ 

Trinity River, TX 

Niagara River, 

Lake Ontario 

Mean 

(pg/kg) 

123 

140 

0.02 

277 

25.8 

65.7 

13 

385 

212 

2047* 

768 

no data 

16.3 

2.4 

2.4 

171 

13.7 

0.01 

7.1 

1800 

2.9 

9.4 

5.28 

4.5 

0.32 

14.3 

Range 

(pgfkg) 

2-830 

900-4500 

0.002-0.038 

98-540 

0.7-330 

47.7-121.5 

5.7-28.5 

200-620 

3-2035 

6.6-6700 

115-3212 

1-190 

0.25-53.6 

1.1-7.4 

0.6-7.3 

20-580 

0.15-64.0 

no data 

3.9-18 

900-4500 

0.3-6.4 

0.15-20 

0,7;,9.3 

0.7-10 

0.1-0.7 

6.9-33 

Sampling 

method Ref. 

core 1 

grab 2 

grab 3 

grab 4 

no data 5 

suspended 6 

sediment 

core 7 

dredge 8 

grab 9 

grab 10· 

grab 2 

grab 11 

grab 12 core 7 

core 7 

core 13 

grab 12 

grab 3 

suspended 

sediment 

grab 

core 

grab 

grab 

grab 

grab 

suspended 

sediment 

6 

2 

7 

12 

12 

14 

12 

6 

* Concentrations for mean and range in units of milligrams per kilogram ** Denotes total DDT (includes all DDT. DDE, and ODD isomers) 

1. Hochreiter (1982) 8. Scott (1980) 

2. Choi and Chen (1976) 9. Stainken and Rollwagen (1979) 

3. Glooschenko and others (1981) 10. Nadeau and Davis (1976) 

4. Heit and others (1984) 11. Routh (1972) 

5. Mowrer and others (1977) 12. Qasim and others (1980) 

6. Oliver and Charlton (1984) 13. Wood and others (1986) 

7. Ricci and others (1983) 14. Smith and others (1987) 

(Brown and others, 1985). The data demonstrate a distinct decline in aqueous and lipid concentrations of PCB's 

with time. 

Organochlorine residues in sediments have also 

shown a general decline over time. Gilliam (1985) 

reported the results of the U.S. Geological Survey-U.S. 

Environmental Protection Agency Pesticide Monitoring 

Network, which monitored selected pesticides in the water 

and bottom sediments of more than 150 river sites across 

the United Statr:s. The number of detections of chlorinated insecticides in bed sediment decreased steadily 

from 1976 through 1978, but increased slightly in 1979 

(fig. 13). In the monitoring network, 36 station-chemical 

combinations exhibited statistically significant trends in 

pesticide concentration; 29 trends decreased with time. 

Five of the seven increasing trends occurred in the Black 

River in Kingstree, S.C., and most of these trends involved 

either DDD, DDE, DDT, chlordane, or dieldrin. 

Polychlorinated Biphenyls and Chlorinated Insecticides 33 
Table 10. Detectable residues of polychlorinated biphenyls (PCB's) and organochlorine 

insecticides in biota 

[~J.g/kg, micrograms per kilogram. Dash indicates no data available] 

Mean Maximum 

Compound Locality 

Aquatic 

organism (~g/kg) (~gfkg) Reference 

PCB Schuylkill River American eel 3,700 8,100 1 

Reading, PA 

PCB Milwaukee River, Red horse 6,630 2 

WI sucker 

PCB New Bedford Mussel 17,000 3 

Harbor 

PCB Hudson River Atlantic 37,500 98,200 4 

tomcod liver 

PCB New York Harbor Lobster 230 410 5 

PCB Utah Lake Carp 115 6 

PCB Long Island 'Windowpane 1,500 2,300 7 

* Sound flounder liver 

t-DDT Atlantic Ocean Deep sea fish 7,100 27,000 8 

t-DDT Milwaukee River, Red horse 150 2 

'WI sucker 

t-DDT Lake Superior Lake trout 3,250 9,780 9 

p,p' -DDE New England Brook trout 18 34 10 

lakes 

Dieldrin Atlantic Ocean Deep sea fish 20 30 8 

Dieldrin New England Brook trout 4 7 10 

lakes 

a-HCH New England Brook trout 11 19 10 

lakes 

{J-HCH Milwaukee River, Red horse 60 2 

WI sucker 

Lindane Lake Superior Lake trout 190 9 

Heptachlor Lake Superior Lake trout 1,630 2,770 9 

Heptachlor Milwaukee River, Red horse 20 2 

epoxide WI sucker 

Chlordane Schuylkill River, American eel 180 270 1 

Reading, PA 

Endrin Schuylkill River, American eel 9 19 1 

Reading, PA 

Aldrin Lake Superior Lake trout 120 220 9 

Endosulfan Milwaukee River, Red horse 20 2 

WI sucker 

Toxaphene Louisiana lakes Spotted gar 410 800 11 

* Denotes total DDT (includes all DDT, DDE, and DDD isomers) 

1. Barker (1984) 7. Greig and others (1983) 

2. DeVault (1985) 8. Barber and Warlen (1979) 

3. Farrington and others (1983) 9. Parejko and others (1975) 

4. Klauda and others (1981) 10. Haines (1983) 

5. Roberts and others (1982) 11. Winger and Andreasen (1985) 

6. Smith and others (1974) 

PCB contamination trends in sediment generally 

have been constant or have slowly declined over the last 

10 yr. Ricci and others (1983) collected sediment cores 

from an Iowa reservoir and determined that PCB residues 

were greatest in the deepest sections of the core, indicating that the more recently deposited sediments were less 

contaminated with PCB's. They observed similar trends 

for DDE and DDD. Van Luik (1984 ), on the other hand, 

found no correlation between PCB concentration and core 

depth for sediment cores collected from the Calumet Sag 

Channel (Ill.). Sediment PCB concentrations in the 

Hudson River have gradually decreased since the mid1970's, but much of this decrease can be attributed to 

substantial dredging operations that have removed an 

estimated 500,000 m3 of sediment over a 25-yr period 

(Brown and others, 1985). Also, some of the Hudson 

River's contaminated sediment has been transported into 

Raritan Bay and is believed to be accumulating (Stainken 

and Rollwagen, 1979). Sullivan and others (1983), 

although observing a decline in PCB concentrations in 

fish from the Fox River, Wis., observed no significant 

decline in the sediment PCB concentrations in the river 

34 Manmade Organic Compounds in the Surface Waters of the United States 
0 

§ 1000 

lL 

0 

~ 

< 

a: 

CJ 

a: 

w 

CL 

UJ 

;,: 

< 

a: 

CJ 

0 

a: 

Q 

::iE 

~ 

ai 

0 

Q. 

(Sample number) 

I 8S'If. Upper confidence limit Mean 

Median 

9S'If. Lower confidence limit 

(16) 

~~----._--~~----~----~----~~---t~0.8 

1878 

PCB, IN MICROGRAMS PER LITER 

Figure 12. Relation of summer mean polychlorinated biphenyl (PCB) concentrations in water and lipid-based PCB 

concentrations in fillets of whole yearling pumpkinseed 

collected in September 1977-83 at Stillwater, N.Y. 

(Reprinted with permission from Environmental Science and 

Technology, v. 19, p. 659 (1985). Copyright 1985 by the 

American Chemical Society.] 

from 1976 to 1981. These studies indicate the extreme 

persistence of PCBs and also may indicate that, despite 

the cessation of their production, PCB residues in capacitors and transformers throughout the United States may 

still be contaminating surface waters. 

Regionally, the highest PCB concentrations typically are found near industrial areas. Martel and others 

(1975), in a study of a northern Virginia watershed, 

observed that the highest sediment PCB concentrations in 

the watershed were near the most industrialized areas. 

Similarly, Mowrer and others (1977) found the highest 

PCB residues in sediments and mussels of Puget Sound, 

Wash., in industrialized and heavily populated areas and 

the lowest concentrations in sparsley populated areas 

surrounding the bay. Farrington and others (1983) analyzed mussels from 62 locations in the eastern and western coastal waters of the United States from 1976 to 1978. 

In all 3 yr of the study, mussels from water near urbanized 

areas had PCB concentrations at least an order of magnitude greater than mussels from water near rural areas. 

Nationally, PCB concentrations were generally highest 

along the northeastern coast from Chesapeake Bay to 

Cape Ann, Maine. 

Residues of chlorinated insecticides are not necessarily highest in industrialized regions. Gilliom (1985) 

described pesticide use as being "strongly regional, with 

100 

(I) 

z 

0 

i= 80 

(.) 

w 

~UJ 

ww 

Q..J CL 60 lL~ 

Oce 

a:rn 

Wo 

mo ::iE.- 40 ;:)a: Zw 

WCL 

CJ 

< 

a: 20 w 

> 

< 

(). .·.·.·.·.·.;.·.;.·.:···:···: -:-:-:.·.·.·.·.·.·. 

1975 1976 

............... 

.·.·.·.-.·.·.·.·.· .. -.-.-... ·. 

D Bed ·sediments 

EZJ Water 

1100 Detections are 

possible in each 

100 samples 

No data 

for bed 

sediments 

in 1980 

1977 1978 1979 1980 

Figure 13. Frequency of detection of organochlorine insecticides in water and bed-material samples from stations in the 

U.S. Geological Survey-U.S. Environmental Protection 

Agency Pesticide Monitoring Network, 1975-80. (Reprinted 

from Gilliam, 1985.) 

most use of each chemical occurring in only one or two 

regions of the country." Because compounds such as 

DDT and toxaphene are applied primarily in cottongrowing areas, their residues are likely to be highest in 

these areas. Compounds such as aldrin and chlordane are 

applied principally to corn and, therefore, are expected tobe found in corn-producing areas. However, residues in 

industrial areas also can be high, as evidenced by the data 

of Qasim and others (1980). They determined that aqueous pesticide concentrations in the Trinity River, Tex., 

were highest in the central reach of the river, which 

drained cropland, pasture, and forest. However, sediment 

pesticide concentrations were highest in the upper reaches 

of the river, which drains the metropolitan areas of Dallas 

and Fort Worth. The authors attribute the latter contamination to numerous point-source discharges of 

wastewater-treatment plants. When surface-water contamination can be attributed to runoff from cropland, the 

highest concentrations are likely to be observed after a 

heavy rain (Pfaender and others, 1977). 

PCB's and chlorinated insecticides have severely 

affected our environment, and, despite their drastically 

declining use over the last 10 to 20 yr, their residues are 

likely to persist in the sediment, water, and biota of the 

Nation's surface-water systems for many years to come. 

At present, these compounds are being replaced by less 

persistent organic compounds. New insecticides, such as 

the carbamate and organophosphorus insecticides discussed in the next section, have found increasing use. 

PCB's have similarly been replaced by such compounds 

as phenylxylylethanes, di-isopropylnaphthalenes, and 

isopropylbiphenyls (Addison and others, 1983). 

Polychlorinated Biphenyls and Chlorinated Insecticides 35 
CARBAMATE AND 

ORGANOPHOSPHORUS INSECTICIDES 

Uses, Production, Properties, and 

Regulatory Limits 

In response to the environmental concern over the 

persistence of the chlorinated insecticides, the carbamate 

and organophosphorus insecticides have seen increasing 

use over the last 20 yr because of their relatively short 

half-lives in the environment. In addition, they have been 

increasingly recommended for the control of pests that are 

not specifically controlled by the organochlorine insecticides. Figure 14 presents the structural diagrams of some 

of the more common carbamate and organophosphorus 

insecticides. The octanol-water partition coefficients, solubilities, and vapor pressures of some of these compounds are given in table 11. 

Table 12 presents usage data for some of the 

organophosphorus insecticides (Gilliom, 1985). Of the 

group, malathion and parathion have probably seen the 

most widespread use. The use of parathion, although it is 

relatively nonpersistent in the environment, has decreased 

considerably in recent years because of its high toxicity to 

mammals. The compound inhibits the activation of the 

acetylcholinesterase enzyme in the nervous systems of 

animals (Baum, 1982), and its use has resulted in numerous deaths of humans involved in its manufacture and 

application. Decline in the use of parathion is evidenced 

by its production decrease from 50.6 million lb in 1969 to 

15 million lb in 1971 (Mulla and others, 1981). In 

contrast, malathion is the least toxic of the organoCARBAMATE INSECTICIDES 

fHa 

CH1SfCH•NOCONHCH1 

CH1 

Aldie arb 

OCONHCH1 

00 

Carbar,.l 

H,c, 

,NCOC •NOCONHCH1 

H1C ~ 

I 

CH1 

OlC8111)'1 

(l--,_cH1 ~cH1 OCONHCH1 

Carbofuran 

ORGANOPHOSPHORUS INSECTICIDES 

S H 0 II I II 

CH -0-~-S-f-C-O-CH CH 1 ? CH -~-0-CH CH 

CH1 0 

Malathion 

8 8 11 n (C1 H0 0)2-P-8-CH1-8-P(OC2H1)2 

Ethlon 

Dlulnon 

s 

II 

CH1CH2-o-r-O-CH2CH1 

6 

N02 

Parathion 

Fenltrothlon 

Figure 14. Structural diagrams of selected carbamate and 

organophosphorus insecticides. 

phosphorus insecticides and has been used to protect 

livestock, to control fleas in kennels, and to control other 

household and garden pests (Baum, 1982). Its low toxicTable 11. Experimentally determined octanol-water partition coefficients (l<ow), solubilities, and vapor pressures of selected organophosphorus and carbamate insecticides at 

room temperature 

36 

[ mg!L, milligrams per liter; mm, millimeter. Numbers in parentheses indicate references; dashes indicate 

no data available] 

Compound LogK Solubility Vapor pressure 

ow (mg/L) (mm of Hg) 

-~ Malathion 2.89 (1) 145 (1) 1.25Xl0_ 5 (1) 

Parathion 3.81 (1) 24 (1) 0.57 X 10 (1) 

Dimethylparathion 2.04 (2) 55-60 (2) 

Diazinon 40 (3) _6 

Ethion 2 (3) 1.5 X 10_ 5 (2) 

Fenitrothion 3.38 (2) 30 (2) 5.4 X 10 (2) 

Chlorpyrifos 5.11 (2) 0.4 (2) - Aldie -' arb 1.0 X 10 (2) 

Carbaryl 40 (2) <0.005_5 (2) 

Carbo fur an 320 (2) 2.0 X 10 (2) 

1. Mulla and others (1981) 

2. Verschueren (1983) 

3. Gilliom (1985) 

Manmade Organic Compounds In the Surface Waters of the United States 
Table 12. Detection limits, water-quality criteria, and use data for selected organophosphorus insecticides 

[p.g!L, micrograms per liter; lblyr, pounds per year. From Gilliom, 1985] 

Water-quality 3 Total use, criteria Principal National use on farms 1981 

Detection 't:!Slli uses and ~milt ion lbl:tr~ (mill ion 

limit Rumanquat1c sources lb/yr) 

Chemical (1o'g/L)1 heal th2 lf fe 1966 1971 1976 1982 

Diazinon 0.10 ncl ncl Corn, general 5.6 3.2 1.6 0.3 9.0 

purpose 

Ethion 0.25 ncl nd Citrus fruits 2.0 2.3 nr nr 2.0 

Malathion 0.25 nd 0.1 General purpose 5.2 3.6 2.8 1.6 28 

Dimeth~~ 0.25 ncl nd Cotton, wheat 8.0 28. 23. 11. 20 

parat 1on 

Methyl 0.50 nd nd Not identHied nr nr nr nr 0.1 

trhhion 

Parathion 0.25 ncl 0.04 Wheat, corn, sorghum 8.5 9.5 6.6 4.0 5.0 

Trfthton 0.50 nd nd General purpose nr nr nr nr 0.1 

nr • none reported; na • no data ava1lable. 

1. Detection limits are for water sa;gtes. Bed·sediment detection limits are 10 times greater and in units of micrograms ~r kil~rem (lucas end others 19 0). 2. .s. Environmental Protection Agenct (1977). 3. Data for 1966, Efchers and others ( 970); for 1971A Andrilenas (1974); for 1976, Eichers and others (1978); for 1982, U.S. Department of Agriculture (1983). Data for 19 2 do not include use on livestock. 

ity to animals is attributed to carboxyesterases that 

degrade malathion to nontoxic metabolites in birds, mammals, and resistant insects (Mulla and others, 1981). The 

production of malathion in the United States reached a 

high in 1971 of 30 million lb (Mulla and others, 1981). 

Table 13 presents some usage data for the carbamate insecticides. Carbaryl, first synthesized in 1953 and 

introduced in 1958, is the most widely used of the 

carbamate insecticides and has been shown to control 

more than 150 major pests (Rajagopal and others, 1984). 

Environmental Fate 

As is evidenced from the half-lives in table 14, the 

carbamate and organophosphorus insecticides are among 

the least environmentally persistent pesticides. This is 

primarily attributable to their relatively rapid chemical 

and biological degradation in both soil and surface-water 

systems. 

Several researchers have documented the chemical 

and biological attenuation of the organophosphorus insecticides. Walker (1976) determined that 97 percent of 

malathion added to sterile and nonsterile river water was 

degraded after 18 d incubation in the dark. Parathion was 

more persistent, with only a 16- to 23-percent loss after 

40 d. Microorganisms capable of degrading both compounds were isolated. Sudhakar-Barik and Sethunathan 

(1978) noted the biological hydrolysis of parathion in 

soils, water, and sediments and the formation of a degradation product, p-nitrophenol, within 6 h after the second 

or third additions of the insecticide. Sharom and others 

(1980b) found that 40 percent of the ethion and leptophos 

and essentially 100 percent of the diazinon and mevinphos 

added to natural water samples was degraded within 16 

weeks. In a study of malathion degradation in soil, Gibson 

and Burns (1977) observed malathion half-lives of about 

1 d. When the soil's organic matter was removed, however, 

degradation was three to six times slower. The authors 

hypothesized that the organic matter has a catalytic effect 

(by acting as a stable enzyme) on the degradation of 

malathion. Wolfe and others (1977) modeled the chemical 

degradation of malathion as a function of pH and temperature. Acid degradation was slow (malathion half-life of 4 

yr at pH=4), but alkaline degradation was fast enough to 

be a competitive degradation pathway in the environment 

(malathion half-life of 22 hat pH=8.2). Weinberger and 

others (1982) documented both hydrolytic and photolytic 

decomposition of fenitrothion in aquatic microcosms. 

Eidt and others (1984) applied fenitrothion formulations 

to two Canadian streams and observed that the 

compound's aqueous concentration was rapidly attenuated in both streams. These studies indicate that the 

organophosphorus insecticides are generally susceptible 

to rapid chemical and microbial degradation in natural 

systems. 

Numerous studies also have documented the chemical and biological degradation of the carbamate insecticides. Carbaryl is rapidly hydrolyzed at alkaline pH 

(Wolfe and others, 1978), and it has been biologically 

degraded to 1-naphthol in the laboratory (Rajagopal and 

others, 1984). Aldicarb has similarly been observed to 

have hydrolytic half-lives of 55 and 6 d at pH values of 

Carbamate and Organophosphorus Insecticides 37 
Table 13. Characteristics of selected carbamate insecticides 

Compound Type Uses Trade name(s) 

Aldicarb Oxime carbamate Insects, mice, Temik 

nematodes 

Oxamyl Oxime carbamate Insects, nematodes Vydate 

Arprocarb N-methylcarbamate Insects Baygon, Unden, 

(propoxur) Sendran, Suncide, 

Blatanex 

Carbaryl N-methylcarbamate Insects Sevin 

Carbofuran N-methylcarbamate Soil insects Furadan 

Table 14. Approximate persistence of pesticides in soils 

[From Metcalf, 1972) 

Pesticide Approxidate half-life, 

years 

Lead, arsenic, copper, mercury 10-30 

Chlorinated insecticides 2-4 

Triazine herbicides 1-2 

Benzoic acid herbicides 0.2-1 

Urea herbicides 0.3-0.8 

Chlorinated herbicides 0.1-0.4 

Organophosphorus insecticides 0.02-0.2 

Carbamate insecticides 0.02-0.1 

8.85 and 9.95, respectively (Given and Dierberg, 1985). 

Sharom and others (1980b) determined that essentially 

100 percent of the carbofuran added to sterile and 

nonsterile natural water was degraded after 16 weeks 

incubation. Carbaryl was degraded even more rapidly, with 

almost 100 percent removal after only 4 weeks. Comparison of the sterile-water and nonsterile-water degradation 

data revealed that, for both compounds, chemical degradation appeared to be of primary importance to the 

attenuation, with biological processes playing only a secondary role. Figure 15 compares the 14C02 evolved from 

three ring-labeled 14C-insecticides (parathion, carbaryl, 

and carbofuran) applied to flooded and nonflooded soils 

(Rajagopal and others, 1984). These data in combination 

with the results of Sharom and others (1980b) indicate 

that while the initial chemical or biological transformation 

step of carbaryl or carbofuran may be rapid, ring cleavage 

of the molecules may be slow (less than 3 percent of 

t- 15.0 

z 

w 

0 

a: 

w 

0. 

~ 

c 

w 

> ..l 

0 

> 

w 

C'l 

0 

0 • 

/ 

....... 

..... - - -Parathion 

--Flooded 

--- Nonflooded 

/ 

,.- ,., 

/ Carbofuran 

/ -- .,.., ______ ..-

- _ - -; ,- Carbaryl 

-- / 

/ 

/ 

./ 

/ / 

6 12 18 24 30 

DAYS AFTER 14C-INSECTICIDE ADDITION 

Figure 15. Cumulative 14C02 (percentage of added ringlabeled 14C-insecticide) evolved in a soil planted to rice under 

flooded and nonflooded conditions. [Reprinted with permission from Residue Reviews, v. 93, p. 113 {1984). Copyright 

1984 by Springer-Verlag New York, Inc.] 

ring-labeled 14C evolved as 14C02 for either carbaryl or 

carbofuran after 30 d). 

Sorption to sediments and bioaccumulation generally are not significant fate processes for most of the 

carbamate and organophosphorus insecticides because of 

their relatively high solubilities and low octanol-water 

partition coefficients. However, sediments having high 

organic-matter contents or aquatic organisms having high 

lipid contents may accumulate significant residues if 

38 Manmade Organic Compounds In the Surface Waters of the United States 
aqueous concentrations are high. Aquatic plants exposed 

to 10 ppm fenitrothion for 5 d accumulated between 216 

and 286 ppm of the contaminant on a dry-weight basis 

(Weinberger and others, 1982). These residues decreased 

rapidly, however, as the fenitrothion was chemically and 

biologically degraded in the water column. Sharom and 

others (1980a) observed the sorption of a variety of 

organophosphorus and carbamate insecticides to soils 

and sediments and noted that increased sorption correlated with both decreased solubility and increased soil or 

sediment organic matter. The order of sorption to soil was 

leptophos > ethion > chlorpyrifos >parathion > diazinon > 

carbaryl> carbofuran > mevinphos. Sorption coefficients 

(~)for the sediment (2.8 percent organic carbon) ranged 

from 158 (ethion) to less than unity (mevinphos). Belluck 

and Felsot (1981) llleasured BCF's between 1 and 10 for 

carbofuran and malathion partitioned between water and 

caddisfl y egg masses, indicating that bioaccumulation was 

minimal. 

Bioaccumulation and sorption of the carbamate and 

organophosphorus insecticides, therefore, will vary with 

such factors as compound solubility and aqueous concentration, sediment and dissolved organic matter concentration, and lipid contents of the aquatic organisms. Because 

the partitioning of these solutes into sediment organic 

matter and biological lipid reservoirs are reversible processes, biological and sediment residues are likely to be 

short-lived as a result of the rapid chemical and biological 

attenuation of these compounds in surface-water systems. 

Environmental Distribution 

Despite extensive use of carbamate and organophosphorus insecticides in the United States, they are not 

detected frequently in surface waters. Additionally, 

because they have relatively high solubilities (and therefore a low tendency to partition into organic matter and 

biological lipid reservoirs), their residues are even less 

frequently detected in sediment and biota. 

The U.S. Geological Survey-U.S. Environmental 

Protection Agency Pesticide Monitoring Network monitored the occurrence of several of the organophosphorus 

insecticides at 161 surface-water stations throughout the 

United States from 1975 to 1980 (Gilliom and others, 

1985). Compared with chlorinated insecticide residues, 

which also were monitored, organophosphorus insecticides were only occasionally detected in water samples. 

Methyl parathion, the organophosphorus insecticide used 

most heavily on farms, was detected in only 3 of almost 

3,000 water samples. The low number of detections of 

organophosphorus insecticides was attributed to their 

relatively low persistence and high aqueous detection 

limits (table 12). Diazinon was detected most frequently 

(34 detections in 2,861 water samples), and this was likely 

related to the fact that, of the organophosphorus insecticides included in the analysis, diazinon is the most persistent and has the lowest detection limit. For all the 

organophosphorus insecticides monitorerl, the number of 

sediment detections was less than the number of aqueous 

detections, which agrees with laboratory data demonstrating that the organophosphorus insecticides do not strongly 

partition into sediment organic matter. Because of the 

small number of detections of organophosphorus insecticides by the Pesticide Monitoring Network, no regional 

trends were observed for either water or sediment samples. 

No studies documenting regional or temporal trends 

of carbamate-insecticide contamination in the United 

States were found in the reviewed literature. However, 

because the behavior and persistence of the carbamate 

and organophosphorus insecticides are similar, it is 

expected that contamination of surface waters of the 

United States by these compounds will not be widespread, 

and that any existing residues will rarely persist for more 

than several weeks. Carbamate and organophosphorus 

insecticide residues are expected to be highest in shallow 

ground water beneath agricultural areas and in streams 

draining these areas. 

HERBICIDES 

Uses, Production, Properties, and 

Regulatory Limits 

In contrast to the declining use of insecticides in 

recent years, herbicide use on major crops in the United 

States has been steadily increasing, as is evidenced by the 

data in figure 16 (Gilliom and others, 1985). Herbicides 

also are applied extensively to surface waters to control 

aquatic vegetation that may impede water flow, navigation, 

and recreation. The majority of herbicide use in the 

United States can be attributed to atrazine and 2,4-D. 

From 1971 to 1976, they accounted for 50 percent of all 

herbicide use (Gilliom, 1985). Since 1976, however, their 

share of total herbicide application has decreased slightly. 

Atrazine use, for example, declined from 41 percent of the 

total amount of herbicide applied to corn in 1976 to 33 

percent by 1980 (Gilliom, 1985). 

Detection limits, water-quality criteria, and U.S. 

farm usage data for some of the most common herbicides 

are presented in table 15 (Gilliom, 1985). Table 16 lists 

the quantities of herbicides applied in 11 agricultural 

watersheds in Ontario during 1975 (Frank and others, 

1982). Both tables indicate that atrazine and 2,4-D are the 

most heavily applied herbicides in those areas. 

The herbicides are characterized by high aqueous 

solubilities and low vapor pressures (table 17). As a 

result, they generally do riot bioconcentrate, sorb to 

sediments, or volatilize from solution to an appreciable 

Herbicides 39 
en 

0 

z 

::') 

0 

0.. 

z 

0 

:J 

d 

::E 

~ 

0 cc ~ < ...1 w 

0.. > 0.. cc < w 

en 0.. 

1-

z 

~ 

0 

w 

cc 

<!' 

~ 

w 

> 

~ 

0 

<. 

Figure 16. Trends in national use of herbicides and insecticides on major crops, 1964-82. [Reprinted from Gilliom and 

others, 1985) .] · 

extent. Figure 17 presents the structural diagrams of some 

of the more common terrestrial and aquatic herbicides. 

Based on their structural characteristics, the herbicides 

are broken down into four groups. The organochlorine 

herbicides have phenoxy, acetic, or benzoic acid functional 

groups attached to an aromatic ring. The other three 

groups are the nitrogen-substituted, urea-substituted, and 

dinitroaniline herbicides. Table 18 lists the four groups 

Cl 

CI-QOCH2COOH CIQOCH2COOH 

Cl Cl 

2.•-o 2,4,5-T 

[~:·.,- [c•,-0{)>-c•~::,-

Dlquat Paraquat 

0 Cl 

CI~NH-~-N(CH 2 .rQJ_ CH8 

Cl C2 H5NH NH-?H 

Diuron CH8 

Atrazlne 

H7C3 -N-C3 H7 NH2 02

N*N02 CIQ.CI 

Cl 0 COOH 

CF1 Plcloram 

Trlfluralln 

Sllvax 

Figure 17. Structural diagrams of selected herbicides. 

along with some of their associated herbicides. Paraquat 

and diquat, commonly referred to as the bipyridylium 

herbicides, are highly potent salts that exist in solution as 

divalent organic cations with associated chloride or bromide. 

Environmental Fate 

The herbicides are, in general, nonpersistent environmental contaminants. Their high solubilities greatly 

Table 15. Detection limits, water-quality criteria, and use data for some common herbicides 

[ f.lg/L, micrograms per liter; lb/yr, pounds per year. From Gilliom, 1985] 

\later-quaUty Total use, 

Criteria2 Principal Nat,onal use on farms3 1981 

Detection '~glli uses and 'mH lion lb£~n Cmflllon 

limit Rumanquat1c sources lb/yr) 

Chemical (J'g/L) 1 health life 1966 1971 1976 1982 

2,4·0 0.5 100 nd Uheat, rangeland, 4 31 38 23 60 

2,4,5-T 0.5 10 nd 

eeneral purpose R1ce, rangeland, 0.8 nr nr nr 2.2 

Silvex 0.5 general purpose nd nd Sugarcane, rice, nr nr nr nr 0.4 

Atrazine 0.5 nd nd rangeland 

Corn 24 54 90 76 92 

nr • none reported; na • no data avallab{e. 1. Detection limits are for water samples. Bed-sediment detection limits are 10 times greater and in units of micrograms 

per kiloiram (Lucas and others, 1980). 2. U.S. Envtronmentel Protection Agency (1977). 3. Data for 1966, Eichers end others (1970); for 1971, Andrilenas (1974); for 1976, Efchers and others (1978); for 1982, u.s. Department of Agriculture (1983). 

40 Manmade Organic Compounds in the Surface Waters of the United States 
Table 16. List of and magnitude of application of herbicides in 11 agricultural watersheds 

in Ontario, 1975 

[kg, kilograms. From Frank and others, 1982) 

1-10 kg 

Bromacil 

Monuron 

Nitrofen 

Pic lor am 

1 

11-100 kg 

Amitrole 

Cyprazine 

DCPA 

Diallate 

Dicamba 

Dinitramine 

Dinoseb 

Mecoprop 

Monolinuron 

Naptalam 

Pebulate 

Prometon 

Terbacil 

Amounts applied 

101-1000 kg 

Bentazon 

2.4-DB 

D1phenam1d 

EPTC 

Linuron 

MCPA 

Hetribuzin 

Paraquat 

2.4.5-T 

Trif1uralin 

1001-10.000 kg 

1 

A1achlor {2755) 

Butylate {1708) 

Ch1oramben {1979) 

Cyanazine {2468) 

2.4-D (3098) 

Hetobromuron (1012) 

Simazine (1389) 

Values in parentheses are actual amounts applied in kilograms. 

>10.000 kg 

Atrazine 

(10.570) 

Table 17. Experimentally determined aqueous solubilities 

and vapor pressures of selected herbicides at room temperature Table 18. Structural groupings of selected herbicides 

[mg!L, milligrams per liter; mm, millimeter. Dash indicates no data 

available] 

Solubility Vapor pressure 

Compound (mg/L) (am of Hg) Reference 

2.4-D -' 890 6.0 X 10 ., 1,2 

2,4.5-T 278 6.5 X 10-., 1,2 

Atrazine 70 3.0 X 10- 1 

Simazine 5 _., 1 

Honuron 230 5.0 X lO_e 1 

Diuron 42 3.1 X 10_1 1 

Linuron 75 1.1 x 1o_. 1 

Hono1inuron 735 1.5 X 10 I 2 

Propani1 {Stam F-34) 50-225 9.0 X 1(., 1 

Picloralll 430 6.2 X 10_, 1 

Bene fin .1 3.9 X 10_ 1 3 

Profluralin .1 6.9 X 10 4 3 

Trifluralin 4 2.0 X 10- 1 

Silvex 140 1 

Diquat 700.000 1 

Paraquat miscible 1 

1. Verschueren {1983) 

2. Bailey and others (1968) 

3. Helling (1976) 

reduce solute partitioning into sediment organic matter or 

biological lipid reservoirs. Although their low vapor pressures and high solubilities combine to make volatilization 

an insignificant fate process, many of the herbicides are 

rapidly biodegraded by microbial populations or are chemically attenuated by photolytic or hydrolytic reactions. 

Documented sediment-water sorption coefficients 

(~) for the herbicides are small relative to such compounds as the organochlorine insecticides, indicating that 

Organochlorine (phenoxy, 

acetic. and benzoic acid) 

Ami ben 

2,4-D 

Dalapon 

Dicamba 

DNBP 

Fenac 

MCPA 

Methoxy fenac 

Picloram 

Silvex 

2,4,5-T 

2,3,6-TBA 

TCA 

Tricamba 

Organonitrogen (triazines) 

Atrazine 

Alachlor 

Ametryn 

Atraton 

Prometone 

Prometryne 

Propazine 

Secbumeton 

Simazine 

Simetryn 

Terbutryn 

Dinitroaniline 

herbicides 

Bene fin 

Butralin 

Dinitramine 

Ethalfluralin 

Fluchloralin 

Isopropalin 

Nitralin 

Oryzalin 

Penoxalin 

Profluralin 

Prosulfalin 

Trifluralin 

Substituted urea 

Buturon 

Cycluron 

Diuron 

Fenuron 

Linuron 

Monuron 

Neburon 

Norea (Herban) 

Herbicides 41 
sorption generally is not an important fate process. Bovey 

and others (1974) applied a 1:1 mixture of picloram and 

2,4,5-T at a rate of 1.12 kg!ha to pastureland and then 

analyzed soil and runoff water for both compounds. 

Concentrations in runoff water ( 400 to 800 ppb) were 

significantly higher than concentrations in soil (undetectable to 238 ppb). El-Dib and Aly (1976b) studied the use 

of bentonite clays to remove the herbicides IPC, CIPC, 

linuron, and monuron from contaminated surface waters 

and concluded that the clay minerals would be unable to 

remove significant amounts of the herbicides from solution. Endothall, an aquatic herbicide, was shown to have a 

sorption coefficient of 0.958 in a sediment-water system 

investigated by Reinert and Rodgers (1984). Similarly, 

Muir and others (1980) determined that distribution 

coefficients for fluridone sorption to pond and river 

sediments were relatively low, ranging from 20 to 137. 

Wauchope and Myers (1985) observed atrazine and 

linuron sorption to freshwater sediments to be rapid and 

reversible. Once again, sorption coefficients were low, 

ranging from 1.3 to 18 for atrazine and 1.5 to 24 for 

linuron. 

Although the majority of herbicides do not sorb 

appreciably to sediments, there are some exceptions. The 

dinitroaniline and bipyridylium herbicides have demonstrated a strong tendency to sorb to sediments. Juo and 

Oginni (1978) observed that sorption maximums for 

paraquat sorbed to a variety of tropical soils ranged from 

1,200 to 7,500 mg!kg of soil. Karickhoff and Brown 

(1978) reported similar paraquat concentrations for the 

silt- and clay-size fractions of five natural sediments, with 

the extent of sorption increasing with the sediment 

fraction's cation exchange capacity. Distribution coefficients were greater than 1,000 for the clay fraction of one 

of the sediments. Khan (1974c) noted that paraquat and 

diquat sorb by ion exchange and that their magnitudes of 

sorption are comparable. LaFleur (1979) noted that 

sorption of trifluralin to several model soils was extensive 

and rap,id, with practically 100 percent of the applied 

herbicide being sorbed to the soil within 100 min of batch 

mixing. Therefore, the low solubilities of the dinitroaniline herbicides (relative to the other herbicides) and the 

strongly cationic nature of the bipyridylium herbicides 

cause them to strongly sorb to natural sediments. 

The herbicides are not readily bioaccumulated 

because of their high solubilities and the ability of biota to 

metabolize them easily. Muir and others (1980) observed 

that BCFs for minnows exposed to fluridone in small 

ponds ranged from less than 1 to 64. BCFs for duckweed 

in the ponds were also low, ranging from 19 to 85. 

Hoeppel and Westerdahl (1983) noted only minor accumulations of 2,4-D by fish in a Georgia lake treated with 

the dimethylamine (DMA) and butylethanol (BEE) salts 

of 2,4-D. All fish collected 13 d after application of the 

herbicide contained nondetectable residues of 2,4-D. 

However, fish subjected to the 2,4-D BEE treatment 

initially appeared to accumulate greater amounts of 

2,4-D than fish subjected to the treatment with 2,4-D 

DMA. In a laboratory study documenting the BCF's of a 

variety of pesticides by caddisfly egg masses, Belluck and 

Felsot (1981) observed that monuron had the lowest BCF 

of the group, having a value of approximately unity. 

Monuron's low BCF was attributed to its high solubility 

relative to the other pesticides studied. With regard to the 

herbicides as a group, bioconcentration varies inversely 

with herbicide solubility, with some of the dinitroaniline 

herbicides probably exhibiting the greatest ability to 

bioconcentrate. 

The majority of the herbicides are rapidly degraded 

in surface-water systems and, therefore, are not persistent 

environmental contaminants. Averitt and Gangstad (1976) 

observed the dissipation of 2,4-D in static water as a 

function of water depth, temperature, and time. Using a 

treatment rate of 4.48 kg acid equivalent per hectare, they 

observed a 58-ppb residue decrease for each 0.61 m 

increase in water depth, a 115-ppb residue decrease for 

each 7 °C increase in temperature, and a 53-ppb concentration decrease for every 7-d interval after treatment. 

Delaune and Salinas (1985) found that anaerobic conditions retard microbial degradation of 2,4-D and that rapid 

degradation is most likely to occur in the oxidized surface 

layer of sediments. Complete degradation of2,4-D to C02 

has been reported, with interim soil degradation products 

of 2,4-dichlorophenol and 2,4-dichloroanisole being 

detected (Smith, 1985). Jones and others (1982) documented half-lives of atrazine in an aqueous estuarine 

environment to range from 3 to 12 d, with the major 

metabolite being hydroxyatrazine. In contrast to 2,4-D 

degradation, oxygen levels had little effect on atrazine 

degradation. Humic substances have been shown to inhibit 

base-catalyzed hydrolytic reactions and to accelerate acidcatalyzed hydrolytic reactions (Purdue, 1983). For example, Li and Felbeck (1972) have shown that the hydrolytic 

half-life of atrazine decreases with decreasing pH and 

increasing humic acid concentration. Muir and others 

(1980) reported fluridone half-lives in pond water ranging 

from 4 to 7 d. Similarly, Bovey and others (1974) observed 

that picloram and 2,4,5-T applied to pastureland dissipated rapidly with no evidence of accumulation. 

Some herbicides have demonstrated greater persistence than such compounds as atrazine and 2,4-D. Anderson (1985) determined that residues of chlorsulfuron and 

metsulfuron (two sulfonylurea herbicides applied to winter wheat and barley) were persistent in soils for as long as 

500 d. Their ultimate attenuation was attributed to hydrolysis and biodegradation, the rates of which decreased 

with decreasing temperature. El-Dib and Aly (1976a, 

1976c) investigated hydrolysis and biodegradation of some 

of the phenylamide herbicides. They found that hydrolysis 

was a significant degradation process only at pH values 

42 Manmade Organic Compounds In the Surface Waters of the United States 
Table 19. Herbicide residues in North American surface waters 

[ppb, parts per billion) 

Mean, Maximum, 

Compound Locality ppb ppb Reference 

Atrazine Rhode River, Maryland 0.13 3.3 1 

Atrazine Ontario surface waters 5.5 32 2 

Atrazine Say1orvi11e Reservoir, Iowa .22 .67 3 

2,4-D New Iberia, Louisiana 359 4 

2,4-D Ontario surface waters 5.2 320 2 

2,4,5-T Runoff; Reise1, Texas 26 287 5 

Pic1oram Runoff; Reise1, Texas 10 19 5 

Diuron Pearl Harbor, Hawaii .33 1 6 

A1achlor Saylorville Reservoir, Iowa .08 .41 3 

Simazine Ontario surface waters .37 3.4 2 

Cyanazine Saylorvi11e Reservoir, Iowa .09 .32 3 

1. Wu and others (1980) 

2. Frank and others (1982) 

3. Leung and others (1982) 

above typical environmental levels. Additionally, the 

phenylamide herbicides were not significantly biodegraded by mixed microbial cultures taken from Nile River 

water. Biodegradation did occur, however, when Bacillus 

cereus was added to the aqueous system. Despite these 

examples, the herbicides are relatively nonpersistent contaminants of surface-water environments, with aquatic 

half-lives usually less than 60 d. 

Environmental Distribution 

Because the herbicides do not readily partition into 

sediment organic matter or the lipid reservoirs of aquatic 

organisms, the highest environmental residues of these 

compounds are typically found in the aqueous phase of 

surface-water systems. Bovey and others (1974) observed 

that picloram and 2,4,5-T residues in runoff from pastureland were primarily in the aqueous phase, with only 

small amounts sorbed to soil particles. Green and others 

(1977) noted that diuron residues in Hawaiian estuaries 

were transported primarily in solution. In a study of an 

Iowa reservoir, Leung and others (1982) detected residues of only atrazine, alachlor, and cyanazine in the 

aqueous phase. In samples from more than 150 surfacewater stations across the United States during a 3-yr 

period, Gilliom and others (1985) reported that all herbicide detections were in water samples, with no detec4. 

5. 

6. 

Averitt and Gangstad (1976) 

Bovey and others (1974) 

Green and others (1977) 

tions in bed-sediment samples. Table 19 presents some 

herbicide residues that have been detected in natural 

surface waters in the United States. 

Herbicides enter surface-water systems in a variety 

of ways, including application of aquatic herbicides directly 

to a body of water to control aquatic vegetation. As a 

result, residues of aquatic herbicides will be greatest 

immediately following application. Terrestrial herbicides 

are carried into surface waters by runoff and base flow, 

and, consequently, surface-water residues are highest if a 

heavy rain immediately follows application of the herbicide. During the spraying season, residues are commonly 

the result of misuse of equipment which leads to accidental discharges of the herbicide into surface waters. Frank 

and others (1982) noted that 82 percent of the pesticides 

lost to water in 11 agricultural watersheds in Canada 

during the spraying season could be attributed to careless 

use of the pesticides near surface waters. On an annual 

basis, however, spills accounted for only 22 percent of the 

loss to water, while runoff accounted for 60 percent and 

base flow for 18 percent of the pesticide loss. 

No definitive regional trends in contamination by 

herbicides in U.S. surface waters have been reported in 

the literature, but regional herbicide-contamination trends 

can be expected to parallel use patterns. For example, 

Gilliom and others (1985) noted that stations in the U.S. 

Geological Survey-U.S. Environmental Protection Agency 

Herbicides 43 
Pesticide Monitoring Network where atrazine was detected 

more than once were all located downriver from large 

corn-producing areas in which atrazine is heavily applied. 

The low persistence of the herbicides coupled with relatively high analytical detection limits, however, prevented 

any other determinations of regional trends. 

Aside from the fact that the highest herbicide 

residues in surface-water systems usually occur during 

spraying seasons, there are few data in the literature 

indicating temporal contamination trends in U.S. surface 

waters. Frequency-of-detection data for some of the chlorophenoxy and triazine herbicides monitored in the Pesticide Monitoring Network program were inconclusive 

with regard to temporal trends (Gilliam, 1985). This is 

primarily attributable to the relatively short persistence of 

the herbicides combined with the fact that herbicide data 

at most stations in the network were collected only during 

1976, 1977, and 1978. Atrazine and 2,4-D were the most 

frequently detected herbicides; this can be explained by 

the heavy agricultural use of these two compounds. 

Atrazine also was the most frequently detected herbicide 

in the water of 11 Canadian agricultural watersheds from 

1975 through 1977 (Frank and others, 1982). Although 

the available literature does not reveal any clear temporal 

trends in herbicide detection in the United States, the 

escalating use of herbicides over the last 20 yr (fig. 16) 

indicates that the amount of herbicides in the Nation's 

surface waters is not likely to decrease in the near future. 

PHENOLS 

Uses, Production, Properties, and 

Regulatory Limits 

Phenols are a widely used group of compounds 

characterized by a benzene ring with one or more hydroxyl 

groups. The dihydroxy phenols can have the second 

hydroxyl on the ortho, meta, or para position of the 

benzene ring, and these isomers are commonly referred 

to by their trivial names, catechol, hydroquinone, and 

resorcinol, respectively. Although phenols were first produced (in 1834) by the distillation of coal tar, most current 

production on a commercial scale employs the cumene 

(isopropylbenzene) process, the toluene oxidation process, or the benzene sulfonation process (U.S. Environmental Protection Agency, 1980a). The basic phenol 

molecule can then be amended with chlorine, nitro, or 

methyl groups. In addition to manufacturing processes, 

chlorophenols are produced by the chlorination of treated 

wastewaters and drinking waters containing phenol. In 

solutions containing as little as 10 mg!L phenol and 20 

mg!L chlorine, 2-chlorophenol has been synthesized in a 

1-h period (Callahan and others, 1979). Similar results 

have been reported for di- and trichlorophenol isomers. 

OH 

6 

Phenol 

OH 

CI¢CI 

Cl 

2,4,8-Trlchlorophenol 

¢ 

N01 

4-Nitrophenol 

P-chloro-m-cre•ol 

OH 

oc• 

2-Chlorophenol 

OH 

CI&.CI 

cr¥c• Cl 

Pentachlorophenol 

¢·0, 

N01 

2,4-Dinltrophenol 

4,8-Dinltro-o-cre•ol 

OH ¢r•• 

Cl 

2,4-Diohlorophenol 

2-Nitrophenol 

¢··· 

CH1 

2,4-Dimethylphenol 

Catechol 

Figure 18. Structural diagrams of selected phenolic compounds. 

Structural diagrams of some of the more common phenolic compounds are presented in figure 18. 

Phenol, C6HsOH, is a monohydroxybenzene that 

exhibits the properties of a weak acid (pi<a = 10) owing to 

the delocalization (via resonance) of the negative 

phenoxide ion throughout the benzene ring (Baum, 1982). 

It is this acidity that differentiates phenols from the 

alcohols. Phenol is a colorless, crystalline solid having a 

characteristic taste and odor. It is primarily an industrial 

chemical used as an intermediate in the production of 

other compounds such as phenolic resins, germicides, 

pharmaceuticals, fungicides, dyes, herbicides, plastics 

(such as Bakelite), explosives (such as picric acid or 

2,4,6-trinitrophenol), antiseptic throat lozenges, and skin 

medications. Production data for phenol are presented in 

table 20. 

Established U.S. Environmental Protection Agency 

criteria for phenol state that chronic toxicity to freshwater 

aquatic life will occur at a concentration of 2.56 mg!L, and 

a limit of 3.5 mg!L is imposed as the maximum level 

allowable to ensure a lack of hazardous effects to humans 

(U.S. Environmental Protection Agency, 1980a). In addition, to avoid undesirable tastes and odors in stream 

water, phenol concentrations must not exceed 0.3 mg!L. 

Monochlorophenols are primarily used as intermediates in the production of more highly chlorinated 

44 Manmade Organic Compounds in the Surface Waters of the United States 
Table 20. World production of phenol in 1978 

[From Moore and Ramamoorthy, 1984) 

8 Production 

Country (10 metric tons) 

North America 

Canada so 

Mexico 25 

u.s. 1,589 

South America 

Argentina 12 

Brazil 65 

Far East 

Japan 289 

Other 22 

phenols, and their approximate annual production in the 

United States is 18,800 t-significantly less than the 

39,000-t production of 2,4-dichlorophenol (Moore and 

Ramamoorthy, 1984). This compound is used extensively 

to manufacture the herbicide 2,4-D. Trichlorophenols are 

used as antimicrobial agents and as antimildew agents for 

textiles, and in the manufacture of wood and glue preservatives. More specifically, 2,4,5-trichlorophenol is used to 

manufacture 2,4,5-T and 2,4,6-trichlorophenol is used to 

produce tetrachlorophenol. The combined annual production of tri- and tetrachlorophenols in the United States is 

less than 7,000 t (Moore and Ramamoorthy, 1984). 

Pentachlorophenol (PCP) has probably seen the 

most extensive use of all the chlorophenols, and it is 

manufactured primarily by the organic chemical and 

pesticide industries. PCP was the second most heavily 

used pesticide in the United States in 1977, when approximately 80 million lb were produced (Guthrie and others, 

1984}. As a herbicide in Japan, it accounted for 50 

percent of all herbicide use in the country until its use was 

regulated in 1971 (Murthy and others, 1979). It is also 

used extensively as a wood preservative, as a fungicide and 

bactericide in cooling tower water, and as an additive in 

adhesives, construction materials, textiles, leather, paint, 

paper, and oil-well-drilling mud. 

Annual U.S. production of 2-nitrophenol, 3-nitrophenol, 4-nitrophenol, and 2,4-dinitrophenol has been 

estimated to be 6,800, less than 500, 19,000, and 500 t, 

respectively (Moore and Ramamoorthy, 1984). For. the 

most part, the nitrophenols serve as intermediates in the 

production of pesticides, pharmaceuticals, dyes, explosives, and wood preservatives. 

Only sparse production data were found for some 

of the other phenolic compounds. Moore and Ramamoorthy (1984) noted that annual U.S. production of 

Produc~ion 

Country (10 metric tons) 

Yestern Europe 

Be1gium/Luxemborg 140 

France 125 

\Jest Germany 420 

Italy 280 

United Kingdom 267 

Eastern Europe >360 

Australia 17 

'World Total 3,711 

4-methyl-2,6-dibutyl phenol is about 16,000 t, and that 

annual U.S. production of nonylphenol was about 69,000 

t in 1981. The latter compound's uses include nonionic 

surfactants, phenolic resins and plasticizers, adhesives, 

lube oil additives, stabilizers, and petroleum demulsifiers 

(Verschueren, 1983). 

Table 21 presents some ~ysicochemical data for 

the phenols. In general, they have high solubility and 

relatively low vapor pressures and octanol-water partition 

coefficients. Among the chlorinated phenols, however, 

compound solubility decreases and octanol-water partition coefficient increases with increased chlorination of 

the aromatic ring. PCP, the most highly chlorinated 

phenol, has a solubility of only 14 mg!L and an octanolwater partition coefficient of 5.0. 

Environmental Fate 

Primary fate mechanisms for phenolic compounds 

include biodegradation and photolysis. Sorption to natural 

sediments and bioaccumulation are generally minor, with 

the exception of the highly chlorinated phenols. Volatilization, hydrolysis, and nonphotolytic oxidation are also 

relatively insignificant fate processes. 

Numerous researchers have documented biological 

degradation of the phenols. Studies by Baird and others 

(1974), Hickman and Novak (1984), and Tabak and 

others (1964) reported significant phenolic biodegradation in activated sludge systems. Other researchers (Chu 

and Kirsch, 1972; Knackmuss and Hellwig, 1978; Tabak 

and others, 1981; Banerjee, Howard, and others, 1984) 

have noted biodegradation of phenols by isolated bacterial 

cultures. 

Biodegradation of phenol in surface waters has also 

been observed. Borighem and Vereecken (1978) used in 

Phenols 45 
Table 21. Experimentally determined physicochemical properties of selected phenolic 

compounds at room temperature 

[mg/L, milligrams per liter; Ko,.., octanol-water partition coefficient; pKu, -log acid dissociatron constant. 

From Callahan and others, 1979] 

Compound Solubility 

(mgfL) 

Phenol 93,000 

2-Chlorophenol 28,500 

2,4-Dichloropbenol 4,500 

2,4,6-Trichlorophenol 800 

Pentachlorophenol 14 

2-Nitrophenol 2,100 

4-Nitrophenol 16,000 

2,4-Dinitrophenol 5,600 

2,4-Dimethylphenol 17,000 

p-Chloro-m-cresol 3,850 

4,6-Dinitro-o-cresol 

situ batch studies in both polluted and unpolluted reaches 

of the River Samber in Belgium to determine the effects 

of certain parameters on the degradation of phenol. They 

found that biodegradation rates increased with both 

increased oxygen transfer to the water and increased 

initial phenol concentration. Visser and others (1977) 

used in situ studies in the St. Lawrence River, Canada, to 

quantify phenol biodegradation. During summer, they 

observed that phenol concentrations as high as 125 ppb 

were reduced to less than 10 ppb within 4 h. Phenol 

degradation was accompanied by a concurrent increase in 

the microbial population in the river-water sample. During the coldest part of winter, however, phenol biodegradation was negligible, probably owing to the absence of 

phenol-degrading bacteria. Table 22 lists the bacteria 

present in the St. Lawrence River water during the winter 

and summer months and the maximum phenol levels they 

tolerated after adaptation. Other microorganisms that 

have been shown to degrade phenol include Pseudomonas 

putida, Bacillus stearothennophilus, and yeast such as 

Oospora, Saccaromycetes, Candida, Debarymyces, and 

Trichosporon cutaneum (Callahan and others, 1979). 

Most of the other phenols have been biodegraded 

under laboratory conditions. Using mixed microbial cultures obtained from soils and sediments, Tabak and others 

(1964) measured the biotransformation of 95 percent of a 

K Vapor pressure pK 

ow (torrs) a 

1.46 0.53 10.02 

2.17 2.2 8.52 

2.75 0.12 7.85 

3.38 1.0 5.99 

5.01 0.00011 4. 74 

1. 76 1.0 7.21 

1.91 2.2 7.15 

1.53 4.09 

2.50 0.0621 10.60 

2.95 

2.85 4.35 

variety of parent phenolic compounds in 1 to 10 d (table 

23). Knackmuss and Hellwig (1978) reported biodegradation of a variety mono- and dichlorophenols by a strain 

of 4-chlorophenol-grown cells. With the exception of 3,5-

and 2,4-dichlorophenol, all half-lives were less than 12 d. 

Stanlake and Finn (1982) isolated a variety of PePdegrading bacteria from soil, water, and sewage and noted 

that the length of the lag phase for bacterial growth on 

PCP decreased when pH was raised from 6.8 to 7.8. 

Murthy and others (1979) examined PCP degradation in 

aerobic and anaerobic soils and noted that pentachloroanisole was a primary degradation product of PCP, along 

with 2,3,5,6- and 2,3,4,5-tetrachlorophenols and 2,3,6-

trichlorophenol. Their proposed degradation pathway of 

PCP is presented in figure 19. 

The biodegradability of the chlorophenols differs 

with chlorine substitution. However, increased biodegradability does not always correlate with decreased chlorination of the phenolic molecule. In a study involving a 

variety of phenolic compounds, Alexander and Aleem 

(1961) found that chlorine substitution in the meta position of the aromatic ring increases the ring's resistance to 

biological cleavage. With the exception of these metachlorophenols, the authors observed virtually 100 percent ring 

cleavage of all the chlorophenols in a soil suspension 

within 47 d. Under aerobic, room-temperature conditions, 

46 Manmade Organic Compounds in the Surface Waters of the United States 
Table 22. Phenol bacteria present during the winter and (or) summer season in St. 

Lawrence River water and maximum phenol levels that could be tolerated 

[mWL, milligrams per liter. From Visser and others, 1977] 

Maximum phenol 

levels tolerated 

after adaptation 

Bacterial species Yinter Summer (mg/L) 

Achromobacter eurydice 

A. cyc1oclastes or guttatus 

A. iophagus 

Arthrobacter terregens 

Azotobacter agilis 

Bacillus brevis 

B. cereus 

B .. circulans 

Brevibacterium maris 

B. sulfureum 

Brevibacterium sp. 

Clostridium carnis or tertium 

Flavobacterium aquatile 

F. breve or peregrinum 

F. lutescens 

F. marinum 

F. solare 

Flavobacterium (marinum ?) 

Flavobacterium sp. 

Micrococcus conglomeratus 

Micrococcus (?) 

Mycobacterium phlei 

Paracolobactrum coliforme 

Pseudomonas fragi 

Pseudomonas sp. 

Pseudomonas sp. 

Pseudomonas sp. 

Pseudomonas sp. 

Pseudomonas sp. 

Sarcina f1ava 

S. hansenii 

Banerjee, Howard, and others (1984) observed zeroorder biodegradation kinetics for a variety of phenolic 

compounds in natural water samples. They also stated 

that degradation rates decreased with increased chlorine 

substitution. However, biodegradation studies in their 

second collection of river water indicated that 

2,4-dichlorophenol was more resistant to biodegradation 

than 2,4,5-trichlorophenol, which in turn was more resistant to degradation than PCP. Data indicating that 

decreased phenolic biodegradation rates do not strictly 

correspond to increasing chlorination of the aromatic ring 

have also been reported by Tabak and others (1964), 

Alexander and Aleem (1961 ), and Smith and Novak (1987). 

X 

X 

X 

X 

X 

X 

X 

X 2,400 

X 1,200 

X 1,200 

X 1,200 

X 2,400 

X 2,400 

X 600 

X 600 

X 2,400 

X 150 

X 600 

X 300 

X 600 

X 1,200 

X 600 

X 300 

X 150 

X 600 

X 1,200 

X 1,200 

X 600 

X 600 

X 2,400 

X 600 

X 150 

X 300 

X 600 

X 1,200 

X 2,400 

X 1,200 

X 2,400 

Available data indicate that photolytic oxidation 

may also be a significant fate mechanism for many 

phenols. Phenol can be photolytically degraded to a 

reddish, high-molecular-weight material. Photolytic end 

products of phenol degradation include 4,4' -, 2,4' -, and 

2,2'-dihydroxybiphenyl, hydroquinone, and catechol (Joschek and Miller, 1966). Callahan and others (1979) have 

indicated that the absorption maximums of metalcoordinated phenols can be significantly extended. In the 

near-ultraviolet spectral region, the absorption maximum 

of phenol does not extend beyond 290 nm. However, when 

phenol is complexed with metal cations such as Fe(III), 

the absorption maximum is extended to 600 nm. Similarly, 

Phenols 47 
Table 23. Time required for bacteria to use 95 percent of 

parent substrate 

[From Tabak and others, 1964) 

1 to 2 days 

Phenol 

Catechol 

Resorcinol 

Quinol 

Phloroglucinol 

o-Cresol 

m-Cresol 

p-Cresol 

Compounds degraded in: * 

3 to 6 days 

o-Nitrophenol 

11-Nitrophenol 

p-Nitrophenol 

2,4,6-Trinitrophenol 

(250 ppm) 

2-Chloro-4-nitrophenol 

2,6-Dichloro4-nitrophenol 

m-Chlorophenol (150 ppm) 

p-Chlorophenol 

o-Phenylphenol (100 ppm) 

Thymol (150 ppm) 

7 to 10 days 

2,4-Dinitrophenol 

2,6-Dimethylphenol 

(200 ppm) 

2,4-Dichlorophenol 

(200 ppm) 

2,4,6-Trichlorophenol 

*Initial concentration was 300 parts per million (ppm) 

unless noted in parentheses. 

the aqueous Fe(III) complex with 2-chlorophenol has an 

absorption maximum of about 550 nm, which is significantly greater than the absorption maximum of the chlorophenolate anion (310) or undissociated 2-chlorophenol 

(290 nm). Table 24 details some of the photolytic degradation products of the phenols. 

Based on their relatively high solubilities, low 

octanol-water partition coefficients, and anionic nature, 

most of the phenols do not strongly sorb to sediments or 

bioaccumulate. However, increased chlorination of the 

phenol molecule decreases compound solubility and 

increases octanol-water partition coefficients. Therefore, 

the more highly chlorinated phenols such as PCP or the 

tetrachlorophenols can exhibit significant sorption and 

bioaccumulation in surface-water systems. 

CIOOCHa Cl 

c•Yc• ~entaoh~:roanlaole 

c·oOH ~·oOH ~·oOH Cl 

Cl Cl CI~CI ~CI Cl 2,3,5,6- 2,3,6-

PCP ""- Tetrachlorophenol Trlchlorophenol "' OH 

c.o:: 

Cl 

2,3,4.5-Tetrachtorophenol 

Figure 19. Structural diagrams showing degradation pathway of pentachlorophenol (PCP) in soils. [Reprinted with 

permission from the Journal of Environmental Science and 

Health, v. 814, p. 13 (1979). Copyright 1979 by Marcel 

Dekker, Inc.] 

Table 25 lists experimentally determined sorption 

coefficients (Kd, equal to the equilibrium solute concentration in sediment divided by its equilibrium aqueous 

concentration) of nonionized chlorophenols sorbed to 

four different natural sediments (Schellenberg and others, 

1984). The authors determined that in natural waters of 

Table 24. Photolysis of phenols and their degradation products 

[nm, nanometer; ppm, parts per million. From Callahan and others, 1979] 

Coq'X)Und light source Experimental conditions 

Phenol 254 Ml Aqueous, oxygenated solutions 

2-Chlorophenol 313 Ml Aqueous alkali 

4-Chlorophenol 254 ard >290 m Aqueous solution with cyanide 

2,4-Dichlorophenol Solar radiation Aerated aqueous solution 

2,4,6·Trichlorophenol In presence of electron acceptor 

Pentachlorophenol Sunlight 2X solution of NePCP 

4-Nitrophenol Sunlight Aqueous solution of 200 ppm 

48 Manmade Organic Compounds in the Surface Waters of the United States 

Products 

4 4 1 ·, 2,4•-, and 2,2•-dihydroxybiphenyl; hydroqu1none; catechol 

Catechol 

4- cyanopheno l 

Cgmplete disappearance after 10 days; insoluble, dark, tarry substance prOduced 

2,6-dichlorophenoxyl semiquinone radical 

anion 

50% degradation after 10 days; products: chloranilic acid, tetrachlororesorcinol, and 

several complex chlorinated benzoquinones 

Degraded in 1·2 month period to hydroquinone, 4-nitrocatechol, and a dark, acidic polymer 
Table 25. Experimentally determined sorption coefficients 

of nonionized chlorinated phenols for different natural sorbents 

(From Schellenberg and others, 1984] 

Sorption coefficient for sorbent * 

Compound 1 2 3 4 

2,3-DCP 21 12 5 <1 

2,4-DCP 25 17 6 <1 

2,4,6-TCP 78 34 9 <1 

2,4,5-TCP 165 74 20 <1 

3,4,5-TCP 276 115 31 <1 

** 2,3,4,6-TeCP 568 200 52 <3 

2,3,4,5-TeCP 1237 362 105 7 

** PCP 3670 930 200 <13 

*1. Lake sediment; organic carbon content (OC) - 9.4% 

2. River sediment; OC- 2.6% 

3. Aquifer material; OC - 0.84% 

4. Aquifer material; OC - 0.03% 

**sorption coefficients for compound calculated from 

chromatographic data. 

low ionic strength (less than 103 mol/L) and of pH not 

more than one log unit greater than the phenolic 

compound's pi<a, sorption of the phenolate anion can be 

ignored. The data indicate that sorption coefficients 

increase with increasing chlorination of the phenolic compound and with increased sediment-organic-matter content. The less chlorinated phenols are not sorbed significantly to any of the soils. In soils having low organicmatter content, even sorption of PCP will not be 

substantial. Similarly, Scott and others (1982) observed 

that phenol did not sorb strongly to soils, with sorption 

coefficients ranging from 0.57 to 1.19. For the case of the 

higher chlorinated phenols such as PCP, sorption will 

likely be greatest in acidic soils and sediments having high 

organic-matter contents. 

Similar to sediment sorption, phenols are not readily 

bioaccumulated in surface-water systems, but increased 

accumulation has been correlated with increased chlorination of the phenol molecule (Biddinger and Gloss, 

1984). Few bioaccumulation data were found in the 

reviewed literature for the majority of the phenols, with 

the exception of PCP. Callahan and others (1979) cite 

several studies that found that PCP was accumulated to 

concentrations well beyond its aqueous concentration. An 

eel exposed to pulp-mill discharge bioaccumulated an 

aqueous PCP concentration of 3 p.g/L to a tissue concentration of 3,000 p.g!L. Similarly, a 72-h exposure to 100 

p.g!L PCP resulted in a 900-fold concentration increase of 

the compound in goldfish, indicating that PCP can be 

bioaccumulated in substantial amounts. For the other 

phenols, bioaccumulation is likely to be an insignificant 

fate process owing to the compounds' generally low 

octanol-water partition coefficients. 

Volatilization, nonphotolytic oxidation, and hydrolysis are not significant fate processes for the phenols. 

2-Chlorophenol has the highest vapor pressure of the 

phenols, but its high solubility prevents any substantial 

volatilization losses. Phenol can be oxidized in aqueous 

solution by molecular oxygen, but the more highly chlorinated phenols are resistant to oxidation under typical 

environmental conditions (Moore and Ramamoorthy, 

1984). The phenols also are resistant to hydrolysis, 

because the high negative charge density of the aromatic 

ring makes the covalent bonds relatively stable (Moore 

and Ramamoorthy, 1984). 

Environmental Distribution 

Phenols may enter surface-water systems directly as 

components of industrial effluents and indirectly as transformation products of other compounds. For example, 

pulp-mill effluents commonly contain elevated concentrations of phenolic compounds (Landner and others, 1977), 

and the chlorination of wastewaters can convert phenol to 

monochlorophenol, dichlorophenol, or even trichlorophenol (Callahan and others, 1979). Table 26 lists the 

concentrations of phenol and some of its chlorinated 

derivatives in municipal and industrial wastewater. 

Because of their generally high solubilities, the 

phenols can be found in relatively high aqueous concentrations in natural surface waters. Sheldon and Hites 

(1979) detected isomers ofbromochlorophenol, dibromophenol, dichlorobromophenol, and dibromochlorophenol · 

in water of the Delaware River near Philadelphia. 

Yasuhara and others (1981) reported extremely high 

concentrations of PCP in the waters of the Hayashida 

River in Japan. Concentrations ranged from 306 to 895 

p.g!L and were attributed to the unregulated discharge of 

leather-tannery wastewaters. Visser and others (1977) 

observed that ambient phenol concentrations in the waters 

of the St. Lawrence River, Canada, were approximately 10 

to 15 p.g!L. Borighem and Vereecken (1978) noted that 

phenol levels in polluted stretches of the River Samber in 

Belgium were considerably higher, ranging from 0.2 to 0.7 

mg!L. In a study of water and sediment quality in the 

Netherlands, Wegman and van den Broek (1983) detected 

all the isomers of mono-, di-, tri-, tetra-, and pentachlorophenol in water samples from the Ijssel River near 

Kampen. Maximum compound concentrations ranged 

from 0.02 (2,3,4,5-tetrachlorophenol) to 3.4 p.g!L (3-

chlorophenol). Despite these detections, the ability of 

Phenols 49 
Table 26. Concentration of phenol and some of its chlorinated derivatives in municipal and industrial wastewater 

[p.g/L, micrograms per liter. From Moore and Ramamoorthy, 1984) 

Compound Concentration Source 

(pg/L) 

Phenol 0.03-20 Treated sewage, European 

cities 

200-3,016,000 3 petroleum refineries (USA) 

3200 2 coal gasifacation plants 

(USA) 

38,000-1,240,000 Coke plant (USA) 

10-300 Specialty chemical plant 

(USA) 

2,4-DCP 220 Chemical plant (Vancouver) 

<0.1 Vood preservation plant 

(British Columbia) 

<0.1 Landfill leachate (Vancouver) 

2,6-DCP 220 Chemical plant (Vancouver) 

2.4 Yood preservation plant 

(British Columbia) 

1.2-5.6 Landfill leachate (British 

Columbia) 

2,4,5-TCP <0.05 Treated sewage, 4 plants ( (Vancouver) 

0.5-2400 2 chemical plants (Vancouver) 

<0.05 Wood preservation plant 

(British Columbia) 

0.05-2 Landfill leachate (Vancouver) 

2,4,6-TCP <0.05-1 Treated sewage, 4 plants 

(Vancouver) 

<0.05-3120 3 chemical plants (Vancouver) 

0.5-1 Yood preservation plant 

(British Columbia) 

0.4-1 Landfill leachate (Vancouver) 

25-115 Pulpmill effluent (Vancouver) 

2,3,4,6-TeCP 0.6-28 Treated sewage, 4 plants 

(Vancouver) 

1.2-8270 4 wood preservation plants 

(British Columbia) 

0.3-166 3 chemical plants (Vancouver) 

0.2-0.8 Landfill leachate (Vancouver) 

PCP 0.5-4. 7 Treated sewage, 4 plants 

(Vancouver) 

0.25-1.3 Treated sewage, 6 cities 

(Ontario) 

0.05-2760 Wood preservation plants 

(British Columbia) 

5,400,000 Chemical plant (Philadelphia) 

0.6-42 Landfill leachate (Vancouver) 

phenolic compounds to be biologically and photolytically 

attenuated makes them relatively nonpersistent contaminants. This is confirmed by data from the U.S. Environmental Protection Agency, which reported that of 110 

raw-water samples taken throughout the United States, 

phenol was detected in only 2 (U.S. Environmental Protection Agency, 1980a). 

Sediment residues of phenols rarely exceed the 

micrograms-per-kilogram range. However, higher residues have been reported occasionally. For example, Van 

Luik (1984) reported that phenol concentrations in the 

sediments of the Calumet Sag Channel (Ill.) ranged from 

1 to 6.9 mglkg. Pierce and others (1977) documented 

sediment PCP concentrations in a small stream ranging 

from 1.1 to 1.3 mglkg as a result of an accidental 

discharge of the compound. Concentrations typical of a 

phenol-contaminated surface-water system were reported 

by Wegman and van den Broek (1983). They determined 

that 100 percent of the sediment samples collected from 

Lake Ketelmeer in the Netherlands were contaminated 

with 2,5-dichlorophenol,. 2,3,5- and 2,4,5-trichlorophenol, 

2,3,4,5- and 2,3,4,6-tetrachlorophenol, and PCP. The 

respective median concentrations were 6.3, 2.4, 6.4, 0.9, 

1.7, and 8.4 ~glkg. The authors also noted that areas of 

highest phenol contamination were near industrial areas. 

Biological phenolic residues also have been 

reported. DeVault (1985) found that fish from harbors 

and tributary mouths of the Great Lakes contained PCP 

concentrations ranging from < 0.08 to 4.52 mglkg. In 

Lake Ontario, Niimi and Cho (1983) found lower PCP 

concentrations in a variety of sampled fish; mean residue 

levels ranged from 0.3 to 24 ~glkg. Rudling (1970) 

analyzed fish from a lake into which PCP had been 

discharged and determined that concentrations in the 

tissues of pike, perch, and eel were 0.2, 0.15, and 3.0 

mglkg, respectively. The elevated PCP concentrations 

found in the eel were attributed to the relatively high lipid 

content of eels. Similarly, Landner and others (1977) 

analyzed perch and northern pike caught in the vicinity of 

a pulp mill. Phenolic residues were concentrated in the 

liver fat of the sampled fish, with mean concentrations of 

2,4,6-trichlorophenol, trichloroguaiacol, and tetrachloroguaiacol of 1.2, 5.0, and 4.4 mglkg fat, respectively. In 

general, biological residues of phenolic compounds 

depend on aqueous concentrations, biological lipid content, and the solubility of the phenolic compound. 

HALOGENATED ALIPHATIC AND 

MONOCYCLIC AROMATIC HYDROCARBONS 

Uses, Production, and Properties 

The aliphatic hydrocarbons are straight -chain molecules composed of only hydrogen and carbon. The 

saturated aliphatic hydrocarbons, commonly referred to 

as the alkanes or paraffin hydrocarbons, are molecules in 

which each carbon atom is bonded covalently to four 

other atoms by single bonds. The general formula for 

alkanes is CnHln+l· Because of double and triple carbon-

. carbon bonds, unsaturated aliphatic hydrocarbons contain 

fewer hydrogen atoms per carbon atom than the alkanes. 

The family of unsaturated aliphatic hydrocarbons containing carbon-carbon double bonds are called the alkenes, or 

olefins. The family of compounds containing carboncarbon triple bonds are called the alkynes, or acetylenes. 

The general formulas for the alkenes and alkynes are 

CnH2 n and CnH2 n_ 2 , respectively. The industrial halogenation of the aliphatic hydrocarbons results in the production of many toxic chemicals that are widely used as 

solvent degreasers, refrigerants, and fumigants and in 

dry-cleaning operations and the pharmaceutical industry. 

50 Manmade Organic Compounds in the Surface Waters of the United States 
Benzene, first isolated in 1825 by Michael Faraday, 

exhibits the common structural feature of the aromatic 

hydrocarbons. The benzene molecule consists of six carbon atoms bound together in a cyclic structure. Each 

carbon atom also is bound to one hydrogen atom. X-ray 

diffraction analysis of benzene indicates that all the 

carbon-carbon bonds have a length of 1.39 A, intermediate to the length of a carbon-carbon single bond (1.54 A) 

and a carbon-carbon double bond (1.34 A) (Baum, 1982). 

The molecule's stability and intermediate carbon-carbon 

bond length are attributable to a resonating electron 

structure within the molecule. Benzene and other aromatic compounds having halogens or other functional 

groups substituted for hydrogen atoms have found wide 

industrial use. Many of these compounds are highly toxic 

and, therefore, pose a serious threat to surface-water 

quality. Figure 20 gives the structural diagrams of some of 

the more common monocyclic aromatic hydrocarbons 

(MAH's), together with some of the halogenated aliphatic 

hydrocarbons. 

Table 27 presents U.S. production data for the 

halogenated aliphatic and MAH's. Benzene, toluene, 

xylene, 1,2-dichloroethane, vinyl chloride, ethyl benzene, 

carbon tetrachloride, and nitrobenzene are among the 

compounds that have seen the largest production over the 

last 20 yr. Although production of some of the compounds 

in table 27 has slowed since 1979, the data indicate that 

production of both the halogenated aliphatic and MAH's 

was significantly greater in 1981-82 than production of 

these compounds from 1965 to 1969. Table 28 documents 

H 

I CI-C.;.H 

I 

Cl 

Dlchloromethane 

Vinyl chloride 

6 

Toluene 

6 

Ethyl benzene 

Cl H 

I I 

H-C-C-H 

I I Cl H 

1.1-Dichloroethane 

Tetrachloroethylene 

o-Xylene 

6 

Chlorobenzene 

Figure 20. Structural diagrams of selected halogenated 

aliphatic and monocyclic aromatic hydrocarbons. 

Table 27. U.S. production of selected halogenated aliphatic and monocyclic aromatic hydrocarbons 

[In metric tonsX 104.• From Moore and Ramamoorthy, 1984] 

Conpound 1961·65n 1966·70n 1971·75n 1976 1977 1978 1979 1980 1981 1982 

Chloromethane 5.5 14.9 20.8 17.1 21.6 20.6 21.0 16.4 18.4 16.0 

Dichloromethane 6.9 14.5 19.0 24.4 21.7 25.9 28.7 25.5 26.8 23.8 

Chloroform 3.6 9.1 11.6 13.2 13.7 15.8 16.1 16.0 18.3 13.5 

Carbon tetrachloride 21.8 36.5 45.9 38.9 36.7 33.4 32.4 32.1 32.9 nd 

Chloroethane 26.6 23.1 28.1 30.4 27.8 24.5 26.4 17.9 14.7 13.1 

1,2-dichloroethane 82.1 234.8 379.2 364.6 498.7 498.9 534.9 503.7 452.3 nd 

1,1,1-trichloroethane nd 14.0 21.9 28.6 28.8 29.2 32.5 31.4 27.8 25.0 

Chlorodifluoromethane nd nd nd 7.7 8.1 9.3 9.6 10.3 11.4 nd 

Trichloroftuoromethane 5.6 10.4 13.6 11.6 9.7 8.8 7.6 10.3 7.4 nd 

Dichlorodifluoromethane 9.6 15.1 19.9 17.8 16.2 14.8 13.3 13.4 14.7 nd 

Vinyl chloride 63.8 142.1 223.1 257.5 271.5 315.8 289.7 293.2 311.7 nd 

Trichloroethylene 13.9 24.5 18.8 14.3 13.5 13.6 14.5 12.1 11.7 8.1 

Tetrachloroethylene 14.2 27.0 32.3 16.3 27.8 32.9 35.1 34.7 31.3 nd 

Bromoform 0.6 0.9 1.2 nd 1.6 1.6 nd nd nd nd 

Ethylene dibromide nd 13.8 14.0 9.1 11 .1 10.4 13.0 8.8 7.6 nd 

Benzene 62.6 104.8 125.8 142.5 143.5 148.8 167.2 200.7 133.9 107.0 

Toluene 39.1 70.2 87.5 99.9 101.8 105.4 101.0 101.7 85.6 105.2 

Xylene 31.9 44.9 53.2 72.2 81.1 84.5 97.2 90.8 88.2 73.0 

Monochlorobenzene 25.0 25.0 17.2 14.9 14.8 13.4 14.7 12.8 12.9 nd 

1,2·dichlorobenzene 1.9 2.7 2.7 2.2 2.1 1.9 2.6 2.2 2.3 nd 

1[4-dichlorobenzene 3.2 2.9 2.9 1. 7 3.0 1.9 3.8 3.4 3.3 nd 

A kyl benzenes 28.3 29.4 23.6 nd 23.6 23.8 28.4 40.6 24.2 nd 

2,4 + 2,6·dinitrotoluene nd 12.6 18.6 18.0 nd 29.8 31.2 nd 22.8 nd 

Ethyl benzene 106.1 184.7 246.9 261.7 376.9 380.3 383.1 346.5 354.3 299.7 

Nitrobenzene 9.7 19.1 20.2 18.6 25.0 26.1 43.2 27.7 40.8 nd 

* units are X 107 For benzene, toluene, and xylene, gallons. •• Annual average nd • no data available 

Halogenated Aliphatic and Monocyclic Aromatic Hydrocarbons 51 
Table 28. Uses of selected halogenated aliphatic and monocyclic aromatic hydrocarbons 

[From Verschueren, 1983] 

COCJl)OUnd 

Chloromethane 

Dichloromethane 

Chloroform 

Carbon tetrachloride 

1,2·dichloroethane 

Vinyl chloride 

Trichloroethylene 

Tetrachloroethylene 

Ethylene dibromide 

Trichlorofluoromethane 

Benzene 

Toluene 

Xylene 

Chlorobenzene 

1,2·dichlorobenzene 

1,4-dichlorobenzene 

1,2,4-trichlorobenzene 

2,6-dinitrotoluene 

Ethyl benzene 

Nitrobenzene 

Branobenzene 

* · manutactur1ng. 

Uses 

Mfg. of silicones, tetraethyllead, synthetic rubber and methyl cellulose; refrigerant mfg.; mfg. of organic chemicals (methylene chloride, chloroform, etc.); mfg. of fumigants; extractant 

Paint stripping and solvent degreasing; mfg. of aerosols, photographic film, synthetic fibers; refrigerant; fumigant; solvent; textile and leather coatings; pharmaceutical; spotting agent Mfg. of fluorocarbon refrigerants and propellants and plastics; mfg. of anesthetics and ~anmaceuticalsc primary source for chlorodifluoromethane; fumigant; solvent; sweetener 

Mfg. of fire ext1nguishers; dry cleaning operations; mfg. of refrigerants, aerosols and 

propellants; mf~. of chlorofluoromethanes; extractant; solvent; veterinary medicine; metal 

degreasing; fum1gant; chlorinating organic compounds Mfg. of vinyl chloride; mfg. of tetraethyllead; intenmediate insecticidal fumigant (~achtree borer, japanese beetle, root·knot nematodes); tobacco flavoring; constituent in paint~ varnish and finish removers, metal degreaser, soaps and scouring cOffipounds Mfg. or PQlyvinyl chlorides and copolymers; adhesives for plastics Dry cleaning operations; metal degreasing; solvents for fats, greases, waxes{· solvents for 

greases and waxes from cotton, wool, caffeine for coffee; solvent for cellu ose ester and 

ethers; solvent for dyeing; refrigerant and heat exchange liquid; organic synthesis; fumigant 

Dry cleaning operations; ~t3l degreasing; solvents for fats, greases, waxes, rubber, gums, caffeine from coffee; remove soot from industrial boilers; mfg. of paint removers Scavenger for lead in gasoline( grain and fruit fumigant, general solvent, waterproofing preparations, organic synthes1s, insecticide 

Mfg. of aerosol sprays{· mf~. of commercial refrigeration equi~nt; blowing agent for 

polyurethane foams; c eamng compounds; solvent; fire extin~uisher 

Mfg. of styrene, phenol, detergent~l organic chemicals, pest1cides, plastics and resins, synthetic rubber, aviation fuel, pnarmaceuticals, dye, explosives, flavors and perfumes, paints end coatings; nylon intcnncdiatcs; food processing( photographic chemicals 

Mfg. of benzene derivat1ves, caprolactam( saccharin, medic1nes, dyes, perfumes, TNT; solvent recovery plants· component gaso ine; solvent for paints and coatings, gums, resins, rubber and viny{ organosols; diluent and thinner in nitrocellulose lacquers Mfg. phthalic acid and anhydride; mfg. of terepthalic acid for polyester; solvent recovery plants; specialty chemical manufacture; mfg. 1sophthal1c acid, av1ation gasoline, protective coatings mfg.; solvent for alkyd resins, lacquers, enamels, rubber cements; dye mfg. 

Solvent recovery plants; intermediate in dyestuffs mfg.; mfg. aniline, insecticide, phenol, chloronitrobenzene · Mfg. of 3,4·dichloroaniline; solvent; dye mfg.;fumigant and insecticide; metal polishes; industrial odor control · Mfg. moth repellents, mfg. air deodorizers, mfg. dyes and intenmediates; pharmaceuticals mfg.; soil fumigant; pesticide Solvent in chemical manufacturing; dyes and intenmediates; dielectric fluid; synthetic transformer oils; lubricants; insecticides 

Mfg. of TNT; urethane polymers, flexible and rigid foams and surface coatings, dyes; organic synthesis Styrene mfg.; acetophenone mfg.; solvent; asphalt constituen~; naphtha constituent 

Mfg. aniline and dyestuffs; solvent recovery plants; mfg. ruooer chemicals, drugs, photographic chem1cals; refining lubricants oils; solvent in TNT production; solvent for 

cellulose eth~rs; cellulose acetate mfg. Solvent (fats, waxes, or resins); intermediates in specialty organic chmeicals synthesis; additive to motor o1l and fuels 

the numerous uses of the halogenated aliphatic and 

MAH's. 

at 25 oc and stirred at 200 rpm, all five compounds 

(methylene chloride, chloroform, 1,1,1-trichloroethane, 

trichloroethylene, and tetrachloroethylene) demonstrated 

50 percent volatilization in less than 30 min and 90 

percent volatilization in less than 90 min. Addition of 

other material (clay, limestone, sand, salt, peat moss, and 

kerosene) to the water had little effect on rate of volatilization, indicating that sorptive processes were unable to 

compete with solute evaporation. Furthermore, the 

authors determined that chemical degradation half-lives 

for the five compounds ranged from 6 to 18 mo. Table 30 

presents the times for 50 and 90 percent removal of the 

five compounds from water, along with the evaporation 

times of 22 other chlorinated aliphatic hydrocarbons as 

determined by Dilling and others (1975). 

With the exception of the phenols discussed previously, the halogenated aliphatic and MAH's are charac· 

terized by high vapor pressures, high solubilities, and low 

octanol-water partition coefficients. Table 29 quantifies 

these physicochemical parameters. 

Environmental Fate 

The primary fate of most of the halogenated ali· 

phatic and MAH's in surface-water systems is volatilization followed by photolytic degradation in the atmosphere. 

Dilling and others (1975) studied volatilization, sorption, 

and chemical degradation of five chlorinated aliphatic 

compounds at aqueous concentrations of 1 ppm. In water 

52 Manmade Organic Compounds In the Surface Waters of the United States 
Table 29. Experimentally determined octanol-water partition coefficients (Kaw), solubilities, and vapor pressures of selected halogenated aliphatic and monocyclic aromatic 

hydrocarbons at room temperature 

[mg!L, milligrams per liter; mm, millimeters. Dash indicates no data available] 

Log Kow Solubility Vapor pressure 

Compound (mg/L) (mm of Hg) Refs. 

Dichloromethane 20,000 349 1 

Chloroform 1.97 7,950 160 1,2 

Carbon tetrachloride 2.64 800 90 1,2 

1,2-Dichloroethane 8,450 61 1,3 

1,1,1-Trichloroethane 1,360 100 1,3 

1,1,2,2-Tetrach1oroethane 3,230 5 1,3 

1,2-Dibromoethane 3,520 11 1,3 

Vinyl chloride 2,700 2,660 4,1 

Trichloroethylene 2.29 1,100 57.9 5 

1,1,2,2-Tetrachloroethy1ene 2.60 400 14 1,2 

Bromoform 3,190 5.6 1 

1,2-Dichloropropane 3,570 42 1,3 

Benzene 2.13 1,780 76 1,2,6 

Toluene 2.69 470 22 1,2 

o-Xylene 2.77 175 5 1 

Chlorobenzene 2.84 448 8.8 1,2 

1,2-Dich1orobenzene 3.38 148 1.5 1,3 

1,4-Dichlorobenzene 3.38 79 0.6 1,2 

1,2,4-Trichlorobenzene 4.02 19 1,7 

Ethyl benzene 3.15 152 7 1,7 

Nitrobenzene 1.85 1,900 0.15 1 

F1uorobenzene 2.27 1,540 2 

Bromo benzene 2.99 446 3.3 1,2 

Iodobenzene 3.25 340 2 

1. Verschueren (1983) 

2. Chiou and others (1977) 

3. Chiou and others (1979) 

4. Dilling (1977) 

5. Callahan and others (1979) 

6. Mackay {1977) 

7. Chiou, Porter, and Schmedding (1983) 

Other researchers similarly have noted the importance of volatilization. Helz and Hsu (1978) determined 

that evaporative losses were the primary cause of the 

attenuation of halogenated aliphatic hydrocarbons discharged into the Back River estuary in Maryland. Hill and 

others (1976) found that the rate of bulk exchange of 

gaseous vinyl chloride between water and air is about 

twice that of oxygen, and they concluded that volatilization 

probably is the most important process in its distribution. 

Within an 8-d period, Jensen and Rosenberg (1975) 

observed 80 percent removal of trichloroethylene from an 

open aquarium kept in the light. Concentrations of chloroform, 1,1,1-trichloroethane, and tetrachloroethylene 

were decreased by 40 to 50 percent over the same period. 

When the same compounds were incubated in closed 

systems (both light and dark), compound disappearance 

typically was less than 5 percent, indicating that solute 

volatilization was the most important fate process. Similarly, Wakeham and others (1983) stated that volatilization appeared to be the major process responsible for the 

removal of a variety of aromatic and chlorinated aromatic 

hydrocarbons from seawater. 

Using theoretical considerations, Cadena and others 

(1984) presented a model to predict the aqueous concentrations of volatile organic contaminants (compounds 

having Henry's law constants greater than 1.6X 10-4 

atm-m3/mol) in streams as a function of traveltime and 

stream-reaeration coefficient. The derived relationship is 

given by 

where C/Ci is the ratio of the concentration of the 

contaminant at traveltime t to its initial concentration, K2 

is the stream reaeration constant, and K!K' is the ratio of 

the overall transfer coefficient for the volatile contaminant 

Halogenated Aliphatic and Monocyclic Aromatic Hydrocarbons 53 
Table 30. Evaporation times for 50 and 90 percent removal 

of chlorinated aliphatic hydrocarbons [1 milligram per liter] at 

25 oc, 200 revolutions per minute of stirring, and a depth of 

6.5 centimeters 

(From Dilling andothers, 1975) 

Time (minutes) for 

evaporation from water 

Compound 50 percent 

Chloromethane 27 

Dichloromethane 19 

Chloroform 20 

Carbon tetrachloride 29 

Chloroethane 21 

1,1-Dichloroethane 22 

1,2-Dichloroethane 29 

1,1,1-Trichloroethane 20 

1,1,2-Trichloroethane 21 

1,1,1,2-Tetrachloroethane 43 

1,1,2,2·Tetrachloroethane 56 

Pentachloroethane 48 

Hexachloroethane 45 

Vinyl chloride 26 

1,1-Dichloroethylene 22 

cis·Dichloroethylene 18 

trans-Dichloroetbylene 24 

Trichloroethylene 21 

Tetrachloroethylene 25 

1,1,2,3-Tetrachloropropane 51 

1,2,2,3-Tetrachloropropane 47 

3-Chloropropene 27 

2-Chloropropene 29 

1-Chloropropene 16 

2,3-Dichloropropene 20 

1,3-Dich1oropropene 31 

(cis and trans) 

1,2,3-Trichloropropene 49 

90 percent: 

91 

60 

68 

97 

79 

109 

96 

65 

102 

>120 

>120 

>140 

>120 

96 

89 

64 

83 

63 

86 

>120 

>120 

89 

110 

59 

68 

98 

>140 

to the overall transfer coefficient of a tracer compound. 

The value of K2 is given by 

where v is stream velocity (m/s) and L is mean stream 

depth. The value of~ can be corrected for temperature 

by multiplying K2 at 20 oc by (1.0238)T-20 , where T is 

stream temperature in degrees Celsius. Ratios of K/K' are 

presented in table 31 for a variety of halogenated aliphatic 

and MAH's. 

Cadena and othets (1984) verified their model by a 

pilot study using benzene and chlorobenzene at a variety 

of concentrations and reaeration rates. Figure 21 presents 

the results of their model applied to a hypothetical contamination problem in the Rio Grande River in New 

Mexico. The concentrations of both methylene chloride 

and dichlorobenzene are predicted to be below 10 percent 

of their initial concentration within 60 km from the point 

of discharge. The authors state that other halogenated 

aliphatic and MAH's also are likely to volatilize at similar 

rates, inasmuch as methylene chloride and dichlorobenzene represent the extremes in volatility of the compounds tested in this study. 

Once volatilized from solution, many halogenated 

aliphatic and MAH's are degraded photolytically in the 

atmosphere. In laboratory experiments simulating atmospheric conditions, photolysis half-lives for vinylidene chloride, cis- and trans-dichloroethylene, trichloroethylene, 

and vinyl chloride ranged from 5 to 12 h under bright 

sunlight and in the presence of nitric oxide (Dilling and 

others, 1976). Half-lives for chlorobenzene, tetrachloroethylene, and 1,1,2-trichloroethane were slightly longer, 

ranging from 20 to 40 h. Methylene chloride and 1,1,1-

trichloroethane, however, were resistant to photolysis, 

with respective half-lives of > 250 and > 1,700 h reported. 

While the aqueous photodecomposition of many of the 

halogenated aliphatic and MAH's will not be considerable because of rapid volatilization, atmospheric photodecomposition of these compounds may be their ultimate 

fate (Hill and others, 1976; Helz and Hsu, 1978). 

Analysis of available data suggests that some of the 

halogenated aliphatic and MAH's can be biologically and 

(or) chemically degraded in surface-water systems, but 

Table 31. Ratios of overall transfer coefficients for selected 

aliphatic and monocyclic aromatic hydrocarbons (K) to overall transfer coefficients for oxygen (K') 

[From Cadena and others, 1984] 

Compound 

1,4-Dichlorobenzene 

1,2-Dichlorobenzene 

1,3-Dichlorobenzene 

Chlorobenzene 

1,1,2-Trichloroethane 

Tetrachloroethylene 

Carbon tetrachloride 

Benzene 

Trichlorofluoromethane 

Chloroform 

1,1-Dichloroethane 

1,2-Dichloropropane 

1-, 2-Dt-cbl·oroethane 

Dichlorodifluoromethane 

Acrylonitrile 

Chloroethane 

Dichloromethane 

Vinyl chloride 

Methylene chloride 

K/K' 

0.45 

.46 

.46 

.49 

.49 

.50 

.51 

.53 

.54 

.55 

.55 

.57 

.sa-

.ss 

.59 

.61 

.62 

.66 

.73 

54 Manmade Organic Compounds In the Surface Waters of the United States 
1: 

~~ DISTANCE. IN KILOMETERS :.L 

1.0 (l)r---1;.;.0:.___....;:2;,.;;0 __ .:,30r----r--....;,:..--6~0-.... 

0 

..... 

0 

Ol_~--~~-----~==~~==~~ 0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6 

TIME. IN DAYS 

Figure 21. Graph of predicted volatilization rates of methylene chloride and dichlorobenzene in the Rio Grande River, 

N. Mex. C/C1 is the ratio of the concentration of the contaminant at travel timet to its initial concentration. [Reprinted with 

permission from Journal of the Water Pollution Control Federation, v. 56, p: 460 (1984).] 

degradation rates are slow compared with volatilization 

rates for these compounds. Helz and Hsu (1978) noted 

that many halogenated aliphatic compounds were rapidly 

volatilized from a Maryland estuary. However, they also 

observed nonconservative downstream mixing of the contaminants during winter, when the river was ice covered 

and substrate evaporation was minimal. They concluded 

that some biological or chemical degradation also was 

occurring. Wakeham and others (1983) also reported that 

volatilization was the primary fate process for aromatic 

hydrocarbons in seawater, but they noted that summer 

biodegradation of these compounds can also be an important process. In closed systems that were incubated in the 

dark, Jensen and Rosenberg (1975) observed slow degradation of 1,1,2-trichloroethane and 1,1,2,2-tetrachloroethylene. Other compounds, such as vinyl chloride, 

are highly resistant to biological or chemical degradation 

in natural surface waters (Hill and others, 1976). 

Because of their high solubilities and low octanolwater partition coefficients, the majority of the halogenated aliphatic and MAR's will not strongly partition 

into sediment organic matter or biological lipid reservoirs. 

However, some of the chlorinated benzenes have been 

observed to bioascumulate. KOnemann and van Leeuwen 

(1980) found that the increasing octanol-water partition 

coefficients of the chlorobenzenes correlated with 

increased BCFs. Guppies exposed to aqueous chlorobenzene concentrations exhibited lipid-normalized BCF's 

ranging from 1,800 (1,4-dichlorobenzene) to 72,000 

(1,2,3,5-tetrachlorobenzene). Oliver and Niimi (1983) 

observed a similar bioconcentration trend for the chlorobenzenes accumulated by rainbow trout, with BCFs as 

high as 4,100 reported for 1,3,5-trichlorobenzene. Lower 

BCF's were reported for the less chlorinated benzenes. 

Sediments rich in organic matter also can accumulate 

appreciable amounts of tri- and tetrachlorobenzenes. 

Chiou, Porter, and Schmedding (1983) reported organicmatter-normalized sorption coefficients of 500 and 190 

for 1,2,4-trichlorobenzene and 1,4-dkhlorobenzene partitioned between water and a Woodburn silt loam soil 

having 1.9 percent organic-matter content. In contrast, 

the distribution coefficient for chlorobenzene was only 50. 

Based on their tendency to rapidly volatilize from 

solution, aliphatic and MAR's are not persistent environmental contaminants in surface-water systems. Secondary 

fate processes include biological and chemical degradation, but these rates are slow relative to evaporative loss 

rates. With the possible exception of some of the highly 

chlorinated benzenes, sorption and bioaccumulation are 

not important fate mechanisms. 

Environmental Distribution 

Aliphatic and MAR's are natural components of oil 

deposits and commonly find their way into surface waters 

as a result of discharges from refineries, waste oil disposal, and accidental spills. Municipal wastewater discharges also hav(f been recognized as sources of aliphatic 

and MAH discharges (Barrick, 1982). Chlorinated aliphatic hydrocarbons are commonly found in chlorinated 

water and wastewater effluents and in industrial discharges. Chlorinated benzenes may enter the environment as a result of their use as solvents, heat-transfer 

fluids, flame retardants, or chemical intermediates or as 

waste products of the electro-industry (Jan and Malnersic, 1980). 

Point discharges have contributed large amounts of 

aliphatic and MAR's to surface waters. Barrick (1982) 

determined that primary sewage effluent from Seattle 

discharges an average of 475 Mg/yr of aliphatic hydrocarbons into central Puget Sound. Eganhouse and Kaplan 

(1982) similarly noted that southern California treatment 

plants discharge hydrocarbons into coastal waters at an 

annual rate of 17,400 Mg. Brooks and others (1981) 

documented concentrations of aliphatic hydrocarbons with 

5 to 14 carbon atoms to be as high as 400 p.g!L as the 

result of an oil well blowout in the Gulf of Mexico. 

Middleditch and Basile (1980) have reported that oil 

platforms in the northwest portion of the gulf routinely 

discharge 200 g/d of alkanes. Halogenated aliphatic and 

MAR's such as methylene chloride (3 to 8 ppm), toluene 

(13 to 20 ppm), and chlorobenzenes (0.50 ppm) have 

been detected in the effluents of chemical-manufacturing 

plants (Jungclaus and others, 1978; Sheldon and Hites, 

1979). Because of the volatility of the halogenated aliphatic and MAR's, however, these effluent concentrations 

are usually rapidly reduced to much lower levels. 

Halogenated Aliphatic and Monocycllc Aromatic Hydrocarbons 55 
Because of the large quantities of halogenated 

aliphatic and MAR's produced annually in the United 

States, these compounds are ubiquitous contaminants in 

surface waters. Their aqueous concentrations, however, 

are usually in the parts-per-billion range as a result of 

their tendency to volatilize into the atmosphere. Higher 

concentrations usually are limited to specific discharge 

points. While Brooks and others (1981) reported volatile 

hydrocarbon residues as high as 400 p.g!L in the vicinity of 

an oil-well blowout, residues 6 and 12 mi down plume 

were reduced to 63 and 4 p.g/L, respectively. Oliver and 

Nicol (1982) reported low concentrations of chlorobenzenes in the Great Lakes, with the highest mean concentration being observed for 1,4-dichlorobenzene ( 45 ng!L) 

in Lake Ontario. Parts-per-trillion residues have also 

been reported by Schwarzenbach and others (1979) in 

Lake Zurich, Switzerland. Concentrations of 1,4-dichlorobenzene and tetrachloroethylene in the lake's upper basin 

were 10 and 40 1

ng/L, respectively. 

With the possible exception of some of the chlorobenzenes, biological residues of the aliphatic and the 

MAR's are generally in the parts-per-billion range. Ferrario and others (1985) detected a variety of chlorinated 

aliphatic and MAR's in clams and oysters from Lake 

Pontchartrain, La. Detectable concentrations ranged from 

0.04 ppb for chlorobenzene to 310 ppb for 1,1,1-

trichloroethane. Jan and Malnersic (1980) reported total 

chlorobenzene residues in fish from Yugoslavian surface 

waters to range from 1.8 ppm (expressed on a fat basis) 

near industrial areas to 0.2 ppm near agricultural and 

wooded areas. Oliver and Nicol (1982) also quantified the 

concentrations of a variety of chlorobenzenes in trout 

from the Great Lakes. Again, residues were in the 

parts-per-billion range, varying from 0.05 (1,2,3,5-tetrachlorobenzene) to 12 ppb (1,2,3,4-tetrachlorobenzene ). 

Similar to biological residues, sediment residues of 

halogenated aliphatic and MAR's are typically in the 

parts-per-billion range. Chlorobenzene residues in the 

surficial sediments of Lake Ontario have been reported to 

range from 1 to 320 ppb (Oliver and Nicol, 1982). Lower 

concentrations were reported for Lakes Superior, Huron, 

and Erie. Chlorobenzenes sorbed to suspended sediments 

in the Niagara River also had concentrations in the 

parts-per-billion range, with the highest concentration 

reported for 1,4-dichlorobenzene (160 ppb) (Oliver and 

Charlton, 1984). Benzene (21 ppb) and chloroform (18 

ppb) had the highest concentrations of a variety of chlorinated aliphatic and MAR's detected in the sediments of 

Lake Pontchartrain, La. (Ferrario and others, 1985). In a 

study of the effects of an industrial discharge on a stream, 

Jungclaus and others (1978) found that effluent toluene 

and methylene chloride concentrations of 13 to 20 and 3 to 

8 ppm, respectively, caused no detectable residues of 

either compound in the stream sediment. 

Few data were found in the reviewed literature with 

regard to spatial or temporal contamination trends in the 

United States; however, based on production and use data, 

some general conclusions can be drawn. The production 

of halogenated aliphatic and MAR's increased significantly until the late 1970's, when production of many of 

the compounds either slowed or declined (table 27). 

Therefore, temporal trends in surface-water contamination are likely to mirror production. Sediment-core data 

from the Great Lakes confirm this hypothesis. Oliver and 

Nicol (1982) reported that chlorobenzene residues in 

Great Lakes sediments increased rapidly from about 1940 

(6- to 7-cm sediment core depth) to the early 1970's (1- to 

2-cm sediment core depth). Surficial sediment residues 

(0- to 1-cm sediment core depth), however, were generally 

lower than residues at a 1- to 2-cm depth, paralleling the 

decreased production of chlorobenzenes in the late 1970's. 

Despite this decline, surficial sediment residues were still 

higher than sediment residues at the 2- to 3-cm depth 

(1965-71). 

Halogenated aliphatic and MAR's are found principally in industrial and municipal wastewater effluents, 

and, as a result, significant environmental residues of 

these contaminants are most commonly found near industrial areas (Jan and Malnersic, 1980). Surface waters 

draining agricultural or forested areas are less likely to 

have significant residues of halogenated aliphatic and 

MAR's than are densely populated and industrialized 

areas. Nonhalogenated aliphatic and MAR's also are 

commonly found in coastal waters near oil-drilling and 

oil-production areas because of their presence in natural 

oil. 

PHTHALATE ESTERS 

Uses,· Production, and Properties 

The phthalate esters are derivatives of phthalic acid 

(the ortho isomer of benzene dicarboxylic acid) and are 

characterized by the fo11owing structure: 

0 

0=8-o-R 

C-0-R 

II 

0 

where R is either an alkyl or aryl group. Structural 

diagrams of some of the phthalate esters are presented in 

figure 22. 

The phthalate esters are manufactured by the oxidation of either naphthalene or o-xylene to produce 

phthalic anhydride, which is subsequently combined with 

the desired alcohol (Peakall, 1975). Phthalate esters are 

used extensively as plasticizers to produce useful products 

from polymers of vinyl chloride, propylene, ethylene, and 

56 Manmade Organic Compounds In the Surface Waters of the United States 
O

·w 

c-0-CHa 

C-O-CH1 II 

0 

o-Dimethyl phthalate 

~ c-o-c4H, 

c-o-c4 H8 II 

0 

DI-n-butyl phthalate 

Bis(2-ethylhexyl) phthalate 

0 ?. C-O-C2 H 1 

. ~-O-C 1 

0 

Dlethyl phthalate 

0 ?. 

c-o-c8H17 

~-o-c 17 

0 

Dl-n-octyl phthalate 

Butyl benzyl phthalate 

Figure 22. Structural diagrams of some common phthalate 

esters. 

styrene. Commonly, their contribution to the weight of the 

plastic is second only to the weight of the polymer itself 

(Mathur, 1974). Uses of plastics include home construction, appliances, furnishings, automobiles, apparel, and 

food containers and wrappings. Table 32 documents the 

relative use of phthalate esters for these various products 

in the United States, along with the small proportion of 

nonplasticizer use. The amount of solid waste generated 

from these uses has been estimated to be in excess of 12 

billion kg annually in the United States (Mathur, 1974). 

The manufacture of phthalate esters in the United 

States has increased rapidly over the last 25 yr. Production 

from 1961 to 1970 increased at an average annual rate of 

10 percent (Peakall, 1975). Total annual production of 

phthalate esters in the United States reached nearly 

5 X 108 kg in 1978 and has continued to increase ( Giam 

and others, 1978). World production has been estimated 

to be three to four times the U.S. production, and some of 

the leading foreign manufacturers of phthalate esters 

include Japan, the U.S.S.R., and Germany (Peakall, 

1975). Of the various esters of phthalic acid, bis(2-

ethylhexyl) phthalate is one of the most widely used, with 

158.8 million kg produced in the United States in 1970 

(Mathur, 1974). As a result of their large-scale production, phthalates are frequently identified as contaminants 

in the water, sediment, and biota of surface-water systems. 

The phthalate esters are relatively stable, colorless 

liquids of medium viscosity and low volatility. Table 33 

presents some of the physicochemical properties of the 

phthalate esters. 

Table 32. Uses of phthalate esters in the United States 

[From Peakall, 1975] 

Percentage 1 

of total use 

A As plasticizers 

Building and construction 

lUre and cable 

Flooring 

Swimming pool liners 

Miscellaneous 

Home furnishings 

Furniture upholstery 

Yall coverings 

Houseware 

Miscellaneous 

Cars (upholstery, tops, etc.) 

Wearing apparel 

Food wrapping and closures 

Medical tubing and intravenous bags 

Total as plasticizers 

21 

17 

2 

4 

10 

4 

3 

6 

13 

8 

3 

3 

94 

B As nonplasticizers 

Pesticide carriers 

Oils 

Insect repellent 

Total as nonplasticizers 6 

1

A dash indicates that exact data are not available. 

Environmental Fate 

The reviewed literature contained only sparse information on the environmental fate of the phthalate esters, 

and most studies were limited to bis(2-ethylhexyl) phthalate. Based on the available data, however, biodegradation, 

sorption, and bioaccumulation appear to be important 

fate-determining processes. . 

Phthalate esters have been metabolized by both 

isolated and mixed bacterial cultures, undergoing primary 

and ultimate biodegradation by enzyme-catalyzed hydrolysis. Of the group, bis(2-ethylhexyl) and di-n-octyl phthalate have been determined to be the most resistant to 

biodegradation, but data indicate that even these two 

compounds can be at least partially degraded (Tabak and 

others, 1981). Mathur and Rouatt (1975) isolated a 

bacterium, Senutia marcescens Bizio, capable of using 

both bis(2-ethylhexyl) phthalate and di-n-octyl phthalate 

as sole sources of carbon and energy. The primary 

degradation product of these phthalates appeared to be 

a-phthalic acid. Fairbanks and others (1985) observed the 

degradation of bis(2-ethylhexyl) phthalate to C02 in 

sludge-amended soils. Half-lives ranged from 8 to 72 d, 

and after 146 d, 76 to 93 percent of the phthalate ester 

was degraded to C02 • A strain of Enterobacter aerogenes, 

isolated from the inside of plastic tubing used as an outlet 

for deionized water, degraded dimethyl phthalate as a sole 

carbon source (Perez and others, 1977). The culture 

Phthalate Esters 57 
Table 33. Solubilities and vapor pressures of selected phthalate esters 

[m~ milligrams per liter; mm, millimeters] 

Compound 

Dimethyl phthalate 

Diethyl phthalate 

Di-n-butyl phthalate 

Di-n-octyl phthalate 

Solubility 

(mg/L) 

0 

5000 (20 C) 

0 

896 (25 C) 

0 

13 (20 C) 

0 

3 (25 C) 

0 

Vapor pressure 

(mm of Hg) Reference 

0 

0.01 (20 C) 1 

0 

.05 (70 C) 1 

0 

.10 (115 C) 1 

0 

.20 (150 C) 1 

_, 0 

Bis(2-ethylhexyl) phthalate 0.4 (25 C) 2 X 10 (20 C) 1 

Butyl benzyl phthalate 2.9 

1. Callahan and others (1979) 

2. Verschueren (1983) 

degraded a dimethyl phthalate concentration of 1,000 ppm 

to less than 400 ppm in 41 d. Engelhardt and Wallnofer 

(1975) determined that a variety of microorganisms were 

capable of degrading di-n-butyl, di-n-octyl, bis(2-

ethylhexyl), and dimethyl phthalates and suggested that 

mixed microbial populations can readily metabolize 

phthalate esters. Their proposed pathway for the degradation of di-n-butyl phthalate by different soil microorganisms is presented in figure 23. 

As a result of their low solubilities, some of the 

phthalate esters strongly partition into the lipid reservoirs 

of aquatic organisms and sediment organic matter. Table 

34 documents BCFs for bis(2-ethylhexyl), di-n-octyl, and 

di-n-butyl phthalates for a variety of aquatic organisms 

(Sanborn and others, 1975; Callahan and others, 1979). 

The data indicate that all three compounds exhibit significant bioaccumulation. The BCFs range from 130 to 

100,000, with variations reflecting differences in organism 

lipid content, compound solubility, and duration of exposure. Few sorption data are available for the phthalate 

esters. However, Matsuda and Schnitzer (1971) presented data indicating that a variety of dialkyl phthalates 

can partition into dissolved fulvic acid, significantly 

enhancing their apparent solubilities. Using batch sorption studies with solution/soil ratios ranging from 5 to 300, 

Fairbanks and others (1985) noted that as much as 87 

percent of the bis(2-ethylhexyl) phthalate added to centrifuge tubes was sorbed to the glass. ·Of the amount 

remaining, 96 to 99 percent was sorbed to the soil, 

suggesting that significant sorption to sediments may 

occur in natural systems. Because phthalate esters have 

been frequently detected in surface-water sediments and 

biota, it is likely that both bioaccumulation and sorption 

will influence their environmental fate. 

_e 0 

8.6 X 10 (20 C) 1,2 

Table 34. Selected bioconcentration factors (BCF's) of 

phthalate esters in aquatic systems 

[From Callahan and others, 1979, and Sanborn and others, 1975] 

Phthalate ester 

Exposure 

Aquatic organism time (days) 

Bis(2-ethylhexyl) Fish 

Bis(2-ethylhexy1) Sowbug 

Bis(2-ethylhexyl) Fathead minnow 

Bis(2-ethylhexyl). Scud 

Bis(2-ethylhexyl) Snail 

Bis(2-ethylhexyl) Mosquito larvae 

Di-n-octyl 

Di-n-octyl 

Di-n-octyl 

Di-n-butyl 

Di-n-butyl 

Daphnia 

Fish 

Algae 

Waterflea 

Scud 

14 

28 

14 

33 

33 

33 

14 

14 

BCF 

130 

230 

800 

13,400 

21,000 

100,000 

2,600 

9,400 

28,500 

5,000 

6,700 

Nonbiological hydrolysis, volatilization, and photolysis do not appear to be significant fate-determining 

processes for the phthalate esters. Estimated hydrolytic 

half-lives range from 3.2 yr for dimethyl phthalate to 

2,000 yr for bis(2-ethylhexyl) phthalate (Callahan and 

others, 1979). Evaporative losses of phthalate esters from 

surface waters also are minor because of their relatively 

low vapor pressures. Fairbanks and others (1985), for 

example, detected no volatilization of bis(2-ethylhexyl) 

phthalate applied to three sludge-amended soils. Finally, 

58 Manmade Organic Compounds in the Surface Waters of the United States 
~ c-o-c4 H8 

c-o-c4 H8 II 

0 

a,b,c 1 

further 

degradation 

Figure 23. Structural diagrams showing degradation of 

di-n-butyl phthalate (DBP}, mono-n-butyl phthalate, and 

phthalic acid by different microorganisms: (a) organisms 

isolated with DBP as carbon source for enrichment, (b) stock 

cultures, (c) bacteria isolated from phthalic acid enrichment 

medium. [Reprinted with permission from Bulletin of Environmental Contamination and Toxicology, v. 13, p. 344 (1975). 

Copyright 1975 by Springer-Verlag New York, Inc.] 

although no data were found on the photodecomposition 

of phthalate esters in aquatic systems, it is unlikely that 

they are degraded photolytically since they do not possess 

significant absorption maximums in the ultraviolet and 

visi~le regions of the electromagnetic spectrum (Callahan 

and others, 1979). 

Environmental Distribution 

Reported environmental levels of anthropogenically 

produced phthalate esters may be inaccurate because of 

the possible natural origins of these compounds and 

sample contamination by plastic sample containers. There 

is strong evidence that phthalate esters have natural as 

well as anthropogenic origins. Ortho-phthalic acid has 

been detected in wood and in the oxidation products of 

lignin, and as the product of the chemical and biological 

oxidation of compounds, such as naphthalene and its 

derivatives (Mathur, 1974). Phthalides also appear to be 

common constituents of plants, with phthalates serving as 

intermediates in biochemical pathways. Mathur (1974) 

has stated that, because phthalates and aliphatic alcohols 

are present in plants, it is reasonable to assume that 

phthalate esters are synthesized naturally in the environment. Additionally, phthalate esters in plastic sample 

containers can easily contaminate natural water, sediment, or biological samples. For example, phthalate esters 

have been found in extracts of the lining of plastic bags 

and in uncontaminated sand shaken in the bag (Mathur, 

1974). As a result, reported environmental residues of 

phthalate esters may not simply reflect the anthropogenic 

origins of this group of compounds. 

Peakall (1975) has ranked the uses of phthalate 

esters according to likelihood of escape into the environment. Nonplasticizer uses, such as pesticide carriers, 

insect repellent, cosmetics, fragrances, and oils, have the 

highest likelihood of escape, because the phthalate ester 

is not encased in a polymer matrix. Phthalate ester 

plasticizers used directly in contact with water pose the 

second-greatest risk of escaping to the environment and 

include swimming-pool liners, garden hoses, and medical 

uses. Other plasticizer uses that involve only contact with 

air are the least likely to result in phthalate ester loss to 

the environment. Additionally, phthalate esters have been 

reported in industrial discharges. Sheldon and Hites 

(1979) reported concentrations of dibutyl, butyl benzyl, 

and bis(2-ethylhexyl) phthalate in an industrial effluent of 

50, 40, and 200 ppb, respective} y. 

Phthalate esters have been detected in numerous 

coastal- and inland-water samples. Giam and others 

(1978) reported concentrations of di-n-butyl phthalate as 

high as 0.471 ppb and concentrations of bis(2-ethylhexyl) 

phthalate as high as 0.316 ppb in water samples from the 

Gulf of Mexico. Both of these compounds also have been 

detected in the Tama River near Tokyo, Japan, with 

concentrations ranging from 0.4 to 6.8 ppb (Morita and 

others, 1974). Jungclaus and others (1978) reported even 

higher concentrations in a small, freshwater river; concentrations of bis(2-ethylhexyl) phthalate ranged from 1 to 50 

ppb, and concentrations of di-n-octyl phthalate ranged 

from 1 to 20 ppb. Sheldon and Hites (1979) also reported 

the presence of aqueous phthalate esters, with concentrations in the Delaware River near Philadelphia as high as 

0.6 ppb for butyl benzyl phthalate and 1.0 ppb for 

bis(2-ethylhexyl) phthalate. Similarly, maximum bis(2-

ethylhexyl) phthalate concentrations of 1.98 and 3.10 ppb 

Phthalate Esters 59 
were reported by Thuren (1986) in water samples from 

the Rivers Svartan and Ronnebyan, respectively, in southern Sweden. 

Sediment residues of phthalate esters also have 

been reported frequently, with detected concentrations 

usually ranging from a few parts per billion to a fraction of 

a part per million. In a study of the Chester River in 

Maryland, Peterson and Freeman (1984) reported maximum sediment residues of diethyl, diallyl, diisobutyl, 

dibutyl, dihexyl, bis(2-ethylhexyl), and di-n-octyl phthalate 

of 44, 5.5, 27 .6, 900, 6.4, 4,800, and 62 ppb, respectively. 

The maximum concentrations of dibutyl phthalate ( 69 

ppb) and bis(2-ethylhexyl) phthalate (248 ppb) found in 

sediment samples from the Gulf Coast and Mississippi 

River Delta were slightly lower (Giam and others, 1978). 

Much higher sediment concentrations were reported by 

Jungclaus and others {1978) and Van Luik (1984). The 

former study detected both bis(2-ethylhexyl) and di-noctyl phthalate at concentrations ranging from 250 to 

56,000 ppb in sediment samples collected downstream 

from a chemical manufacturing plant's discharge. The 

study by Van Luik (1984), which analyzed sediments from 

the Calumet Sag Channel (Ill.), determined that the 

average sediment concentration of bis(2-ethylhexyl) 

phthalate was 29,800 ppb, with a high value of 63,770 ppb. 

Laboratory studies have indicated that phthalate 

esters can bioaccumulate, and these results have been 

confirmed in field studies identifying significant residues 

of phthalate esters in aquatic organisms. Despite aqueous 

bis(2-ethylhexyl) phthalate concentrations of 0.316 ppb or 

less, biota from the Gulf of Mexico contained concentrations ranging from 1 to 135 ppb (Giam and others, 1978). 

De Vault (1985) detected high concentrations of both 

bis(2-ethylhexyl) and di-n-butyl phthalate in two fish from 

Cl 

0 

Cl 

CH30H 

Cl Na OH 

Cl 

Cl 

the Menominee and Kinnickinnic Rivers in Wisconsin. 

Unfortunately, no aqueous concentrations were reported. 

The ubiquity of phthalate esters is evidenced by the data 

of Stalling and others (1973), who detected residues of 

bis(2-ethylhexyl) phthalate in fish collected throughout 

the United States. Concentrations ranged from 0.2 to 10 

ppm and appeared highest near industrialized areas. 

POLYCHLORINATED DIBENZO-P-DIOXINS 

Uses, Production, and Properties 

The polychlorinated dibenzo-p-dioxins (PCDD's) 

are characterized by the following structure: 

nc•-0::)()-ncl 

Chlorination of the molecule can produce a total of 75 

chlorinated species, ranging from monochlorodibenzo-pdioxin to octachlorodibenzo-p-dioxin. 

Although not deliberately manufactured and having 

no commercial use, PCDD's are produced inadvertently 

during the manufacture of2,4,5-trichlorophenol (2,4,5-T) 

from 1,2,4,5-tetrachlorobenzene (fig. 24 ). Because 2,4,5-

T is used in the production of a variety of pesticides along 

with tetra- and pentachlorophenol, PCDD's have been 

found as contaminants in many of these compounds. 

Table 35 documents the detected residues of tetra-, hexa-, 

hepta-, and octachlorodibenzo-p-dioxin in several pestiCl 

2 ,4,5-Trlchlorophenol 

Cl 

2,4,5-T 1,2,4,5-Tetrachlorobenzene ~ CIOOOCI Cl 0 Cl 

TCDD 

Figure 24. Structural diagrams showing formation of 2,3,7,8-tetrachlorodibenzo-p-dioxin 

(TCDD) as a byproduct in the synthesis of 2,4,5-trichlorophenoxyacetic acid (2,4,5-T}. 

[Reprinted with permission from Journal of Environmental Quality, v. 7, p. 172 (1973). 

Copyright 1973 by the American Society of Agronomy, Crop Science Society of America, 

and Soil Science Society of America.) 

60 Manmade Organic Compounds in the Surface Waters of the United States 
Table 35. Chlorinated dibenzo-p-dioxins found in commercial pesticides 

[From Helling and others, 1973] 

Chlorodibenzo-p-dioxin 

detected SaDJRl~~ 

Pesticide Tetra- Hexa- Hepta- Octa- Contaminated Tested 

2,4,5-T ** ** 23 42 

Silvex * 1 7 

2,4-D * 1 24 

2,4-DB 0 3 

2,4-DEP 0 2 

Dich1orprop 0 1 

Erbon ** 1 1 

Sesone * 1 1 

Dicamba 0 8 

Ch1orophenols 

Tri- * * * 4 6 

Tetra- **   3 3 

Penta- **  ** ** 10 11 

Others ** ** * 5 22 

* Concentration of at least one sample was between 0.5 and 10 parts 

per million (ppm) 

** Concentration of at least one sample was greater than 10 ppm 

Concentration was less than the 0.5 ppm detection limit 

DMPA, ronnel, and tetradifon had chlorodioxin residues; chloroneb, 

hexachlorophene, nematocide, and nitrofen had no detectable 

residues 

cides that were manufactured between 1950 and 1970. 

Although the extent of contamination is likeiy to vary with 

manufacturer, one company produced 2,4,5-T from 1966 

to 1968 that typically contained 10 ppm of 2,3,7,8-

tetrachlorodibenzo-p-dioxin (TCDD). Of 42 samples of 

2,4,5-T tested, 23 contained detectable residues of 

PCDD's. One analyzed sample of pentachlorophenol 

contained nearly 1,000 ppm of both hepta- and 

octachlorodibenzo-p-dioxin (Helling and others, 1973). 

More recent commercial production of such compounds 

as 2,4,5-T has resulted in much lower residues, with 

TCDD concentrations usually less than 0.1 ppm (Isensee 

and Jones, 1975). 

PCDD's are extremely toxic compounds. They were 

first identified as environmental contaminants in 1970 as 

a result of a report by the Office of Science and Technology noting that mice given a relatively large dose of the 

herbicide 2,4,5-T during early pregnancy produced an 

abnormally high number of deformed offspring (Helling 

and others, 1973). Analysis of the 2,4,5-T given to the 

mice revealed that it contained approximately 27 ppm of 

TCDD. Subsequent studies have shown that PCDD's 

have very high mammalian toxicities. Table 36 reports the 

single oral doses of some PCDD's required to kill 50 

percent of the exposed animals (LD50 ). TCDD appears to 

be the most toxic of the PCDD's, with an LD50 of just 2 

Jtg/kg for guinea pigs. TCDD LD50 values of 10 Jtg/kg for 

rabbits and 0.6 Jtg/kg for guinea pigs have also been 

documented (Isensee and Jones, 1975). 

Some of the less severe effects of PCDD exposure 

to humans include changes in skin color, skin rashes, 

excessive hair growth, tingling and numbness of the arms 

Polychlorinated Oibenzo-p-dioxins 61 
Table 36. Single oral doses of dioxins necessary to kill 50 

percent of exposed animals (LD50) 

(p.g/kg bw, micrograms per kilogram of body weight. From Kriebel, 

1981) 

Dibenzo-p-dioxin 

2,3,7,8-

tetrachloro1,2,3,7,8-

pentachloro1,2,3,4,7,8-

hexachloroAnimal 

Guinea pig 

Rat (male) 

Rabbit 

Mouse 

Guinea pig 

Mouse 

Guinea pig 

Mouse 

LDso 

(pg/kg bw) 

2 

22 

115 

284 

3 

338 

73 

825 

and legs, and liver damage (Kriebel, 1981). TCDD exposure has also been linked to certain cases of birth defects. 

Pregnant women living near Alsea, Oreg., experienced a 

significant increase in the number of spontaneous abortions during the months of June and July from 1972 to 

1977. From March to June of each of these years, 

herbicides, including 2,4,5-T, were routinely applied to 

nearby forests (Kriebel, 1981). Similarly, many Vietnam 

veterans claim that PCDD's in Agent Orange (approximately a 1:1 mixture of 2,4,5-T and 2,4-D) applied to 

Vietnamese forests have caused birth defects in their 

children (Kriebel, 1981). These examples indicate that 

even the low residues of PCDD's found in certain herbicides may cause adverse health effects owing to the high 

toxicity of the PCDD's. 

Inasmuch as PCDD's are not deliberately manufactured, only gross estimates of production can be made. 

Based on a combined annual production figure of 30,000 

t for 2,4,5-T, and tri-, tetra-, and pentachlorophenol in the 

United States, and assuming that the production of each 

kilogram of the technical grade formulation results in the 

production of 1 mg of PCDD, then approximately 30 kg 

are produced each year from these sources. Similarly, 

more than 40,000 t of Agent Orange were sprayed over 

the jungles of Southeast Asia, with the formulation having 

a mean TCDD concentration greater than 1.9 mg!kg 

(Crosby and Wong, 1977). 

Although PCDD's are produced as byproducts in 

the manufacture of trichlorophenol, their properties and 

environmental behavior more closely resemble those of 

the chlorinated insecticides. As a group, they are sparingly soluble, lipophilic compounds of relatively long 

environmental persistence. Increased chlorination correlates with decreased solubility and increased lipophilicity. 

Table 37. Bioconcentration factors (BCF's) of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) for a variety of aquatic organisms 

[From Isensee and Jones, 1975] 

Organism Exposure time, days BCF 

Algae 30 2,000-18,600 

Duckweed 30 1,200-5,000 

Snails 30 1,400-47,100 

Daphnids 30 7,800-48,000 

Mosquito fish 3 1,000-63,300 

Catfish 6 2,000-27,900 

TCDD, which has been the subject of most of the 

research on PCDD's, has a solubility of about 3 to 5 ppb 

(Biddinger and Gloss, 1984) and an octanol-water partition coefficient of 4.24X_l06 (Marple and others, 1986). 

Environmental Fate 

Almost all the published data on the fate of the 

PCDD's is limited to TCDD. As a result, any conclusions about the environmental fate of the PCDD's are at 

best qualitative. 

As a result of their relatively low solubilities, the 

PCDD's are likely to strongly sorb to sediments and to 

bioaccumulate in aquatic organisms. Chemical and biological degradation processes appear very slow, making 

the PCDD's persistent environmental contaminants. Few 

data are available regarding the volatilization of PCDD's 

from surface-water systems. 

Extremely low concentrations of PCDD's can cause 

significant residues in surface-water biota. By contaminating aquatic microcosms with TCDD sorbed to sediment, Isensee and Jones (1975) found that BCF's for a 

variety of organisms ranged from 1,000 to 63,300. After 

30 d of exposure, equilibrium aqueous concentrations of 

TCDD as low as 50 pg!L caused TCDD concentrations in 

daphnids to be as great as 2.4 p.g!L (corresponding to a 

BCF of 48,000). Other bioconcentration data from their 

study are summarized in table 37. Callahan and others 

(1979) have reported that mosquito larvae bioaccumulated TCDD to concentrations 2,800 to 9,200 times greater 

than the aqueous concentration. BCF's determined for 

brine shrimp and silversides were 1,600 and 54, respectively. Unfortunately, no bioaccumulation data were found 

for the other PCDD's. Based on the strong tendency of 

TCDD to bioconcentrate, however, it is likely that other 

PCDD's of equal or greater chlorination will also be 

readily bioaccumulated. 

62 Manmade Organic Compounds In the Surface Waters of the United States 
Available data based on studies of TCDD indicate 

that the PCDD's strongly partition into sediment organic 

matter. The solubility of TCDD (3 to 5 ppb) is approximately equal to the solubility of DDT, which has been 

documented as having sorption coefficients (~) greater 

than 104 and organic carbon/organic matter normalized 

sorption coefficients (K.,c,om) greater than 106 (table 7). 

Helling and others (1973) have noted that TCDD applied 

to fields coincident with 2,4,5-T application will not 

undergo significant leaching by runoff. Only transport 

due to soil erosion would be likely. Ward and Matsumura 

(1978) confirmed the conclusions of Helling and others 

(1973) in a study employing anaerobic sediment-water 

microcosms dosed with radio-labeled TCDD. They found 

that typically 93 to 96 percent of the recovered radioactivity was from the sediment and speculated that much of 

the remaining TCDD was associated with dissolved 

organic matter. Similar results also were reported by 

Isensee and Jones (1975), who observed that 85 to 99 

percent of the TCDD originally added to a model ecosystem remained in the sediment after a 39-d period. 

PCDD's appear to be highly resistant to both 

chemical and biological degradation. In a laboratory 

sediment-water system incubated under anaerobic conditions, the half-life of TCDD was found to be approximately 600 d (Ward and Matsumura, 1978). Similarly, 

Helling and others (1973) reported that TCDD applied to 

moist soils exhibited half-lives of approximately 1 yr in the 

laboratory. TCDD half-lives in soil greater than 10 yr also 

have been documented (Moore and Ramamoorthy, 1984). 

There is some evidence that the PCDD's can be 

chemically degraded by photolysis; however, photolytic 

decomposition rates will probably be unable to compete 

with sorption and bioaccumulation rates in surface-water 

systems. Crosby and Wong (1977) demonstrated that 

TCDD dissolved in organic solvents and exposed to 

natural sunlight on leaves and glass plates was rapidly 

photodegraded, with half-lives of less than 6 h. The 

organic solvent was believed to serve as a hydrogen donor, 

because in its absence photolytic degradation was greatly 

reduced. Longer half-lives were observed for TCDD on 

soils, and this was attributed to upper soil particles 

shading lower soil particles. The soil's protective effect 

against the photolytic degradation of TCDD also has been 

noted by Plimmer (1978). In surface-water systems, the 

rapid sorption of TCDD to sediments greatly reduces the 

amount of light reaching the TCDD molecules, thereby 

reducing the rate of photodecomposition. This is confirmed by an experiment by Ward and Matsumura (1978), 

who compared the degradation of TCDD in sedimentwater microcosms incubated in the dark with identically 

prepared microcosms incubated in the light. After 39 d, 

the average recovery of TCDD from the "dark" microcosms was 93.29 percent and the average recovery of 

TCDD from the "light" microcosm was 89.45 percent, 

indicating the lack of a significant photolytic influence. In 

aquatic systems, it is likely that PCDD's will be rapidly 

sorbed to sediments or bioaccumulated by aquatic organisms. 

Environmental Distribution 

Because PCDD's are byproducts of any manufacturing process that involves the production or use of 

trichlorophenol, they are common I y discharged to surface 

waters in effluents from chemical manufacturing plants, 

leather tanneries, kraft -pulp mills, \mod-processing plants, 

or sewage-treatment plants (Moore and Ramamoorthy, 

1984). For example, VanNess and others (1980) analyzed 

a series of industrial effluents and determined that 5 of 

the 20 samples contained detectable amounts of TCDD, 

with concentrations ranging up to 100 ppt. PCDD's also 

enter the environment in conjunction with the application 

of pesticides, as discharges from chemical disposal sites, 

and from fly ash from municipal incinerators. Table 38 

gives estimates of the input of TCDD's into the Canadian 

environment from a variety of sources. 

There is some evidence that PCDD's can be produced by the combustion of trichlorophenol or its derivatives. Small quantities of TCDD have been produced by 

burning 2,4,5-T under conditions simulating a forest fire 

(Kriebel, 1981), and it has been estimated that as much as 

1 p.g of PCDD's will be produced for every square meter 

of forest that burns immediately after a herbicide treatment. Additionally, 2,4,6-trichlorophenol applied to dried 

leaves and allowed to burn resulted in a residual TCDD 

concentration of 2,100 ppm, about 100,000 times greater 

than the TCDD concentration in the original aliquot of 

trichlorophenol (Kriebel, 1981). 

Upon entering surface-water systems, PCDD's 

strongly partition from water into sediments and biota. As 

a result, aqueous residues of PCDD's are generally below 

detectable limits, even with a significant input of PCDD's 

to the system. Despite low aqueous concentrations, sediment and biological residues of PCDD's have been 

reported in the parts-per-trillion to parts-per-billion range. 

VanNess and others (1980) analyzed soil samples 

from a trichlorophenol-manufacturing plant. Eighty-five 

percent of the soil samples contained detectable residues 

of TCDD, with concentrations ranging from < 20 ppt to 

600 ppb. O'Keefe and others (1984) detected a TCDD 

concentration of 10 ppt in a sediment sample collected 

from the lower Hudson River (N.Y). Two other samples 

from the lower Hudson, however, contained TCDD concentrations of less than 7.3 ppt. Kriebel (1981) summarized the results of a study by Dow Chemical Company 

that reported PCDD residues in soils from a variety of 

locations in the United States, including soils from the 

grounds of one of Dow's own chemical-manuf?cturing 

plants. The ranges of the detectable concentrations are 

Polychlorinated Dibenzo-p-dioxins 63 
Table 38. Estimated annual inputs of polychlorinated dibenzo-p-dioxins to the Canadian 

environment from major sources 

[From Moore and Ramamoorthy, 1984) 

Annual inputs, in kilograms 

British 

Source Columbia Prairies Ontario Quebec Maritimes 

Pentachlorophenol 570 

usage 

2,4-D usage .3 

Precipitated fly < .1 

ash 

Airborne fly ash < .1 

Total 570 

presented in table 39. Most of the detectable concentrations were less than 3 ppb, with the exception of soil 

samples taken from the Dow plant, which were highly 

contaminated with isomers of tetra-, hexa-, hepta-, and 

octachlorodibenzo-p-dioxin. Based on these data, contaminated soils carried into surface waters by runoff can be a 

significant source of sediment TCDD residues. 

Because of the high toxicity of PCDD's, and particularly TCDD, numerous researchers have documented 

biological residues of these compounds in edible fish. As 

a result of the heavy application of Agent Orange to South 

Vietnamese forests, carp and catfish sampled downstream 

from the sprayed areas contained average TCDD residues 

ranging from 70 to 810 ppt (Kriebel, 1981; Moore and 

Ramamoorthy, 1984). TCDD residues as high as 57 ppt 

have even been detected in the fat of some Vietnam 

veterans (Kriebel, 1981). Table 40 summarizes the 

detected residues of TCDD in a variety of fish from both 

fresh and marine waters. Although few studies have 

reported on the more highly chlorinated dioxins, they 

have also been detected in aquatic organisms. Concentrations as high as 388 (oct a-), 304 (hepta-), 203 (hexa-), and 

80 (penta-) ppt have been found in carp from the Niagara 

River (Moore and Ramamoorthy, 1984). 

POLYCYCLIC AROMATIC HYDROCARBONS 

Uses, Production, and Properties 

The polycyclic (or polynuclear) aromatic hydrocarbons (PAR's) are a large group of environmentally important compounds. Although little is known about their 

360 350 110 130 

8.9 < .1 < .1 < .1 

< .1 1.6 5.1 < .1 

1.5 2.2 2.2 .6 

370 354 117 131 

toxicities, several of the PAR's have been identified as 

carcinogens or mutagens. As a group, the PAR's are 

structurally characterized by two or more fused-ring 

compounds based on benzene, with properties intermediate between benzene and olefinic hydrocarbons (Harrison 

and others, 1975). Two aromatic rings are considered to 

be fused when carbon atoms from each ring are shared. 

The environmentally significant PAR's have from two 

(naphthalene) to seven (coronene) rings. Structural diagrams of some of these compounds are given in figure 25. 

PAR's originate from both natural and anthropogenic sources; however, it is believed that their occurrence in 

aquatic systems is due primarily to anthropogenic inputs 

(Cossa and others, 1983). PAR's are produced mainly by 

high-temperature ( > 700 °C) pyrolytic reactions such as 

municipal incineration. or forest fires. Their residues also 

have been detected in fossil fuels and wood-preservative 

sludges (table 41). 

A few of the simpler PAR's are produced commercially. Naphthalene is used in several chemical manufacturing processes to produce moth balls, pesticides, fungicides, dyes, wetting agents, synthetic resins, cutting fluids, 

solvents, and lubricants (Verschueren, 1983). Commercial 

production of petroleum naphthalene in the United States 

was 4.9X104 tin 1976, 4.7X104 tin 1980, and 6.5X1~ t 

in 1981 (Moore and Ramamoorthy, 1984). The production of chloronaphthalenes (used in automobile capacitors 

and as oil additives for fabric dyeing and engine cleaning) 

is much less, with U.S. production probably less than 270 

t per year (Moore and Ramamoorthy, 1984). Commercial 

uses of other PAR's such as anthracene, acenaphthene, 

and phenanthrene are small and limited primarily to dye 

and plastics manufacturing (Verschueren, 1983). Pyrolitic 

64 Manmade Organic Compounds in the Surface Waters of the United States 
Table 39. Residues of chlorodibenzo-p-dioxins (micrograms per kilogram) in soil samples 

from various locations in the United States 

[From Kriebel, 1981] 

Range of detectable concentrations 

Location Tetra- Hexa- Hepta- OctaIn and around 

Dow Chemical plant 1-120 7-280 70-3200 490-20,000 

Metropolitan area .005-.03 .03-.3 .1-3 .4-22 

Urban area nd .03-1.2 .03-2 .05-2 

Rural area nd nd 0.05 0.2 

nd - no sample exceeded the limit of detection 

Table 40. Mean and maximum concentrations of detected residues of 2,3,7,8-

tetrachlorodibenzo-p-dioxin in a variety of freshwater and marine fish 

[ppt, parts per trillion] 

Aquatic Mean Maximum 

Locality organism (ppt) (ppt) Reference 

Rhode Island coast Striped bass 3.4 5.0 1 

Little Neck Bay, Long Striped bass 16.3 39 1 

Island, NY 

Newark Bay, NJ Striped bass 37.8 67 1 

Lower Hudson River Striped bass 61.4 120 1 

Chesapeake Bay, MD Striped bass 2.8 3.5 1 

Great Lakes Yellow perch 3.8 4.3 2 

Great Lakes Brown bullhead 6.0 8.6 2 

Great Lakes White sucker 3.0 4.0 2 

Great Lakes Channel catfish 15.5 17.7 2 

Great Lakes American eel 19.8 38.5 2 

Great Lakes Rainbow smelt 20.0 32.9 2 

Lake Ontario Lake trout 61.2 3 

East coast of Canada Ocean haddock 3.6 3 

Lake Huron Lake trout 30.4 3 

Lake Ontario Rainbow trout 32.3 3 

Michigan rivers Channel catfish 157 695 4 

Michigan rivers Carp 55 153 4 

Michigan rivers White sucker 11 21 4 

Lake Ontario Brown trout 14 162 4 

Lake Ontario Brown bullhead 4 5 4 

Lake Erie Coho salmon 1.4 1.9 4 

1. O'Keefe and others (1984) 

2. Ryan and others (1984) 

3. Ryan and others (1983) 

4. Moore and Ramamoorthy (1984) 

Polycyclic Aromatic Hydrocarbons 65 
00 

Naphthalene Anthracene 

Fluoranthene 

Phenanthrene 

Pyrena Benzo(a)pyrene 

lndeno( 1.2.3-cd)pyrene Coronene 

Figure 25. Structural diagrams of selected polycyclic aromatic hydrocarbons. 

sources are probably the major input of PAH's to the 

environment. It has been estimated, on the basis of data 

from 1966 to 1969, that the global emission of 

benzo(a)pyrene from anthropogenic sources is about 4.6 

million kglyr (Edwards, 1983). Table 42 lists the sources 

and approximate amounts of benzo(a)pyrene emitted to 

the atmosphere. 

In general, the PAH's are compounds having relatively low solubilities and vapor pressures and high 

octanol-water partition coefficients. These properties vary 

with the number of compound rings, as indicated in table 

43. As the number of rings increases, compound solubility 

and vapor pressure generally decrease, whereas the 

octanol-water partition coefficient increases. Acute toxicity is greatest for the low-molecular-weight PAH's, such as 

naphthalene, and increases with alkyl substitution (Van 

Luik, 1984). Benzo(a)pyrene is perhaps the most highly 

carcinogenic compound of the group, but other compounds, such as benzo(b )fluoranthene, benzoG)fluoranthene, benzo(a)anthracene, chrysene, and indeno(1,2,3-

cd)pyrene, also have been recognized as carcinogens 

(Harrison and others, 1975). 

Environmental Fate 

The PAH's are persistent surface-water contaminants. Because of their low solubilities and high octanolwater partition coefficients, they strongly partition from 

water into biota and particulate and dissolved organic 

matter. As a result, even low aqueous concentrations can 

result in environmentally significant sediment and biological residues. There is some evidence that PAH's can be 

photolytically degraded in the water column, but the 

importance of this process in natural systems is likely to 

be small because of the strong sorption of PAH's to 

sediments. The ultimate fate of the PAH's probably is 

sorption to sediments followed by slow biodegradation. 

Other transport and transformation processes, such as 

volatilization and hydrolysis, are not important fate processes. 

Several laboratory studies have investigated the 

sorption of PAH's to sediments. Gardner and others 

(1979) exposed sand and salt marsh sediment to crude oil 

containing small amounts of anthracene, fluoranthene, 

benzo(a)anthracene, and benzo(a)pyrene. Sorption of all 

four compounds was extensive, with the sediments having 

the highest organic-matter content sorbing the greatest 

amount of PAH's. PAH concentrations as high as 10 

mg!kg also have been reported in rapid-sand-filter solids 

(Harrison and others, 1975). In a study comparing anthracene sorption to organic matter (yeast cells) and mineral 

particulate matter (powdered calcite), Herbes (1977) 

observed that the particulate organic matter rapidly 

sorbed the anthracene in solution, with equilibrium 

reached in minutes. In the presence of a 250-mgiL 

yeast-cell concentration, 72 percent of an initial anthracene concentration of 0.02 ~giL was sorbed by the 

organic matter. In contrast, the concentration of calcite 

required to sorb just 30 percent of the initial anthracene 

concentration was 300 giL, indicating that anthracene is 

sorbed by partitioning into organic matter and that adsorption to mineral surfaces is relatively insignificant. Hassett 

and others (1980) also noted the importance of organic 

matter in the sorption of dibenzothiophene to several soils 

and sediments. Laboratory-measured sorption coefficients ranged from 5.8 to 388. 

The sediment -water partition coefficients normalized for organic-carbon content (~c) for a variety of the 

PAH's can be calculated from solubility data using the 

empirical relation derived by Banwart and others (1982): 

log ~c= -0.686 log S+4.273, 

where S is compound solubility, in milligrams per liter. 

The relation was developed from data from Karickhoff 

66 Manmade Organic Compounds In the Surface Waters of the United States 
Table 41. Residues of polycyclic aromatic hydrocarbons (PAH's) in oils, gasoline, bitumen, and sludge 

[In parts per million except for wood preservative sludge, which is in grams per liter of raw sludge; km, kilometer. Dash indicates no data available. 

From Verschueren, 1983] 

Residue in 

Wood preserv- Motor oil 

Coqx>und Gasoline Crude oil Bitunen ative sludge Fresh Used Used ~5,000 km} ~10 000 km} 

Anthracene 1.55·2.59 

Benzo(a)anthracene 0.04-0.27 1.7-2.3 0.13-0.86 5.18 

Benzo(b)fluoranthene 0.16·3.9 <1.0 0.40-1.60 0.08 45-82 56-141 

Benzo(j)fluoranthene 0.009 <1.0 0.31 

Benzo(k)fluoranthene 0.009 <1.3 0.34-1.41 

Benzo(g,h,i)perylene 0.32·9 <1.6 1.37·5.50 0.12 109-208 153·289 

Benzo(a)pyrene 0.13-8.28 0.4·2.8 0.02·0.1 83.2·162 110-242 

Chrysene 0.05-2.96 6.9-17.5 1.64·5.14 0.56 86·190 129-237 

Coronene 0.06·1.9 o.oo 25·37 28·63 

Dibenz(a,h)anthracene 0.167 0.07 

Fluoranthene 0.70·7.56 2.9·5.0 26.47 0.11 109·173 129-270 

Fluorene 6.61 

Jndeno(1,2,3·cd)pyrene 0.04·2.88 0.03 34·59 47·83 

Phenanthrene 15.7-20.5 26·70 

Pyrene 1.55-17.1 3.5-4.5 0.17·0.80 24.2 0.29 286·450 330·700 

and others (1979) and Banwart and others (1980), along 

with original sorption data for three nitrogen-heterocyclic 

compounds. Table 44 presents the calculated values of 

~c for some of the PAH's along with experimentally 

determined Koc values for additional PAH's from 

Karickhoff and others (1979). The ~c values indicate 

that a sediment sample having only 0.1 percent organiccarbon content may have an equilibrium benzo(a)pyrene 

concentration of 8.6 X Hf X 0.001, or 860 times greater 

than the compound's concentration in the surrounding 

water. Means and others (1980, 1982) also determined 

~c values for several PAH's sorbed to 14 sediment and 

soil samples. Mean values for the 14 samples ranged from 

517 ( anthracene-9-carboxylic acid) to 1.8 X 106 (3-

methylcholanthrene ). These data indicate the strong tendency of benzo(a)pyrene and other PAH's of similarly low 

solubilities to partition into sediment organic matter. 

In addition to their tendency to sorb to sediments, 

PAR's also accumulate in the lipid reservoirs of aquatic 

organisms, with BCF's generally increasing as compound 

solubility decreases. This observation is confirmed by the 

data of Southworth and others (1978b ), who studied the 

bioaccumulation of seven PAH's by Daphnia pulex. They 

determined that the octanol-water partition coefficients of 

the PAH's were good estimators of BCF's, with 

laboratory-measured BCFs ranging from 100 for naphthalene to 10,000 for benzo(a)anthracene. Casserly and 

others (1983) also found that BCF's increase with 

decreasing solubility. Exposure of the alga Selenast1Um 

capricornutum to aqueous residues of naphthalene, phenanthrene, and pyrene resulted in BCFs of 12,600, 24,000, 

and 36,300, respectively. BCF's as great as 10,000 also 

have been documented for PAH's bioaccumulated by fish 

(Mackay, 1982a). Table 45 lists BCF's determined for a 

variety of PAH's. For a given compound, differences in 

BCF's determined for different organisms are at least 

partially attributable to differences in the lipid contents of 

the organisms. For a given compound, BCF's will increase 

with increasing lipid content of the aquatic organism. 

In aqueous systems, PAH's are photodegraded by 

reactions with oxygen resulting in the formation of quinones (Smith and others, 1978). Although several laboratory experiments have documented short photolytic 

half-lives for PAH's in distilled water samples, photodecomposition in natural systems may be relatively insignificant owing to the strong partitioning of PAH's into 

particulate and dissolved organic matter. Smith and others 

(1978) measured photolytic half-lives of several hours for 

benzo(a)anthracene and benzo(a)pyrene in oxygensaturated distilled water exposed to sunlight. However, 

when half-lives were similarly determined in natural 

waters or in distilled water containing dissolved humic 

Polycyclic Aromatic Hydrocarbons 67 
Table 42. Estimated annual benzo(a)pyrene (BaP) emissions to the atmosphere 

[kg, kilogram. Values in parentheses are percentages of the total 

emission. From Edwards, 1983) 

Sources 

Heating and power 

Industrial processes 

Enclosed incineration 

Coal refuse burning 

Forest and agricultural fires 

Other refuse burning 

Trucks and buses 

Automobiles 

TOTAL 

• Concentration (kg X 10 ) 

United States Global 

0.43 (36.8)* 2.36 (51.6)* 

.18 (15.4) ** .95 (20.8>** 

• 03 (2.6) .09 (2.0) 

.31 (26.5) .62 (13.6) 

.13 (11.1) .38 (8.3) 

.07 (6.0) .13 (2.8) 

.01 (0.8) . 03 (0.7) 

.01 (0.8) .01 (0.2) 

1.17 4.57 

* 91 percent from coal and 8 percent from wood 

** 99 percent from coke production 

acids, photolysis was significantly retarded. The low 

absorbance of the natural water used indicated that light 

screening alone pro~ably was not responsible for the 

differences in photolysis rates. Partition of the two PAH's 

into the dissolved organic matter, however, may have 

altered the reactivity of the molecules and thereby 

reduced the photolysis rates. Photolytic degradation of 

PAH's may be an important fate process only when the 

PAH's are present in oil slicks or in shallow, low-turbidity 

waters with low dissolved-organic-matter content (Callahan and others, 1979) . 

Several studies have documented the biodegradation of PAH's.· Low-molecular-weight PAH's, such as 

naphthalene, acenaphthene, and acenaphthylene, have 

been rapidly degraded in laboratory cultures, with greater 

than 90 percent degradation of initial concentrations of 5 

to 10 mg!L occurring within 7 d (Tabak and others, 1981 ) . 

PAH's of higher molecular weights, such as fluoranthene, 

benzo(a)anthracene, chrysene, benzo(a)pyrene, and 

anthracene, however, are much more resistant to biodegradation. For example, Smith and others (1978) were 

unable to develop a microbial culture capable of degrading benzo(a)pyrene or benzo(a)anthracene when they 

were present as sole carbon sources. Figure 26 presents a 

proposed degradation pathway for phenanthrene by soil 

pseudomonads that is typical for three-ring PAH's (Harrison and others, 1975). 

Table 43. Experimentally determined octanol-water partition coefficients (l<ow), solubilities, and vapor pressures of selected polycyclic aromatic hydrocarbons (PAH's) 

[mg!L, milligrams per liter; mm, millimeters. From Callahan and others, 1979] 

Nwnber of Log K0 w Solubi1ity0 Vapor pressure 0 

PAH rings (mg/L at 25 C) (mm of Hg at 20 C) 

Naphthalene 2 3.37 34.4 0.0492 

-" Phenanthrene 3 4.46 1.00 6.8 X 10 

Anthracene -" 3 4.45 0.045 1.95 X 10 

_7 

Pyrene 4 nd 0.14 6.85 X 10 

_v 

Benzo(a)anthracene 4 5.61 0.014 5 X 10 

_11 

Benzo(k)fluoranthene 4 nd nd 9.59 X 10 

Chrysene 4 5.61 0.002 nd 

_o* Benzo(a)pyrene 5 6.04 0.0038 5 X 10 

Benzo(ghi)pery1ene 6 nd 0.00026 nd 

* - vapor pressure determined at 25°C. 

nd - no data available. 

68 Manmade Organic Compounds In the Surface Waters of the United States 
Table 44. Calculated and experimental sediment-water 

sorption coefficients normalized for organic carbon (K0 J for a 

variety of polycyclic aromatic hydrocarbons (PAH's) 

There is evidence that PAH's are slowly biodegraded after they have been sorbed to sediments. Gardner 

and others {1979) exposed fine sand, medium sand, and 

marsh sediment to crude oil 'enriched with benzo(a)pyrene, benzo(a)anthracene, anthracene, and fluoranthene. 

Based on initial PAH concentrations ranging from 7.6 to 

73.9 p.gjg dry sediment, the percent degradation per week 

of the four PAH's ranged from 0.84 to 3.5 percent, with 

benzo( a)pyrene having the slowest degradation rate. The 

authors also noted that the presence of a polychaete 

worm, Capitella capitata, enhanced the degradation of the 

PAH's either by improving the sediment conditions for 

microbial degradation or by metabolically degrading the 

PAH's themselves. In a similar study, Varanasi and others 

{1985) determined that over a 4-week period, PAH's, 

including benzo(a)pyrene, present in sediments were not 

significantly degraded. However, benzo(a)pyrene was 

metabolized by both shrimp and fish· exposed to the 

contaminated sediment. These studies indicate that a 

possible fate of PAH's in surface-water systems is sorption to sediments followed by slow biodegradation. 

* Solubility K 

PAH (milligrams per liter) oc 

Naphthalene a 34.4 1.3 X 10 (E) 

Phenanthrene .. 1.00 2.3 X 10 (E) 

Fluoranthene .. .265 4.7 X 10 (C) 

Pyrena .. .14 8.4 X 10 (E) 

Anthracene .045 2.6 X 10 .. (E) 

Benzo(a)anthracena .014 3.5 X 10 s (C) 

Benzo(a)pyrene I .0038 8.6 X 10 (C) 

Chrysene • .002 1.3 X 10 (C) 

Benzo(ghi)pery1ena .00026 • 5.4 X 10 (C) 

C- Calculated according to Banwart and others (1982). 

E - Experimental values are from Karickhoff and others 

(1979). 

* - All solubility data are from Callahan and others (1979) 

except for f1uoranthene, which is from Harrison and 

others (1975). Based on their relatively low vapor pressures (table 

43), PAH's are not expected to volatilize from water at a 

significant rate, and this observation has been confirmed 

in laboratory studies. Southworth {1979) predicted the 

Table 45. Bioconcentration factors (BCF's) for selected polycyclic aromatic hydrocarbons 

[Dash indicates no data available] 

Exposure 

Compound Organism time (days) BCF Reference 

Naphthalene Danhnia l!ulex 1 131 1 

Naphthalene s. caEricorm!tY:m 1 12,500 2 

Naphthalene Fish 427 3 

Ch1oronaphthalenes Grass shrimp 3 50-300 4 

Acenaphthene Fish 389 3 

Phenanthrene DaEhnia l!ulex 1 325 1 

Phenanthrene s. cal!ricornutum 1 23,800 2 

Phenanthrene Fish 2,630 3 

2-Ch1orophenanthrene Fish 4,266 3 

2-Methy1phenanthrene Fish 3,020 3 

Anthracene Dai!hni~ l!ulex 1 917 1 

Anthracene Fish 912 3 

9-Methylanthracene );!a:Qhni,a :QUlex 1 4,583 1 

9-Methylanthracene Fish 4,571 3 

Pyrene Da:Qhnia :Qulex 1 2,702 1 

Pyrene s. caEricornutum 1 36,300 2 

Pyrene Fish 2,692 3 

Fluorene Fish 1,288 3 

Benzo(a)anthracene DaEhnis :Qulex 1 10,109 1 

Benzo(a)anthracene Fish 10,000 3 

Benzo(a)pyrene Fish 22 4 

Benzo(a)pyrene Snail 3,000 4 

Pery1ene QaRhnii.l :Qulex 1 7,191 1 

1. Southworth and others (1978b) 3. Mackay (1982a) 

2. Casserly and others (1983) 4. Biddinger and Gloss (1984) 

Polycyclic Aromatic Hydrocarbons 69 
Phenanthrene 

o8-~-~7 ! 

OH 

-oa 

1 ..... ··•····• .... :< 

~OH 

~OH 

hCH:OzH 

~OH 

a-Hydroxymuconlc aemialdehyde 

Figure 26. Structural diagrams showing proposed degradation pathway of phenanthrene by soil pseudomonads. 

[Reprinted with permission from Water Research, v. 9, p. 337 

(1975). Copyright 1975 by Pergamon Press ltd.] 

volatilization rates of several PAH's and determined that 

volatilization would not be a significant attenuation 

method in streams with depths greater than 1 m and 

average velocities less than 1 m/s. In a stream 1.0 m deep 

at 25 °C, predicted volatilization half-lives of benzo(a)-

anthracene and benzo(a)pyrene were greater than 100 h 

for current and wind velocities up to 1 and 4 m/s, 

respectively. Similarly, volatilization half-lives ofbenzo(a)-

pyrene and benzo(a)anthracene in aqueous solutions 

stirred at a fast rate have been determined to be 22 and 89 

h, respectively (Smith and others, 1978). The presence of 

sediments will likely increase these experimental half-lives 

as a result of PAH sorption. 

The structure of PAH's, for the most part, prevents 

hydrolytic degradation, a:nd this process is not expected to 

be an important fate process in surface-water systems 

(Callahan and others, 1979). 

Environmental Distribution 

PAH's enter surface-water systems in a variety of 

ways, including atmospheric deposition, surface runoff 

and soil leaching, industrial discharges, and municipal 

wastewater effluents. Heit and others (1984) presented 

data indicating that fluoranthene is atmospherically deposited in remote lakes in Rocky Mountain National Park. 

Similarly, Heit (1985) determined that atmospheric deposition is the prime source of PAH's for Cayuga Lake, 

N.Y. Runoff into the Delaware River estuary between 

Trenton and Philadelphia has been found to contain both 

particulate and dissolved quantities of petroleum hydrocarbons, many of which were identified as PAH's (Whipple 

and Hunter, 1979). Naphthalene concentrations ranging 

from 1 to 4 mg!L have been reported by Jungclaus and 

others (1978) in the wastewater of a specialty chemicals 

manufacturing plant. Barrick (1982) estimated that 

approximately 1 t per year of 3- to 7-ring PAH's are 

discharged to Puget Sound from Seattle's primary municipal wastewater effluents. Likewise, Eganhouse and 

Kaplan (1982) determined that southern California's 

wastewater-treatment plants annually discharge more than 

17,400 t tons of aliphatic and aromatic hydrocarbons to 

coastal waters. 

The origin of PAH residues in surface waters 

commonly can be determined by the ratio of the concentration of parent hydrocarbons to their methyl isomers. 

PAH's of petroleum origin consist of homologous series 

in which the methyl and higher alkyl-substituted isomers 

great I y outnumber the parent compounds (Pancirov and 

Brown, 1977). In contrast, PAH's of pyrolitic origin (such 

as the combustion of fossil fuel or wood) contain predominantly the parent hydrocarbons (Pancirov and Brown, 

1977). Therefore, PAH's in a water, sediment, or biota 

sample having a high concentration of pyrene relative to 

methylpyrene are likely to be of pyrolytic origin, while 

PAH's in samples having a high methylpyrene concentration relative to pyrene are likely a result of direct contamination from petroleum or petroleum-derived products. 

The limitation of this reasoning is the assumption that 

both the parent hydrocarbon and its methyl isomer have 

the same aquatic fates. If, for example, pyrene is biodegraded at a faster rate than methylpyrene, the initial 

parent/methyl isomer ratio will be upset. 

Because of their relatively low solubilities, PAH's 

are usually found only at low aqueous concentrations in 

surface waters. Sheldon and Hites (1979), for example, 

measured aqueous methylnaphthalene concentrations of 

0.02 and 0.2 p.g!L at two locations in the Delaware River 

estuary. Similarly, Stainken and Frank (1979) detected 2-

to 5-ring PAH's at concentrations ranging from 2 to 20 

p.g!L in the bottom waters of Raritan Bay (N.J.). Downstream from a chemical manufacturing plant's effluent, 

Jungclaus and others (1978) reported low concentrations 

of naphthalene ( 6 to 10 J.t.g/L) in river water. Harrison and 

others (1975) have also reported low aqueous concentrations of PAH's. Their maximum reported concentration 

70 Manmade Organic Compounds in the Surface Waters of the United States 
Table 46. Sediment residues of polycyclic aromatic hydrocarbons 

(p.glkg, micrograms per kilogram. Dash indicates no data available) 

Mean Max. 

Compound Locality (pg/kg) (pgjkg) Reference 

Benzo(a)pyrene Adirondack acid lake 409 690 1 

region, NY 

Benzo(a)pyrene Yestern U.S. lakes 38,000 305,000 2 

Benzo(a)pyrene Lake George, NY 30 3 

Benzo(a)pyrene Casco Bay, ME 122 805 4 

Phenanthrene Adirondack acid lake 239 324 1 

region, NY 

Phenanthrene Lake George, NY 150 3 

Pyrene Lake George, NY 150 3 

Pyrene Calumet-Sag Channel 6,300 10,000 5 

Fluoranthene Lake George, NY 160 3 

Fluoranthene Rocky Mountain lakes 165 375 6 

P~rylene Adirondack acid lake 443 605 1 

region, NY 

Perylene Lake George 1 NY 16 3 

Perylene Cayuga Lake 1 NY 207 510 7 

Benzo(g,h, i)- Adirondack acid lake 830 1,356 1 

perylene region, NY 

Benzo(k)fluor- Casco Bay, ME 77.8 193 4 

anthene 

Anthracene Adirondack acid lake 27 32 1 

region, NY 

Benzo(a)anthracene Lake George, NY 30 3 

Dibenzo(a,c/a,h)- Cayuga Lake. NY 54 120 7 

anthracene 

Indeno(c,d)pyrene Adirondack acid lake 805 1,294 1 

region, NY 

Coronene Adirondack acid lake 495 801 1 

region, NY 

Acenaphthene Casco Bay, ME 

Chrysene/ Cayuga Lake, NY 

tripheny1ene 

1. Heit and others (1981) 

2. Heit {1979) 

3. Heit and others (1980) 

4. Larsen and others {1983) 

for any of the PAH's in several European rivers was only 

0.76 ~giL. 

As a result of their generally low aqueous solubilities and high octanol-water partition coefficients, PAH's 

will strongly partition into sediment-associated organic 

matter, and, consequently, sediments act as a reservoir for 

PAH's. Although aqueous PAH concentrations in surface 

waters are typically in the parts-per-trillion to low partsper-billion range, sediment PAH concentrations can be 

several orders of magnitude higher. For example, Jungclaus and others (1978) detected significant residues of 

phenanthrene (0.2 to 25 ppm), fluorene (2 to 10 ppm), 

acenaphthylene (0.2 to 5 ppm), and pyrene (0.5 to 75 

ppm) in the sediment of a river downstream from a 

chemical manufacturing plant's discharge. Aqueous res293 1,150 4 

176 605 7 

5. Van Luik {1984) 

6. Heit and others (1984) 

7. Heit (1985) 

idues of these same compounds, however, were not 

detected. Table 46 presents documented residues of a 

variety of PAH's in sediment samples collected from U.S. 

surface waters. Benzo(a)pyrene concentrations greater 

than 300 ppm have been detected in the sediments of 

Lake Hansen, Los Angeles, and numerous researchers 

have reported sediment PAR concentrations ranging from 

10 ppb to 10 ppm. 

In addition to strongly partitioning into sediment 

organic matter, PAR's also accumulate in the lipid reservoirs of aquatic organisms. Table 47 presents mean and 

maximum reported concentrations of PAR's in aquatic 

organisms and illustrates that residues greater than 400 

ppb have been reported. The data in table 4 7 from Brown 

and Pancirov (1979) are concentrations of PAR in fish 

Polycyclic Aromatic Hydrocarbons 71 
Table 47. Residues of polycyclic aromatic hydrocarbons in biota 

[#lg/kg, micrograms per kilogram. Dash indicates no data available] 

Aquatic Mean Max. 

Compound Locality organism (pgfkg) (pg/kg) Reference 

Pyrene Baltimore Canyon Summer flounder 2.7 4.1 1 

Pyrene Manhasset Neck, NY Mussels 370 2 

Pyrene Lake George, NY Mussels 90 3 

Benzo(a)pyrene Lake Erie Fish stomachs 7.1 15.5 4 

Benzo(a)pyrene u.s. east coast Shellfish 1.3 3.0 5 

Methylpyrene Baltimore Canyon Summer flounder 1.9 2.7 1 

lndeno(l,2,3- Lake Erie Fish stomachs 15 15 4 

c,d)pyrene 

Fluoranthene Biloxi, MS Mussels 90 2 

Fluoranthene Lake George, NY Mussels 60 3 

Benzo(a)- Baltimore Canyon Summer flounder 0.7 1.1 1 

anthracene 

Benzo(a)- u.s. east coast Shellfish 2.0 8.0 5 

anthracene 

Dibenz(a,h)- Lake Erie Fish stomachs 2.3 4.1 4 

anthracene 

Chrysenefbenz- Kinnickinnic River, Carp 460 6 

(a) anthracene tJI 

Chrysene Lake Erie Fish stomachs 13.1 26.1 4 

Naphthalene Biloxi, MS Mussels 120' 2 

Phenanthrene Lake George, NY Mussels 60 3 

Perylene Lake George, NY Mussels 300 3 

Fluorene Lake Erie Fish stomachs 15.4 15.4 4 

1. Brown and Pancirov (1979) 4. Maccubbin and others (1985) 

2. Farrington and others (1983) 5. Pancirov and Brown (1977) 

3. Heit and others (1980) 6. DeVault (1985) 

from the Baltimore Canyon area off the coast of New 

Jersey and Delaware. Despite the lack of any oil or gas 

production in the area, biological residues of PAR's were 

still found at significant concentrations (up to 4.1 ppb ), 

indicating their strong tendency to bioaccumulate. 

Although no studies have addressed temporal contamination trends of PAR's in surface waters across the 

Nation, several localized studies have reported data indicating that PAH residues in sediments have been increasing over time. In a study of two New York lakes, Heit and 

others {1981) analyzed sediment cores for a variety of 

PAR's. They determined that with the exception of perylene, anthropogenically derived PAH concentrations 

decreased with increasing sediment depth until a background concentration was reached at a depth at which 

sediment was estimated to be about 30 yr old. Hurtt and 

Quinn (1979) examined PAH residues in sediment cores 

collected from Narragansett Bay, R.I., and also discovered 

that PAH residues steadily decreased with sediment core 

depth, with concentrations stabilizing at a depth of 20--25 

em. Likewise, Heit and others {1984) observed that the 

highest fluoranthene concentrations in Rocky Mountain 

National Park occurred in the surficial sediments of four 

lakes. Lower concentrations were observed in deeper 

sediments. Finally, Bates and others {1984) noted that 

PAH concentration changes in the sediments of the central 

main basin of Puget Sound, Wash., paralleled the urbanization of Seattle. 

Several studies have also investigated regional PAH 

contamination trends and determined that the highest 

PAH residues are typically found near highly populated, 

industrialized areas. Larsen and others {1983) reported 

strongly regional PAH contamination trends in the sediments of Casco Bay, Maine. In general, total PAH concentrations in the bay's upper and middle sections were 

less than 900 ppb. Sediment samples from stations near 

the metropolitan area of Portland, however, had total 

PAH concentrations in excess of 2,000 ppb. Heit {1979) 

noted similar results after collecting and analyzing sediment samples from a variety of western lakes for residues 

of benzo(a)pyrene. The highest concentrations of the 

carcinogenic PAH were found in Lake Hansen, which is 

located in the urban environment of Los Angeles. Farrington and others (1983) sampled mussels from 62 

locations along the Atlantic and Pacific coasts of the 

United States and analyzed them for PAH residues. 

Again, mussels sampled from stations near urban areas 

had elevated PAH concentrations relative to mussels from 

stations near rural, sparsely populated areas. 

72 Manmade Organic Compounds In the Surface Waters of the United States 
SUMMARY 

This report critically reviews the scientific literature 

with regard to the processes influencing the occurrence 

and distribution of manmade organic compounds in 

surface-water systems. The distribution and concentration 

of organic compounds in surface-water systems are 

af~ect~ by sorption, bioaccumulation, photolysis, hydrolysis, biodegradation, and volatilization. 

In aqueous systems, nonionic organic compounds 

are sorbed by partitioning between water and the 

sediment-associated organic matter. Adsorption does not 

contribute significantly to sorption owing to the competitive advantage polar water molecules have for surface 

adsorption sites. Sediments having high organic-matter 

contents exhibit the greatest sorption capacity for nonionic 

organic compounds, and compounds having low aqueous 

solubilities exhibit the strongest tendency to partition into 

sediment organic matter. Ionic organic compounds may 

be able to compete successfully with polar water molecules for surface adsorption sites and, therefore, may be 

sorbed by an adsorption process. 

Organic compounds can similarly partition into the 

lipid reservoirs of aquatic organisms, and available data 

indi~te that the extent of bioaccumulation depends primanly on an organism's lipid content rather than its 

P?sition ~n. th~ aq~atic food chain. The popular concept of 

biOmagmficatiOn m aqueous systems is not well supported 

by the scientific literature. 

Photolysis, hydrolysis, and biodegradation act to 

r~uce the ~ncentratio? of an organic compound by 

either a chemical or a biological transformation. Volatilization also reduces the solute's concentration in the 

surface-water system, but does so by physically partitioning the solute into the atmosphere. The rates of all four 

processes are commonly modeled by first-order kinetics 

in which the rate of transformation and (or) transport i~ proportional to the solute's concentration in the water. 

Both the compound and environmental parameters that 

influence these processes are discussed. 

Data regarding production and use, environmental 

fate, and environmental distribution ·are presented for 

eight groups of anthropogenic organic compounds. 

Polychlorinated biphenyls (PCB's) and chlorinated insecticides, despite almost complete cessation of their production more than 10 yr ago, remain at significant concentrations in the sediments and biota of surface waters across 

the United States and around the world. Their low 

solubilities cause them to strongly partition into particulate organic matter and biological lipid reservoirs. They 

are extremely persistent organic contaminants and are 

highly resistant to biodegradation, photolysis, and hydrolysis: Although their vapor pressures are relatively low, 

theu unusually high activities in water permit them to be 

volatilized from solution and transported atmospherically. 

In natural systems, aqueous concentrations of PCB's and 

c~l~rinated insecticide~ are typically low (in the parts-pertnlhon range), but sediment and biological concentrations 

in the parts-per-million range have been reported. Regionally, the highest PCB residues are commonly found in 

waters near highly populated, industrial areas while the 

highest chlorinated insecticide residues are oft~n found in 

surface waters draining agricultural areas. Surface-water 

residues of PCB's and chlorinated insecticides in the 

United States appear to be slowly decreasing, and this is 

attributable to North American bans on production and 

use of most of these compounds. 

Organophosphorus and carbamate insecticides have 

seen widespread use in recent years as replacements for 

many of the chlorinated insecticides. They are readily 

degraded by both chemical and biological processes, and, 

as a result, they are among the least environmentally 

persistent pesticides. Their generally high aqueous solubilities prevent significant partitioning into organic phases, 

and, therefore, the organophosphorus· and carbamate 

insecticides do not tend to accumulate in sediments or 

biota. When present in surface-water systems, their highest concentrations are typically found in aqueous samples, 

but these residues typically do not persist for prolonged 

periods. 

Herbicides, characterized by high aqueous solubilities and low vapor pressures, have been increasingly 

applied to major crops in the United States, w~th atrazine 

and 2,4-D being the most extensively used. Similar to the 

organophosphorus and carbamate insecticides the herbi- . , c1des are relatively nonpersistent environmental contaminants that are readily biodegraded or chemically attenuated by photolytic or hydrolytic reactions. Herbicide 

bioaccumulation and sorption to sediments is minor, with 

~he possi~l~ exception of the dinitroaniline and bipyridylmm herbicides. In natural systems, herbicide residues are 

generally found in the aqueous phase. Aquatic herbicides 

are applied directly to water bodies, and their residues are 

therefore greatest immediately following application. Terrestrial herbicides are transported to surface waters primarily in the dissolved phase of surface runof~ and, 

consequently, residues of these herbicides are usually 

greatest if a heavy rain immediately follows application of 

the herbicide. In view of the increasing use of herbicides 

in the United States, it is unlikely that herbicide residues 

in surface waters will decrease in the near future. 

Phenols are a widely used group of compounds 

manufactured primarily by the organic chemical and 

pesticide industries. As a group, the phenols are of 

generally high solubility. However, increasing chlorination 

of the molecule results in decreased solubility and 

increased octanol-water partition coefficients. Phenols are 

attenuated primarily by biodegradation and photolysis. 

Sorption to sediments is generally minor, inasmuch as the 

relatively high solubility of phenols prevents substantial 

Summary 73 
partitioning of the solute into sediment-associated organic 

matter and the negative charge of particle surfaces minimizes adsorption of phenolate anions. Sediments having 

high organic-matter contents, however, have been observed 

to sorb appreciable amounts of the highly chlorinated 

phenols because of their lower solubilities. Similarly, 

phenols are not readily bioaccumulated, with the exception of the highly chlorinated phenols such as pentachlorophenol. Phenols commonly enter surface waters in 

industrial or municipal effluents, and chlorination of 

these effluents can convert phenol to mono-, di-, or even 

trichlorophenol isomers. Despite their large-scale production, phenolic residues in surface waters usually are not 

found at concentrations greater than 1 ppm. 

Halogenated aliphatic and monocyclic aromatic 

hydrocarbons are industrial chemicals that have a variety 

of uses, including solvent degreasers, refrigerants, and 

styrene and dye manufacturing. They are highly soluble in 

water, and their 1

primary fate in natural waters is volatilization followed by atmospheric photolysis. While many 

of these compounds have been biologically and chemically degraded in the laboratory, the degradation rates are 

generally not fast enough to compete successfully with 

volatilization from the water body. Sorption to sediments 

and bioaccumulation are not important fate processes, 

with the possible exception of the chlorinated benzenes. 

Because of the large quantities of halogenated aliphatic 

and monocyclic aromatic hydrocarbons produced annually, they are ubiquitous surface-water contaminants. Sediment, biological, and aqueous environmental residues, 

however, rarely exceed the parts-per-billion range. Spatially, halogenated aliphatic and monocyclic aromatic 

hydrocarbons are typically found near industrial and 

highly populated areas. Temporally, there is some evidence that environmental residues have decreased slightly 

since the late 1970's, paralleling a similar decrease in 

production of these compounds in the United States. 

Phthalate esters represent a relatively new class of 

environmental pollutants. They are extensively used as 

plasticizers to produce useful products from polymers of 

vinyl chloride. As a result of their low solubilities and 

nonionic nature, they tend to strongly partition into particulate organic matter and biological lipid reservoirs. 

Therefore, aqueous residues exceeding 50 ppb rarely 

have been reported, whereas sediment and biological 

residues exceeding 1 ppm have been reported. Reported 

environmental levels of phthalate esters should be viewed 

with caution, however, because there is strong evidence 

that they may be produced naturally. Additionally, environmental samples can be contaminated with phthalate 

esters by the use of plastic sampling containers. The 

ultimate fate of phthalate esters in surface waters appears 

to be biodegradation. 

Although not deliberately manufactured and having 

no commercial use, polychlorinated dibenzo-p-dioxins 

(PCDD's) enter the environment as impurities in compounds whose production involves the use of 2,4,5-

trichlorophenol. These compounds include many pesticides and both tetra- and pentachlorophenol. Although 

the quantity of PCDD's produced each year is probably 

very small, they are extremely persistent contaminants 

that have been shown to be highly toxic to laboratory 

animals (particularly 2,3,7 ,8-tetrachlorodibenzop-dioxin). PCDD's strongly sorb to sediments and soils, 

and erosion of soil treated with PCDD-contaminated 

pesticides is a likely input of PCDD's to surface-water 

systems. PCDD's also strongly partition into biological 

lipid reservoirs, and aqueous concentrations as low as 50 

pg!L have resulted in biological concentrations as great as 

2.4 p.g!L. There is some evidence that PCDD's can be 

degraded photolytically; however, the majority of PCDD's 

in surface waters will probably be unavailable for photolysis owing to their rapid sorption to sediments. The 

highest PCDD residues in surface-water systems are 

found in the sediments and biota, with reported concentrations typically in the parts-per-trillion to parts-perbillion range. 

Polycyclic aromatic hydrocarbons (PAH's) are produced by high-temperature pyrolytic reactions, such as 

municipal incineration, forest fires, and fuel combustion. 

Their residues also have been detected in fossil fuels and 

wastewater sludges. Important environmental-fate processes include sorption and bioaccumulation, and the 

ultimate fate of PAH's in surface-water systems may be 

sorption to sediments followed by slow biodegradation. 

Photolytic degradation of PAR's has been demonstrated 

in laboratory studies, but in natural systems it is uncertain 

if photolysis is a significant fate process owing to the rapid 

sorption of PAH's to sediments. Because of their low 

solubilities, aqueous residues of PAR's in natural systems 

are typically low relative to sediment and biological concentrations. Based on sediment core data, PAH residues 

in surface-water systems appear to be increasing with 

time, and concentrations seem to parallel industrialization 

both temporally and spatially. 

SELECTED REFERENCES 

Abdullah, M.I., Ringstad, 0., and Kveseth, NJ., 1982, Polychlorinated 

biphenyls in the sediments of the Inner Oslofjord: Water, Air, and 

Soil Pollution, v. 18, p. 485-497. 

Acree, J., Beroza, M., and Bowman, M.C., 1963, Codistillation of DDT 

with water: Journal of Agricultural Food Chemistry, v. 11, p. 

278-283. 

Adams, R.S., 1973, Factors influencing soil adsorption and bioactivity 

of pesticides: Residue Reviews, v. 47, p. 1-45. 

Addison, R.F., Paterson, S., and Mackay, Donald, 1983, The predicted 

environmental distribution of some PCB replacements: Chemosphere, v. 12, no. 6, p. 827-834. 

Agatwal, H.C., Mittal, P.K., Menon, K.B., and Pillai, M.K.K., 1986, 

DDT residues in the river Jamuna in Delhi, India: Water, Air, and 

Soil Pollution, v. 28, p. 89-104. 

74 Manmade Organic Compounds In the Surface Waters of the United States 
Agatwal, H.C., Saxena, D.M., and Lal, Rup, 1982, Accumulation and 

metabolism of DDT and its metabolites by Tetrahymena: Water, 

Air, and Soil Pollution, v. 18, p. 441-447. 

Akhavein, A.A., and Linscott, D.L., 1968, The dipyridilium pesticides: 

Paraquat and diquat: Residue Reviews, v. 23, p. 98-140. 

Alexander, Martin, 1985, Biodegradation of organic chemicals: Environmental Science and Technology, v. 18, no. 2, p. 106-111. 

Alexander, M., and Aleem, M.I.H., 1961, Effect of chemical structure 

on microbial decomposition of aromatic herbicides: Agricultural 

and Food Chemistry, v. 9, no. 1, p. 44-47. 

Allison, L.E., 1965, Organic carbon, in Black, C., Evans, D., White, J., 

Ensminger, L., and Clark, F., eds., Methods of soil analysis, pt. 2: 

Madison, Wis., American Society of Agronomy, p. 1367-1378. 

Almgren, Mats, Grieser, Franz, Powell, J.R., and Thomas, J.K., 1979, 

A correlation between the solubility of aromatic hydrocarbons in 

water and micellar solutions, with their normal boiling points: 

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Organochlorine residues in adipose tissue of residents of the 

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Banerjee, Sujit, Howard, P .H., Rosenberg; A.M., Dombrowski, A.E., 

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Banerjee, Sujit, Sugatt, R.H., and O'Grady, D.P., 1984, A simple 

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Banerjee, Sujit, Yalkowsky, S.H., and Valvani, S.C., 1980, Water 

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Selected References 75 
Birmingham, B.C., and Colman, Brian, 1983, Potential phytotoxicity of 

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Bovey, R.W., Burnett, Earl, Richardson, Clarence, Merkle, M.G., 

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Bowman, B.T., and Sans, W.W., 1985a, Partitioning behavior of 

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Boyle, E.A., Edmond, J.M., and Sholkovitz, E.R., 1977, The mechanism of iron removal in estuaries: Geochimica et Cosmochimica 

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Bradley, J.R., Sheets, T.J., and Jackson, M.D., 1972, DDT and 

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Brooks, J.M., Wiesenburg, D.A., Burke, R.A., and Kennicutt, M.C., 

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Buser, Hans-Rudolf, 1986, Polybrominated dibenzofurans and 

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aromatic solutes: Environmental Science and Technology, v. 17, 

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Canton, J.H., Greve, P.A., Slooff, W., and van Esch, G.J., 1975, 

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Canton, J.H., van Esch, G.J., Greve, P.A., and van Hellemond, 

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Caron, G., Suffet, I.H., and Belton, Tom, 1985, Effect of dissolved 

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Carranza de Storani, M.M., Rosas, S.B., and Ghittoni, N.E., 1986, 

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Casserly, D.M., Davis, E.M., Downs, T.D., and Guthrie, R.K., 1983, 

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Chapman, B.M., 1982, Numerical simulation of the transport and 

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76 Manmade Organic Compounds in the Surface Waters of the United States 
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Chiou, C.T., and Freed, V.H., 1977, Environmental Science and 

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Chiou, C.T., Freed, V.H., Schmedding, D.W., and Kohnert, R.L., 1977, 

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Chiou, C.T., Kohnert, R.L., Freed, V.H., and Tonkyn, R.G., 1983, 

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Chiou, C.T., Malcolm, R.L., Brinton, T.I., and Kile, D.E., 1986, Water 

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Chiou, C.T., and Manes, M., 1980, On the validity of the codistillation 

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Chiou, C.T., Peters, L.J., and Freed, V.H., 1979, Physical concept of 

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---1981, Science, v. 213, p. 684. 

Chiou, C.T., Porter, P.E., and Schmedding, D.W., 1983, Partition 

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Chiou, C.T., Porter, P.E., and Shoup, T.D., 1984, Environmental 

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Chiou, C.T., Schmedding, D.W., Block, J.H., and Manes, M., 1982, 

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Chiou, C.T., Schmedding, D.W., and Manes, Milton, 1982, Partitioning 

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Chiou, C.T., and Shoup, T.D., 1985, Soil sorption of organic vapors 

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Chiou, C.T., Shoup, T.D., and Porter, P.E., 1985, Mechanistic roles of 

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Christman, R.F., and Gjessing, E.T., 1983, Priorities in humic research, 

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Clark, J.R., Patrick, J.M., Moore, J.C., and Forester, Jerrold, 1986, 

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Clayton, J.R., Pavlou, S.P., and Breitner, N.F., 1977, PCB's in coastal 

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Coates, J.T., and Elzerman, A.W., 1986, Desorption kinetics for 

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Cole, R.I-1., Frederick, R.E., Healy, R.P., and Rolan, R.G., 1984, 

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Connolly, J.P., Armstrong, N.E., and Miksad, R.W., 1983, Adsorption 

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---1984b, Comparison of the carcinogenic risks from fish vs. 

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Conrad, Ralf, and Seiler, Wolfgang, 1985, Characteristics of abiological carbon monoxide formation from soil organic matter, humic 

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Cossa, D., Picard-Berube, M., and Gouygou, J.P., 1983, Polynuclear 

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Criddle, C.S., McCarty, P.L., Elliott, M.C., and Barker, J.F., 1986, 

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Czuczwa, J.M., McVeety, B.D., and Hites, R.A., 1984, Polychlorinated 

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Dahchour, A, Bitton, G., Coste, C.M., and Bastide, J., 1986, Degradation of the herbicide propanil in distilled water: Bulletin of 

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Dao, T.H., Lavy, T.L., and Dragun, J., 1983, Rationale of the solvent 

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Davani, B., Ingram, J., Gardea, J.L., and Eiceman, G.A., 1986, Organic 

compounds in soils and sediments from unlined waste disposal 

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Davis, J.A., 1982, Adsorption of natural dissolved organic matter at 

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Selected References 77 
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Delfino, JJ., 1979, Toxic substances in the Great Lakes: Environmental Science and Technology, v. 13, no. 12, p. 1462-1468. 

DePinto, J.V., Young, T.C., and Martin, S.C., 1980, Aquatic sediments: 

Journal of the Water Pollution Control Federation, v. 52, no. 6, p. 

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---1981, Aquatic sediments: Journal of the Water Pollution 

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DeVault, D.S., 1985, Contaminants in fish from Great Lakes harbors 

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Dhamotharan, S., Gulliver, J.S., and Stefan, H.G., 1981, Unsteady 

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Dietrich, A.M., Millington, D.S., and Christman, R.F., 1983, Specific 

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Dietrich, W.E., 1982, Settling velocity of natural particles: Water 

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Dilling, W.L., 1977, Interphase transfer processes. II. Evaporation 

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Dilling, W.L., Bredeweg, C.J., and Tefertiller, N.B., 1976, Organic 

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Science and Technology, v. 10, no. 4, p. 351-356. 

Dilling, W.L., Tefertiller, N.B., and Kallos, G.J., 1975, Evaporation 

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trichloroethane, trichloroethylene, tetrachloroethylene, and other 

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Di Salvo, L.H., Guard, H.E., and Hunter, Leon, 1975, Tissue hydrocarbon burden of mussels as potential monitor of environmental 

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Di Toro, D.M., and Horzempa, L.M., 1982, Reversible and resistant 

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Di Toro, D.M., Horzempa, L.M., Casey, M.M., and Richardson, W., 

1982, Reversible and resistant components of PCB adsorptiondesorption; Adsorbent concentration effects: Journal of Great 

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Di Toro, D.M., Jeris, J.S., and Ciarcia, Daniel, 1985, Diffusion and 

partitioning of hexachlorobiphenyl in sediments: Environmental 

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Norris, L.A., 1981, The movement, persistence, and fate of the 

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Selected References 85 
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Parejko, Ronald, Johnston, Raymond, and Keller, Robert, 1975, 

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Pfaender, F.K., Shuman, M.S., Dempsey, Herbert, and Harden, C.W., 

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Pierce, R.H., Brent, C.R., Williams, H.P., and Reeves, S.G., 1977, 

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86 Manmade Organic Compounds in the Surface Waters of the United States 
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Powers, C.F., Sanville, W.D., Stay, F.S., and Schuytema, G.S., 1977, 

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Presley, B.J., and Trefry, J.H., 1980, Sediment-water interactions and 

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Ramamoorthy, S., 1985, Competition of fate processes in the bioconcentration of lindane: Bulletin of Environmental Contamination 

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Rao, P.S.C., Green, R.E., Balasubramanian, V., and Kanehiro, Y., 

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Reinert, K.H., and Rodgers, J.H. Jr., 1984, Influence of sediment types 

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Reinert, R.E., 1972, Journal of the Fish Research Board of Canada, v. 

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Reutergardh, L., 1980, Chlorinated hydrocarbons in estuaries, in 

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Ricci, E.D., Hubert, W.A., and Richard, J.J., 1983, Organochlorine 

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Rice, C.P., Eadie, B.J., and Erstfeld, K.M., 1982, Enrichment of PCBs 

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Richardson, E.M., and Epstein, E., 1971, Retention of three insecticides on different size soil particles suspended in water: Soil 

Science Society of America Proceedings, v. 35, p. 884-887. 

Rihan, T.I., Mustafa, H.T., Caldwell, George, Jr., and Frazier, Leroy, 

1978, Chlorinated pesticides and heavy metals in streams and 

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Riley, G.A., 1963, Organic aggregates in seawater and the dynamics of 

their formation and utilization: Limnology and Oceanography, v. 

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Roberts, A.E., Hill, D.R., and Tifft, E. C., Jr., 1982, Evaluation of New 

York Bight lobsters for PCBs, DDT, petroleum hydrocarbons, 

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Rodgers, P.W., and Swain, W.R., 1983, Analysis of PCB loading trends 

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Rosales, M.T.L., and Escalona, R.L., 1983, Organochlorine residues in 

organisms of two different lagoons of northwest Mexico: Bulletin 

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Rosales, M.T.L., Escalona, R.L., Alarcon, R.M., and Zamora, V., 1985, 

Organochlorine hydrocarbon residues in sediments of two different lagoons of Northwest Mexico: Bulletin of Environmental 

Contamination and Toxicology, v. 35, p. 322-330. 

Rossmann, R., and Siebel, E., 1977, Surficial sediment redistribution 

by wave energy; Element-grain size relationships: Journal of 

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Routh, J.D., 1972, DDT residues in Salinas River sediments: Bulletin 

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Rudling, Lars, 1970, Determination of pentachlorophenol in organic 

tissues and water: Water Research, v. 4, p. 533-537. 

Ryan, J.J., Lau, P.Y., Pilon, J.C., Lewis, Dave, McLeod, H.A., and 

Gervais, Adrian, 1984, Incidence and levels of 2,3,7,8-tetrachlorodibenzo-p-dioxin in Lake Ontario commercial fish: Environmental Science and Technology, v. 18, no. 9, p. 719-721. 

Ryan, J.J., Pilon, J.C., Conacher, H.B.S., and Firestone, David, 1983, 

Interlaboratory study on determination of 2,3,7,8-tetrachlorodibenzo-p-dioxin in fish: Journal of the Association of 

Official Analytical Chemists, v. 66, no. 3, p. 700-707. 

Selected References 87 
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Sabljic, Aleksandar, and Protic, Miroslava, 1982, Relationship bet\oveen 

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Sabourin, T.D., Stickle, W.B., Michot, T.C., Villars, C.E., Garton, 

D.W., and Mushinsky, H.R., 1984, Organochlorine residue levels 

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Saeger, V.W., Hicks, 0., Kaley, R.G., Michael, P.R., Mieure, J.P., and 

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Salim, R., and Cooksey, B.G., 1981, The effect of centrifugation on the 

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Sanborn, J.R., Metcalf, R.L., Yu, Ching-Chieh, and Lu, Po-Y\J"lg, 

1975, Plasticizers in the environment: The fate of di-n-octyl 

phthalate (DOP) in two model ecosystems and upatke and 

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Sanders, H.O., and Chandler, J.H., 1972, Biological magnification of a 

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Sanville, W.O., Lauer, W.L., Stay, F.S., Schuytema, G.S., and Powers, 

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92 Manmade Organic Compounds In the Surface Waters of the United States 

* U.S. GOVERNMENT PRINTING OFFICE:1988-201-985/74145