U.S. Department of the Interior 

U.S. Geological Survey

Transport, Behavior, and Fate of Volatile 

Organic Compounds in Streams

U.S. Geological Survey Professional Paper 1589

science for a chanuinu world
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Edited by: Mary A. Kidd

Manuscript preparation by: Joy K. Monson
Transport, Behavior, and Fate of Volatile 

Organic Compounds in Streams

ByR.E. RATHBUN

U.S. GEOLOGICAL SURVEY PROFESSIONAL PAPER 1589

UNITED STATES GOVERNMENT PRINTING OFFICE, WASHINGTON : 1998
U.S. DEPARTMENT OF THE INTERIOR 

BRUCE BABBITT, Secretary

U.S. GEOLOGICAL SURVEY 

Thomas J. Casadevall, Acting Director

The use of firm, trade, and brand names in this report is for 

identification purposes only and does not constitute 

endorsement by the U.S. Government.

Library of Congress Cataloging-in-Publications Data

Rathbun, R.E.

Transport, behavior, and fate of volatile organic compounds in 

streams / by R.E. Rathbun.

p. cm. - (Geological Survey professional paper; 1589) 

Includes bibliographical references. 

1. Water quality. 2. Volatile organic compounds. 

3. Environmental chemistry. I. Title. II. Series. 

GB1205.R37 1998 98-2685 

628.1 ' 68--dc21 CIP

ISBN 0-607-89393-1

For sale by U.S. Geological Survey, Information Services

Box 25286, Federal Center

Denver, CO 80225
CONTENTS

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

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

The National Water-Quality Assessment Program......................... 1

Purpose and Scope .......................................................................... 2

Target Volatile Organic Compounds .............................................. 2

Processes Affecting the Transport, Behavior, and Fate of 

Volatile Organic Compounds in Streams.......................................... 4

Volatilization................................................................................... 4

The Two-Film Model ............................................................. 6

Water-Film and Air-Film Coefficients and Resistances......... 8

Reference-Substance Concept ................................................11

Henry's Law Constant ............................................................12

Techniques for Measuring the Henry's

Law Constant........................................................... 15

Experimental Values of the Henry's

Law Constant........................................................... 16

Equations for Calculating the Henry's

Law Constant...........................................................16

Calculated Values of the Henry's

Law Constant...........................................................27

Temperature Dependence of the Henry's

Law Constant...........................................................28

Distribution of Henry's Law Constants for

the Target Analytes..................................................32

Prediction of Volatilization Coefficients................................35

Dependence of the Volatilization Half-Life on Water

Velocity, Flow Depth, and Water Temperature.................40

Absorption...............................,.......................................................43

Dry Deposition with Particles .........................................................44

Wet Deposition................................................................................44

Experimental Values of the Gas Scavenging Ratio................45

Calculated Values of the Gas Scavenging Ratio ....................45

Estimation of the Importance of Wet Deposition...................46

Microbial Degradation ....................................................................48

Classifications of Degradability .............................................48

Characteristics of the Degradation Process ............................49

Kinetics of the Degradation Process.......................................50

Comparison of Laboratory and Environmental

Degradation Studies...........................................................51

Microbial Degradation of the Target Analytes.......................53

Prediction of Microbial Degradation Coefficients .................56

Estimation of the Importance of Microbial Degradation........57

Sorption...........................................................................................59

Fundamentals of the Sorption Process....................................60

Sorption Coefficients..............................................................60

Factors Affecting the Sorption Process ..................................61

Solids Concentration Effect....................................................63

Kinetics of the Sorption Process.............................................64

III
IV CONTENTS

Solids, Soils, and Sediments...................................................65

Sorption to Other Sorbents .....................................................67

Experimental Sorption Coefficients .......................................68

Prediction Equations for Sorption Coefficients........................ 72

Estimation of the Importance of Sorption ................................74

Recent Developments............................................................... 77

Hydrolysis......................................................................................... 79

Hydrolysis Half-Lives for the Target Analytes........................ 81

Estimation of the Importance of Hydrolysis............................. 88

Aquatic Photolysis ............................................................................ 89

Kinetics of the Aquatic Photolysis Process.............................. 91

Photolysis Half-Lives for the Target Analytes......................... 94

Estimation of the Importance of Aquatic Photolysis................ 96

Oxidation and Chemical Reaction .................................................... 99

Oxidation Half-Lives for the Target Analytes........................ 100

Estimation of the Importance of Oxidation............................ 104

Chemical Reaction.................................................................. 105

Bioconcentration and Aquatic Toxicblogy..................................... 105

Kinetics of the Bioconcentration Process............................... 107

Octanol-Water Partition Coefficients for the

Target Analytes.................................................................. 110

Bioconcentration Factors for the Target Analytes.................. 112

Prediction Equations for the Bioconcentration Factor........... 116

Estimation of the Importance of Bioconcentration................. 116

Aquatic Toxicology ................................................................118

Advection and Dispersion............................................................... 127

Vertical Mixing.......................................................................127

Lateral Mixing........................................................................ 128

Longitudinal Mixing............................................................... 129

Estimation of the Importance of Longitudinal Dispersion..... 130

Information Needs................................................................................ 133

Volatilization................................................................................... 133

Absorption....................................................................................... 134

Wet Deposition ...............................................................................134

Microbial Degradation.................................................................... 135

Sorption........................................................................................... 135

Hydrolysis.......................................................................................135

Aquatic Photolysis ..........................................................................136

Oxidation.........................................................................................136

Bioconcentration............................................................................. 136

Summary and Conclusions ................................................................... 136

References............................................................................................. 140
CONTENTS V

FIGURES

1. Schematic diagram showing the processes affecting the transport, behavior, and fate of volatile organic

compounds in streams................................................................................................................................................ 5

2. Schematic diagram showing the two-film model...................................................................................................... 6

3—19. Graphs showing:

3. Percentage resistance to volatilization in the water film as a function of the Henry's law constant

at 25 degrees Celsius........................................................................................................................................... 10

4. Frequency distribution of the Henry's law constants at 25 degrees Celsius for the target analytes................... 33

5. Percentage resistance in the water film for the volatilization of three target analytes as a function

of the water-film and air-film mass-transfer coefficients at 25 degrees Celsius ................................................ 34

6. Oxygen-absorption (reaeration) coefficient (K2 ) in per day at 20 degrees Celsius as a function of 

water velocity and flow depth for the equations of Owens and others (1964), Churchill and others 

(1962), and O'Connor and Dobbins (1958)........................................................................................................ 37

7. Variation of the volatilization half-life with water velocity for three target analytes at 20 degrees

Celsius with flow depth as a parameter.............................................................................................................. 41

8. Variation of the volatilization half-life with flow depth for three target analytes at 20 degrees

Celsius with water velocity as a parameter......................................................................................................... 42

9. Variation of the volatilization half-life with water temperature for three target analytes .................................. 43

10. Concentrations of three target analytes in a stream resulting from gas scavenging from the

atmosphere as a function of flow depth and water temperature ......................................................................... 47

11. Fraction of the concentration of naphthalene remaining as a function of distance downstream for

volatilization and for volatilization and microbial degradation combined at 20 degrees Celsius...................... 59

12. Fraction of the organic compound sorbed to sediment as a function of the sediment concentration

for three target analytes....................................................................................................................................... 76

13. Fraction of naphthalene and benzene removed by volatilization and photolysis as a function of flow

depth at 20 degrees Celsius................................................................................................................................. 98

14. Ratio of the fraction of four target analytes removed by oxidation to the fraction removed by

oxidation and volatilization as a function of flow depth..................................................................................... 105

15. Frequency distribution of the logarithm base 10 of the octanol-water partition coefficient for the

target analytes ..................................................................................................................................................... 111

16. Fraction of hexachlorobutadiene removed by volatilization and bioconcentration as a function

of flow depth....................................................................................................................................................... 118

17. Concentration of a conservative dispersant and the nonconservative dispersants dichloromethane 

and diisopropyl ether as a function of time for a deep, low-velocity stream characteristic of the 

O'Connor-Dobbins(1958) equation................................................................................................................... 131

18. Concentration of a conservative dispersant and the nonconservative dispersants dichloromethane 

and diisopropyl ether as a function of time for an intermediate-depth, high-velocity stream 

characteristic of the Churchill and others (1962) equation................................................................................. 132

19. Concentration of a conservative dispersant and the nonconservative dispersants dichloromethane 

and diisopropyl ether as a function of time for a shallow, medium-velocity stream characteristic 

of the Owens and others (1964) equation........................................................................................................... 133
VI CONTENTS

TABLES

1. Volatile organic compound target analytes............................................................................................................... 2

2. Experimental and interpolated values of Henry's law constants at 20° Celsius and 25° Celsius for

43 target analytes....................................................................................................................................................... 17

3. Henry's law constants excluded from the results presented in table 2 because they differed substantially

from the average constants........................................................................................................................................ 26

4. Constants in the equation for the temperature dependence of the Henry's law constants of 39 target analytes ...... 29

5. Evaluation of equations 51 and 52 for predicting the temperature dependence of the Henry's law constant.......... 33

6. Water-film and air-film reference-substance parameters for the target analytes...................................................... 38

7. Prediction of the volatilization coefficients and distances downstream required for 90-percent loss

by volatilization of three target analytes................................................................................................................... 40

8. Experimental values of gas scavenging ratios for 11 target analytes........................................................................ 46

9. Estimated rate coefficients for anaerobic microbial degradation of six target analytes with estimated 

half-lives for aerobic microbial degradation between 6 months (K'^0.0038 day ) and 1 year 

(K'd=0.0019 day" 1 )................................................................................................................................................... 55

10. Estimated and experimental rate coefficients for anaerobic microbial degradation of 15 target analytes 

with estimated half-lives for aerobic microbial degradation between 4 weeks (K'd=0.025 day" 1 ) and 

6 months (K'd=0.0038 day" 1 ).................................................................................................................................... 56

11. Estimated rate coefficients for anaerobic microbial degradation of eight target analytes with estimated 

half-lives for aerobic microbial degradation between 7 days (K'^0.099 day" ) and 4 weeks 

(K'd=0.025 day' 1 ).....................................................................................................^^ 56

12. Estimated and experimental rate coefficients for aerobic and anaerobic microbial degradation of

13 target analytes....................................................................................................................................................... 57

13. Classification of sediments according to grain size.................................................................................................. 66

14. Threshold values of the ratio of the mineral fraction to the organic-carbon fraction at which mineral

adsorption becomes important for three types of organic compounds ..................................................................... 67

15. Experimental sorption coefficients for 27 target analytes......................................................................................... 69

16. Equations for predicting sorption coefficients of volatile organic compounds........................................................ 73

17. Predicted sorption coefficients for 28 target analytes............................................................................................... 75

18. Literature values of half-lives for hydrolysis of 30 target analytes and estimates of the potential for

hydrolysis of 19 target analytes................................................................................................................................. 82

19. Fractions removed by volatilization and hydrolysis and total fraction removed for nine target analytes

with hydrolysis half-lives of 548 days or less........................................................................................................... 90

20. Photolysis half-lives fornine target analytes............................................................................................................ 94

21. Thirty-three target analytes for which aquatic photolysis was not considered significant....................................... 95

22. Eighteen target analytes for which no information was available on aquatic photolysis ......................................... 96

23. Experimental values of the half-lives for oxidation of 12 target analytes in water by the hydroxy radical............. 101

24. Calculated values of the half-lives for oxidation of 39 target analytes in water....................................................... 102

25. Experimental values of the octanol-water partition coefficient for 40 target analytes............................................. 110

26. Calculated values of the octanol-water partition coefficient for 15 target analytes.................................................. 111

27. Experimental values of the bioconcentration factor based on the whole body weight for various

organisms for 26 target analytes ............................................................................................................................... 112

28. Experimental values of the bioconcentration factor based on the lipid weight for various organisms

for seven target analytes............................................................................................................................................ 115

29. Comparison of whole-weight and lipid-weight bioconcentration factors................................................................. 116

30. Minimum concentrations of 40 target analytes resulting in 50-percent mortality of various aquatic organisms..... 120

31. Minimum concentrations of 31 target analytes that affected at the 50-percent level various functions of

aquatic organisms...................................................................................................................................................... 124
CONTENTS VII

CONVERSION FACTORS

Multiply By To obtain

nanometer (nm)

millimeter (mm)

centimeter (cm)

meter (m)

kilometer (km)

meter per meter (m/m)

meter per second (m/s)

meter per second per second (m/s/s)

meter per day (m/day)

day per meter (day/m)

square meter (m2 )

square meter per second (m2/s)

square meter per day (irr/day)

cubic meter (m3 )

cubic meter per second (m3 /s)

liter (L)

gram (g)

gram per square meter (g/m2 )

gram mole per square meter per second (g mol/nr/s)

gram mole per square meter per day (g mol/m2/day)

photon per square centimeter per second (photon/cnr/s)

nanogram per liter (ng/L)

microgram per liter (ng/L)

milligram per liter (mg/L)

kilogram per liter (kg/L)

gram mole per cubic meter (g mol/m3 )

gram per liter (g/L)

gram millimole (g mmol)

gram micromole (g jamol)

gram per gram mole (g/g mol)

nanogram per kilogram (ng/kg)

gram per kilogram (g/kg)

milliliter per gram (mL/g)

liter per kilogram (L/kg)

cubic centimeter per gram mole (cm3 /g mol)

liter per gram mole per centimeter (L/g mol/cm)

pascal (Pa)

pascal cubic meter per gram mole (Pa m3 /g mol)

pascal cubic meter per gram mole per kelvin (Pa m3/g mol/K.)

kilojoule per gram mole (kJ/g mol)

kilojoule per gram mole per kelvin (kJ/g mol/K)

3.937x10~8

3.937xlO~2

3.937x10-'

3.281

6.214x10"'

1.00

3.281

3.281

3.281

3.048x10-'

10.76

10.76

10.76

35.31

35.31

3.531xlQ-2 

2.205x10"3 

2.048x1Q-4 

2.048xlO"4 

2.048x1O"4 

9.29xl02 

6.243x10-' 

6.243x10~8 

6.243x10"5 

6.243x10' 

6.243x10~5 

6.243x10~2 

2.205x10~6 

2.205x10~q 

1.00

l.OOxlO-' 2 

l.OOxlQ-3 

1.60xlO"2 

1.60xlQ-2 

1.60xlQ-2 

4.88xl02 

l.OlxlO5 

1.58x10-' 

8.78xlO~2 

l.OSxlO2 

6.02x10'

inch

inch

inch

foot

mile

foot per foot

foot per second

foot per second per second

foot per day

day per foot

square foot

square foot per second

square foot per day

cubic foot

cubic foot per second

cubic foot

pound

pound per square foot

pound mole per square foot per second

pound mole per square foot per day

photon per square foot per second

pound per cubic foot

pound per cubic foot

pound per cubic foot

pound per cubic foot

pound mole per cubic foot

pound per cubic foot

pound mole

pound mole

pound per pound mole

pound per pound

pound per pound

cubic foot per pound

cubic foot per pound

cubic foot per pound mole

cubic foot per pound mole per foot

standard atmosphere

standard atmosphere cubic foot per pound mole

standard atmosphere cubic foot per pound mole per degree rankine

kilocalorie per pound mole

kilocalorie per pound mole per degree rankine

Temperature in kelvins (K) may be converted to degrees Celsius (°C) using 

°C= K-273.15

Temperature in kelvins (K) may converted to degrees rankine (°R) using 

°R = (K-273.15)(1.8) + 491.7

Temperature in kelvins (K) may be converted to degrees Fahrenheit (°F) using 

°F = (K-273.15) (1.8)+ 32.0

TRANSPORT, BEHAVIOR, AND FATE OF 

VOLATILE ORGANIC COMPOUNDS IN STREAMS

By R.E. Rathbun

ABSTRACT

Volatile organic compounds (VOCs) 

are compounds with chemical and physical 

properties that allow the compounds to move 

freely between the water and air phases of the 

environment. VOCs are widespread in the 

environment because of this mobility. Many 

VOCs have properties making them suspected 

or known hazards to the health of humans and 

aquatic organisms. Consequently, understanding 

the processes affecting the concentration and 

distribution of VOCs in the environment is 

necessary.

The transport, behavior, and fate of 

VOCs in streams are determined by combinations of chemical, physical, and biological 

processes. These processes are volatilization, 

absorption, wet and dry deposition, microbial 

degradation, sorption, hydrolysis, aquatic photolysis, oxidation, chemical reaction, bioconcentration, advection, and dispersion. The relative 

importance of each of these processes depends 

on the characteristics of the VOC and the stream.

The U.S. Geological Survey National 

Water-Quality Assessment Program selected 

55 VOCs for study. This report reviews the 

characteristics of the various processes that 

could affect the transport, behavior, and fate 

of these VOCs in streams.

INTRODUCTION

Many organic compounds have chemical and 

physical properties that allow the compounds to move 

freely between the water and air phases of the environment. These compounds commonly are called volatile

organic compounds (VOCs). VOCs in general have 

low molecular weights, high vapor pressures, and 

low-to-medium water solubilities. Many VOCs 

are toxic and are the focus of Federal regulations 

related to water quality (Leahy and Thompson, 

1994). Consequently, VOCs were designated for a 

national synthesis study under the National WaterQuality Assessment (NAWQA) Program of the 

U.S. Geological Survey (USGS).

THE NATIONAL WATER-QUALITY 

ASSESSMENT PROGRAM

The USGS began the NAWQA Program in 

1991. The objectives of this program are to describe 

current water-quality conditions for a large part of the 

water resources of the United States, to define longterm trends or the absence of trends in water quality, 

and to identify and describe the primary factors 

affecting the observed water-quality conditions 

and trends (Hirsch and others, 1988). The NAWQA 

Program consists of two parts. These are investigations of the water quality of river basins and 

aquifer systems and national synthesis assessment" 

(Leahy and Thompson, 1994). The synthesis assessments focus on high-priority water-quality issues. 

Initial national synthesis assessments were on 

the occurrences of nutrients and pesticides in 

surface water and ground water. In 1994, a national 

synthesis project on VOCs was initiated because 

of the widespread occurrence of these compounds 

in the waters of the environment and the lack of 

information on factors related to their occurrence 

and behavior. The objectives of the VOC project 

are to determine the occurrence of VOCs in surface 

water and ground water, to determine trends or 

absence of trends in the concentrations, to identify 

probable sources of VOCs, and to describe the primary 

processes affecting the concentrations of VOCs in 

waters of the United States.
TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

PURPOSE AND SCOPE

The purpose of this report is to present a review 

of the characteristics of the various processes that 

affect the transport, behavior, and fate of VOCs in 

streams. Among these processes, volatilization is the 

one likely to be most important for VOCs. Other 

processes such as absorption, dry deposition with 

particles, wet deposition, microbial degradation, 

sorption, hydrolysis, aquatic photolysis, oxidation, 

chemical reaction, and bioconcentration may be 

important for some compounds under certain environmental conditions. Also, all VOCs in streams are 

subject to advective and dispersive mixing processes. 

Fundamentals of these processes are presented, and 

where appropriate, estimates of first-order rate coefficients for streams are given. This report emphasizes 

specific VOCs selected for study by the NAWQA 

Program. These target VOCs cover a wide range of 

chemical and physical properties. Consequently, much 

of this discussion also will be directly applicable to 

other VOCs.

TARGET VOLATILE 

ORGANIC COMPOUNDS

The NAWQA Program selected 55 target 

analytes for study. Details of the procedure used 

in the selection of these analytes have been described 

previously (J.S. Zogorski, U.S. Geological Survey, 

written commun., 1994). Briefly, the procedure began

with 125 compounds regulated under the Tafe 

Drinking Water Act and the Clean Water Act that 

could be considered to be volatile on the basis of 

their chemical and physical properties. This list was 

reduced to only VOCs that occurred frequently in 

natural waters, or were either known or pc ssible 

human carcinogens, or were of non-cancer humanhealth concerns, or were likely to have detrimental 

effects on the health of aquatic ecosystems. A 

few compounds were included because of special 

uses and/or environmental effects, such as those 

compounds used as additives in reformulated automotive gasolines, those suspected of deteriorating the 

ozone layer, and those which can undergo significant 

bioconcentration.

The 55 VOCs selected as target analytes are 

listed in table 1. The International Union of Pure and 

Applied Chemistry (IUPAC) names of the compounds 

will be used in this report. Also given in table 1 are 

alternative names or abbreviations sometimes used in 

the literature for 34 of the analytes. The Chemical 

Abstracts Services (CAS) numbers and the USGS 

parameter codes are presented for identification 

purposes. The two dimethylbenzene compounds 

(1,3-dimethylbenzene and 1,4-dimethylbe-izene) have 

the same USGS parameter code because these isomers 

cannot be separated by the analytical technique used 

by the USGS National Water-Quality Laboratory. 

Consequently, concentrations under this parameter 

code are for the sum of the concentrations of these two 

compounds.

Table 1 . Volatile organic compound target analytes

[CAS, Chemical Abstracts Services; USGS, U.S. Geological Survey; IUPAC, International Union of Pure and Applied Chemistry]

CAS 

number

74-87-3 

75-09-2 

67-66-3

56-23-5

74-83-9 

75-25-2

75-27-4

1 24-48-1

75-00-3 

75-34-3 

107-06-2

USGS 

parameter 

code

34418 

34423 

32106

32102

34413 

32104

32101

32105

34311 

34496 

32103

IUPAC 

name

Halogenated alkanes

chloromethane 

dichloromethane 

trichloromethane

tetrachloromethane

bromomethane 

tribromomethane

bromodichloromethane

chlorodibromomethane

chloroethane 

1 , 1 -dichloroethane 

1 ,2-dichloroethane

Alternative 

name

methyl chloride 

methylene chloride 

chloroform

carbon tetrachloride

methyl bronide 

bromoform

ethyl chlorHe 

ethylidenedichloride 

ethylenedichloride
INTRODUCTION

Table 1. Volatile organic compound target analytes—Continued

[CAS, Chemical Abstracts Services; USGS, U.S. Geological Survey; IUPAC, International Union of Pure and Applied Chemistry]

CAS 

number

71-55-6

79-00-5

67-72-1

106-93-4

78-87-5

96-18-4

96-12-8

75-69-4

75-71-8

76-13-1

75-01-4

75-35-4

1 56-59-2

156-60-5

79-01-6

127-18-4

593-60-2

10061-01-5

10061-02-6

87-68-3

71^13-2

100-^2-5

91-20-3

108-88-3

100-41^

103-65-1

98-82-8

104-51-8

95-47-6

108-38-3

1 06-42-3

95-63-6

108-90-7

95-50-1

541-73-1

106-46-7

87-61-6

120-82-1

1634-04-4

637-92-3

994-05-8

108-20-3

107-02-8

107-13-1

USGS 

parameter 

code

34506

34511

34396

77651

34541

77443

82625

34488

34668

77652

39175

34501

77093

34546

39180

34475

50002

34704

34699

39702

34030

77128

34696

34010

34371

77224

77223

77342

77135

85795

85795

77222

34301

34536

34566

34571

77613

34551

78032

50004

50005

81577

34210

34215

IUPAC 

name

Halogenated alkanes — Continued

1,1,1 -trichloroethane

1 , 1 ,2 -trichloroethane

hexachloroethane

1 ,2-dibromoethane

1 ,2-dichloropropane

1 ,2,3-trichloropropane

1 ,2-dibromo-3-chloropropane

trichlorofluoromethane

dichlorodifluoromethane

l,l,2-trichloro-l,2,2-trifluoroethane

Halogenated alkenes

chloroethene

1,1-dichloroethene

cis- 1 ,2-dichloroethene

trans-l ,2-dichloroethene

trichloroethene

tetrachloroethene

bromoethene

cis- ] ,3-dichloropropene

trans-l ,3-dichloropropene

hexachlorobutadiene

Aromatic hydrocarbons

benzene

styrene

naphthalene

Alkyl benzenes

methylbenzene

ethylbenzene

n-propylbenzene

iso-propylbenzene

n-butylbenzene

1 ,2-dimethylbenzene

1 ,3-dimethylbenzene

1 ,4-dimethylbenzene

1 ,2,4-trimethylbenzene

Halogenated anomalies

chlorobenzene

1 ,2-dichlorobenzene

1 ,3-dichlorobenzene

1 ,4-dichlorobenzene

1 ,2,3-trichlorobenzene

1,2,4-trichlorobenzene

Ethers

methyl tertiary-butyl ether

ethyr tertiary-butyl ether

tertiary-amyl methyl ether

diisopropyl ether

Others (aldehydes and nitriles)

2-propenal

2-propenenitrile

Alternative 

name

methyl chloroform

EDB

DBCP

Freon 11,CFC11

Freon 12, CFC12

Freon 113, CFC113

vinyl chloride

TCE

perchloroethylene, PCE

vinyl bromide

vinyl benzene

toluene

cumene

o-xylene

m-xylene

p-xylene

o-dichlorobenzene

m-dichlorobenzene

p-dichlorobenzene

MTBE

ETBE

TAME

DIPE

acrolein

acrylonitrile
TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

PROCESSES AFFECTING THE

TRANSPORT, BEHAVIOR, AND

FATE OF VOLATILE ORGANIC

COMPOUNDS IN STREAMS

The transport, behavior, and fate of VOCs 

in streams are determined by the effects of combinations of various chemical, physical, and biological 

processes. These processes are depicted schematically 

in figure 1. Among these processes, volatilization is 

the one most important for VOCs. However, in some 

instances, the other processes shown in figure 1 may 

affect the fate of VOCs in streams.

Some of these processes such as volatilization, 

absorption, wet and dry deposition, bioconcentration, 

and sorption involve movement of VOCs between 

different parts of the environment. Other processes 

such as microbial degradation, hydrolysis, aquatic 

photolysis, oxidation, and chemical reaction 

result in changes of the chemical compounds to 

other compounds. Complete mineralization by 

microbial degradation results in the conversion 

of the compounds to water and carbon dioxide or 

methane, depending on whether the degradation 

is under aerobic or anaerobic conditions. Transport 

within the same water parcel or to other parts of the 

stream system occurs by advective and dispersive 

transport of the dissolved VOCs with the water 

and movement of sorbed VOCs with the suspended 

particulates.

The relative importance of the various processes 

shown in figure 1 depends on the characteristics of the 

VOC and the stream under consideration. In the case 

of microbial degradation, previous occurrences of the 

VOC in the stream reach also may affect the ability of 

the bacterial community in the stream to degrade the 

compound. In subsequent sections of this report, each 

of these processes is discussed, and experimental 

values of first-order rate coefficients for the processes 

are presented when appropriate.

The term "'stream" will be used in this report 

to describe all natural open-channel flows, regardless 

of size. In some examples, the terms "deep, lowvelocity," "intermediate-depth, high-velocity," and 

"shallow, medium-velocity" will be used to describe 

qualitatively streams with these combinations of flow 

depth and water velocity. Tidally affected streams are 

not considered in this report.

VOLATILIZATION

Volatilization is the movement of a VOC 

from the bulk water phase of a stream across the 

water/air interface into the air. Volatilization of 

organic solutes from water is a first-order process. 

A first-order process is one in which the rate of 

reaction is directly proportional to the concentration 

of the species undergoing the process (Tirsley, 

1979). Thus, the rate of volatilization, dC/dt, is 

given by

dC "dt = Kv (C-Ce ) (1)

where

C is the concentration (g mol/m3 ) of the VOC

in the water, 

t is time (days),

Kv is the volatilization coefficient (day~ ), and 

Ce is what the concentration (g mol/m ) of the 

VOC in the water would be if the water 

were in equilibrium with the partial 

pressure of the VOC in the air. 

In many environmental situations, t>e VOC 

concentrations in the air above the stream surface 

are negligible with respect to the concentrations 

in the water because of rapid dispersion in the 

atmosphere. Consequently, Ce in equation 1 can 

be neglected, and the equation can be integrated 

to give

C = C0 exp(-Kv t) (2)

where

C0 is the concentration (g mol/m3 ) of the VOC

in the water at time zero.

Times are commonly expressed as half-lives, where 

the half-life is the time required for the concentration 

to be reduced by volatilization to one-half of the initial 

concentration. It follows from equation 2 that the halflife, t0 5v (days), is given by

0.693

I0.5v (3)
WET AND DRY 

DEPOSITION

VOLATILIZATION, 

ABSORPTION SUNLIGHT

ADVECTION

OXIDATION AND 

CHEMICAL REACTION

AQUATIC PHOTOLYSIS

HYDROLYSIS

MICROBIAL DEGRADATION BIOCONCENTRATION

DEPOSITION 

(AGGRADATION)

SORPTION (SUSPENDED 

SEDIMENT)

RESUS PENSION 

(DEGRADATION)

BOTTOM SEDIMENT

1. Schematic diagram showing the processes affecting the transport, behavior, and fate of volatile organic compounds in streams.
TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

Times can be converted to distances along a 

stream using the relationship

L =

tU

1.16x10

-5 (4)

where

L is the distance (m) along the stream, 

U is the mean water velocity (m/s), and

1.16x 10 is the days-to-seconds conversion factor.

Concentration as a function of distance is obtained by

combining equations 2 and 4 to give

C = C0 exp|(-Kv )(1.16xlO~5 )tJ . (5)

Equation 5 is somewhat simplified in that it neglects 

the effect of dispersion.

THE TWO-FILM MODEL

Equation 5 requires a value of the volatilization 

coefficient to permit prediction of the varation of the 

concentration of a VOC with distance along a stream. 

Various models are available for this purpose, one 

of them being the two-film model (Lewis and 

Whitman, 1924). This model assumes uniformly 

mixed water and air phases separated by thin films 

of water and air in which mass transfer is only by 

molecular diffusion. Figure 2 is a schematic representation of the two-film model depicting transfer from 

the water to the air phase. This figure indicates a 

constant concentration in the bulk water phase, a 

concentration gradient across the water film, a discontinuity at the interface, a concentration gradient across 

the air film, and a constant concentration in the bulk 

air phase. The discontinuity at the interface is because 

concentrations in water are usually expressed on a

w 

o

I

Q

BULK AIR PHASE WITH 

UNIFORM CONCENTRATION

AIR FILM

WATER FILM

BULK WATER PHASE WITH 

UNIFORM CONCENTRATION

CONCENTRATION 

Figure 2. Schematic diagram showing the two-film model.
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

mass-per-unit-volume basis and concentrations in air 

are usually expressed on a partial pressure basis. The 

interface between the water and air films is assumed to 

have no resistance to mass transfer. A dynamic steadystate is assumed with equilibrium at the interface. 

The two-film model is a conceptual model in that the 

existence of a coherent stagnant film over the entire 

surface of a flowing stream seems unlikely.

The rate of mass transfer over the flow depth of 

a stream is generally large compared with the rate of 

mass transfer into or out of the stream. This difference 

in rates results in almost no gradient in concentration 

of dissolved solutes over most of the flow depth 

and a large gradient in concentration near the surface. 

Consequently, most of the resistance to mass transfer 

is in the films at the water/air interface as conceptualized in the two-film model (fig. 2). A discussion of 

resistances in the films and several examples of 

measurements of dissolved oxygen in streams, 

which provide experimental evidence in support of 

this phenomenon, have been presented previously 

(Bennett and Rathbun, 1972).

The basic equation of the two-film model 

can be developed as follows. Flux equations for the 

transfer of a solute through the water and air films 

(fig- 2) are

TSL =

TSL =

ka(Pi~P)

RT

(6)

(7)

where

Nw and Na are the mass fluxes through the water and

air films (g mol/m2/day); 

Cj is the concentration of the solute at the

interface on the water side (g mol/m3 ); 

kw is the mass-transfer coefficient for the

water film (m/day); 

ka is the mass-transfer coefficient for the air

film (m/day); 

Pi is the partial pressure of the solute at the

interface on the air side, in pascals; 

p is the partial pressure of the solute in 

the uniformly mixed air phase, in 

pascals;

R is the ideal gas constant (Pa m /g mol/K); and

T is the temperature, in kelvins. 

With the assumption of steady-state conditions, 

the mass fluxes through the films must be equal. 

Consequently,

N =

k.(Pi-P) 

RT = kw (C-C) (8)

where

N is the mass flux (g mol/irr/day). 

Equation 8 contains the concentration and partial 

pressure of the solute at the interface. These parameters, however, are difficult, if not impossible, to 

measure.

To eliminate these parameters, the flux 

equations are written in terms of overall mass-transfer 

coefficients and overall concentration-difference 

driving forces. The result is

N =

Kao(Pe-P) 

RT

= Kwo (C-Ce ) (9)

where

Kao is the overall mass-transfer coefficient b?sed

on the air phase (m/day), 

Kwo is the overall mass-transfer coefficient b?sed

on the water phase (m/day), and 

pe is the partial pressure, in pascals, of the 

solute in the air in equilibrium with the 

water concentration C.

Solute concentrations in environmental situations are 

almost always dilute enough such that ideal solution 

laws apply. Therefore, Henry's law can be used to 

describe the equilibrium conditions. This law is 

defined by

= HC (10)

(11)

where 

H is the Henry's law constant of the solute 

(Pa m3 /g mol).
TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

Because equilibrium conditions are assumed at the 

interface, it follows that

Pi = H C: . (12)

By algebraic manipulation of equations 6, 7, 

8, 10, 11, and 12 and comparison of the results with 

equation 9, it follows that

1

K,

RT 

Hk. (13)

Equation 13 is the basic equation of the two-film 

model, and it indicates that the overall mass-transfer 

coefficient for volatilization of a solute from a stream 

depends on the mass-transfer coefficients for the water 

and air films and the Henry's law constant of the solute.

The relationship between the volatilization 

coefficient, Kv, of equation 1 and the overall masstransfer coefficient based on the water phase, Kwo, 

can be developed as follows. The rate of change of the 

concentration with time, dC/dt, and the mass flux, N, 

are related by

VdC

A dt

= N (14)

where

V is the volume of the water phase (m ), and 

A is the water-surface area (m2).

If the water surface is reasonably flat,

—- = Y

where

Y is the water depth (m). 

Comparing equations 1 and 9 and considering 

equations 14 and 15 gives

KWO = KV Y. (16)

WATER-FILM AND AIR-FILM 

COEFFICIENTS AND RESISTANCES

In the general situation, the film coefficients, 

kw and ka are difficult, if not impossible, to measure 

because they involve the concentration and partial

pressure of the solute at the interface. In two limiting 

cases, however, kw and ka can be measured directly. 

The first limiting case involves solutes with large 

values of the Henry's law constant such that the 

second term on the right-hand side of equation 13 is 

negligible with respect to the first. For ther^e solutes,

(17) K,

Because the overall mass-transfer coefficient, Kwo, 

can be measured directly, the water-film coefficient, 

kw, also can be measured for these solutes. The 

condition under which equation 17 applies is called 

water-film-controlled mass transfer, which means 

that virtually all the resistance to mass trarsfer is 

in the water film. Thus, volatilization in this case is 

controlled by mixing conditions in the water. Many of 

the VOCs of environmental significance have Henry's 

law constants such that most of the resistance to 

volatilization is in the water film.

The second limiting case involves solutes with 

small values of the Henry's law constant such that 

the first term on the right-hand side of equation 13 is 

negligible with respect to the second term. For these 

solutes,

RT

K, (18)

Because the overall mass-transfer coefficient, Kwo, 

can be measured directly, the air-film coefficient, ka, 

also can be measured for these solutes. The condition 

under which equation 18 applies is called air-filmcontrolled mass transfer, which means that virtually 

all the resistance to mass transfer is in the air film. 

Thus, volatilization in this case is controlled by 

mixing processes in the air. Solutes with Fenry's law 

constants small enough that equation 18 applies will 

have low potential to volatilize from water and, 

therefore, most likely would not be classifed as 

VOCs.

Some solutes, however, will have Henry's law 

constants intermediate between the values of these 

limiting cases. Consequently, resistances cf both the 

water and air films will be important. The resistances 

of the films are controlled primarily by the thicknesses
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

of the films that, in turn, are controlled primarily 

by mixing conditions within the bulk water and air 

phases. Mixing within the water phase of streams is 

primarily the result of shear stresses at the bottom and 

banks of the stream. Mixing within the air phase above 

streams is primarily the result of wind shear. In some 

cases, shear stresses within one phase may affect 

mixing in the other phase. In a wide, shallow stream 

with low banks and a low water velocity, wind shear 

may significantly affect mixing in the water phase. 

An example of such an effect was unexpectedly high 

oxygen-absorption coefficients attributed to wind 

effects on a wide, shallow reach of the Rock River 

(Grant, 1978). Similarly, mixing within the water 

phase may have some effect on the air-film resistance. 

An example of such an effect is the observation that 

water-evaporation rates, which are in theory controlled 

completely by the air-film resistance, were larger 

for a canal than for a lake with comparable 

windspeeds (Jobson, 1980).

The relative importance of the resistances of 

the water and air films to volatilization of solutes 

from streams can be estimated from equation 13. 

The reciprocal of a mass-transfer coefficient is the 

resistance to that mass transfer. Therefore, equation 13 

can be written as

r = (19)

where

r is the resistance to mass transfer (day/m), and the t, 

w, and a subscripts indicate the total, water-film, and 

air-film resistances, respectively. It follows that the 

percentage resistance in the water film is given by

100 r. 100

1 +

RTk (20)

Thus, the percentage of the total resistance to mass 

transfer that is in the water film depends on the 

relative magnitudes of the mass-transfer coefficients 

for the water and air films and the Henry's law 

constant of the VOC.

The percentage resistance in the water 

film as a function of the Henry's law constant is 

presented in figure 3 for a temperature of 25°C.

Three combinations of water-film and air-film 

mass-transfer coefficients were used to cover the 

range of coefficients that might be expected for 

streams. Minimum and maximum water-film and airfilm coefficients were paired to obtain the maximum 

possible range of percentage resistances. The third 

combination was for values of the two film coefficients assumed appropriate for average conditions 

in a stream.

Values of 0.3, 3, and 6 m/day were used 

for the water-film coefficients. These values cove*- 

the range of water-film coefficients estimated for 

benzene, trichloromethane, dichloromethane, and 

methylbenzene for 12 streams of the United States 

(Rathbun and Tai, 1981). Values of 400, 800, and 

1,200 m/day were used for the air-film coefficients. 

These values cover the range of air-film coefficients 

determined for evaporation of water from a canal 

in southern California (Jobson and Sturrock, 1979; 

Jobson, 1980), the open ocean (Liss and Slater, 1974), 

and three lakes in the United States (Smith and otl ers, 

1981).

Figure 3 indicates that as the Henry's law 

constant increases, the percentage resistance in the 

water film increases, and the increase is fastest for 

the smallest water-film coefficient. This observation 

is consistent with a thicker water film for the lowe^ 

water-mixing intensity, as indicated by the low waterfilm coefficient. Conversely, for the highest watermixing intensity, as indicated by the highest waterfilm coefficient, the percentage resistance in the 

water film approaches 100 percent more slowly.

Stated another way, for a specific Henry's 

law constant, the percentage resistance in the 

water film is largest for the combination of the small 

water-film coefficient and the large air-film coefficient. Conversely, the percentage resistance in the 

water film is smallest for the combination of the 

high water-film coefficient and the small air-film 

coefficient. Percentage resistances for the average 

film coefficients are intermediate between these 

extremes.

Figure 3 indicates that for solutes with Henry's 

law constants larger than 300 Pa m3/g mol, at least 

90 percent of the resistance to volatilization will 

be in the water film for all three combinations of film 

coefficients. Conversely, for solutes with Henry's law 

constants smaller than about 0.07 Pa m3/g mol, less 

than 10 percent of the resistance to volatilization will
10 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

kw = 0.3 METERS/DAY; ka = 1,200 METERS/DAY

_—— kw = 6.0 METERS/DAY: ka = 400 METERS/DAY

3.01 0.1 1 10 100 1,000 10,000 50,000

HENRY'S LAW CONSTANT, IN PASCALS CUBIC METER PER GRAM MOLE 

Figure 3. Percentage resistance to volatilization in the water film as a function of the Henry's law constant at 25 degrees Celsius.

be in the water film for the three combinations of film 

coefficients. For solutes with intermediate values of 

the Henry's law constant, both resistances will be 

significant for volatilization from water.

A useful lower limit is the Henry's law constant 

of water at 25°C, which is 0.057 Pa m3/g mol. Solutes 

with constants less than this value are considered 

nonvolatile because the tendency to partition into the 

air phase is less than that of water. Consequently, the 

concentration of the solute would actually increase 

with time because the water would evaporate faster 

than the solute would volatilize.

The previous discussion used air-film coefficients that were about two orders of magnitude larger 

than the water-film coefficients. The air-film coefficients are larger that the water-film coefficients 

because molecular diffusion in gases is faster than 

molecular diffusion in liquids. In the case of no mixing 

in the bulk water and air phases, mass transfer is solely 

by molecular diffusion and the flux equations (eqs. 6 

and 7) become

D,

(21)

Pi-P

RT (22)

where

D is the molecular-diffusion coefficient of the solute 

being transferred (m2/day), 5 is the thickness of the 

film (m), and the w and a subscripts refer to the water 

and air phases. Comparing equations 6 ard 21 and 

equations 7 and 22, it follows that

(23)

and

(24)

Equations 23 and 24 indicate that the film 

coefficients are proportional to the moleculardiffusion coefficients to the 1.0 power ftr the 

no-mixing case. Other theoretical values for this
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

11

exponent range from 0.5 for the penetration model 

(Danckwerts, 1951) to a value for the film-penetration 

model that depends on the mixing intensity (Dobbins, 

1956). The film-penetration model exponent varies 

from 0.5 for a high-mixing intensity to 1.0 for 

the low-mixing intensity characteristic of the 

molecular-diffusion limit of the film model. Experimental values of the exponent generally range 

between 0.50 and 0.67, and these are the values 

generally used to express dependence of the masstransfer coefficient on the molecular-diffusion 

coefficient.

Molecular-diffusion coefficients of benzene in 

water and air can be used to illustrate the differences 

between water-film and air-film coefficients. In air, 

benzene has a molecular-diffusion coefficient at 25°C 

of 0.805 m2/day (Lugg, 1968), and in water, the value 

is 9.50x10"5 m2/day (Tominaga and others, 1984). 

Taking ratios of these values and raising the ratio 

to the 0.50 power and the 0.67 power results in 

values of 92 and 430, respectively. These numbers 

indicate that the air-film coefficients for benzene are 

likely to exceed the water-film coefficients by these 

factors, depending on the value assumed for the 

power dependence of the film coefficient on the 

diffusion coefficient. The large range is indicative 

of the sensitivity to the value of the exponent 

defining the dependence of the film coefficient on the 

molecular-diffusion coefficient. Finally, these values 

are qualitatively consistent with the ratios of the film 

coefficients used as representative values in figure 3.

REFERENCE-SUBSTANCE CONCEPT

The two-film model (eq. 13) describes the 

relative importance of the water-film and air-film 

resistances to volatilization of organic solutes from 

water. The model does not, however, give any 

information on the magnitudes of the water-film and 

air-film mass-transfer coefficients. These coefficients 

can be estimated using the reference-substance 

concept.

The reference-substance concept assumes 

that the film mass-transfer coefficients for an organic 

solute are directly proportional to the film masstransfer coefficients for a reference substance and 

that the proportionality constants are independent 

of mixing conditions within the water and air phases. 

These relationships are

worg

^aorg

= <*> kwref

Laref

(25)

(26)

where O and *F are the water and air referencesubstance constants, the w and a subscripts 

indicate the water and air phases, and the org and 

ref subscripts indicate the organic solute and the 

reference substance, respectively. Values of the 

reference-substance constants are measured under 

controlled conditions in the laboratory. If the 

reference substances then are selected so that there is 

information on the values of the film coefficients for 

these substances for streams, equations 25 and 26 can 

be used to estimate film coefficients for organic 

solutes for the same streams.

Oxygen commonly is used as the reference 

substance for the water film for two reasons. First, the 

Henry's law constant for oxygen at 20.0°C, computed 

from the water solubility data of Benson and Krause 

(1980), is 74,900 Pa m3/g mol. From figure 3, it is 

apparent that, for practical purposes, all the resistance 

to the absorption of oxygen by water is in the wate^ 

film for all three combinations of film coefficients. 

Consequently, equation 17 applies, and the water-film 

mass-transfer coefficient for oxygen absorption can be 

measured directly.

The second reason for using oxygen as the 

reference substance for the water film is that considerable information exists on the oxygen-absorption 

coefficient, commonly called the reaeration coefficient, for streams. Tracer techniques (Tsivoglou and 

others, 1968; Kilpatrick and others, 1989) exist for 

measuring the oxygen-absorption coefficient, and 

numerous equations (Rathbun, 1977) exist for 

predicting the oxygen-absorption coefficient as a 

function of the hydraulic and physical characterises 

of streams.

Equation 25 has been verified in the laboratory 

for 25 organic solutes (Rathbun, 1990), most of them 

VOCs commonly used and observed in the environment. These verifications were done by measuring 

simultaneously the rates of absorption of oxygen 

and the desorption of the organic solute from wate*" 

in a tank stirred at a constant rate and maintained 

at a constant temperature. The absorption coefficient 

for oxygen and the desorption coefficient for the 

organic solute changed as the mixing rate was varied
12 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

from one experiment to the next, but the ratio of 

the coefficients was constant, within experimental 

error, thus verifying the underlying assumption 

of equation 25. Experiments at different temperatures also showed that the ratio was independent of 

temperature.

Water commonly is used as the reference 

substance for the air film because there is no liquidfilm resistance in the evaporation of a pure liquid; 

consequently, evaporation of a pure liquid is 

controlled 100 percent by the air-film resistance. Little 

information exists on the evaporation of water from 

streams. However, equations have been developed 

(Rathbun and Tai, 1983) for predicting the air-film 

coefficient as a function of windspeed and temperature 

for evaporation of water from a canal, which can be 

used as a reasonable approximation for streams.

Equation 26 has been verified in the laboratory 

for 1,2-dibromoethane and water (Rathbun and Tai, 

1986). This verification was done by measuring, under 

identical experimental conditions, the evaporation 

rates of pure 1,2-dibromoethane and pure water as a 

function of windspeed. These measurements showed 

that the mass-transfer coefficients for both of these 

compounds increased as the windspeed increased, 

but that the ratio of the coefficients was constant, 

within experimental error, thus verifying equation 26. 

Evaporation measurements for 23 other organic 

solutes and water (Rathbun, 1990) permit computation 

of the reference-substance parameter for these solutes; 

these measurements, however, were limited to a single 

windspeed.

The large number of VOCs in use and detected 

in the waters of the environment precludes the experimental measurement of the reference-substance 

parameters for all these VOCs. However, prediction 

equations have been developed (Rathbun, 1990). 

Theoretical, semiempirical, and empirical approaches 

were used. Theoretical approaches were based on 

Graham's law and the mass-transfer models discussed 

previously. Graham's law (Glasstone, 1946) states 

that the diffusion coefficient of a solute is inversely 

proportional to the square root of the molecular 

weight. The mass-transfer models indicate that the 

mass-transfer coefficient of a solute is dependent on 

the molecular-diffusion coefficient raised to a power, 

usually between 0.50 and 0.67. Semiempirical 

approaches consisted of using the mass-transfer 

models with empirical equations for predicting the 

molecular-diffusion coefficients of solutes in water

and air as a function of molecular weight and the 

molal volume at the normal boiling temperature. 

Empirical approaches consisted of computing loglog fits of the experimental values of the referencesubstance parameters as a function of nu^ecular 

weight and the molal volume.

Various combinations of these approaches 

resulted in five predictive equations for tH waterfilm reference-substance parameter and four 

predictive equations for the air-film referencesubstance parameter. Equations based on the 

empirical models had the smallest mean rbsolute 

percentage errors. These equations were

O = 2.52V -0.301 

m

= 4.42 M

-0.462

(27)

(28)

where

Vm is the molal volume at the normal boiling

temperature (cm3/g mol), and 

M is the molecular weight (g/g mo1 ). 

The molal volume can be computed from the 

molecular structure by using the LeBas procedure 

(Reid and others, 1987).

HENRY'S LAW CONSTANT

The two-film model and figure 3 indicate that 

the Henry's law constant is important in determining 

the partitioning of organic solutes between water 

and air. Various defining equations for the Henry's 

law constant have been presented in the literature. 

These defining equations have resulted in Henry's 

law constants with a variety of dimensional units and 

also a nondimensional form. Also, in some studies, a 

Henry's law constant equal to the reciprocal of the one 

commonly used has been defined. Consecmently, care 

should be exercised in using Henry's law constants 

from the literature. The following paragraphs present 

a discussion of these various forms of the Henry's law 

constant.

Henry's law applies to dilute solutions and 

describes the distribution of a solute between water 

and air phases at equilibrium. If these phases are in 

equilibrium, then the fugacities, f, must be equal:

f. = f« (29)
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

13

For the air phase,

(30)

where

y is the mole fraction of the solute in 

the air,

F is the fugacity coefficient, and

Pt is the total pressure, in pascals. 

The fugacity coefficient, F, characterizes the degree of 

nonideality in the air phase. For the low concentrations 

characteristic of environmental situations and environmental pressures, ideal mixtures can be assumed, and 

F will equal 1.0. By Dalton's law, the product of the 

mole fraction y and the total pressure, Pt, is the partial 

pressure of the solute, p. Consequently, it follows that

(31)

For the water phase,

fw = y mf fr (32)

where

Y is the activity coefficient of the solute

in water, 

mf is the mole fraction of the solute in

water, and

fr is the reference fugacity. 

The activity coefficient characterizes the degree of 

non-ideality between the solute and water, and it is 

defined on the convention that when the mole fraction, 

mf, is 1.0, the activity coefficient is 1.0. It follows 

from equation 32 that the reference fugacity is the 

fugacity of the pure solute in the liquid state at the 

system temperature. Thus, for typical environmental 

pressures, this reference fugacity is the vapor pressure 

of the pure solute. Therefore, combining equations 29, 

31, and 32 gives

p = Y mf p (33)

where 

P° is the vapor pressure, in pascals, of the solute 

at the system temperature.

Henry's law commonly is written

p = HC (34)

where H, the Henry's law constant, is the proportionality constant between the concentration of the 

solute in the water, C, and the partial pressure, p, 

of the solute in the air above the water. Combining 

equations 34 and 33 gives

Ymf p° (35)

Equation 35 indicates that the Henry's law constant 

has units of pressure divided by concentration, for 

example, Pa m /g mol.

However, the Henry's law constant has not 

always been defined in this manner. In early masstransfer studies (Lewis and Whitman, 1924; Sherwood 

and Holloway, 1940; Whitney and Vivian, 1949) and 

in some textbooks (Prutton and Maron, 1951; Rich, 

1973), the relation C = Hp was used, so that the 

Henry's law constant was the reciprocal of the one 

defined by equation 34, which is the one now used in 

most studies. Also, in some studies where distribution 

coefficients were determined (Hine and Mookerjee, 

1975; loffe and Vitenberg, 1982) and in studies wHre 

Bunsen coefficients were determined (Park and others, 

1982; Warner and Weiss, 1985; Bu and Warner, 

1995), the coefficients are the reciprocal of the 

commonly used Henry's law coefficient defined by 

equation 34.

Various forms of equation 35 have appeared in 

the literature. The solute mole fraction, mf, is given by

mf =

CV v s + M l w Vs

(36)

where

Vs is the volume of the water (m3 ), and 

Mw is the gram moles of water per cubic meter

(g mol/m3 ).

For dilute aqueous solutions, the MWVS term in the 

denominator of equation 36 is much larger than th^ 

CVS term, and it follows that

w 10/18

(37)
14 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

where 

10/18 is the gram moles of water per

cubic meter. 

Combining equations 37 and 35 gives

H = yp (38)

Equation 38 was presented by Mackay and Shiu 

(1975).

An expression for the activity coefficient of a 

solute in a dilute aqueous solution can be developed as 

follows. Consider an aqueous solution of the solute in 

equilibrium with the pure organic solute. At equilibrium, the fugacities are equal, or

and

f = f A org A aq

mforgYorg fr = mfaqYaq fr

(39)

(40)

where the org and aq subscripts indicate the organic 

and aqueous phases. If the solubility of water in the 

organic compound is small, it follows that

mfore org =1.0 and yore =1.0 org (41)

and

Yaq ~~

mf, (42)

aq

where 

mfaq is the mole fraction solubility of the

organic solute in water. 

Assuming again that the aqueous solution 

is dilute, which is true for most environmental 

conditions, it follows that

18 S

106 M (43)

where 

S

M

is the solubility of the organic solute

(mg/L), and 

is the molecular weight (g/g mol).

Combining equations 42, 43, and 38 gives

(44)

Equation 44 commonly is used to compute Henry's 

law constants from solubility and vapor-pressure data 

for the solute of interest (Mackay and Shiu, 1981).

Another form of the Henry's law constant 

that commonly is used in the literature is the nondimensional Henry's law constant. This constant is 

obtained by dividing by RT, where R is the ideal gas 

constant (8.314 Pa m3/g mol/K) and T is the absolute 

temperature, in kelvins. Equation 44 thus becomes

_ 0.120 Mp° (45)

where

Hd is the nondimensional Henry's law constant. 

Equation 45 with the numerical constant replaced by 

16.04 to accommodate vapor pressures in mm Hg was 

used by Dilling (1977).

The defining equation for the Henry's law 

constant (eq. 34) also can be converted to the 

nondimensional form by dividing the partial pressure, 

p, by RT. The result is

TI _ P/RT _ gmol/m" Hd -

C

(46)

gmol/m"

Equation 46 indicates that the nondimenr''onal 

Henry's law constant is the ratio of the air concentration to the water concentration at equilibrium. Thus, 

the nondimensional Henry's law constant provides a 

direct measure of the partitioning of the rolute 

between the water and air phases.

Finally, the Henry's law constant is sometimes 

expressed on a mole fraction basis, or

= Hmf mf (47)

where Hmf has units of pressure, usually pascals. 

References containing Henry's law constants with 

units of pressure have constants on a mole traction 

basis, for example, Keeley and others (1988). For 

dilute aqueous solutions, equation 37 applies, and 

combining this equation with equation 47 gives
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

15

P =

Hmf C 

M,, (48)

Comparing equations 48 and 34 indicates that

H =

Hmf

M,

(49)

TECHNIQUES FOR MEASURING 

THE HENRY'S LAW CONSTANT

Several techniques are available for measuring 

Henry's law constants. The multiple equilibration 

(ME) technique (McAuliffe, 1971; Munz and Roberts, 

1986) depends on the determination of the concentration of the organic compound in either the gas or 

liquid phase after repeated equilibrations of an inert 

gas with an aqueous sample containing the organic 

compound of interest. McAuliffe (1971) used 

measurements of the gas-phase concentration, and 

Munz and Roberts (1986) used measurements of the 

liquid-phase concentration. This technique assumes 

that the gas and liquid phases are in equilibrium at the 

end of each equilibration step.

The batch-stripping (BS) technique (Mackay 

and others, 1979) involves stripping the organic 

compound of interest from a water solution by 

bubbling an inert gas through the water. The Henry's 

law constant is determined from the slope of a plot of 

the liquid or gas concentration as a function of time. 

An advantage of the BS technique is that only relative 

concentrations are needed. This technique assumes 

that the concentration of the solute in the gas phase 

leaving the stripping chamber is in equilibrium with 

the liquid phase. The BS technique has been used for 

poly chlorinated biphenyl compounds (PCBs) with 

Henry's law constants as small as 91 Pa m3/g mol at 

25°C (Dunnivant and others, 1988).

The equilibration cell (EC) technique (Leighton 

and Calo, 1981; Tancrede and Yanagisawa, 1990) 

depends on the measurement of the concentrations 

of the organic compound in the air and water phases 

of a mixture in a closed container. This technique 

assumes that the air and water phases are homogeneous, 

in equilibrium, and that there are no temperature 

gradients. Nonattainment of equilibrium is most likely 

to be important for those compounds with the largest 

Henry's law constants. Equilibrium techniques such as 

the EC technique are generally applicable only to 

compounds with Henry's law constants larger than

250 Pa nr/g mol at 25°C because of the limited 

sizes of the headspace and water sample and limited 

analytical sensitivities (Fendinger and Glotfelty, 1988).

The equilibrium-partitioning-in-closed-systems 

(EPICS) technique (Lincoff and Gossett, 1984) is 

based on the measurement of the gas phase concentrations of the organic compound of interest in two 

closed bottles that contain different water volumes. 

This technique has the advantage that only relative 

rather than absolute concentrations are necessary. 

The EPICS technique assumes that the gas and liquid 

phases are in equilibrium. The coefficient of variation 

of Henry's law constants measured with the EPICS 

technique increased rapidly for constants less than 

about 500 Pa m3/g mol at 25°C (Gossett, 1987). 

Consequently, the technique is not useful for low 

values of the Henry's law constant.

The wetted-wall-column (WWC) technique 

(Fendinger and Glotfelty, 1990) involves equilibrating 

the organic compound of interest in a thin water f 1m 

on the wall of a vertical column with a concurrent 

flow of air. The organic compound can be introduced 

either as a vapor for air-to-water transfer or in solution 

for water-to-air transfer. The technique was used 

(Fendinger and Glotfelty, 1988) for three pesticides 

with Henry's law constants in the range from 0.2 to 

0.6xlO~4 Pam3/gmol.

The static-headspace (SH) technique (Robbins 

and others, 1993) involves determining the relative 

headspace concentration of one or more organic 

compounds of interest from aliquots of the same 

solution in three closed containers with different 

headspace-to-liquid volume ratios. This technique 

has the advantage that absolute concentrations are not 

needed, and the composition of the liquid matrix also 

is not needed. The essential assumptions are that the 

compositions of the solutions in the three containers 

are the same and that equilibrium is attained befor? the 

headspace samples are collected. The SH technique 

has been applied to the measurement of Henry's law 

constants as small as 53.5 Pa m3 /g mol at 25°C 

(Robbins and others, 1993).

The multiple-injection-interrupted-flow (MIIF) 

technique (Keeley and others, 1988) is a headspaoe 

technique based on comparative analyses of bottles 

containing vapor only with bottles containing the 

organic compound of interest and water. This 

technique has the advantage that little manipulation 

of the samples is required. However, the technique is 

limited to compounds with appreciable vapor pressure.
16 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

EXPERIMENTAL VALUES OF 

THE HENRY'S LAW CONSTANT

Experimental values of the Henry's law 

constant in Pa m3/g mol for 43 of the 55 target 

analytes (table 1) are presented in table 2 for temperatures of 20°C and 25°C. Some of these values were 

experimentally determined at these temperatures; 

others are interpolated values calculated at these 

temperatures from Henry's law constant-versustemperature relations. These interpolated values 

were used whenever the experimental temperatures 

were not at precisely 20°C or 25°C. The temperature 

relations were not used to extrapolate outside the 

reported temperature range of each study. The 

temperature relations are discussed in more detail 

in the section on the temperature dependence of the 

Henry's law constant.

When two or more experimental values existed 

at a temperature, an arithmetic mean Henry's law 

constant was computed. These mean values are given 

in table 2. Experimental values for four compounds 

were excluded from table 2 because they differed 

substantially from other values and were believed to 

be in error. These values are summarized in table 3. 

Three of the six excluded values are from McConnell 

and others (1975), and it was previously noted 

(Leighton and Calo, 1981) that some of their results 

appeared to be in error. The large differences shown in 

table 3 indicate the difficulties of measuring accurate 

values of the Henry's law constant.

Only two values were available for hexachlorobutadiene at 20°C, and those two values differ 

substantially (table 2). The larger of the two values 

was from McConnell and others (1975). Because 

values from this study were large for three of the 

other compounds for which values were excluded 

(table 3) and because the one value at 25°C (table 2) 

is intermediate between the 20°C values, the smaller 

value may be more representative of the true value 

at 20°C. There is, however, insufficient basis to 

exclude the 20°C value of McConnell and others 

(1975).

EQUATIONS FOR CALCULATING 

THE HENRY'S LAW CONSTANT

Experimental values of the Henry's law constant 

were not available for 12 of the target analytes. 

Constants for these compounds can be computed from 

equations 44 and 38. Mackay and Shiu (1981) used 

equation 44 for their extensive compilation of Henry's

law constants. The development of relatively simple 

gas-chromatographic techniques (Tse and others, 1992; 

Li and others, 1993) for the measurement cf activity 

coefficients has resulted in the use of equation 38 to 

compute Henry's law constants from these activity 

coefficients and vapor-pressure data.

Errors from using these equations result from 

failure to meet the assumptions in the derivations 

discussed previously and from errors in vapor 

pressures, solubilities, and activity coefficients. The 

equations assume that the mutual solubilities of the 

organic compound and water are small. This, this 

assumption is most valid for hydrophobic compounds 

and less valid for the more water-soluble compounds. 

Errors in vapor-pressure data in the literature have 

been noted, for example, by Hoffman (19P4) and 

Rathbun and Tai (1987). Nirmalakhandan and Speece 

(1988) noted that errors in vapor-pressure data could 

range from 6 to 200 or 300 percent, with the error 

increasing as the molecular weight increases and the 

vapor pressure decreases. Also, vapor-pressure data 

often are measured at high temperatures that require 

extrapolation to temperatures appropriate for environmental conditions, and such extrapolation could 

introduce errors.

Errors in solubility data in the literature also 

have been noted, for example, by Mackay and Shiu 

(1981), Bharath and others (1984), and Rathbun 

and Tai (1988). Many of these errors result from the 

shake-flask method used in early solubility determinations. This method involves mixing an excess of the 

organic solute with water, followed by separation of 

the phases and analysis of the water phase for the 

organic compound. This method, however, frequently 

results in the dispersion and retention of compound 

molecules in the water by the process of accommodation (Peake and Hodgson, 1966; 1967) that are not 

truly in solution. Consequently, the measured solubilities are larger than the true solubilities, with this 

source of error likely to be most important for the least 

soluble compound. The generator column technique 

(May and others, 1978) now used for solubility 

determinations does not have this problerr and also 

eliminates several other sources of error common in 

the shake-flask method. Nirmalakhandan and Speece 

(1988) indicated that experimental errors in solubility 

data are likely to be in the range of ±3 to 5 percent. 

They also noted that, unlike vapor-pressure data, the 

temperature dependence of water solubilities has not 

been studied extensively.
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

17

Table 2. Experimental and interpolated values of Henry's law constants at 20° Celsius and 25° Celsius for 43 target analytes

[USGS, U.S. Geological Survey; °C, degrees Celsius; H, Henry's law constant, in pascals cubic meter per gram mole; e, experimental 

value; i, interpolated value calculated from the Henry's law constant-temperature relationship; NS, not specified; EPICS, equilibrium 

partitioning in closed systems; EC, equilibrium cell; BS, batch stripping; ME, multiple equilibration; MIIF, multiple injection interrupted 

flow; SH, static headspace; WWC, wetted-wall column; av, arithmetic mean]

uses

_ . Temperature parameter Compound *o H e, i

code ( C)

34418 chloromethane 20

20

20

25

25

34423 dichloromethane 20

20

20

20

20

20

20

25

25

25

25

25

25

32106 trichloromethane 20

20

20

20

20

20

20

20

20

20

20

25

25

25

25

25

25

816

739

665

20°C av = 740

957

845

25°C av = 901

301

263

229

201

173

188

247

20°C av = 229

324

291

260

215

323

300

25°C av = 286

283

317

337

313

332

312

305

279

301

336

294

20°C av = 310

398

428

401

399

363

384

e

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e

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e

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e

e

i

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e

e

e

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e

e

e

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e

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i

Experimental „ , 

, . Reference technique

NS

NS

EPICS

NS

EPICS

NS

EC

EPICS

BS

EPICS

EPICS

EPICS

EC

EPICS

BS

EPICS

BS

EPICS

NS

EC

EPICS

BS

BS

BS

ME

EPICS

ME

EPICS

EPICS

EC

EPICS

BS

BS

EPICS

ME

Glew and Moelwyn-Hughes (1953)

McConnell and others (1975)

Gossett (198 7)

Glew and Moelwyn-Hughes (1953)

Gossett (198 7)

McConnell and others ( 1 975)

Leighton and Calo (1981)

Lincoff and Gossett ( 1 984)

Lincoff and Gossett (1984)

Gossett (198 7)

Yurteri and others (1987)

Ashworth and others (1988)

Leighton and Calo (1981)

Lincoff and Gossett (1984)

Lincoff and Gossett (1984)

Gossett (198 7)

Warner and others (1987)

Ashworth and others (1988)

McConnell and others (1975)

Leighton and Calo (1981)

Lincoff and Gossett (1984)

Lincoff and Gossett (1984)

Lalezary and others (1984)

Nicholson and others (1984)

Munz and Roberts (1986)

Gossett (1987)

Munz and Roberts (1987)

Ashworth and others (1988)

Dewulf and others (1995)

Leighton and Calo (1981)

Lincoff and Gossett (1984)

Lincoff and Gossett (1984)

Nicholson and others (1984)

Gossett (1987)

Munz and Roberts (1987)
18 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

Table 2. Experimental and interpolated values of Henry's law constants at 20° Celsius and 25° Celsius for 43 target analytes—Continued

[USGS, U.S. Geological Survey; °C, degrees Celsius; H, Henry's law constant, in pascals cubic meter per gram mole; e, experimental 

value; i, interpolated value calculated from the Henry's law constant-temperature relationship; NS, not specified; EPICS, equilibrium 

partitioning in closed systems; EC, equilibrium cell; BS, batch stripping; ME, multiple equilibration; MIIF, multiple injection interrupted 

flow; SH, static headspace; WWC, wetted-wall column; av, arithmetic mean]

USGS 

„ , Temperature „ parameter Compound v, H e, i

code

32 1 06 trichloromethane (Continued) 25

25

25

32102 tetrachloromethane 20

20

20

20

20

20

20

20

20

20

25

25

25

25

25

25

25

25

25

34413 bromomethane 20

25

32104 tribromomethane 20

20

25

25

25

32101 bromodichloromethane 20

25

25

32 1 05 chlorodibromomethane 20

20

343

427

378

25°C av = 391

2,220

2,140

2,450

2,030

2,380

2,390

2,350

2,280

2,350

2,000

20°C av = 2,260

2,750

2,440

3,050

3,020

3,060

2,990

2,980

2,880

2,560

25°C av = 2,860

532

631

44.2

43.9

20°C av = 44.0

61.1

54.1

53.9

25°C av = 56.4

157

212

215

25°C av = 214

83.4

104

20°C av = 93.7

e

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e

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e

e

Experimental n . 

. . Reference technique

BS

EPICS

EPICS

NS

EC

BS

ME

ME

ME

EPICS

EPICS

EPICS

EPICS

EC

ME

ME

EPICS

BS

EPICS

BS

EC

EPICS

NS

NS

BS

ME

BS

ME

BS

BS

BS

BS

BS

EPICS

Wamer and others (1987)

Ashworth and others ( 1 988)

Dewulf and others (1995)

McConnell and otl <?rs (1975)

Leighton and Calo (1981)

Munz and Roberts (1982)

Hunter-Smith and others (1983)

Munz and Roberts (1986)

Munz and Roberts (1987)

Gossett(1987)

Yurteri and others (1987)

Ashworth and others (1988)

Dewulf and others (1995)

Leighton and Calo (1981)

Hunter-Smith and others (1983)

Munz and Roberts (1987)

Gossett(1987)

Wamer and others (1987)

Ashworth and others (1988)

Ashworth and others (1988)

Tancrede and Yanagisawa ( 1 990)

Dewulf and others (1995)

Glew and Moelwy-r-Hughes (1953)

Glew and Moelwy-r-Hughes (1953)

Nicholson and others ( 1984)

Munz and Roberts (1987)

Nicholson and others (1984)

Munz and Roberts (1987)

Wamer and others (1987)

Nicholson and others (1984)

Nicholson and others (1984)

Warner and others (1987)

Nicholson and others (1984)

Ashworth and others (1988)
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

19

Table 2. Experimental and interpolated values of Henry's law constants at 20° Celsius and 25° Celsius for 43 target analytes—Continued

[USGS, U.S. Geological Survey; °C, degrees Celsius; H, Henry's law constant, in pascals cubic meter per gram mole; e, experimental 

value; i, interpolated value calculated from the Henry's law constant-temperature relationship; NS, not specified; EPICS, equilibrium 

partitioning in closed systems; EC, equilibrium cell; BS, batch stripping; ME, multiple equilibration; MIIF, multiple injection interrupted 

flow; SH, static headspace; WWC, wetted-wall column; av, arithmetic mean]

USGS 

parameter 

code

32105

34311

34496

32103

34506

_ . Temperature Compound *"0

chlorodibromom ethane (Continued) 25

25

25

chloroethane 20

20

25

25

1 , 1 -dichloroethane 20

20

20

25

25

25

25

1 ,2-dichloroethane 20

20

20

20

25

25

25

25

25

1,1,1-trichloroethane 20

20

20

20

20

20

20

20

20

20

H

116

79.3

120

25°C av = 105

946

1,110

20°C av = 1,030

1,130

1,230

25°C av = 1,180

435

570

390

20°C av = 465

551

552

633

502

25°C av = 560

92.3

127

149

78.3

20°Cav = lI2

155

111

143

63.6

99.2

25°Cav=114

1,540

1,510

1,100

1,340

1,370

1,360

1,340

1,570

1,480

1,190

20°C av = 1,380

e, i

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i

e

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i

Experimental 

technique

BS

BS

EPICS

EPICS

EPICS

EPICS

EPICS

EPICS

EPICS

EPICS

EPICS

BS

EPICS

EPICS

NS

EC

EPICS

EPICS

EC

BS

EPICS

BS

EPICS

EC

BS

ME

EPICS

BS

ME

EPICS

EPICS

EPICS

EPICS

Reference

Nicholson and others ( 1 984)

Warner and others ( 1 987)

Ashworth and others (198P)

Gossett(1987)

Ashworth and others (198P)

Gossett(1987)

Ashworth and others (1988)

Gossett(1987)

Ashworth and others (1988)

Dewulf and others 1995)

Gossett(1987)

Warner and others (1987)

Ashworth and others ( 1 988 )

Dewulf and others (1995)

McConnell and others (1975)

Leighton and Calo ( 1981)

Ashworth and others (1988)

Dewulf and others (1995)

Leighton and Calo (1981)

Warner and others (1987)

Ashworth and others (1988)

Ashworth and others (1988)

Dewulf and others (1995)

Leighton and Calo (1 98 1 )

Munz and Roberts (1982)

Hunter-Smith and others ( 1 983)

Lincoff and Gossett (1984)

Lincoff and Gossett (1984)

Munz and Roberts (1987)

Gossett (198 7)

Yurteri and others (1987)

Ashworth and others (1988)

Dewulf and others (1995)
20 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

Table 2. Experimental and interpolated values of Henry's law constants at 20° Celsius and 25° Celsius for 43 target analytes—Continued

[USGS, U.S. Geological Survey; °C, degrees Celsius; H, Henry's law constant, in pascals cubic meter per gram mole; e, experimental 

value; i, interpolated value calculated from the Henry's law constant-temperature relationship; NS, not specified; EPICS, equilibrium 

partitioning in closed systems; EC, equilibrium cell; BS, batch stripping; ME, multiple equilibration; MIIF, multiple injection interrupted 

flow; SH, static headspace; WWC, wetted-wall column; av, arithmetic mean]

USGS 

parameter 

code

34506

34511

34396

77651

34541

77443

34488

_, , Temperature Compound *

1,1,1 -trichloroethane 25

25

25

25

25

25

25

25

25

25

1 , 1 ,2-trichloroethane 20

20

25

25

25

hexachloroethane 20

20

20

25

25

25

1,2-dibromoethane 20

25

1,2-dichloropropane 20

20

25

25

25

1,2,3-trichloropropane 20

25

trichlorofluoromethane 20

20

20

H

1,970

1,320

1,710

1,740

1,690

1,720

1,760

1,780

1,780

1,520

25°C av = 1,700

67.0

75.0

20°C av = 71.0

82.5

92.2

113

25°C av = 95.9

290

285

599

20°C av = 391

393

998

846

25°C av - 746

61.8

65.9

224

193

20°C av = 208

286

286

362

25°Cav = 311

28.2

34.3

7,770

7,400

8,150

20°C av = 7,770

e, i

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e

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e

i

e

i

e

e

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i

e

Experimental 

technique

EC

ME

EPICS

BS

EPICS

ME

EPICS

BS

SH

EPICS

EC

EPICS

EC

EPICS

BS

ME

ME

EPICS

ME

BS

EPICS

EPICS

EPICS

EC

EPICS

EC

BS

EPICS

EC

EC

ME

EC

EPICS

Reference

Leighton and Calo ( 1981)

Hunter-Smith and others (1983)

Lincoff and Gosser (1984)

Lincoff and Gosser ( 1 984)

Gossett(1987)

Munz and Roberts (1987)

Ashworth and others (1988)

Ashworth and othe~s (1988)

Robbins and others (1993)

Dewulf and others (1995)

Leighton and Calo (1981)

Ashworth and othe~s (1988)

Leighton and Calo (1981)

Ashworth and othe-s (1988)

Ashworth and othe-s (1988)

Munz and Roberts ( 1 986)

Munz and Roberts ( 1 987)

Ashworth and othe-s (1988)

Munz and Roberts (1987)

Warner and others (1987)

Ashworth and othe-s (1988)

Ashworth and othe-s (1988)

Ashworth and othe-s (1988)

Leighton and Calo (1981)

Ashworth and othe-s (1988)

Leighton and Calo (1981)

Warner and others (1987)

Ashworth and othe-s (1988)

Leighton and Calo (1981)

Leighton and Calo (1981)

Hunter-Smith and others (1983)

Warner and Weiss (1985)

Ashworth and others (1988)
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

21

Table 2. Experimental and interpolated values of Henry's law constants at 20° Celsius and 25° Celsius for 43 target analytes—Continued

[USGS, U.S. Geological Survey; °C, degrees Celsius; H, Henry's law constant, in pascals cubic meter per gram mole; e, experimental 

value; i, interpolated value calculated from the Henry's law constant-temperature relationship; NS, not specified; EPICS, equilibrium 

partitioning in closed systems; EC, equilibrium cell; BS, batch stripping; ME, multiple equilibration; MllF, multiple injection interrupted 

flow; SH, static headspace; WWC, wetted-wall column; av, arithmetic mean]

USGS 

. _, , Temperature ¥¥ parameter Compound v, H e, i

code ( '

34488 trichlorofluoromethane (Continued) 25

25

25

25

25

34668 dichlorodifluoromethane 20

20

20

25

25

25

77652 1,1,2-trichloro- 1,2,2- 20 

trifluoroethane

20

25

25

39175 chloroethene 20

20

25

25

34501 1,1-dichloroethene 20

20

20

25

25

25

25

25

77093 c/s-l,2-dichloroethene 20

20

20

9,790

9,040

9,230

5,910

10,200

25°C av = 8,830

40,600

27,000

26,600

20°C av = 31,400

37,200

32,900

32,500

25°C av = 34,200

24,800

25,300

20°C av = 25,000

32,300

32,500

25°C av = 32,400

2,200

2,200

20°C av = 2,200

2,660

2,690

25°C av = 2,680

2,850

2,090

2,210

20°C av = 2,380

3,700

2,590

1,520

2,620

2,840

25°C av = 2,650

299

441

365

20°C av = 368

e

i

i

e

e

e

i

i

e

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e

i

e

i

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e

i

e

i

i

e

i

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e

e

e

i

e

e

Experimental _ , 

, . Reference technique

EC

ME

EC

BS

EPICS

NS

EC

ME

EC

EC

ME

EPICS

EC

EPICS

EC

EPICS

EPICS

EPICS

EPICS

EC

EPICS

EPICS

EC

EPICS

BS

EPICS

BS

EPICS

EPICS

EPICS

Park and others (1982)

Hunter-Smith and others ( 1 983)

Warner and Weiss ( 1 985)

Warner and others ( 1 987)

Ashworth and others (1988)

McConnell and others (1975)

Warner and Weiss (1985)

Munz and Roberts ( 1 987)

Park and others (1982)

Wamer and Weiss (1 985)

Munz and Roberts (1987)

Ashworth and others (1988)

Bu and Warner (1995)

Ashworth and others (1988)

Bu and Warner (1995)

Gossett(1987)

Ashworth and others ( 1 988)

Gossett(1987)

Ashworth and others (1988)

Leighton and Calo (1981)

Gossett(1987)

Ashworth and others ( 1 988)

Leighton and Calo (1981)

Gossett(1987)

Warner and others (1987)

Ashworth and others (1988)

Ashworth and others (1988)

Gossett(1987)

Yurteri and others (1987)

Ashworth and others (1988)
22 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

Table 2. Experimental and interpolated values of Henry's law constants at 20° Celsius and 25° Celsius for 43 target analytes—Continued

[USGS, U.S. Geological Survey; °C, degrees Celsius; H, Henry's law constant, in pascals cubic meter per gram mole; e, experimental 

value; i, interpolated value calculated from the Henry's law constant-temperature relationship; NS, not specified; EPICS, equilibrium 

partitioning in closed systems; EC, equilibrium cell; BS, batch stripping; ME, multiple equilibration; MIIF, multiple injection interrupted 

flow; SH, static headspace; WWC, wetted-wall column; av, arithmetic mean]

USGS 

. „ , Temperature 

parameter Compound 0 H e, i

code ( }

77093 C/.S- 1,2-dichloroethene (Continued) 25

25

34546 fra«5-l,2-dichloroethene 20

20

20

25

25

25

39180 trichloroethene 20

20

20

20

20

20

20

20

20

20

25

25

25

25

25

25

25

25

25

25

25

34475 tetrachloroethene 20

20

20

20

20

380

460

25°C av = 420

731

914

868

20°C av = 838

929

539

958

25°C av = 809

890

757

999

688

111

793

716

1,050

853

654

20°Cav = 818

984

1,030

881

984

1,190

1,040

944

1,030

768

1,050

864

25°C av = 979

2,000

1,230

1,530

1,320

1,190

i

e

i

e

e

i

e

e

e

i

e

i

i

i

i

e

e

i

i

i

i

e

e

i

i

e

e

e

i

e

i

e

i

i

Experimental „ _ 

. . Reference technique

EPICS

EPICS

EPICS

EPICS

EPICS

EPICS

BS

EPICS

NS

EC

BS

EPICS

BS

ME

EPICS

EPICS

EPICS

EPICS

EC

EPICS

BS

EPICS

BS

ME

EPICS

EPICS

EC

SH

EPICS

NS

EC

BS

EPICS

BS

Gossett(1987)

Ashworth and others (1988)

Gossett(1987)

Yurteri and others ( 1 987)

Ashworth and others ( 1 988)

Gossett(1987)

Warner and others ( 1 987)

Ashworth and others (1988)

McConnell and others ( 1 975)

Leighton and Calo (1981)

Munz and Roberts (1982)

Lincoff and Gossen (1984)

Lincoff and Gossen ( 1 984)

Munz and Roberts ( 1 987)

Gossett(1987)

Yurteri and others ( 1 987)

Ashworth and others ( 1 988)

Dewulf and others (1995)

Leighton and Calo (1981)

Lincoff and Gossen ( 1 984)

Lincoff and Gossen ( 1 984)

Garbarini and Lion (1985)

Warner and others ( 1 987)

Munz and Roberts ( 1 987)

Gossett(1987)

Ashworth and others ( 1 988)

Tancrede and Yamgisawa (1990)

Robbins and others ( 1993)

Dewulf and others (1995)

McConnell and others (1975)

Leighton and Calc (1981)

Munz and Roberts (1982)

Lincoff and Gossett( 1984)

Lincoff and Gossett (1984)
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

23

Table 2. Experimental and interpolated values of Henry's law constants at 20° Celsius and 25° Celsius for 43 target analytes—Continued

[USGS, U.S. Geological Survey; °C, degrees Celsius; H, Henry's law constant, in pascals cubic meter per gram mole; e, experimental 

value; i, interpolated value calculated from the Henry's law constant-temperature relationship; NS, not specified; EPICS, equilibrium 

partitioning in closed systems; EC, equilibrium cell; BS, batch stripping; ME, multiple equilibration; MUF, multiple injection interrupted 

flow; SH, static headspace; WWC, wetted-wall column; av, arithmetic mean]

USGS 

_, , Temperature parameter Compound *1

code ( '

34475 tetrachloroethene (Continued) 20

20

20

20

20

25

25

25

25

25

25

25

25

25

25

25

39702 hexachlorobutadiene 20

20

25

34030 benzene 20

20

20

20

20

25

25

25

25

25

25

25

25

25

34696 naphthalene 20

25

H

1,440

1,330

1,300

1,430

1,120

20°C av = 1,390

1,610

1,770

1,550

1,860

1,760

2,910

1,730

1,880

1,360

1,790

1,480

25°C av = 1,790

2,510

429

1,040

441

501

746

458

387

20°C av = 507

562

552

612

562

585

535

588

535

483

25°C av = 557

36.6

48.9

e, i

i

i

e

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i

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i

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e

e

e

e

e

i

e

e

e

i

i

e

e

i

e

i

i

e

e

e

e

e

i

e

e

Experimental 

technique

ME

EPICS

EPICS

EPICS

EPICS

EC

EPICS

BS

ME

EPICS

BS

EPICS

BS

EC

SH

EPICS

NS

BS

BS

EC

ME

EPICS

EPICS

EPICS

BS

EC

ME

BS

MIIF

EPICS

BS

SH

EPICS

EPICS

BS

Reference

Munz and Roberts (1987)

Gossett(1987)

Yurteri and others ( 1 987)

Ashworth and others (1988)

Dewulf and others (1995)

Leighton and Calo ( 1 98 1 )

Lincoff and Gossett ( 1 984 )

Lincoff and Gossett (1984)

Munz and Roberts ( 1 987)

Gossett (198 7)

Wamer and others ( 1 987)

Ashworth and others (1988)

Ashworth and others (1988)

Tancrede and Yanagisawa (1990)

Robbins and others ( 1 993)

Dewulf and others ( 1 995)

McConnell and others (1975)

Oliver (1985)

Wamer and others ( 1 987)

Leighton and Calo (1981)

loffe and Vitenberg ( 1 982)

Yurteri and others (1987)

Ashworth and others (1988)

Dewulf and others (1995)

Mackay and others (1979)

Leighton and Calo (1981)

loffe and Vitenberg ( 1 982)

Warner and others ( 1 987)

Keeley and others (1988)

Ashworth and others (1988)

Ashworth and others (1988)

Robbins and others (1 993)

Dewulf and others (1995)

Yurteri and others (1987)

Mackay and others (1979)
24 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

Table 2. Experimental and interpolated values of Henry's law constants at 20° Celsius and 25° Celsius for 43 target analytes—Continued

[USGS, U.S. Geological Survey; °C, degrees Celsius; H, Henry's law constant, in pascals cubic meter per gram mole; e, experimental 

value; i, interpolated value calculated from the Henry's law constant-temperature relationship; NS, not specified; EPICS, equilibrium 

partitioning in closed systems; EC, equilibrium cell; BS, batch stripping; ME, multiple equilibration; MIIF, multiple injection interrupted 

flow; SH, static headspace; WWC, wetted-wall column; av, arithmetic mean]

USGS 

parameter Compound 

code

34696 naphthalene (Continued)

340 1 0 methyl benzene

34371 ethylbenzene

77224 n-propylbenzene

77135 1,2-dimethylbenzene

Temperature 

<°C)

25

25

25

20

20

20

20

20

25

25

25

25

25

25

25

25

25

20

20

25

25

25

25

25

20

25

20

20

20

25

25

25

H c. i

56.0 e

44.6 Q

74.4 e

25°C av = 56.0

519 i

530 i

595 e

562 e

440 i

20°C av = 529

673 e

642 i

682 i

647 e

601 e

605 e

651 e

674 e

563 i

25°C av = 638

609 e

509 i

20°C av = 559

854 e

653 e

798 e

793 e

681 i

25°C av = 756

893 e

1,090 e

426 e

480 e

322 i

20°C av = 409

493 e

534 e

426 i

25°C av = 484

Experimental 

technique

BS

BS

WWC

EC

ME

EPICS

EPICS

EPICS

BS

EC

ME

EPICS

BS

MIIF

EPICS

SH

EPICS

EPICS

EPICS

BS

BS

EPICS

SH

EPICS

EPICS

EPICS

EPICS

EPICS

EPICS

EPICS

SH

EPICS

Reference

South worth (1979)

Mackay and others (1982)

Fendinger and Glotfelty ( 1 990)

Leighton and Calo ( 1981)

loffe and Vitenbe-g (1982)

Yurteri and others ( 1 987)

Ashworth and others (1988)

Dewulf and other? (1995)

Mackay and others ( 1 979)

Leighton and Calo (1981)

loffe and Vitenbe-g (1982)

Garbarini and Lion (1985)

Warner and other? (1987)

Keeley and other? (1988)

Ashworth and others (1988)

Robbins and othe-s (1993)

Dewulf and other? ( 1 995)

Ashworth and others ( 1 988)

Dewulf and others (1995)

Mackay and others (1979)

Warner and other? ( 1 987)

Ashworth and others (1988)

Robbins and othe-s (1993)

Dewulf and others ( 1 995)

Ashworth and others (1988)

Ashworth and others (1988)

Yurteri and others (1987)

Ashworth and others (1988)

Dewulf and others (1995)

Ashworth and others (1988)

Robbins and othe-s (1993)

Dewulf and others (1995)
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

25

Table 2. Experimental and interpolated values of Henry's law constants at 20° Celsius and 25° Celsius for 43 target analytes—Continued

[USGS, U.S. Geological Survey; °C, degrees Celsius; H, Henry's law constant, in pascals cubic meter per gram mole; e, experimental 

value; i, interpolated value calculated from the Henry's law constant-temperature relationship; NS, not specified; EPICS, equilibrium 

partitioning in closed systems; EC, equilibrium cell; BS, batch stripping; ME, multiple equilibration; MIIF, multiple injection interrupted 

flow; SH, static headspace; WWC, wetted-wall column; av, arithmetic mean]

USGS 

parameter 

code

85795

85795

77222

34301

34536

34566

_ . Temperature Compound ^

1,3-dimethylbenzene 20

20

20

25

25

25

1 ,4-dimethylbenzene 20

20

25

25

25

1,2,4-trimethylbenzene 20

chlorobenzene 20

20

20

25

25

25

25

25

1 ,2-dichlorobenzene 20

20

25

25

25

1,3-dichlorobenzene 20

20

25

25

H

430

606

481

20°C av = 506

547

754

615

25°C av = 639

654

456

20°C av = 555

754

762

605

25°C av = 707

475

269

319

346

20°Cav = 311

328

382

315

398

365

25°C av = 358

119

170

20°C av = 144

193

197

159

25°C av = 183

179

298

20°C av = 238

266

289

25°C av = 278

e, i

i

e

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e

e

i

e

i

e

e

i

e

e

e

e

e

e

e

e

e

e

e

e

e

Experimental 

technique

ME

EPICS

EPICS

ME

EPICS

EPICS

EPICS

EPICS

EPICS

BS

EPICS

EPICS

EC

EPICS

EPICS

EC

BS

BS

BS

EPICS

BS

EPICS

BS

BS

EPICS

BS

EPICS

BS

EPICS

Reference

loffe and Vitenberg (1982)

Ashworth and others (1988)

Dewulf and others (1995)

loffe and Vitenberg (1982)

Ashworth and others ( 1 988)

Dewulf and others (1995)

Ashworth and others (1988)

Dewul f and others (1995)

Ashworth and others ( 1 988)

Ashworth and others (1988)

Dewul f and others (1995)

Yurteri and others (1987)

Leighton and Calo ( 1981)

Yurteri and others (1987)

Ashworth and others (1988)

Leighton and Calo (1981)

Mackay and others (1979)

Mackay and Shiu ( 1981)

Warner and others (1987)

Ashworth and others ( 1 988)

Oliver (1985)

Ashworth and others (1988)

Mackay and Shiu (1981)

Warner and others ( 1 987)

Ashworth and others (1988)

Oliver (1985)

Ashworth and others (1988)

Warner and others ( 1 987)

Ashworth and others (1988)
26 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

Table 2. Experimental and interpolated values of Henry's law constants at 20° Celsius and 25° Celsius for 43 target analytes—Continued

[USGS, U.S. Geological Survey; °C, degrees Celsius; H, Henry's law constant, in pascals cubic meter per gram mole; e, experimental 

value; i, interpolated value calculated from the Henry's law constant-temperature relationship; NS, not specified; EPICS, equilibrium 

partitioning in closed systems; EC, equilibrium cell; BS, batch stripping; ME, multiple equilibration; MIIF, multiple injection interrupted 

flow; SH, static headspace; WWC, wetted-wall column; av, arithmetic mean]

USGS 

parameter 

code

34571

77613

34551

78032

^ . Temperature Compound r

1,4-dichlorobenzene 20

20

20

25

25

25

1,2,3-trichlorobenzene 20

25

1 ,2,4-trichlorobenzene 20

20

25

25

methyl tertiary-butyl ether 25

H

149

190

262

20°C av = 200

240

276

321

25°C av = 279

88.7

127

119

185

20°C av = 152

144

195

25°C av = 170

64.3

«> i

e

e

e

e

e

e

e

e

e

e

e

e

e

Experimental 

technique

BS

EPICS

EPICS

BS

BS

EPICS

BS

BS

BS

EPICS

BS

EPICS

SH

Reference

Oliver (1985)

Yurteri and other? (1987)

Ashworth and others (1988)

Mackay and Shiu (1981)

Warner and other? (1987)

Ashworth and oth-rs (1988)

Oliver (1985)

Mackay and Shiu (1981)

Oliver (1985)

Ashworth and others (1988)

Warner and other? (1987)

Ashworth and others (1988)

Robbins and others (1993)

Table 3. Henry's law constants excluded from the results presented in table 2 because they differed substantially from the average 

constants

[USGS, U.S. Geological Survey; °C, degrees Celsius; H, Henry's law constant, in pascals cubic meter per gram mole; av, arithmetic mean]

USGS 

parameter 

code

32106

34506

34506

34488

34501

34501

Compound

trichloromethane

1,1,1 -trichloroethane

1,1,1 -trichloroethane

trie h lorofl uoromethane

1 , 1 -dichloroethene

1,1-dichloroethene

Temperature

20

20

25

20

20

20

H

Table 2 

av

310

1,380

1,700

7,770

2,380

2,380

Excluded 

value

546

3,430

499

81,200

15,200

7,530

Referer<!e

Munz and Roberts (1982)

McConnell and others (1975)

Warner and others (1987)

McConnell and others (1975)

McConnell and others (1975)

Yurteri and others (1987)
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

27

CALCULATED VALUES OF 

THE HENRY'S LAW CONSTANT

Details regarding the calculation of the 

Henry's law constant for the 12 target analytes 

for which experimental values were not available 

are given in the following paragraphs. Some of 

the values are for 20°C, others are for 25°C. 

The temperature selected usually depended upon 

the temperature at which solubility data were 

available.

1. Bromoethene: No solubility data were available 

for this compound (Mackay and others, 

1993). The logarithm of the octanol-water 

partition coefficient, log Kow, for this 

compound was estimated from the log Kow 

value for chloroethene from Hansch and Leo 

as cited by Mackay and others (1993). The 

estimation procedure described by Lyman 

(1990a) was used. This procedure consisted 

of subtracting the value for the fragment factor 

for attachment of chlorine to a vinyl carbon 

atom and adding the corresponding factor 

for bromine. The result was a log Kow value 

for bromoethene of 1.52. Solubilities computed 

from equations 2-9 and 2-19 of Lyman (1990b) 

and this estimated log Kow were 10,100 and 

11,700 mg/L, respectively. Using the geometric 

mean of these values as suggested by Lyman 

(1990b), the estimated solubility for bromoethene at 25°C was 10,900 mg/L. Bromoethene 

has a boiling point of 15.8°C (Mackay and 

others, 1993). Consequently, it is a gas at 25°C. 

Solubilities for gases should be considered as 

under 1.0 atmosphere of the pure gas (Roberts, 

1984) rather than the corresponding vapor 

pressure of the pure gas. Thus, the Henry's 

law constant can be calculated from the 

estimated solubility and a pressure of 1.0 

atmosphere. The resultant Henry's law constant 

was 994 Pa m3/g mol.

2. 1.2-dibromo-3-dichloropropane: No data were 

available for this compound in Mackay and 

others (1993). Verschueren (1983) reported 

a vapor pressure of 107 Pa at 21°C and a 

solubility of 1,000 mg/L at room temperature, resulting in a calculated Henry's law 

constant of 25.3 Pa m3/g mol. Howard 

(1991) reported a vapor pressure of 77.3 Pa

at 20°C and a solubility of 1,230 mg/L, 

resulting in a calculated Henry's law constrnt 

of 14.8 Pa m3/g mol.

3. c/5-1.3-dichloropropene: Weber and others (1981) 

reported a vapor pressure of 4,000 Pa at 20°C, 

and Billing (1977) reported a solubility of 

2,700 mg/L at 20°C for this compound, 

resulting in a calculated Henry's law constrnt 

of 165Pam3/gmol.

4. trans- 1.3-dichloropropene: Weber and others 

(1981) reported a vapor pressure of 2,900 Fa 

at 20°C, and Dilling (1977) reported a solubility 

of 2,800 mg/L at 20°C for this compound, 

resulting in a calculated Henry's law constc it 

of 115Pam3 /gmol.

5. iso-propylbenzene: Four vapor pressures for 

this compound at 25°C (Mackay and others, 

1992a) ranged from 605 to 613 Pa and 

averaged 610 Pa. Eight water solubilities at 

25°C (Mackay and others, 1992a) ranged from 

48.3 to 80.4 mg/L and averaged 60.2 mg/L. The 

calculated Henry's law constant based on tl^se 

average values was 1,220 Pa m3/g mol. Li and 

others (1993) measured the infinite dilutionactivity coefficient of iso-propylbenzene in 

water at 25°C, and the result was 1.02xl05 . 

Using this value in equation 38, the calculated 

Henry's law constant was 1,120 Pa m /g mol.

6. n-butylbenzene: A vapor pressure of 137 Pa at 

25 °C was reported for this compound by 

Mackay and others (1992a). Solubility values 

determined using the shake-flask method at 

25 °C ranged from 11.8 to 50.5 mg/L (Mackay 

and others, 1992a). Three values determined 

using the generator column technique were all 

13.8 mg/L (Mackay and others, 1992a). Using 

this latter solubility, the calculated Henry's 

law constant was 1,340 Pa m3 /g mol. Using 

the infinite dilution-activity coefficient of 

n-butylbenzene in water (Li and others, 1993) 

and equation 38, the calculated Henry's law 

constant was 1,240 Pa m3/g mol. Using 

vapor pressures from Stull (1947) and 

solubility values from Owens and others 

(1986), calculated values were 1,100 and 

1,520 Pa m3/g mol at 20° and 25°C.
28 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

7. styrene: The vapor pressure for this compound 

at 25 °C calculated from an equation from 

Jordan (1954) was 864 Pa. Solubility values 

at 25°C (Mackay and others, 1993) ranged 

from 160 to 330 mg/L and averaged 264 mg/L. 

The calculated Henry's law constant was 

340 Pa m3/g mol.

8. ethyl tertiary-butyl ether: No data were available 

for this compound in either Mackay and others 

(1993) or Verschueren (1983). The vapor 

pressure for this compound was estimated 

from a logarithm vapor pressure-versus-normal 

boiling temperature regression for seven ethers 

using vapor-pressure data from Stull (1947). 

The boiling temperature of ethyl tertiary-butyl 

ether is 73.1°C (Weast, 1985), and boiling 

temperatures for the seven ethers were such that 

they bracketed the boiling temperature of ethyl 

tertiary-butyl ether. The vapor pressure at 20°C 

estimated by this procedure was 13,100 Pa. The 

water solubility at 20° was 12,000 mg/L (Evan 

and Edlund, 1936). The calculated Henry's law 

constant was 112 Pa m3 /g mol.

9. tertiary-amyl methyl ether: No data were available 

for this compound in either Mackay and others 

(1993) or Verschueren (1983). The normal 

boiling temperature for tertiary-amyl methyl 

ether is 86.3°C (Weast, 1985). The vapor 

pressure at 20°C estimated using the same 

relationship as used for ethyl tertiary-butyl 

ether was 7,760 Pa. The water solubility at 

19.9°C was 11,000 mg/L (Stephenson, 1992). 

The calculated Henry's law constant was 

72.1 Pam3/gmol.

10. diisopropyl ether: An experimental Henry's law 

constant for this compound was determined 

at 23°C using the batch stripping procedure 

(Nielsen and others, 1994). The resultant value 

of 209 Pa m3 /g mol was adjusted to values 

of 167 and 243 Pa m3 /g mol at 20° and 25°C 

using vapor-pressure data from Stull (1947), 

solubility data from Stephenson (1992), and 

equation 44.

11. 2-propenal: A logarithm vapor pressure-versusreciprocal absolute temperature relationship 

was computed for vapor-pressure data for this 

compound from Stull (1947). The vapor 

pressure calculated from this relationship at 

20°C was 29,400 Pa. The water solubility at

20°C was 208,000 mg/L (Verschueren, 1983). 

The calculated Henry's law constant was 

7.92 Pa m3 /g mol.

12. 2-propenenitrile: A logarithm vapor pressureversus-reciprocal absolute temperature relationship was computed for vapor-pressure data for 

this compound from Stull (1947). 7he vapor 

pressure at 25°C calculated from tHs relationship was 14,600 Pa. The water solubility at 

about 25°C was 80,000 mg/L (Lyman, 1990b). 

The calculated Henry's law constant was 

9.72 Pa m3 /g mol.

TEMPERATURE DEPENDENCE O^ 

THE HENRY'S LAW CONSTANT

The temperature dependence of the Henry's 

law constant usually is expressed with a form of the 

Arrhenius equation. The equation used is

loge H = A- B

(50)

where 

A and B are constants; and

T is the absolute temperature, in kelvins. 

Values of the A and B constants for 39 of the target 

analytes are given in table 4. For 30 of the target 

analytes, two or more temperature-dependence studies 

were available, and constants for each of these studies 

are given in table 4. The correlation coefficients also 

are given in table 4 to provide a measure of the degree 

of fit of the data to equation 50 and a measure of 

the internal consistency of the data for each isotherm. 

The trichloromethane isotherm of Tancrele and 

Yanagisawa (1990) was excluded because the 

predicted value at 25°C differed by more than three 

standard deviations of the mean value given in table 2.

Correlation coefficients were available 

for 114 isotherms. Of these isotherms, 94 had 

correlation coefficients greater than or equal to 

0.980, indicating good fits of the experimental 

Henry's law constants to equation 50. Only two 

correlation coefficients were less than O.SOO. These 

were the 0.876 value for hexachloroethane and the 

0.681 value for 1,2-dichlorobenzene (Ashworth and 

others, 1988). These investigators attributed the low 

value for 1,2-dichlorobenzene to a strong dependence 

of the activity coefficient on temperature for aromatic 

compounds in water and, therefore, deviations from 

the temperature dependence expressed by equation 50.
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

29

Table 4. Constants in the equation for the temperature dependence of the Henry's law constants of 39 target analytes

[USGS, U.S. Geological Survey; A, B, constants in the temperature dependence equation; °C, degrees Celsius; R, correlation coefficient; EPICS, equilibrium 

partitioning in closed systems; BS, batch stripping; nd, not determined]

USGS 

parameter Compound 

code

34418 chloromethane

34423 dichloromethane

32106 trichloromethane

32102 tetrachloromethane

3 44 1 3 bromomethane

3 2 1 04 tribromomethane

32101 bromodichloromethane

32105 chlorodibromomethane

34311 chloroethane

34496 1 , 1 -dichloroethane

32103 1,2 -dichloroethane

34506 1,1,1-trichloroethane

A

16.68

20.79

17.86

19.73

20.56

18.24

20.01

19.40

20.08

20.49

23.43

20.28

21.35

22.94

20.63

22.63

18.57

22.28

22.78

21.27

22.57

22.41

17.32

23.13

19.68

22.83

22.23

26.15

17.51

15.80

20.17

17.01

20.99

16.83

10.16

18.49

22.09

17.95

21.74

21.50

21.07

21.29

18.88

18.12

21.74

B

2,931

4,190

3,602

4,191

4,472

3,836

4,268

3,998

4,180

4,322

5,200

4,274

4,608

5,030

4,382

4,385

3,211

4,250

4,404

3,951

4,363

4,341

3,248

5,670

4,679

5,210

5,210

6,373

3,124

2,580

4,131

3,137

4,404

3,513

1,522

4,141

4,324

3,207

4,262

4,186

4,061

4,130

3,399

3,169

4,299

Temperature 

range 

(°C)

4.1-39.9

10.3-34.6

1.9-24.9

10-30

10-30

9.6-34.6

10-30

1.9-24.9

10-30

10-30

10-30

10-30

9.6-34.6

10-30

2.0-25.0

1.0-27.2

5-33

10-30

10.0-34.6

10-30

25-47.2

2.0-25.0

5.0-40.1

10-30

10-30

10-30

10-30

10-30

10.3-34.6

10-30

9.6-34.6

10-30

2.0-25.0

1.0-27.2

10-30

2.0-25.0

1.0-26.1

5-33

10-30

10-30

10-30

9.6-34.6

10-30

25-50

2.0-25.0

R

0.999

.995

.984

.994

.994

.971

.994

.996

.994

.997

.999

.997

.998

.998

.997

.997

nd

.998

.998

.998

.952

.986

.998

1.000

.999

.999

1.000

.956

1.000

.992

.997

.996

.983

.998

.937

.975

.998

nd

.999

.999

1.000

.997

.999

.985

.996

Reference

Glew and Moelwyn-Hughes (1953)

Gossett(1987)

Leighton and Calo ( 1981)

Lincoff and Gossett (1984) (EPICS)

Lincoff and Gossett (1984) (BS)

Gossett (198 7)

Ashworth and others ( 1988)

Leighton and Calo (1981)

Lincoff and Gossett (1984) (EPICS)

Lincoff and Gossett ( 1 984) (BS)

Nicholson and others (1984)

Munz and Roberts (1987)

Gossett (198 7)

Ashworth and others (1988)

Dewulf and others (1995)

Leighton and Calo (1981)

Hunter-Smith and others (1983)

Munz and Roberts (1987)

Gossett (1987)

Ashworth and others (1988)

Tancrede and Yanagisawa (1990)

Dewulf and others (1995)

Glew and Moelwyn-Hughes (1953)

Nicholson and others (1984)

Munz and Roberts (1987)

Nicholson and others (1984)

Nicholson and others (1984)

Ashworth and others (1988)

Gossett (1987)

Ashworth and others (1988)

Gossett (198 7)

Ashworth and others (1988)

Dewulf and others (1995)

Leighton and Calo (1981)

Ashworth and others (1988)

Dewulf and others (1995)

Leighton and Calo (1981)

Hunter-Smith and others (1983)

Lincoff and Gossett (1984) (EPICS)

Lincoff and Gossett (1984) (BS)

Munz and Roberts (1987)

Gossett (1987)

Ashworth and others (1988)

Robbins and others (1993)

Dewulf and others (1995)
30 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

Table 4. Constants in the equation for the temperature dependence of the Henry's law constants of 39 target analytes—Continued

[USGS, U.S. Geological Survey; A, B, constants in the temperature dependence equation; °C, degrees Celsius; R, correlation coefficient; EPICS, equilibrium 

partitioning in closed systems; BS, batch stripping; nd, not determined]

USGS 

parameter 

code

34511

34396

77651

34541

77443

34488

34668

77652

39175

34501

77093

34546

39180

34475

Compound

1 , 1 ,2-trichloroethane

hexachloroethane

1 ,2-dibromoethane

1 ,2-dichloropropane

1 ,2,3-trichloropropane

trichlorofluoromethane

dichlorodifluoromethane

1,1,2-trichloro- 1,2,2-

trifluoroethane

chloroethene

1 , 1 -dichloroethene

cis- 1 ,2-dichloroethene

irons-} ,2-dichloroethene

trichloroethene

tetrachloroethene

A

16.65

20.85

24.88

15.27

17.23

20.02

21.37

15.07

16.77

18.05

22.13

21.01

22.18

21.91

21.18

24.97

18.89

17.67

23.52

20.38

17.65

20.01

16.69

20.92

16.86

22.29

23.47

21.23

22.67

23.02

19.38

18.73

22.92

23.05

23.08

24.65

22.85

22.43

24.01

22.18

26.10

19.50

24.05

B

3,647

4,843

5,637

2,550

3,876

4,282

4,708

3,438

4,070

2,665

3,875

3,513

3,515

3,432

3,243

4,348

3,281

2,931

4,564

3,734

2,907

4,196

3,143

4,198

2,964

4,592

4,929

4,308

4,690

4,821

3,702

3,510

4,856

4,857

4,679

5,119

4,622

4,443

4,931

4,368

5,566

3,580

4,993

Temperature 

range

(°C)

2.5-26.1

10-30

10-30

10-30

10-30

1.9-24.9

10-30

13.6-24.9

26.5-45.0

5-33

0.85-40.8

10-30

10-30

0.85-40.8

10-30

0.04-39.6

10.3-34.6

10-30

2.5-26.1

10.0-34.6

10-30

10.3-34.6

10-30

10.0-34.6

10-30

1.0-26.1

10-30

10-30

10-30

9.6-34.6

10-30

25-50

25-47.2

2.0-25.0

1.0-26.1

10-30

10-30

10-30

9.6-34.6

10-30

25-47.2

25-45

2.0-25.0

R

0.996

.984

.987

.876

.963

.998

.906

.944

.989

nd

.997

.999

.999

.997

.965

.997

.994

.985

.981

.997

.987

.990

.987

.997

.992

.998

.996

.990

.998

.998

.999

.989

.942

.994

.997

.997

1.000

.998

.998

.993

.926

.987

.999

Reference

Leighton and Calo (1981)

Ashworth and others ( 1 988)

Munz and Roberts (1987)

Ashworth and others ( 1 988)

Ashworth and others ( 1 988)

Leighton and Calo (1981)

Ashworth and others (1988)

Leighton and Calo (1981)

Tancrede and Yanagnawa (1990)

Hunter-Smith and others (1983)

Warner and Weiss (1985)

Ashworth and others (1988)

Munz and Roberts (1987)

Warner and Weiss (1985)

Ashworth and others (1988)

Bu and Warner (1995)

Gossett(1987)

Ashworth and others (1988)

Leighton and Calo (1981)

Gossett(1987)

Ashworth and others (1988)

Gossett(1987)

Ashworth and others (1988)

Gossett(1987)

Ashworth and others (1988)

Leighton and Calo (1981)

Lincoff and Gossett (1984)(EPICS)

Lincoff and Gossett (1984)(BS)

Munz and Roberts (1987)

Gossett (1987)

Ashworth and others (1988)

Robbins and others (1993)

Tancrede and Yanagisawa (1990)

Dewulf and others (1995)

Leighton and Calo (1981)

Lincoff and Gossett (1984) (EPICS)

Lincoff and Gossett (1984) (BS)

Munz and Roberts (1987)

Gossett (1987)

Ashworth and others (1988)

Tancrede and Yanagisawa (1990)

Robbins and others (1993)

Dewulf and others (1995)
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

31

Table 4. Constants in the equation for the temperature dependence of the Henry's law constants of 39 target analytes—Continued

[USGS, U.S. Geological Survey; A, B, constants in the temperature dependence equation; °C, degrees Celsius; R, correlation coefficient; EPICS, equilibrium 

partitioning in closed systems; BS, batch stripping; nd, not determined]

USGS

parameter Compound A

code

34030 benzene 19.45

18.22

17.06

18.92

19.22

34010 methylbenzene 18.90

21.31

16.66

17.86

20.81

34371 ethylbenzene 23.45

22.19

23.60

77224 n-propylbenzene 19.36

77135 1,2-dimethylbenzene 17.07

17.68

22.40

85795 1,3-dimethylbenzene 20.37

17.81

20.89

85795 1 ,4-dimethylbenzene 18.46

22.88

34301 chlorobenzene 17.28

15.00

34536 1,2-dichlorobenzene 10.01

34566 1,3-dichlorobenzene 14.41

34571 1,4-dichlorobenzene 14.90

34551 1,2,4-trichlorobenzene 18.89

78032 methyl tertiary-butyl ether 30.06

This hypothesis is supported by the fact that

B

3,918

3,518

3,194

3,768

3,887

3,707

4,408

3,024

3,382

4,317

4,994

4,624

5,092

3,681

3,220

3,398

4,872

4,193

3,337

4,315

3,520

4,912

3,424

2,689

1,422

2,564

2,720

4,028

7,721

1 ,3-dichlorobenzene and 1 ,2,4-trichlorobenzene

also had relatively low correlation coefficients

(table 4). Conversely, chlorobenzene and 1,4-

dichlorobenzene in the same study had high

correlation coefficients.

When several isotherms are available for a

specific target analyte, the most logical choice is

to use the one with the best fit as indicated by the

largest correlation coefficient. However, consideration

also should be given to the temperature range.

Equation 50 assumes that the enthalpy of vaporization

Temperature

range R Reference

(°C)

1.0-27.2 0.998 Leighton and Calo ( 1 98 1 )

1 0-30 .998 loffe and Vitenberg ( 1 982)

1 0-30 .984 Ashworth and others ( 1 988)

25-50 .990 Robbins and others ( 1 993)

2.0-25.0 .996 Dewulf and others (1995)

1.0-23.0 .996 Leighton and Calo (1981)

1 0-30 .999 loffe and Vitenberg ( 1 982)

10-30 .991 Ashworth and others (1988)

25-50 .994 Robbins and others ( 1 993)

2.0-25.0 .996 Dewulf and others (1995)

10-30 1 .000 Ashworth and others ( 1 988)

25-40 1 .000 Robbins and others ( 1 993)

2.0-25.0 .989 Dewulf and others (1995)

10-30 . 99 8 Ashworth and others (1988)

10-30 .983 Ashworth and others (198 8)

25-50 .961 Robbins and others (1993)

2.0-25.0 .960 Dewulf and others (1995)

1 0-30 .98 1 loffe and Vitenberg (1 982)

1 0-30 .999 Ashworth and others ( 1 988)

2.0-25.0 .978 Dewulf and others (1995)

1 0-30 .994 Ashworth and others ( 1 988)

2.0-25.0 .966 Dewulf and others (1995)

1.0-23.0 .986 Leighton and Calo (1981)

10-30 .982 Ashworth and others (1988)

10-30 .681 Ashworth and others (1988)

10-30 .922 Ashworth and others (1988)

10-30 .970 Ashworth and others (1988)

10-30 .904 Ashworth and others (1988)

25-50 .981 Robbins and others (1993)

for the volatilization process is constant over the

temperature range of interest. This is a reasonable

assumption for the relatively short temperature ranges

indicated in table 4. Extrapolation outside these

temperature ranges, however, could lead to error.

The enthalpy of vaporization, contained in the

constant B of equation 50, decreases as the temperature increases (Reid and others, 1987). Consequently,

extrapolation of an isotherm beyond the experimental

temperature range will most likely result in predicted

Henry's law constants that are larger than the true

values. Such errors can be minimized by selecting an
32 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

isotherm with a temperature range that brackets the 

temperature at which the Henry's law constant is 

being predicted.

Isotherms were not available for 16 of the target 

analytes. To estimate Henry's law constants for these 

analytes for temperatures other than the 20°C or 25°C 

considered in table 2 and the text, a procedure based 

on equation 44 can be used. Writing equation 44 for a 

base temperature, bT in kelvins, and the temperature at 

which the Henry's law constant is to be estimated, T in 

kelvins, and taking ratios gives

pT SbT

pbT S1

(51)

where

HT and HbT are the Henry's law constants

(Pa m3/g mol) at the temperature 

of interest and at the base temperature, respectively; 

PJ and pt,T are the corresponding vapor pressures

of the pure solute; and 

ST and SbT are the corresponding water solubilities.

Vapor pressure data as a function of temperature 

are available for numerous compounds in several 

standard references such as Stull (1947), Jordan 

(1954), and Wilhoit and Zwolinski (1971). The 

temperature dependence of the water solubility has 

been studied much less extensively. However, some 

information exists, for example, in Stephen and 

Stephen (1963). This difference in effort is perhaps 

indicative of the fact that the vapor pressure is much 

more dependent on temperature than is the water 

solubility. If solubility data as a function of temperature are not available, equation 51 can be approximated as

PT —— 

PbT

(52)

Equations 51 and 52 were evaluated by 

calculating the Henry's law constant at 10°C for 

six chlorinated aliphatic VOCs and for benzene, and 

comparing the results with the mean values computed 

from the isotherm constants given in table 4. The 

arithmetic means of the Henry's law constants at 25°C 

from table 2 were used as the base values. These 

particular compounds were selected because solubility 

data as a function of temperature were available for

the chlorinated compounds in Stephenson (1992) and 

for benzene in May and others (1983). Vapor-pressure 

data for all compounds were from Stull (1947).

The results of the evaluation are presented 

in table 5. Comparisons are generally good, with 

the largest errors for 1,2-dichloroethane and 

1,2-dichloropropane. The larger errors for these 

two compounds could be the result of the fact that 

there were few Henry's law constants available. 

Also, there was considerable variability among the 

four values for 1,2-dichloroethane at 25°C (table 2), 

with values ranging from 63.6 to 155 Pa m3/g mol. 

Conversely, the three values available for 1,2- 

dichloropropane agreed reasonably well. The results 

in table 5 indicate that the predictive abilities of 

equations 51 and 52 are comparable, with equation 51 

having a slightly smaller overall average absolute 

percent difference. This limited test indicates that 

equations 51 and 52 can be used to predict the temperature dependence of the Henry's law constant when 

isotherm data are not available.

DISTRIBUTION OF HENRY'S LAW 

CONSTANTS FOR THE TARGET ANALYTES

Henry's law constants at 25°C for tH 

target analytes ranged from 9.72 Pa m /g mol for 

2-propenenitrile to 34,200 Pa m3/g mol ftr dichlorodifluoromethane. The median was 420 Pa m /g mol for 

cis- 1,2-dichloroethene.

The distribution of the Henry's law constants is 

presented in figure 4. The largest number of Henry's 

law constants (33) was within the class between 100 

and 1,000 Pa m3/g mol, and 46 analytes had constants 

larger than 100 Pa m3/g mol. The percentage of 

the total resistance to volatilization that is in the 

water film can be computed from equatior 20. 

Calculation of this percentage resistance for the 

average film coefficients of 3 m/day for the water 

film and 800 m/day for the air film indicates that 

90 percent of the total resistance will be in the water 

film for a solute with a Henry's law constant of 

83.7 Pa m3 /g mol. A line corresponding to this 

Henry's law constant is plotted in figure 4; this line 

includes one value from the 10 to 100 Pa m /g mol 

class. Consequently, 47 of the 55 target analytes 

have 90 percent or more of the resistance to volatilization in the water film for average environmental 

conditions, indicating that mixing within the water 

phase will control volatilization for most c f the target 

analytes.
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

33

Table 5. Evaluation of equations 51 and 52 for predicting the temperature dependence of the Henry's law constant

[USGS, U.S. Geological Survey; H 10, Henry's law constant at 10 degrees Celsius, in pascals cubic meter per gram mole; eq., equation; 

iso mean, mean Henry's law constant calculated from the isotherms in table 4; |av|, average absolute percent difference between eq. 51 

or eq. 52 and the isotherm mean]

USGS 

parameter 

code

34423

32106

32102

32103

34541

34475

34030

Compound

dichloromethane

trichloromethane

tetrachloromethane

1,2-dichloroethane

1 ,2-dichloropropane

tetrachloroethene

benzene

HIQ Percent difference

eq.51

136

170

1,420

54.6

155

717

257

eq. 52

150

196

1,380

54.5

147

762

258

iso mean

135

179

1,350

83.7

124

747

289

|av| =

eq.51

0.74

-5.03

5.19

-34.8

25.0

^.02

-11.1

12.3

eq. 52

11.1

9.50

2.22

-34.9

18.5

2.01

-10.7

12.7

•x>

30

CO

[if 25

>-

— J

2

h- 20

LLJ

O

CC

£

£'5

CC 

LLJ

CO

= 10

5

°0

I

-

LU 

O

K

CO

LU

CC

5

_1 

U.

CC

LU

<

•g

LU 

O

CC 

LU 

Q.

O

~

LU

o

K

CO

LU

CC

2

C

CC

LU

<£

1

O

CC 

LU 

Q.

s

I ———— •

1

LU 

0

K

CO

LU

CC

5

Lz

CC

sI*

1

1

————— r

. ......

-

•

-

*

•

-

j

•

-

-

"

-

-

1 1 10 100 1,000 10,000 100,000 1,000,000

HENRY'S LAW CONSTANT, IN PASCALS CUBIC METER PER GRAM MOLE 

Figure 4. Frequency distribution of the Henry's law constants at 25 degrees Celsius for the target analytes.
34 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

Conversely, about 50 percent of the resistance 

to volatilization of 2-propenenitrile and 2-propenal 

will be in the air film for the same average environmental conditions. Between these limits, there are 

six target analytes for which mixing conditions in 

both the air and water phases are likely to be important 

in controlling volatilization. These six are the 

compounds with the higher molecular weights 

and/or the higher water solubilities. They are, in 

increasing order of Henry's law constant, 1,2- 

dibromo-3-chloropropane, 1,2,3-trichloropropane, 

naphthalene, tribromomethane, methyl tertiary-butyl 

ether, and 1,2-dibromoethane.

Deviation from the average environmental 

conditions can either increase or decrease the relative 

importance of the two film resistances to volatilization. Percentage resistance in the water film as a 

function of the water-film mass-transfer coefficient 

is shown in figure 5 for a temperature of 25°C for 

three of the target analytes, with the air-film masstransfer coefficient as a parameter. The compound 

2-propenenitrile has the lowest Henry's law constant 

of any of the target analytes, and figure 5 indicates that 

a significant fraction of the resistance to volatilization 

is in the air film for this compound. As the waterfilm mass-transfer coefficient increases from the 

zero condition of molecular diffusion only, the 

percentage resistance in the water film decreases 

rapidly and, correspondingly, the resistance in the 

air film increases. As would be expected, the 

air-film resistance for a specific water-film masstransfer coefficient is least for the highest air-film 

mass-transfer coefficient. For most conditions, 

there are significant resistances to volatilization 

of 2-propenenitrile in both the water and air films.

Conversely, for tetrachloromethane, the lines 

for the three air-film mass-transfer coefficients are 

virtually indistinguishable from each other and from 

the 100-percent resistance line. Thus, volatilization 

of tetrachloromethane is water-film controlled, with 

virtually all the resistance in the water film. Methyl 

tertiary-butyl ether is intermediate between these 

extremes, with both films offering resistance to

Figure 5. Percentage resistance in the water film for the volatilization of three target analytes as a function of the water-film and 

air-film mass-transfer coefficients at 25 degrees Celsius; Henry's 

law constants are 9.72, 64.3, and 2,860 pascals cubic meter per 

gram mole for 2-propenenitrile, methyl tertiary-butyl ether, and 

tetrachloromethane, respectively.

100

80

60

40

20

Z 0 

LU

^100

LU

Q.

. 2-PROPENENITRILE

AIR-FILM COEFFICIENT

111

80

60

40

LL| 20 

111

S 100 

Q_

80

60

40

20

400 METERS/DAY 

800 METERS/DAY

———— 1,200 METERS/DAY

METHYL TERTIARY-BUTYL ETHEF

AIR-FILM COEFFICIENT

400 METERS/DAY 

800 METERS/DAY

____ 1,200 METERS/DAY

TETRACHLOROMETHANE

AIR-FILM COEFFICIENT

400 METERS/DAY 

800 METERS/DAY

_—— 1,200 METERS/DAY

12345

WATER-FILM MASS TRANSFER 

COEFFICIENT, IN METERS PEP DAY
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

35

volatilization, but with the water film having more 

than 60 percent of the total resistance for the three 

combinations of film coefficients considered.

The change in the dependence of the percentage 

resistance on the film coefficients as the Henry's law 

constant increases is evident in figure 5. However, 

only six target analytes have Henry's law constants 

lower than methyl tertiary-butyl ether. Consequently, 

dependence of the percentage resistance on the film 

coefficients for most of the target analytes will be as 

indicated in figure 5 for methyl tertiary-butyl ether 

and tetrachloromethane. This again emphasizes the 

importance of the water-film mass-transfer coefficient 

for most of the target analytes.

PREDICTION OF 

VOLATILIZATION COEFFICIENTS

Volatilization coefficients and the corresponding half-lives (eq. 3) for the target analytes for a 

specific stream can be predicted from equation 16, 

the two-film model (eq. 13), the Henry's law constant 

(table 2 and/or table 4), and the reference-substance 

concept (eqs. 25 and 26). Use of the referencesubstance concept requires a value of the oxygenabsorption (reaeration) coefficient and an estimate of 

the mass-transfer coefficient for water evaporation for 

the stream of interest.

The reaeration coefficient either can be 

measured using tracer techniques or predicted from 

the hydraulic and physical properties of the stream. If 

an accurate value of the volatilization coefficient for 

the target analyte is needed, then the reaeration coefficient should be measured using a tracer technique. 

Tracer measurements of the reaeration coefficient 

require much more effort than using empirical 

equations from the literature to predict the reaeration 

coefficient. However, the measured coefficient is 

much more likely to be an accurate representation of 

the oxygen-absorption characteristics of the stream. 

The radioactive tracer technique (Tsivoglou and 

others, 1968) uses krypton as the tracer for oxygen and 

tritium to adjust the krypton concentrations for the 

effects of dispersion and dilution. The hydrocarbon 

tracer technique (Kilpatrick and others, 1989) uses 

propane as the tracer for oxygen and rhodamine-WT 

dye as the dispersion-dilution tracer. Both steady-state 

and slug-injection versions of the hydrocarbon tracer 

technique are available.

If only an estimate of the volatilization 

coefficient for the target analyte is needed, then 

the reaeration coefficient can be estimated for the 

stream of interest by using one of the many reaerationcoefficient prediction equations from the literature. 

Several reviews of reaeration prediction equations 

have been done—for example, Bennett and Rathbun 

(1972); Bansal (1973); Zogorski and Faust (1973); 

Wilson and Macleod (1974); Covar (1976); and 

Rathbun (1977). The general consensus of these 

reviews is that prediction equations work reasonably 

well for the data upon which each equation is bas^d. 

However, when used on other streams or other reaches 

of the same stream with hydraulic and physical 

properties outside the range of the data base used in 

equation development, large prediction errors may 

result. Thus, a prediction equation should be selected 

that is based on hydraulic and physical data covering 

the same range as the data for the stream of interest.

To assist in selecting the appropriate equation, 

a chart of the form of Covar (1976) can be used. 

This chart has three areas of predicted reaeration 

coefficients as a function of flow depth and water 

velocity, with the areas corresponding to the equations 

of Owens and others (1964) for shallow flows, 

Churchill and others (1962) for intermediate-depth 

flows, and O'Connor and Dobbins (1958) for deen 

flows. The approach used here differs somewhat 

from that used by Covar (1976) with regard to the 

equation for shallow flows. Covar used the equation 

of Owens and others (1964) based on a data base that 

included the data of Churchill and others (1962). 

These data, however, are for depths ranging from 

0.646 to 3.48 m; thus, the shallow-flow data of O^ens 

and others (1964) are biased toward deeper flows by 

the inclusion of the Churchill and others (1962) data. 

A more appropriate shallow-flow equation is that of 

Owens and others (1964) based solely on their data, 

with depths ranging from 0.119 to 0.744 m. This 

equation is

K 2° = 6.92 U°'73 Y-L

(53)

where 

K220 is the base e reaeration coefficient at 20°C

(day' 1 ),

U is the water velocity (m/s), and 

Y is the flow depth (m). 

Water velocities of the data base for equation 53 

ranged from 0.040 to 0.558 m/s.
36 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

The intermediate-depth equation from Churchill 

and others (1962) is

K,M = 5.01 (54)

Equation 54 is based on a data base with water velocities ranging from 0.564 to 1.52 m/s and flow depths 

from 0.646 to 3.48 m.

The deep-flow equation from O'Connor and 

Dobbins (1958) is

K 2° = 3.93 U

°- 5° -1

(55)

Equation 55 was verified using a data base with water 

velocities ranging from 0.058 to 1.28 m/s and flow 

depths ranging from 0.274 to 11.3 m.

A chart showing various combinations of water 

velocity and flow depth appropriate for each of these 

three equations is presented in figure 6, with the limits 

of the data bases for each of the three equations 

indicated by rectangles. Figure 6 indicates that the 

reaeration coefficient, K2, can vary from approximately 100 day" 1 to 0.05 day" 1 , depending on the 

flow depth. Figure 6 can be used to determine which 

of equations 53, 54, or 55 should be used to compute 

reaeration coefficients for specific combinations of 

water velocity and flow depth. The K2 lines for the 

O'Connor and Dobbins equation (eq. 55) match 

reasonably well with the lines for the Churchill and 

others equation (eq. 54), with the slope of the lines 

being somewhat higher for the latter equation. The 

lines for the O'Connor and Dobbins equation (eq. 55) 

do not match as well with the lines for the Owens and 

others equation (eq. 53). Such disagreement, however, 

is common in predicting reaeration coefficients for 

streams. If the water-velocity and flow-depth values 

fall outside the three rectangles in figure 6, two 

options are available. The first is to use whichever 

of equations 53, 54, or 55 is closest to the desired 

values and extrapolate, with the understanding that 

prediction errors are likely to be larger when an 

equation is used outside the range of the data upon 

which the equation was based. The second option 

is to search the literature—for example, Rathbun 

(1977)—for an equation based on data that encompass 

the values of interest. In general, however, water 

velocities and flow depths likely to be measured

in streams of the United States will be within the 

ranges of water velocities and flow depth? covered 

in figure 6.

Equations 53, 54, and 55 give reaeration coefficients at 20°C. To adjust to other temperatures, the 

temperature correction of Elmore and West (1961) can 

be used. This correction has the form

K/ = K2 20 (1.0241)0 - 20 (56)

where 

K.2® is the reaeration coefficient (day" 1 ) at

temperature 0 (°C).

The mass-transfer coefficient for evaporation 

of water from a stream can be estimated from the 

equation (Rathbun and Tai, 1983)

k/6' 1 = 416+156 Va (57)

where

ka is the mass-transfer coefficient at 26.1 °C 

(m/day), and

Va is the windspeed (m/s). 

Any temperature could have been used as the base 

temperature. The temperature of 26.1 °C was chosen 

because this was the average water temperature of the 

data base used to establish the dependence of the gasfilm coefficient on windspeed. The mass-transfer 

coefficient at 26.1°C can be adjusted to the desired 

temperature by using the equation (Rathbun and Tai, 

1983)

26.1 V = k_exp [0.00934(0-26.1)] (58)

where

i, 0

is the mass-transfer coefficient (m/day) at

temperature 0 (°C).

Reference-substance parameters, O and VF, for 

the water- and air-film coefficients can be computed 

from equations 27 and 28. Equation 27 recuires the 

molal volume at the normal boiling temperature. 

This volume can be computed from the molecular 

structure using the LeBas procedure (Reid and others, 

1987).
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

37

15

3 

2.5

O)

£

LU

g 0.8 

^ 0.7-

o i ° 6

0.5 

0.4

0.3 

0.25

0.2 - 

0.15 -

0.1 -

0.09 -

0.08 -

J

O'CONNOR AND DOBBINS (1958) -

OWENS AND OTHERS (1964)

0.07

0.03 0.04 0.05 0.07 0.1 0.2 0.3 0.4 0.5 0.7 1

WATER VELOCITY, IN METERS PER SECOND

Figure 6. Oxygen-absorption (reaeration) coefficient (K.2) in per day at 20 degrees Celsius as a function of water velocity and 

flow depth for the equations of Owens and others (1964), Churchill and others (1962), and O'Connor and Dobbins (1958).
38 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

The reference-substance parameters for the

target analytes are presented in table 6. The water-film

reference-substance parameter, O, ranged from 0.506 

for hexachlorobutadiene to 0.774 for chloromethane.

This range indicates that the water-film mass-transfer

coefficient for the volatilization of the target analytes

from a stream will range from about 51 to 77 percent

of the reaeration coefficient for absorption of oxygen

by the same stream.

The air-film reference-substance parameter, 

¥, ranged from 0.338 for hexachlorobutadiene to

0.722 for chloromethane. This range indicates that 

the air-film mass-transfer coefficient for the volatilization of the target analytes from a stream will range

from about 34 to 72 percent of the mass-transfer

coefficient for evaporation of water from the same

stream. However, the relative importance of these two

mass-transfer coefficients is dependent on the Henry's 

law constant; as discussed previously, many of the 

target analytes have Henry's law constants large enough 

that the air- film mass-transfer coefficient will contribute

little to the overall volatilization coefficient. If VOCs

other than the 55 target analytes (table 1) prove to be

important in some streams, the reference-substance

parameters can be computed from equations 27

and 28.

Table 6. Water-film and air-film reference-substance

parameters for the target analytes

[USGS, U.S. Geological Survey; O, water-film reference-substance 

parameter; *P, air-film reference-substance parameter]

USGS

parameter Compound <D> T 

code

34418 chloromethane 0.774 0.722

34423 dichloromethane .697 .568

32106 trichloromethane .645 .485

32102 tetrachloromethane .607 .432

34413 bromomethane .763 .539

32104 tribromomethane .631 .343

32101 bromodichloromethane .640 .419

32105 chlorodibromomethane .636 .375

34311 chloroethane .694 .645

34496 1,1-dichloroethane .643 .529

32103 1,2-dichloroethane .643 .529

34506 1,1,1-trichloroethane .605 .461

34511 1,1,2-trichloroethane .605 .461

34396 hexachloroethane .530 .354

77651 1,2-dibromoethane .633 .393

Table 6. Water-film and air-film reference-substance

parameters for the target analytes — Continued

[USGS, U.S. Geological Survey; O, water-film reference-substance 

parameter; *P, air-film reference-substance parameter]

USGS

parameter Compound O 

code

34541 1,2-dichloropropane 0/03

77443 1,2,3-trichloropropane .574

82625 l,2-dibromo-3-chloropropane .568 

34488 trichlorofluoromethane .05

34668 dichlorodifluoromethane .670

77652 1,1,2-trichloro- 1,2,2- .583 

trifluoroethane

39175 chloroethene .716

34501 1,1-dichloroethene .659

77093 c/5-l,2-dichloroethene .659

34546 fra«s-l,2-dichloroethene .659 

39180 trichloroethene .617 

34475 tetrachloroethene .585

50002 bromoethene .709

34704 cis-l,3-dichloropropene .615

34699 fra«s-l,3-dichloropropene .615

39702 hexachlorobutadiene .506

34030 benzene .638

77128 styrene .578

34696 naphthalene .560

34010 methylbenzene .599

34371 ethylbenzene .569

77224 n-propylbenzene .544

77223 iso-propylbenzene .544

77342 n-butylbenzene .524

77135 1 ,2-dimethylbenzene .569

85795 1,3-dimethylbenzene .569

85795 1,4-dimethylbenzene .569

77222 1,2,4-trimethylbenzene .544

34301 chlorobenzene .601

34536 1,2-dichlorobenzene .572

34566 1 ,3-dichlorobenzene .572

34571 1 ,4-dichlorobenzene .572

77613 1,2,3-trichlorobenzene .548

34551 1,2,4-trichlorobenzene .548

78032 methyl tertiary-butyl ether .583

50004 ethyl tertiary-butyl ether .556

50005 tertiary-amyl methyl ether .556

81577 diisopropyl ether .556

34210 2-propenal .712

34215 2-propenenitrile .698

V

0.498

.440

.354 

.455

.482

.394

.654

.534

.534

.534 

.464 

.417

.510

.502

.502

.338

.590

.517

.470

.547

.512

.484

.484

.460

.512

.512

.512

.484

.499

.441

.441

.441

.400

.400

.558

.521

.521

.521

.688

.706
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

39

The procedure for estimating volatilization 

coefficients and the corresponding half-lives can 

be demonstrated best with an example. Consider 

a wastewater containing tribromomethane, 1,2,4- 

trichlorobenzene, and chloromethane that is 

discharged into a stream which has a water velocity 

of 0.307 m/s and a flow depth of 0.557 m. Estimate 

the distance downstream at which the concentrations 

will be reduced to 10 percent of the initial concentration for windspeeds of 0.2, 3.0, and 6.0 m/s. The water 

and air temperatures are both 16.5°C. Steps in the 

estimation procedure are:

1. Estimate the reaeration coefficient. For a water 

velocity of 0.307 m/s and a flow depth of 

0.557 m, figure 6 indicates that these values 

fall within the ranges of both the equation 

of Owens and others (1964) and the equation 

of O'Connor and Dobbins (1958). From 

equation 53, the calculated coefficient K220 

is 8.14 day" 1 and from equation 55, the 

calculated coefficient K220 is 5.24 day"1 . 

Use the smaller value because this will result 

in lower estimates of the mass-transfer coefficients and, therefore, possibly conservative 

estimates of the distances required for the 

concentrations to be reduced to 10 percent 

of the initial concentrations.

2. Adjust the estimated reaeration coefficient 

from 20°C to 16.5°C. From equation 56, 

K2 16-5 = 4.82 day" 1 .

3. Convert the reaeration coefficient to a film coefficient for the reference substance, oxygen, 

kwref' by multiplying by the flow depth. The 

equation used for this conversion is analogous 

to equation 16 and can be written kwref = K2 Y. 

The result is 2.68 m/day.

4. Calculate the water-film mass-transfer coefficients, kw, for the three compounds being 

considered in this example. Use equation 25 

and the water-film reference-substance 

parameters, <D, from table 6. The results 

for the three compounds are presented in 

table 7.

5. Calculate the air-film mass-transfer coefficient for 

evaporation of water from equation 57. The 

results are 447, 884, and 1,350 m/day for the 

three windspeeds.

6. Adjust these air-film mass-transfer coefficient? 

from the base temperature of 26.1°C to 16.5°C 

using equation 58. The results from this calculation are 409, 808, and 1,230 m/day.

7. Calculate the air-film mass-transfer coefficients 

for the three compounds being considered ir 

this example by using equation 26 and the airfilm reference parameters, *F, from table 6. ""he 

results for the three compounds for the three 

windspeeds are given in table 7.

8. Calculate the Henry's law constants at

16.5°C, using equation 50 and the constants 

from table 4. The two sets of constants 

for chloromethane give values of 558 and 

707 Pa m3/g mol. Use the average of 

632 Pa m3/g mol. The two sets of constants 

for tribromomethane give values of 35.0 

and 34.0 Pa m3/g mol. Use the average of 

34.5 Pa m3 /g mol.

9. Calculate the overall mass-transfer coefficients

based on the water phase by using the values in 

table 7 and equation 13. The results are given in 

table 7.

10. Calculate the volatilization coefficients, Kv, from 

equation 16 and the flow depth. The results are 

given in table 7.

11. Calculate the distances required for the concentration to be reduced to 10 percent of the initial 

concentration by using equation 5. The results 

are given in table 7.

The results in table 7 indicate that substantial 

distances are required for the concentration to be 

reduced to 10 percent of the initial concentration, 

even for chloromethane, which is a gas at normal 

environmental temperatures. The influence of 

the Henry's law constant is evident, with longer 

distances needed for tribromomethane, which has 

the lowest Henry's law constant of the three VOCs 

considered in this example. The distance required 

for the concentration of each VOC to be reduced 

to the target percentage decreased as the windspeed 

increased, with the effect of windspeed decreasing as 

the Henry's law constant increased. The windspeec1 

had appreciable effect on the volatilization of tribromomethane, less effect on 1,2,4-trichlorobenzene, 

and little effect on chloromethane. These observations 

are in agreement with previous discussions of the 

relationships among windspeed, Henry's law constant, 

and volatilization.
40 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

Table 7. Prediction of the volatilization coefficients and distances downstream required for 90-percent loss by volatilizat:on of three 

target analytes

[USGS, U.S. Geological Survey; kw, water-film mass-transfer coefficient, in meters per day; ka, air-film mass-transfer coefficient, in 

meters per day; H, Henry's law constant, in pascals cubic meter per gram mole; Kwo , overall mass-transfer coefficient based on the water 

phase, in meters per day; Kv, volatilization coefficient, in reciprocal days; L, distance, in kilometers]

USGS 

parameter Compound kw 

code

34418 chloromethane 2.07

3 2 1 04 tribromomethane 1 .69

34551 1 ,2,4-trichlorobenzene 1.47

ka

295

583

888

140

111

422

164

323

492

H Kwo

632 2.02

2.04

2.05

34.5 .917

1.19

1.32

146 1.28

1.37

1.40

Kv

3.63

3.66

3.68

1.65

2.14

2.37

2.30

2.46

2.51

L

16.8

16.7

16.6

37.0

28.5

25.8

26.6

24.8

24.3

DEPENDENCE OF THE VOLATILIZATION

HALF-LIFE ON WATER VELOCITY, FLOW

DEPTH, AND WATER TEMPERATURE

Procedures similar to those in the previous 

example can be used to estimate the effects of water 

velocity, flow depth, and water temperature on the 

volatilization half-life. The volatilization half-life 

is the time required for the concentration of a VOC 

to be reduced to one-half by volatilization. The halflives for tribromomethane, trichloromethane, and 

dichlorodifluoromethane are presented in figures 7, 8, 

and 9. These compounds represent low, middle, and 

high values of Henry's law constants (table 2) for the 

target analytes (fig. 4).

Figure 7 shows the effect of water velocity 

on the half-lives predicted from the equations of 

Owens and others (1964), Churchill and others 

(1962), and O'Connor and Dobbins (1958). These 

are equations 53, 54, and 55, and they were used in 

conjunction with equations 13, 25, 26, 16, and 3 to 

predict the volatilization half-lives. Values in figure 7 

are for a water temperature of 20°C. An air-film masstransfer coefficient of 800 m/day at a base temperature 

of 26.1 °C was assumed, and this value was adjusted to 

a value of 756 m/day at 20°C by using equation 58. 

Computations for each of the three equations were 

for the ranges of water velocities and midpoints of the

depth ranges of the data used to develop the equations. 

These depths were 0.4 m for the Owens and others 

(1964) equation, 2.0 m for the Churchill and others 

(1962) equation, and 6.0 m for the O'Connor and 

Dobbins (1958) equation.

The half-lives of the three VOCs decrease as 

the water velocity increases (fig. 7), which is consistent with much of the resistance to volatilization of the 

target analytes being in the water film. Consequently, 

as the water velocity increases, turbulent mixing 

within the water phase increases, the water-film 

thickness decreases, the mass-transfer coefficient 

for the water film increases, the volatilization coefficient increases, and the half-life decrease?- Halflives and the dependence of the half-life en water 

velocity are nearly the same for trichloromethane 

and dichlorodifluoromethane, despite the fact that 

the Henry's law constant for dichlorodifluoromethane 

(31,400 Pa m3 /g mol at 20°C) is about 100 times 

larger than the Henry's law constant for trichloromethane (310 Pa m3/g mol at 20°C). Figure 3, 

however, indicates that the Henry's law constants 

for both of these compounds are large enough that 

most of the resistance to volatilization is in the water 

film. Consequently, half-lives for a given flow depth 

and water temperature depend mostly on water 

velocity because the second term on the rght-hand
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

41

I

111

1.6

1.4

0.8

0.6

'•4o

0.5

0.4

1

Z 

tf

0.3

0.2

0.1

O'CONNOR AND DOBBINS (1958) 

FLOW DEPTH = 6 METERS

TRIBROMOMETHANE

TRICHLOROMETHANE

——— DICHLORODIFLUOROMETHANE

0.2 0.4 0.6 0.8 1.0 1.2 1.4

CHURCHILL AND OTHERS (1962) 

FLOW DEPTH = 2 METERS

——— TRIBROMOMETHANE

.......... TRICHLOROMETHANE

——— DICHLORODIFLUOROMETHANE

0.6 0.8 1.0 1.2 1.4 1.6

OWENS AND OTHERS (1964)

FLOW DEPTH = 0.4 METER

——— TRIBROMOMETHANE

.......... TRICHLOROMETHANE

——— DICHLORODIFLUOROMETHANE

0 0.1 0.2 0.3 0.4 0.5 0.6 

WATER VELOCITY, IN METERS PER SECOND

side of equation 13 is negligible with respect to the 

first term. Because the Henry's law constant for tritromomethane is relatively small (44.0 Pa m3/g mol 

at 20°C), the air film contributes some resistance to 

volatilization. The half-lives for all three compounds 

decrease considerably as the flow depth decreases, 

indicating that VOCs will persist much longer in de^p, 

slow-flowing streams than in shallow, swift-flowing 

streams.

Figure 8 presents half-lives for the three target 

analytes as a function of flow depth for water velocities of 0.3 m/s for the equation of Owens and others 

(1964), 1.0 m/s for the equation of Churchill and 

others (1962), and 0.7 m/s for the equation of 

O'Connor and Dobbins (1958). These velocities 

were selected because they were near the midpoint? 

of the velocity ranges of the data used to develop the 

equations. Half-life computations were over the flowdepth range appropriate for each of the equations.

Figure 8 indicates that the half-life increases as 

the flow depth increases. This again is consistent with 

much of the resistance to volatilization of the target 

analytes being in the water film. Consequently, for a 

specific water velocity, the effect of turbulent mixing 

on the water-film thickness decreases as the flow 

depth increases because of the greater water volume, 

the film thickness increases, the mass-transfer coef~- 

cient for the water film decreases, the volatilization 

coefficient decreases, and the half-life increases. TH 

half-lives for trichloromethane and dichlorodifluoromethane are similar for the three equations; the 

half-life for tribromomethane is longer. As for 

water velocity, the Henry's law constants for 

trichloromethane and dichlorodifluoromethane 

are such that virtually all the resistance to volatilization is in the water film, whereas the air film contri' x - 

utes some resistance to the volatilization of 

tribromomethane.

Figures 7 and 8 indicate that flow depth 

probably has more of an effect on the half-life than 

water velocity. This observation is consistent with the 

previous observation (Rathbun, 1977) that, with or\° 

exception, the exponent on flow depth was larger tl an

Figure 7. Variation of the volatilization half-life with water 

velocity for three target analytes at 20 degrees Celsius with flc w 

depth as a parameter; Henry's law constants are 44.0, 310, and 

31,400 pascals cubic meter per gram mole for tribromometharie, 

trichloromethane, and dichlorodifluoromethane, respectively.
42 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

16

14

12 

CO 

< 10

O'CONNOR AND DOBBINS (1958) 

VELOCITY = 0.7 METER PER SECOND 

———— TRIBROMOMETHANE

.......... TRICHLOROMETHANE

2.5

2.0

CO 

Q1.5

z

LJJ 

LJL 

_l 

LL 1.0

X 

0.5

CHURCHILL AND OTHERS (1962) 

VELOCITY = 1 METER PER SECOND

———— TRIBROMOMETHANE

.......... TRICHLOROMETHANE

———— DICHLORODIFLUOROMETHANE

0.35

0.30

0- 25

0.20

TJ 0.15 

LL

0.10

0.05

OWENS AND OTHERS (1964) 

VELOCITY = 0.3 METER PER SECOND

TRIBROMOMETHANE 

TRICHLOROMETHANE

——— DICHLORODIFLUOROMETHANE

10 12

0 0.2 0.4 0.6

FLOW DEPTH, IN METERS

0.8

the exponent on water velocity in semiempirical 

and empirical prediction equations for the reaeration 

coefficient.

Figure 9 presents half-lives as a function of 

water temperature. These half-lives were computed 

using equations 56 and 58 to define the temperature 

dependencies of the reaeration coefficient and the 

air-film mass-transfer coefficient. The constants 

from table 4 were used to define the temperature 

dependence of the Henry's law constants. Computations were for the combinations of water velocities 

and flow depths used for figures 7 and 8.

The results in figure 9 indicate that the half-lives 

for all three compounds decrease as the temperature 

increases. Such a decrease is consistent with increased 

molecular movement as the temperature increases, 

with correspondingly higher volatilization coefficients 

and, therefore, shorter half-lives. Half-lives for trichloromethane and dichlorodifluoromethane are similar, 

for reasons discussed previously. Half-lives for tribromomethane are considerably longer because the 

Henry's law constant for this compound is in the range 

where the air film contributes to the overall resistance 

to volatilization.

Results in figures 7, 8, and 9 for the three 

target analytes indicate that the volatilization halflives decrease as the temperature increases and as 

the flow depth decreases. The half-lives also generally 

decrease as the water velocity increases, with the 

effect dependent on the Henry's law constant of the 

VOC of interest. Graphs similar to figures 7, 8, and 9 

can be prepared for other VOCs as needed and/or for 

other water-velocity and flow-depth conditions.

This analysis of the effect of water velocity, 

flow depth, and water temperature on the half-lives of 

VOCs in streams is based on the two-film model. This 

model assumes thin films of water and air at the airwater interface in which mass transfer is Vy molecular 

diffusion alone. This is only a conceptual model, but it 

does provide a framework for analysis of the volatilization process and predictions are qualitatively 

consistent with observations.

Figure 8. Variation of the volatilization half-life vith flow 

depth for three target analytes at 20 degrees Celsius with water 

velocity as a parameter; Henry's law constants are 44.0, 310, and 

31,400 pascals cubic meter per gram mole for tribromomethane, 

trichloromethane, and dichlorodifluoromethane, respectively.
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

43

14

12

LU 8

< 6 I 6

2 

3.0

2.5

^2.0

a

O'CONNOR AND DOBBINS (1958)

VELOCITY = 0.7 METER PER SECOND

FLOW DEPTH = 6 METERS

——— TRIBROMOMETHANE

.......... TRICHLOROMETHANE

——— DICHLORODIFLUOROMETHANE

0.5

0 

0.5

0.4

CHURCHILL AND OTHERS (1962)

VELOCITY = 1 METER PER SECOND

FLOW DEPTH = 2 METERS

——— TRIBROMOMETHANE

.......... TRICHLOROMETHANE

———— DICHLORODIFLUOROMETHANE

1

z

LU

0.3

0.2

0.1

OWENS AND OTHERS (1964)

VELOCITY = 0.3 METER PER SECOND

FLOW DEPTH = 0.4 METER

———— TRIBROMOMETHANE

.......... TRICHLOROMETHANE

———— DICHLORODIFLUOROMETHANE

0 5 10 15 20 25 30 35

WATER TEMPERATURE, IN DEGREES CELSIUS

ABSORPTION

Absorption is the movement of a VOC from the 

bulk air phase above a stream across the air/water 

interface into the water. This process is sometimes 

called condensation or vapor deposition, and occur? 

when the air concentration is greater than the concentration that would be in equilibrium with the water 

concentration. Absorption is the reverse of volatilization, and can be described by the same equations, with 

the sign of the concentration-difference driving force 

reversed. Combining equations 9 and 10 gives

(59)

where

C is greater than p/H for volatilization, and 

p/H is greater than C for absorption. 

The overall mass-transfer coefficient, Kwo, is assumed 

to be the same for both processes, and this coefficient 

can be estimated for VOCs by using procedures 

described previously in the section on volatilization. 

A laboratory study (Sherwood and Holloway, 1940) 

indicated that mass-transfer coefficients for absorption 

and desorption of gases from water were the same, 

within experimental error.

The importance of volatilization as a 

mechanism for the loss of VOCs from streams has 

been firmly established. The importance of the reverse 

process of absorption is less clear. Absorption of 

VOCs would be most important in areas of high air 

concentrations and low water concentrations such that 

the driving force is toward the water. Target analyles 

with small Henry's law constants would be most 

susceptible to absorption because of their preference 

for the water phase. However, demonstration of 

absorption for VOCs may be difficult because of 

generally low ambient air concentrations and the 

corresponding low water concentrations that may 

be below analytical detection limits available at the 

present time. Methods for concentrating air samples 

to lower the concentrations that can be measured a~e 

fairly standard. Concentration methods for water

Figure 9. Variation of the volatilization half-life with water 

temperature for three target analytes; Henry's law constants a* 

20 degrees Celsius are 44.0, 310, and 31,400 pascals cubic meter 

per gram mole for tribromomethane, trichloromethane, and di^hlorodifluoromethane, respectively.
44 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

samples are less well developed. Some work has 

been done with PCBs in lakes. Doskey and Andren 

(1981) estimated that absorption could be a significant 

source of PCBs in Lake Michigan when the air film 

controlled the uptake process. Conversely, they 

estimated that volatilization was important when 

the water film controlled the mass-transfer process. 

The uncertainty in the estimation of the direction of 

transfer was the result of inaccuracies in the Henry's 

law constant for the PCBs, the determination of 

the controlling film resistance, and the resultant 

overall mass-transfer coefficients. Swackhamer 

and others (1988) determined from mass-balance 

calculations that the fluxes of 17 PCB congeners 

in Siskiwit Lake were from the lake to the atmosphere, 

indicating volatilization. The fluxes for four other 

congeners were small or negative. Lakes, however, 

are less dynamic than streams, and demonstration 

of absorption in streams is expected to be more 

difficult.

DRY DEPOSITION WITH PARTICLES

The dry deposition of organic compounds 

associated with particles onto a surface depends on the 

characteristics of the surface, mass-transfer resistance 

in the deposition layer above the surface, particle size 

and concentration, concentrations of the organic 

compound in the air, and meteorological conditions 

(Eisenreich and others, 1981). In simplified form, the 

deposition flux is defined by

= VdCa (60)

where

f\

Nd is the flux (g mol/m /s), 

Vd is the deposition velocity (m/s), and 

Ca is the concentration of the organic compound 

sorbed on the particles in the air 

(g mol/m3 ).

The deposition velocity depends on the surface type, 

the distribution of the particle sizes, and the windspeed 

(Eisenreich and others, 1981) and also the near-surface 

turbulence (Swackhamer and others, 1988). There is 

considerable uncertainty in values of the deposition 

velocity, and this value is difficult to measure because 

of uncertainties in extrapolating from dry collector 

surfaces to water surfaces (Swackhamer and others, 

1988).

The sorption of organic compounds to particles 

in the atmosphere depends on the vapor pressure of 

the compound and the size, surface area, snd organiccarbon content of the particles. Eisenreich and 

others (1981) estimated for the range of eroected 

atmospheric environments that organic compounds 

with vapor pressures greater than 1(T4 mm Hg 

(1.3xlO~2 Pa) should exist almost entirely in the 

gas phase, and those having vapor pressures less 

than 10~8 mm Hg (1.3xl(T6 Pa) should exist almost 

entirely in the particulate phase. Vapor pressures 

at 25°C for the target analytes range from about 

4,200 mm Hg (5.6xl05 Pa) for chloromethane to 

about 7.5x 10~2 mm Hg (10 Pa) for naphthalene. 

Consequently, the target analytes will all b<*. essentially 

100 percent in the gas phase, and dry deposition with 

particles will not be a significant process for these 

compounds.

WET DEPOSITION

It is difficult, if not impossible, to separate 

experimentally the contributions to the removal of 

organic compounds from the air made by wet deposition with particles and gas scavenging. The total 

washout ratio, Wt, is given by (Pankow and others, 

1984)

Wt = (61)

where

Ww g

W

is the gas scavenging ratio, 

p is the particle scavenging ratio, and 

(|> is the fraction of the organic compound in 

the air that is associated with the particulate phase.

Because § is estimated to be small for the target 

analytes, equation 61 reduces to

Wt = Wg . (62)

Gas scavenging occurs when gaseoMS organic 

compounds partition into the precipitation and fall to a 

land or water surface. The gas scavenging ratio, Wg, 

can thus be defined as
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

45

W = (63)

where

Cp and Ca are the concentrations of the organic

compound in the precipitation

and the air.

It has been estimated (Slinn and others, 1978) that 

a falling raindrop reaches equilibrium with trace 

concentrations of organic compounds in the air in 

a fall distance of about 10m. Consequently, it 

is generally assumed that equilibrium exists, and 

equation 63 can be combined with Henry's law 

(eq. 34) and the ideal gas law to give

Wg = (64)

where

H is the Henry's law constant (Pa m3/g mol). 

Thus, Wg is equal to the reciprocal of the nondimensional Henry's law constant as defined by equation 46.

The temperature dependence of the gas 

scavenging ratio can be estimated from equation 64 

and the temperature dependence of the Henry's law 

constant as given by equation 50. Using trichloroethene 

as an example, the constants of Dewulf and others 

(1995) from table 4 yield computed values of the gas 

scavenging ratio at 25°C and 8°C of 2.88 and 7.26, 

respectively. These values indicate that the ratio 

increases as the temperature decreases so that the 

trichloroethene concentration in the precipitation 

increases with decreasing temperature. The exponential 

dependence of the Henry's law constant on temperature 

has a much greater effect on the gas scavenging 

ratio than the explicit dependence on temperature 

in equation 64. The ratio of the gas scavenging ratios 

at 8°C and 25°C for trichloroethene is 2.5, which is 

generally consistent with the ratio of 3 to 6 found 

between experimental values at 8°C for a number of 

organic compounds and values calculated from the 

Henry's law constant at 25°C (Ligocki and others, 

1985).

Gas scavenging ratios for organic compounds 

generally range between 1 and 104 (Eisenreich and 

others, 1981), with the ratio increasing as the Henry's 

law constant decreases (eq. 64). It was estimated 

previously that wet deposition with particulate mattef

was not likely to be important for the target analytes 

because of their very limited tendency to sorb to 

particulate matter in the atmosphere.

The most common forms of precipitation 

are rain and snow, and concentrations of organic 

compounds in these types of precipitation might b^ 

expected to be different because of different temperatures. However, Swackhamer and others (1988) noted 

that the concentrations of PCBs in rain and snow 

in the Siskiwit Lake region were similar, with a few 

exceptions. Conversely, Czuczwa and others (1988) 

noted significant seasonal differences in the concentrations of alkylbenzenes in precipitation, with snow and 

winter rain having the highest concentrations, autumn 

rain having intermediate concentrations, and spring 

and summer rain having the lowest concentrations1 . Possible explanations are seasonal temperature differences that affect the equilibria between the air anc? 

precipitation phases, and seasonal differences in the 

atmospheric mixing heights and atmospheric reactivities that affect the air concentrations. There also a~e 

indications that concentrations of organic compounds 

in fog are enriched relative to other types of precipitation. This enrichment results from the high specific 

surface area of the fog droplets, with adsorption 

of the organic compounds occurring at the air-water 

interface. Goss (1994) calculated an enrichment 

factor of 1.8 for 1,2,4-trichlorobenzene for 8-jum 

droplets.

EXPERIMENTAL VALUES OF THE 

GAS SCAVENGING RATIO

Experimental values of the gas scavenging ratio 

for 11 target analytes are presented in table 8. Mo^t of 

the data are accompanied by values indicating ± one 

standard deviation of the mean value and, therefore, 

the variability in the data. The considerable variability 

in the data is indicative of the experimental difficulties 

in collecting and analyzing air and precipitation 

samples for small concentrations of VOCs.

CALCULATED VALUES OF THE 

GAS SCAVENGING RATIO

Gas scavenging ratios for the other 44 target 

analytes or for other VOCs that might be important in 

certain situations can be estimated using equation 64 

and Henry's law constants from table 2. If ratios ?t 

a temperature other than 20°C or 25 °C are needed
46 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

the Henry's law constant can be computed at the 

appropriate temperature using equation 50 and 

temperature parameters from table 4. If the temperature parameters are not available for the compound 

of interest, Henry's law constants at the appropriate 

temperature can be estimated from equation 51 or 52 

and procedures discussed previously. The same 

procedures can be used to adjust the experimental 

ratios in table 8 to others temperatures.

ESTIMATION OF THE IMPORTANCE 

OF WET DEPOSITION

A mass balance about the surface of a stream 

during a storm event yields

(WJ(RF)(Ca )

(65)

where 

Cww is the concentration (ng/L) of the organic

compound in the stream resulting from gas 

scavenging from the atmosphere during 

the storm,

RF is the amount of rainfall (m),

Ca is the concentration (ng/L) of the compound 

in the air, and

Y is the flow depth (m) of the stream. 

Equation 65 is a simplistic approach and assumes that 

the rainfall occurs instantaneously and is uniformly 

and instantaneously distributed over a stream reach of 

constant width and depth containing no background 

concentrations of the organic compounds of interest. 

Consequently, estimates of gas scavenging and the 

resultant wet deposition are likely to be upt>er bound 

estimates. The relative validity of equation 65 

increases as the intensity of the precipitation event

Table 8. Experimental values of gas scavenging ratios for 11 target analytes

[USGS, U.S. Geological Survey; °C, degrees Celsius; Wg, mean value of the gas scavenging ratio; ±, one standard deviation; nd, not determined]

uses

parameter 

code

39180

34475

34696

34010

34371

77135

85795

77222

34536

34571

34551

Compound

trichloroethene

tetrachloroethene

naphthalene

methylbenzene

ethylbenzene

1 ,2-dimethylbenzene

1 ,3-,l ,4-dimethylbenzene

1 ,2,4-trimethylbenzene

1 ,2-dichlorobenzene

1 ,4-dichlorobenzene

1 ,2,4-trichlorobenzene

Temperature

(°C)

8

8

11

8

8

7

-1

8

7

-1

8

7

-1

8

7

-1

8

7

-1

8

11

8

8

wg

3.7 ±1.3

3.611.1

230 ± nd

250 ± 73

22 ±5

31 ±9

42 ±16

27 ±11

32 ±48

36 ±34

35 ±15

69 + 53

94 + 66

33 ±17

70 + 70

67 + 50

27 + 9

54 ±44

52 ±40

46 ±13

44±nd

39 ±10

66 ±51

Reference

Ligocki and others (1985)

Ligocki and others (1985)

Pankow and others (1984)

Ligocki and others (1985)

Ligocki and others (1985)

Hart and others (1993)

Hart and others (1993)

Ligocki and others (1985)

Hart and others (1993)

Hart and others (1993)

Ligocki and others (1985)

Hart and others (1993)

Hart and others (1993)

Ligocki and others (1985)

Hart and others (1993)

Hart and others (199 3)

Ligocki and others (1985)

Hart and others (1993)

Hart and others (1993)

Ligocki and others (1985)

Pankow and others (1984)

Ligocki and others (1985)

Ligocki and others (1985)
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

47

increases. On the other hand, the more intense the 

event, the less likely equilibrium is achieved between 

the precipitation and the air.

Naphthalene, methylbenzene, and tetrachloroethene were selected as examples to demonstrate the 

possible significance of wet deposition as a source 

of VOCs in streams..These three target analytes were 

chosen because they have been detected in urban 

air (Ligocki and others, 1985). Also, these three 

compounds have Henry's law constants that approximately span the range of Henry's law constants for the 

target analytes.

The concentrations of these three target 

analytes predicted to occur in a stream as result 

of gas scavenging were computed from equation 65. 

The concentrations are presented in figure 10 as a 

function of flow depth for temperatures of 5°C, 15°C, 

and 25°C. Gas scavenging ratios were computed from 

equation 64. The average Henry's law constants at 

25°C from table 2 were used. Henry's law constants at 

5°C and 15°C for methylbenzene and tetrachloroethene were computed from equation 50 and the 

constants of Dewulf and others (1995) as given in 

table 4. Henry's law constants at 5°C and 15°C for 

naphthalene were computed from equation 51 because 

temperature-dependent data were not available for this 

compound. Solubilities were from May and others 

(1983), and vapor pressures were estimated from an 

equation from Jordan (1954). A rainfall of 0.025 m 

was assumed, and air concentrations for the three 

compounds were from Ligocki and others (1985).

The results in figure 10 indicate that the concentrations of the three compounds in the water decrease 

rapidly as the flow depth increases because of the 

increasingly larger volume of water into which the 

scavenged compounds are mixed. The results also 

indicate that the concentrations decrease as the 

Henry's law constant increases, indicating a 

decreasing tendency to be scavenged. Decreasing the 

temperature results in increases of the concentrations 

of the three compounds in the stream water.

Figure 10. Concentrations of three target analytes in a stream 

resulting from gas scavenging from the atmosphere as a function 

of flow depth and water temperature; calculated Henry's law 

constants at 15 degrees Celsius are 19.6, 340, and 831 pascals 

cubic meter per gram mole for naphthalene, methylbenzene, and 

tetrachloroethene, respectively; air concentrations (Ligocki and 

others, 1985) are 0.450, 3.80, and 1.20 nanograms per liter for 

naphthalene, methylbenzene, and tetrachloroethene, respectively.

TETRACHLOROETHENE 

5 DEGREES CELSIUS 

15 DEGREES CELSIUS 

——— 25 DEGREES CELSIUS

METHYLBENZENE 

5 DEGREES CELSIUS 

15 DEGREES CELSIUS

___ 25 DEGREES CELSIUS

=3 35

CO

LU 

CC 

CC 30

I z 25

NAPHTHALENE 

. 5 DEGREES CELSIUS 

, 15 DEGREES CELSIUS

——— 25 DEGREES CELSIUS

2468

FLOW DEPTH, IN METERS
48 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

The concentrations in figure 10 for all three 

compounds are in the low nanogram-per-literconcentration range. These concentrations are 

below the current analytical method reporting limits 

(MRLs) of 50 ng/L for methylbenzene and tetrachloroethene and 200 ng/L for naphthalene (J.S. Zogorski, 

U.S. Geological Survey, written commun., 1996). The 

fact that these concentrations are less than the MRLs 

does not, however, mean that these concentrations are 

not hazardous to humans or aquatic life. Indeed, little 

is known about the effects of long-term exposures 

of humans and aquatic life to low concentrations 

of organic compounds. Consequently, as MRLs are 

lowered, the question of what concentration levels are 

hazardous must be addressed. Conversely, long-term 

exposure to VOCs may be somewhat less of a problem 

than exposure to other nonvolatile organic compounds 

because VOCs generally will not persist in streams for 

long times because of their volatile nature. In the case 

of continuous discharges of VOCs, however, longterm exposures could occur because equilibrium 

concentrations of the VOCs would be established in 

the stream. These equilibrium concentrations would 

be the result of a balance of the processes depicted 

schematically in figure 1.

The concentrations in figure 10 indicate that 

scavenging of VOCs from the atmosphere is not likely 

to be a significant source for streams. However, 

equation 65 is simplistic, and other factors could result 

in both higher and lower concentrations in the stream 

water. For example, rainfall on the impervious 

part of an urban watershed could run off into the 

stream, resulting in higher concentrations than those 

resulting from rain falling directly onto the surface of 

the stream. Flushing of VOCs from the surface also 

could be another possible source of VOCs. However, 

much of this runoff would enter a stream as sheet flow 

over the land surface. Consequently, volatilization to 

the atmosphere could be rapid. Figure 10 indicates 

that the highest concentrations are predicted for the 

shallowest flows. However, volatilization is inversely 

proportional to the flow depth and would be rapid 

for these shallow flows. The relative importance 

of volatilization and scavenging depends on the 

directions and magnitudes of the concentration 

gradients driving each process, and the relative 

importance of these processes has not been studied. 

However, Czuczwa and others (1988) hypothesized 

that the net effect of scavenging on atmospheric

concentrations of alkylbenzenes and tetraohloroethene 

is likely to be minor because of rapid volatilization 

back to the atmosphere. Volatilization from a water 

surface is likely to be slower than volatilization from a 

land surface because of shallower flows over the land 

surface. However, the relative rates of thec e processes 

have not been determined.

MICROBIAL DEGRADATION

Microbes such as bacteria, protozoa, fungi, and 

algae are present in almost all parts of the environment. Microbes use organic compounds as a source of 

food and energy. As a rule of thumb, about one-half of 

the organic carbon assimilated by microbes is used to 

form new biomass, with the other one-half used as an 

energy source for the synthesis process (Gaudy and 

Gaudy, 1980). The energy is obtained by oxidation of 

some of the organic matter that has been rssimilated. 

Consequently, microbial degradation may be defined 

as the metabolism of organic compounds by microorganisms in which the organisms use the compound 

as sources of food and energy to form microbial 

biomass and inorganic and organic end products such 

as carbon dioxide, water, and methane (Larson and 

Cowan, 1995). The term biotransformatkn has been 

used by some researchers, for example, Lanzarone 

and McCarty (1990), who defined this process as 

one in which the structure of the organic compound 

is changed, but the change is something less than 

complete degradation to these simple end products.

The principle of microbial infallibil; ty says 

that microbes can degrade all organic compounds, 

and microbes will eventually develop the ability to 

degrade whatever compounds humans develop 

(Richards and Shieh, 1986). However, the timeframe 

over which this occurs may be long for practical 

purposes, and some compounds may appear 

to be resistant to degradation (Alexander, 1975). 

Consequently, classifications of degradab'lity have 

been developed.

CLASSIFICATIONS OF 

DEGRAD ABILITY

Grady (1985) classified organic compounds as 

biodegradable, persistent, or recalcitrant. Biodegradable, however, is a nonspecific term in that no extent 

of transformation is specified. Transformation may
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

49

not always be beneficial to the environment because 

the transformation can be from a nontoxic compound 

to a toxic compound, or from a readily metabolized 

compound to a persistent or recalcitrant compound. 

A persistent compound is one that does not undergo 

biodegradation under a specific set of conditions, 

whereas a recalcitrant compound is one that is 

inherently resistant to biodegradation (Grady, 1985) 

or one that endures for long time periods in the 

environment because of the inability of the microbes 

to degrade it (Alexander, 1981). However, because of 

the many factors involved in the microbial degradation 

of organic compounds in the environment, recalcitrance may be difficult to prove (Grady, 1985). 

Also, persistence of a compound in the environment 

may result because the compound is actually recalcitrant or because conditions for degradation are 

unsatisfactory.

A somewhat more quantitative measure of 

biodegradability under aerobic conditions is the refractory index (RI), which is defined as the ratio of the 

ultimate biochemical oxygen demand (BODU) to the 

ultimate oxygen demand (UOD) (Hart and Helfgott, 

1975; Helfgott and others, 1977). BODU is the oxygen 

consumed by a heterogeneous community of bacteria 

from domestic sewage during microbial degradation 

of a specific organic compound over an extended 

period of time in the absence of light. UOD is the 

oxygen consumed in the complete wet phase chemical 

oxidation of the compound to carbon dioxide, water, 

and nitrate.

RI values of approximately 1.0 indicate readily 

degradable compounds, with values between 0.3 

and 0.7 indicating partial degradation. RI values 

approaching zero indicate refractory compounds for 

the specific conditions of the measurement, and a 

negative value indicates an inhibitory compound. The 

RI value is negative when the BODU is negative, 

which occurs when the oxygen uptake is greater for 

the bacteria alone than for the mixture of the bacteria 

and the specific organic compound being tested. 

Among VOCs, RIs of 0.2 were determined for 

benzene and gasoline, 0.0 for chloroethene, and 

trichloromethane was inhibitory (Helfgott and others, 

1977).

Biodegradability also has been classified in 

terms of the ratio of the biodegradation half-life 

(BHL) and the retention time (RT) of the organic 

compound in a specific section of the environment

(Larson and Cowan, 1995). Readily biodegradable 

compounds have BHL/RT ratios between 0.1 and 0.3, 

and they do not accumulate or persist in the environment. Practically biodegradable compounds have 

BHL/RT ratios less than or equal to 1.0, and they do 

not accumulate to steady-state concentrations abcve 

the input rate. Once the input is stopped, the concentration decreases rapidly and there is no accumulation 

or persistence in the environment. Slowly biodeg~adable compounds have BLH/RT ratios between 1.0 and 

5.0, and these compounds can persist in the environment after the input is stopped. These compound? 

will not persist indefinitely, however, because finite 

degradation rates exist. Nonbiodegradable compounds 

have BHL/RT ratios greater than 5, and they persist 

more or less indefinitely in the environment (Larson 

and Cowan, 1995).

This discussion of the classification of 

biodegradability indicates that classification is at best 

qualitative. This limitation results from the fact tl at 

microbial degradation of organic compounds in the 

environment is a complex process with many variables 

affecting the process.

CHARACTERISTICS OF THE 

DEGRADATION PROCESS

Biodegradation of organic compounds in tH 

environment is highly system-specific (Grady, 1^5). 

A microbe capable of degrading the compound of 

interest must be present, conditions must exist for 

the necessary enzymes to be synthesized, and oth0* 

system conditions such as nutrient, trace-metal, aid 

oxygen concentrations, pH, and temperature mus 4: all 

be adequate and within acceptable ranges for degradation to occur at a measurable rate.

Microbial degradation of organic compounds is 

catalyzed by enzymes that, in most cases, are highly 

specific for the compound being degraded (Dugan, 

1975). Consequently, a major factor in determining 

the susceptibility of an organic compound to micnbial 

degradation is the time it has been in the environment 

and, therefore, the time available for the necessary 

enzymes to develop (Grady, 1985). Biogenic 

compounds have been in the environment for millions 

of years, and microbes and enzymes generally are 

likely to exist that can effect degradation of these 

naturally-occurring compounds. In comparison, 

many anthropogenic compounds have been in the
50 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

environment for only an instant on the time scale 

of evolution, and microbes and enzymes may 

not exist that can degrade these compounds (Grady, 

1985).

The time required for a microbial population 

to develop the ability to degrade specific organic 

compounds has been called the adaptation time 

(Aelion and others, 1989) or the acclimation time 

(Wiggins and others, 1987). The term adaptation 

will be used in this report. Adaptation assumes 

that the microbial population has the genetic capacity 

for degradation, but the population must adapt to 

the specific organic compound of interest. The time 

required for adaptation is the difference between the 

time when the specific organic compound is first 

introduced into the environment and the time when 

significant degradation of the compound begins. 

The extent of this adaptation time, also called the lag 

period, depends on many factors, including the time 

needed for the enzymes to be induced or to develop. 

Also, there must be sufficient time for small populations of microbes to become large enough to cause 

measurable degradation of the organic compound. 

Other factors that may affect the length of the adaptation time are concentrations of inorganic nutrients, 

preferential use of other organic compounds in the 

system before the use of the compound of interest, 

the time needed for the microbes to adapt to toxins 

or inhibitors present in the system, and predation by 

protozoa on the degrading microbes (Wiggins and 

others, 1987). Lag periods may vary from a few 

hours to days or weeks, depending on the organic 

compound, the microbes, and the system characteristics (Scow, 1990). In the case of recalcitrant 

compounds, adaptation does not occur within the 

timeframe of interest.

Degradation of some organic compounds also 

may occur by the process of cometabolism. This 

process is defined as the degradation of a compound 

that does not provide food or energy for the degrading 

microbes, but the compound is broken down during 

the degradation of other organic compounds in the 

system (Scow, 1990). Cometabolism probably occurs 

because of the lack of specificity of the microbial 

enzymes, and growth of the specific microorganisms 

does not occur (Battersby, 1990). The result is 

that the rate of transformation of a cometabolized 

compound is usually very slow because of the low 

population of organisms, in contrast to more rapid

transformation that occurs with a growing population. 

Compounds with chlorine and nitro substituents 

sometimes are susceptible to cometabolism. An 

example of cometabolism is the aerobic degradation 

of trace concentrations of trichloroethene and 

1,2-dichloroethane with the degradation of methane 

(Lanzarone and McCarty, 1990).

KINETICS OF THE 

DEGRADATION PROCESS

The classifications of biodegradability permit a 

qualitative estimate of the effect of microbial degradation on the concentrations of organic comoounds in 

the environment. These classifications do not, 

however, permit quantitative comparison of the effect 

of microbial degradation with the effects of other 

processes acting on the concentrations of organic 

compounds in the environment, nor do they permit 

mathematical description of the microbial degradation 

process. This can be done, however, by considering 

the kinetics of the degradation process.

The kinetics of the microbial degradation 

process can be described by the Monod eouation, 

which has the form (Herbert and others, 1956)

(66)

where

\JL is the specific growth rate of the microbes

expressed as rate of increase of the

microbe concentration per unit of microbe

concentration,

\jim is the maximum growth-rate cor^tant, 

C is the concentration of the organic

substrate, and

Ks is the saturation constant. 

Analysis of equation 66 indicates that the constant Ks 

is numerically equal to the substrate concentration at 

which the specific growth rate p, is one-hrlf the 

maximum growth rate. Equation 66 assumes that the 

compound being degraded sustains growth and is the 

only source of carbon (Howard and Banerjee, 1984), 

conditions not likely to be true in an environmental 

situation. Also, p. and Ks values apply to a specific 

bacteria, whereas in a mixed population typical of 

the environment, there may be a wide range of p, and 

Ks values.
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

51

Monod (Herbert and others, 1956) also 

indicated that the growth rate of the microbes, dB/dt, 

is a constant fraction, z, of the substrate utilization 

rate, dC/dt. Therefore, it follows that

dB = _ dC 

dt Z dt (67)

where

z is the yield of the reaction, equal to the mass

of microbes formed per mass of substrate

consumed.

Because the growth rate of the microbes is related to 

the specific growth rate by

dB (68)

it follows that

dC 

dt z(Ks

(69)

Equation 69 is the Michaelis-Menton equation, and it 

expresses the rate of change of the substrate concentration as a function of time.

There are, however, several assumptions in 

these equations that may not be completely valid in 

environmental systems with mixed populations of 

microbes and more than one substrate (Scow, 1990). 

First, the constant relation between the growth 

rate of the microbes and the substrate utilization rate 

expressed by equation 67 will most likely not be valid 

for systems containing several substrates because the 

population growth may be partly the result of degradation of substrates other than the one of interest. Also, 

yields may not necessarily be the same for different 

species of microbes. Finally, with the dilute solutions 

common in the environment, a significant amount 

of the energy obtained from the substrate may be 

necessary for maintenance of the population rather 

than growth.

Despite these limitations, equation 69 does 

serve as an empirical basis for describing the 

microbial degradation of organic compounds in 

natural systems. Values of Ks commonly range from 

0.1 to 10 mg/L (Scow, 1990) and consequently are

larger than concentrations of organic compounds 

present in many environmental situations. With this 

consideration, equation 69 reduces to

dC "dt

zK0

= KdBC (70)

where 

Kd is a second-order rate coefficient for

microbial degradation.

Equation 70 indicates that the degradation rate is 

dependent on the concentrations of the microbes 

and the compound being degraded.

Simplification of equation 70 is possible for 

certain situations. If an organic compound is being 

discharged continuously at low concentrations into a 

surface water such that the microbes are adapted to the 

compound, then the biomass concentration will likely 

be relatively constant so that equation 70 become^

dC 

dt

-^ = K'H C (71)

where the degradation process is now first order in 

the substrate concentration with pseudo first-order rate 

coefficient K'^.

Equations 70 and 71 are empirical simplifications of a complex process. Observed changes in 

concentrations as a result of microbial degradation in 

natural systems are likely the sum effects of a number 

of different processes, given the specificity of enzyme 

processes and the number of different organic 

compounds that could be present. However, laboratory 

data for the microbial degradation of three compounds 

indicated good fits to equation 71, and the pseudc 

first-order rate constants were proportional to the 

microbial concentration, as indicated by equation 70 

(Paris and others, 1981).

COMPARISON OF LABORATORY AND 

ENVIRONMENTAL DEGRADATION STUDIES

Various factors affect the microbial degradation 

process, and in the laboratory, many of these factors 

can be controlled to study the fundamentals of the 

process. The coefficient describing microbial degradation is not a fundamental property of an organic 

compound that can be quantified, such as water 

solubility or vapor pressure (Battersby, 1990), but is
52 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

the coefficient describing a highly system-specific 

process (Grady, 1985) and site-specific process within 

a given system (Battersby, 1990). Such specificity is 

the result of differences in various factors such as the 

biomass concentration, population diversity, nutrient 

and trace-metal concentrations, temperature, and pH. 

Other variables such as vitamin and amino acid 

concentrations also may be important in some 

situations (Fewson, 1988). Effects of these variables 

as well as the effect of the organic-compound 

concentration on the degradation process have been 

determined in numerous laboratory studies. These 

studies generally have used high concentrations of 

single substrates in batch cultures of one strain of 

bacteria (Fewson, 1988), and results have been 

valuable in elucidating the fundamentals of the 

microbial degradation process and determining 

degradation products and pathways.

However, because of differences between 

laboratory and environmental conditions, various 

admonitions have been made regarding the application 

of laboratory degradation rate coefficients to field 

situations. Howard and others (1975; 1978) noted that 

the behavior of an organic compound in a laboratory 

test does not precisely predict the behavior of the 

compound in nature, and that laboratory simulations 

of nature are difficult because of varying conditions. 

Scow (1990) stated that using laboratory-derived rate 

coefficients to predict the persistence of an organic 

compound in the environment should be done with 

caution, and Kobayashi and Rittmann (1982) and 

Bedding and others (1983) noted that laboratory 

results cannot always be used to predict biodegradability or biotransformation in natural systems 

or wastewater-treatment processes. Howard and 

Banerjee (1984) concluded that the results of laboratory screening tests should not be applied to natural 

waters containing oligotrophic organisms because 

these organisms are active when organic carbon 

concentrations are low but are inactive when organic 

carbon concentrations are high. They noted that 

the application of screening-test results to environmental situations with easily degradable and resistant 

compounds gave good qualitative comparisons, 

but that application to compounds of intermediate 

degradability resulted in extremely variable comparisons. Smith and Dragun (1984) concluded that the 

high metabolic efficiency of laboratory cultures 

was unlikely in ground-water aquifers. Conversely, 

Battersby (1990) observed that the high concentrations

of organic compounds used in laboratory biodegradability tests can result in underestimation of degradability in all natural waters except the most polluted. 

Similarly, Larson (1983) reported that results from 

biodegradability screening tests did not directly 

correlate with results in environmental tests for all 

compounds tested, but that degradation rates from 

the screening tests tended to underestimate the 

potential for degradation in the environm ent. Also, 

laboratory degradation rates for high compound 

concentrations underestimated the degradation rates 

at more realistic environmental concentrations. 

Boethling and Alexander (1979) cautioned against the 

extrapolation of laboratory coefficients to the environment because environmental concentrations usually 

are much lower than the concentrations used in laboratory experiments, and laboratory degradation rates at 

relatively high concentrations are much 1 igher than 

rates at concentrations likely to be presert in environmental waters. Similarly, Kuhn and others (1985) 

concluded that differences in concentration between 

the laboratory and field situations complicated the 

quantitative comparison of results. In summary, this 

discussion of applying the results of laboratory 

degradation studies to field situations and the differences of opinion expressed in these references indicate 

the difficulty of such applications.

The concentrations in laboratory degradation 

studies usually are higher than the concentrations in 

most environmental situations. The concentration 

of the organic compound can have several effects on 

the degradability in environmental situations. If the 

concentration is too low, then the mechanisms for 

uptake by the cells may be inadequate, the enzymes 

necessary for degradation may not be induced, and 

there may be insufficient energy, even for maintenance of the community population (Grady, 1985; 

Fewson, 1988). The existence of threshoM concentrations below which microbial degradation does not 

occur has been established. Battersby (1990) pointed 

out that natural biodegradable compounds such as 

glucose, glycerol, and lactate have threshold concentrations below which they do not degrade. Typical 

threshold concentrations for many anthropogenic 

compounds are in the range of 0.1 to 1 mg/L 

(Kobayashi and Rittmann, 1982), although the 

threshold concentration may be dependent on 

the compound (Battersby, 1990). Also, E oethling 

and Alexander (1979) noted that the presence of 

other utilizable carbon sources in the system may
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

53

lower the threshold concentration for the compound 

of interest. There also is evidence that organic 

compounds that are usually degradable are persistent 

at low concentrations (Scow, 1990). Conversely, high 

concentrations of the organic compound may be toxic 

or inhibitory to the microbial community (Grady, 

.1985; Fewson, 1988). The optimum concentration 

for microbial degradation is dependent on the characteristics of the organic compound and the microbial 

species, with the possibility of a shift in the order 

and rate of the kinetics as the organic compound is 

depleted during the degradation process (Scow, 1990).

The use of pure cultures in laboratory studies 

also may complicate application of the resultant rate 

coefficients in the environment. Mixed cultures may 

degrade organic compounds that are resistant to pure 

cultures (Smith and Dragun, 1984); consequently, 

lack of degradation in laboratory studies with pure 

cultures is not necessarily a good criterion for 

predicting behavior in the environment (Kobayashi 

and Rittmann, 1982; Grady, 1985). Examples of the 

importance of mixed cultures are cometabolism and 

sequential degradation where initial transformation 

products are degraded by another microbe or series 

of microbes (Bedding and others, 1983).

Other factors contribute to differences between 

laboratory and environmental degradation studies. 

Sediments in the environment are important because 

they affect the concentrations of bacteria and nutrients 

(Boethling and Alexander, 1979). Bacteria attached 

onto sediments often have a higher rate of growth 

and, therefore, a higher rate of degradation of organic 

compounds than free-floating bacteria (Battersby, 

1990). This has been attributed to concentrations 

of nutrients sorbed on the sediments that are higher 

than concentrations in the accompanying water. 

The importance of nutrients was demonstrated 

by Bouwer (1989), who observed that the biotransformation of aromatic compounds in pond 

water increased markedly when acetate, nitrogen, and 

phosphorus were added. Swindol and others (1988) 

observed that the effect of nutrients varied among the 

compounds under study with different samples of 

sediment from the same horizon of an aquifer. They 

found, contrary to expectations, that the addition of 

ammonium sulfate decreased the biotransformation 

of 1,2-dibromoethane and methylbenzene.

Temperature also is important in microbial 

degradation. Growth of microbes has been observed 

from -12°C to 100°C, and organisms are sometimes

classified into three groups according to temperature 

(Scow, 1990). Psychrophiles are active at temperatures 

less than 25°C, mesophiles are active between 25°C 

and 40°C, and thermophiles are active above 40°C. 

Most laboratory studies are done in the temperature 

range from 20°C to 25°C. Consequently, if psychrophilic bacteria are involved and laboratory degradat; on 

coefficients are extrapolated to lower environmental 

temperatures, errors will result because psychrophilic 

bacteria have an optimum growth rate at a temperature 

less than 15°C.

MICROBIAL DEGRADATION OF 

THE TARGET ANALYTES

Microbial degradation of VOCs in surface 

waters has not been studied extensively. Painter 

and King (1985) considered the problem of aerobic 

microbial degradation of VOCs possibly to be nonrelevant, presumably because volatilization rates 

will normally be much higher than microbial degradation rates. Grady (1985) noted that VOCs may appear 

to be recalcitrant because they are lost by volatilization before adaptation can occur and a microbial 

community capable of degrading the compounds 

can develop.

Two types of source conditions for VOCs in 

surface waters are important. The first type is the 

continuous discharge of a wastewater containing 

VOCs into a receiving stream. In this situation, 

microbial degradation of the VOCs may be a significant process because exposure to the VOCs may have 

been long enough that adaptation will have occurred. 

The second type is an instantaneous, short-term 

exposure where the microbial community most likely 

will not have sufficient time to develop the necessary 

enzymes for degradation and, therefore, will not be 

able to adapt to and degrade the VOCs. This type of 

exposure could result from an instantaneous spill, s 

short-term continuous discharge such as might occur 

with a leaking storage tank, stormwater runoff, or 

other types of intermittent nonpoint sources.

There have been some laboratory determinations of the first-order rate coefficients for microbial 

degradation of VOCs. One of the most comprehens: ve 

of these is the work of Tabak and others (1981), 

who determined the degradability of the 96 organic 

compounds on the Environmental Protection Agency 

(EPA) Consent Decree List. Of these compounds, 31 

were target analytes (table 1). The test method was
54 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

a static-culture-flask screening procedure using 

biochemical oxygen demand (BOD) dilution water 

containing 5 mg/L of yeast extract as the synthetic 

medium and concentrations of the test compounds 

of 5 and 10 mg/L. Incubation was at 25°C in the dark 

for 7 days, followed by three weekly subcultures to 

promote adaptation. Settled domestic wastewater was 

used as the microbial inoculum.

The results of Tabak and others (1981) indicated 

that biodegradation of naphthalene was rapid enough 

that it could be a dominant process determining the 

fate of this compound in aquatic systems. Significant 

biodegradation with rapid adaptation was observed for 

benzene and methylbenzene, with chlorobenzene and 

ethylbenzene exhibiting somewhat slower rates of 

adaptation and degradation. A gradual adaptation 

followed by significant biodegradation was 

observed for 1,2-, 1,3-, and 1,4-dichlorobenzene and 

1,3,4-trichlorobenzene. However, subsequent subcultures for these compounds indicated declining 

degradation rates. These declining rates were attributed to possible changes in the original heterogeneous 

microbial population as a result of the subculturing 

process or accumulation of toxic byproducts during 

the metabolism process. The chlorinated aliphatics 

1,1- and 1,2-dichloroethane and 1,1,1 -triehloroethane 

potentially were degradable after a considerable 

adaptation period, whereas the tetrachloroethanes 

might be considered recalcitrant because very 

extensive adaptation periods appeared to be necessary 

for these compounds.

Dichloromethane and tetrachloromethane 

were rapidly degraded, and trichloromethane required 

adaptation. Bromodichloromethane and tribromomethane had significant losses at the end of 

the third subculture period, whereas chlorodibromomethane was recalcitrant for the four 7-day 

incubation periods. Trichloroethene and tetrachloroethene were degraded significantly, with gradual adaptation. The 1,1-dichloroethene compound degraded 

relatively rapidly, with the czs-1,2- and trans-1,2- 

dichloroethenes exhibiting somewhat slower rates 

of degradation. The 1,2-dichloropropane and 1,3- 

dichloropropene compounds were essentially 

biodegradable, with gradual adaptation and 

relatively high degradation rates. Also, hexachlorobutadiene had 100-percent loss at the end of the 

first 7-day incubation period. Both 2-propenal and 

2-propenenitrile showed rapid adaptation and significant biodegradation.

The results presented by Tabak and others 

(1981) were qualitative in nature. Schnoor and others 

(1987) did a quantitative analysis of thes 0; data and 

presented first-order rate coefficients as defined by 

equation 71 for 24 target analytes. Coefficients ranged 

from 0 for 1,1,2-trichloroethane to 0.37 day" 1 for 

1,1,1 -trichloroethane.

These results indicate that most of the VOCs 

were subject to microbial degradation. However, these 

measurements were for idealized conditions in the 

laboratory where extensive steps were taken to 

encourage adaptation to the compounds of interest. 

Also, the concentrations used were 5 and 10 mg/L, 

which are higher than most concentrations likely to be 

measured in the environment.

Other limited data on the microbial degradation of VOCs in surface water exist. Bouwer (1989) 

observed high rates of degradation of aromatic 

VOCs in water from a strip-pit pond formed as a 

result of coal-mining operations, but only after 

amendment with nutrients. Degradation-rate coefficients ranged from 3.6 day for 1,2-dimethylbenzene 

to 6.7 day" 1 for styrene. Without the nuHent 

amendment, significant concentrations of all 

compounds except styrene remained after 43 days 

of incubation. Lee and Ryan (1979) observed degradation of benzene and chlorobenzene in river water, after 

some unspecified adaptation period. Deg^adationrate coefficients were 0.11 day" 1 for benzene and 

0.0045 day" 1 for chlorobenzene. Vaishnav and Babeu 

(1987) determined degradation-rate coefficients for 

benzene and naphthalene in ground wate^, river water, 

and harbor water by using adapted micrcbial seeds 

and media supplemented with nutrients. Benzene 

had degradation-rate coefficients of 0.025 day" 1 

in ground water and 0.044 day" in river water, and 

insufficient concentration changes in the harbor water 

for computation of a rate coefficient. Naphthalene 

showed no significant changes in any of the water 

samples, in contrast to the results of Tab^k and 

others (1981), who concluded that naphthalene 

was readily degradable. Bouwer and others (1981) 

studied the microbial degradation of halogenated 

1- and 2-carbon organic compounds unchr aerobic 

and anaerobic conditions. None of the compounds 

indicated any degradation at initial concentrations 

from 8.8 to 85 (J.g/L after 25 weeks of aerobic 

incubation. Degradation was observed, however, 

in the deoxygenated anaerobic medium in the presence
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

55

of methanogenic bacteria for trichloromethane, 

bromodichloromethane, and chlorodibromomethane, 

but not for trichloroethene or tetrachloroethene.

Anaerobic conditions are not likely to exist in 

the main flow of streams, but may exist in bottom 

sediments. If VOCs exist in the bottom sediments, 

then anaerobic degradation may be a significant fatedetermining process, particularly if the VOCs are 

sorbed to sediments, which will minimize volatilization. Presence of VOCs in bottom sediments has 

been established for a river estuary (Pereira and others, 

1988), a lake (Ferrario and others, 1985), and rivers 

(Gotoh and others, 1992). A study (Van Beelen and 

Van Keulen, 1990) of the anaerobic degradation of 

trichloromethane and benzene in the presence of river 

sediments indicated first-order rate coefficients of 

0.058 day" 1 and 0.27 day" 1 at 10°C and 20°C for 

trichloromethane, but little degradation for benzene.

The lack of rate coefficients for the microbial 

degradation of VOCs in surface water is emphasized 

by the compilation of Howard and others (1991). 

They assembled environmental degradation rate 

coefficients expressed as half-lives for 331 chemicals, 

where the half-life (t0 5md) is the time required for the 

concentration to be reduced by microbial degradation 

to one-half of the initial concentration. Half-lives were 

assembled for a number of biotic and abiotic processes 

including aerobic biodegradation in water and soil and 

anaerobic biodegradation. Half-lives for microbial 

degradation based on grab samples of surface water 

or river die-away studies were preferred to half-lives 

determined from screening tests or field studies under 

controlled conditions. A river die-away test involves 

collecting a sample of river water, adding the organic 

compound of interest, and determining in the laboratory the decrease with time of the concentration of 

the compound (Burns, 1982). Temperature and 

bacterial population densities are maintained at 

ambient levels.

However, data were not available for a substantial number of compounds. In the absence of the 

preferred type of data, screening tests were evaluated, 

and high and low values of the half-lives were 

estimated. When there were few data or conflicting 

data, scientific judgment was used to adjust and/or 

estimate the high and low values of the half-lives. In 

the absence of data for a specific compound, half-lives 

for related compounds were used in some cases.

Of the 331 compounds included in the comp'lation of Howard and others (1991), 45 were VOC 

target analytes (table 1). High and low values of th? 

half-lives for these target analytes for aerobic and 

anaerobic microbial degradation were converted to 

first-order rate coefficients by using

0.693 (72)

0.5md

where

K'd is the pseudo first-order rate coefficient

(day" 1 ) for microbial degradation defined

by equation 71, and 

to.Smd is half-life (days) for microbial degradation

previously defined.

Because of the general lack of the preferred type 

of experimental data and the subsequent scientific 

judgment used in estimating half-lives, the values of 

the half-lives for aerobic microbial degradation fall 

into the general categories of 6 months to 1 year, 

4 weeks to 6 months, and 7 days to 4 weeks. Target 

analytes having these half-lives are presented in 

tables 9, 10, and 11. High and low values of the rate 

coefficients for anaerobic microbial degradation of 

these target analytes are presented in the tables also. 

Half-lives for 13 target analytes did not fall into these 

three general classifications for aerobic degradation, 

and first-order rate coefficients for both aerobic and 

anaerobic degradation for these 13 analytes are 

presented in table 12.

Table 9. Estimated rate coefficients for anaerobic microbial 

degradation of six target analytes with estimated half-lives 

for aerobic microbial degradation between 6 months 

(K'd = 0.0038 day" 1 ) and i year (K'd = 0.0019 day" 1 )

[K'd, pseudo first-order rate coefficient for microbial degradation; LISTS, 

U.S. Geological Survey]

uses

parameter 

code

32102

34511

34488

77652

39180

34475

Compound

tetrachloromethane

1 ,1 ,2-trichloroethane

trichlorofluoromethane

l,l,2-trichloro-l,2,2- 

trifluoroethane

trichloroethene

tetrachloroethene

Anaerobic rate 

coefficient 

(day' 1 )

High

0.099

.00096

.0010

.00096

.0071

.0071

Low

0.025

.000^8

.000 "8

.000"8

.000*2

.000*2
56 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

Table 10. Estimated and experimental rate coefficients for 

anaerobic microbial degradation of 15 target analytes with 

estimated half-lives for aerobic microbial degradation between 

4 weeks (K'd = 0.025 day" 1 ) and 6 months (K'd = 0.0038 day" 1 )

[K'd, pseudo first-order rate coefficient for microbial degradation; USGS, 

U.S. Geological Survey; *, indicates experimental value]

USGS 

parameter 

code

32106

32104

32105

34396

77651

82625

34668

39175

34501

50002

34536

34566

34571

34551

78032

Compound

trichloromethane

tribromomethane

chlorodibromomethane

hexachloroethane

1 ,2-dibromoethane

1 ,2-dibromo-3-chloropropane

dichlorodifluoromethane

chloroethene

1,1-dichloroethene

bromoethene

1 ,2-dichlorobenzene

1 ,3-dichlorobenzene

1 ,4-dichlorobenzene

1 ,2,4-trichlorobenzene

methyl tertiary-butyl ether

Anaerobic rate 

coefficient 

(day' 1 )

High

0.099

.0062

.025

.0062

.35*

.0058

.0062

.0062

.0086

.0058

.0058

.0058

.0062

.0062

.0062

Low

0.025

.000%

.0038

.000%

.046*

.000%

.0010

.000%

.0040

.000%

.000%

.000%

.000%

.000%

.000%

Table 11 . Estimated rate coefficients for anaerobic microbial 

degradation of eight target analytes with estimated half-lives for 

aerobic microbial degradation between 7 days (K'j = 0.099 day" 1 ) 

and 4 weeks (K'd = 0.025 day" 1 )

[K'd, pseudo first-order rate coefficient for microbial degradation; USGS, 

U.S. Geological Survey]

USGS 

parameter 

code

34418

34423

34413

34311

77135

85795

85795

77222

Compound

chloromethane

dichloromethane

bromomethane

chloroethane

1 ,2-dimethylbenzene

1 ,3-dimethylbenzene

1 ,4-dimethylbenzene

1 ,2,4-trimethylbenzene

Anaerobic rate 

coefficient 

(day'1 )

High

0.025

.025

.025

.025

.0038

.025

.025

.025

Low

0.0062

.0062

.0062

.0062

.0019

.0013

.0062

.0062

Analysis of tables 9, 10, 11, and 12 indicates 

that the rate coefficients for anaerobic microbial 

degradation are smaller than the rate coefficients 

for aerobic degradation for all compounds, with the 

exception of several halogenated compounds. High

and low anaerobic coefficients for tetrachloromethane, 

trichloromethane, and 1,2-dibromoethane were larger 

than the corresponding aerobic coefficients. Coefficients for trichloroethene, tetrachloroethene, and 

1,1-dichloroethene were mixed, in that some anaerobic 

coefficients were larger than the corresponding 

aerobic coefficients and some were smaller. Anaerobic 

and aerobic coefficients for chlorodibromomethane 

were the same, within experimental error.

Scientific judgment was required in the estimation of most of the microbial degradation coefficients 

in tables 9, 10, 11, and 12. The only coefficients not 

requiring such judgment were the anaerob; c coefficients for 1,2-dibromoethane and naphthalene and 

the aerobic coefficients for benzene, naphthalene, and 

iso-propylbenzene; these values are marked with an 

asterisk in tables 10 and 12 to differentiate them from 

the estimated values. The small number of experimental values indicates the general lack of suitable 

microbial degradation coefficient data and also that 

the coefficients in tables 9, 10, 11, and 12 and the 

accompanying discussion should be considered very 

qualitative in nature.

All the rate coefficients in tables 9, 10, 11, and 

12 are for an unadapted microbial community. Rate 

coefficients for an adapted microbial community are 

likely to be larger, but it is difficult, if not impossible, 

to estimate how much larger because of th°! great 

number of variables affecting the degradation process. 

Tabak and others (1981), in the work discussed 

previously, used procedures designed specifically to 

develop adapted communities. Rate coefficients from 

their work for aerobic degradation ranged from 0 to 

0.37 day" 1 , and this maximum is within th^ range of 

coefficients in tables 9, 10, 11, and 12.

PREDICTION OF MICROBIAL 

DEGRADATION COEFFICIENTS

Some research has been directed tovard the 

development of structure-activity relationships and 

group contribution calculations that would permit the 

prediction of microbial degradation coefficients in the 

absence of experimental data. Desai and others (1990) 

developed such a model and applied it to f rst-order 

degradation coefficients from the literature for 

11 compounds, one of which was a target analyte 

(n-butylbenzene). The percentage differences between 

the experimental and predicted loge K'd values ranged 

from -3.39 to 27.4 percent, and the percentage
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

57

Table 12. Estimated and experimental rate coefficients for aerobic and anaerobic microbial degradation of 13 target analytes 

[USGS, U.S. Geological Survey; *, indicates experimental value]

USGS 

parameter 

code

34496

32103

34506

34541

34030

77128

34696

34010

34371

77223

34301

34210

34215

Compound

1,1 -dichloroethane

1 ,2-dichloroethane

1 , 1 , 1 -trichloroethane

1 ,2-dichloropropane

benzene

styrene

naphthalene

methylbenzene

ethylbenzene

iso-propylbenzene

chlorobenzene

2-propenal

2-propenenitrile

Aerobic rate 

coefficient 

(day'1 )

High

0.022

.0069

.0050

.0042

.14*

.050

1.39*

.17

.23

.35*

.010

.099

.55

Anaerobic rate 

coefficient 

(day ')

Low

0.0045

.0038

.0025

.00054

.043*

.025

.035*

.032

.069

.087*

.0046

.025

.030

High

0.0054

.0017

.0012

.0010

.0062

.012

.028*

.012

.0039

.087

.0025

.025

.14

Low

0.0011

.00096

.00063

.00013

.00096

.0062

.0027*

.0033

.0030

.022

.0012

.0058

.0075

differences for the experimental and predicted 

K'd values ranged from 10.7 to -44.7 percent. In 

general, these techniques are in the early stages of 

development. Because of the complex nature of the 

microbial degradation process, degradation coefficients predicted by using these procedures should be 

considered as qualitative estimates only, and therefore, 

should be used with caution. Also, much of this work 

has been directed toward the higher molecular weight 

compounds rather than the lower molecular weight 

VOC compounds.

Attempts to develop structure-activity relationships have resulted, however, in certain generalizations regarding the biodegradability of organic 

compounds. Compounds with highly branched 

structures are usually more resistant to degradation 

than the corresponding straight-chain structures 

because branching hinders p-oxidation, which is the 

process by which straight chains usually are degraded 

(Howard and others, 1975). Long-chain molecules 

are degraded more readily than short-chain molecules, 

and unsaturated aliphatic molecules are more readily 

degraded than the corresponding saturated aliphatic 

molecules (Scow, 1990). Substitutions on aromatic 

rings have different effects, with carboxyl and 

hydroxyl groups having a tendency to increase 

degradability and halogens and nitro groups having

a tendency to reduce degradability (Howard and 

others, 1975). Substitutions in the meta position 

of aromatic rings result in increased resistance to 

degradation (Kobayashi and Rittmann, 1982). Ether 

functions are sometimes particularly resistant to 

microbial degradation (Howard and others, 1975). 

Finally, highly oxidized compounds such as halogenated compounds may be resistant to further oxidation 

under aerobic conditions but may be degraded under 

anaerobic conditions (Scow, 1990). This observation 

is consistent with the previous discussion that 

anaerobic degradation coefficients are larger than 

aerobic degradation coefficients for some chlorinated 

aliphatic compounds.

ESTIMATION OF THE IMPORTANCE 

OF MICROBIAL DEGRADATION

Analysis of the rate coefficients for microbial 

degradation presented and discussed previously 

indicates that the largest coefficient for any of the 

target analytes was the value of 1.39 day" 1 for aerobic 

degradation of naphthalene (table 12). This analysis 

excludes from consideration the values of 3.6 day" 

for 1,2-dimethylbenzene and 6.7 day" 1 for styrene 

(Bouwer, 1989) because the degradation character zed 

by these coefficients occurred only after amendment
58 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

of the medium with nutrients. Also, these coefficients 

were for adapted bacteria and, therefore, apply only to 

specific conditions.

The change in concentration of a solute with 

distance due to the effects of multiple first-order 

processes in a stream can be approximated by

C/C0 = exp|~-(L/U)y K:l (73) 

L ,4-, J

where

C0 is the concentration at the source or zero

distance, 

C is the concentration at longitudinal

distance L,

U is the water velocity, and 

Kj is the first-order rate coefficient for process i 

(Falco and Mulkey, 1976; Rathbun and 

others, 1988).

If the active processes are volatilization and microbial 

degradation, equation 73 becomes

0 = exP [-(L/U)(K (74)

where

Kv and K'd are the first-order rate coefficients for 

volatilization and microbial degradation defined previously by equations 1 

and 71, respectively.

Equations 73 and 74 are simplified in that they assume 

a one-dimensional stream system.

Equation 74 was used to compute the change in 

the concentration ratio with distance downstream for 

naphthalene. Computations were done for volatilization alone and for volatilization plus microbial 

degradation by using both the maximum and 

minimum coefficients for microbial degradation from 

table 12. Volatilization coefficients for naphthalene 

were computed using equations 53, 54, and 55 and 

procedures discussed previously. Three sets of 

hydraulic conditions were selected to obtain a large 

range of volatilization coefficients. These conditions 

correspond to the extremes of the three rectangles 

presented in figure 6. For the upper left-hand corner 

of the O'Connor-Dobbins rectangle, a water velocity 

of 0.07 m/s and a flow depth of 10 m were used as 

representative of a deep, low-velocity stream. For 

the lower right-hand corner of the Owens and others

rectangle, a water velocity of 0.56 m/s and a flow 

depth of 0.12 m were used as representative of a 

shallow, medium-velocity stream. For the approximate 

middle of the right-hand edge of the Cmrchill and 

others rectangle, a water velocity of 1.52 m/s and 

a flow depth of 1.5 m were used as representative 

of an intermediate-depth, high-velocity stream. 

Distances were selected so as to have approximately 

10 percent of the initial concentration remaining at the 

downstream end of the stream reach for the maximum 

loss process.

The results are presented in figure 11. 

For the deep, low-velocity stream, microbial 

degradation of naphthalene was significant, with 

even the minimum microbial degradatior coefficient 

contributing substantially to the total los? relative 

to volatilization alone. The contribution of the 

minimum coefficient to the loss of naphthalene in 

the intermediate-depth, high-velocity stream was 

minimal; however, the maximum coefficient indicated 

that microbial degradation was still significant relative 

to volatilization. Finally, results for the shallow, 

medium-velocity stream indicate virtually little contribution from microbial degradation to the total loss 

of naphthalene from such a stream because of large 

losses as a result of volatilization.

The results in figure 11 indicate that microbial 

degradation of naphthalene is most likely to be 

important is streams where losses from volatilization 

are minimal. Volatilization coefficients are inversely 

proportional to the flow depth and directly proportional to the water velocity. Therefore, microbial 

degradation is most likely to be important in deep, 

low-velocity streams, as indicated by the results in the 

O'Connor-Dobbins section of figure 11. Other target 

analytes have rate coefficients for microbial degradation (tables 9—12) smaller than the coefficient for 

naphthalene; therefore, contributions of microbial 

degradation to the loss of these compounds in streams 

is expected to be less than for naphthalen e. Coefficients in tables 9-12 indicate that microbial degradation is most likely to be important for the aromatic 

compounds, 2-propenal, and 2-propenenitrile, and less 

likely to be important for the halogenated alkanes and 

alkenes. These conclusions, however, must be considered highly qualitative because of the very systemspecific and site-specific nature of the microbial 

degradation process. Each situation must be evaluated 

to determine if microbial degradation is important in 

the loss of VOCs from streams.
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

59

1 0

08 _ \

O'CONNOR-DOBBINS (1958)

VELOCITY = 0.07 METER PER SECOND

FLOW DEPTH = 10 METERS

0.6

0.4

0.2

1.0

0.8

0.6

0.4

0.2

C 0

1.0

0.8

0.6

0.4

0.2

\ . —_« ._

VOLATILIZATION

VOLATILIZATION +

MINIMUM DEGRADATION

VOLATILIZATION +

MAXIMUM DEGRADATION

5,000 10,000

CHURCHILL AND OTHERS (1962)

VELOCITY = 1.52 METERS PER SECOND

FLOW DEPTH = 1.5 METERS

VOLATILIZATION

VOLATILIZATION +

MINIMUM DEGRADATION

———_ VOLATILIZATION +

MAXIMUM DEGRADATION

50,000 100,000

OWENS AND OTHERS (1964)

VELOCITY = 0.56 METER PER SECOND

FLOW DEPTH =0.12 METER

VOLATILIZATION

VOLATILIZATION +

MINIMUM DEGRADATION

VOLATILIZATION +

MAXIMUM DEGRADATION

1,000 2,000 3,000 4,000

DISTANCE DOWNSTREAM, L, IN METERS

SORPTION

Streams as depicted schematically in figure 1 

contain sediment. This sediment is present both as 

a suspended phase, which is transported along with 

the water, and as the materials composing the bottom 

and banks of the stream. The suspended and bottom 

sediment phases are not, however, independent of 

each other. There is a continual exchange of sediment 

between these phases, with the direction of net 

exchange dependent on flow conditions. With 

increasing flows, normally there will be degradation 

of the bottom and resuspension of sediments into the 

suspended phase of the stream. With decreasing flows, 

normally there will be aggradation and settling of 

sediments to the bottom. With a constant flow and 

equilibrium conditions, normally there will be a 

continuous exchange of sediment particles between 

the suspended and bottom phases; however, there will 

be no net change in the sediment concentration in the 

water.

Organic compounds such as VOCs sorb to 

sediment particles to varying degrees, depending on 

the characteristics of both the VOC and the sediment. 

Once sorbed, the transport and fate of the VOC in 

a water and sediment system (fig. 1) can be quite 

different from that of a VOC in the water phase alone 

because of the continuous exchange of sediments 

between the water and bottom phases. Also, the 

reactivity of a VOC in the sorbed state may differ 

substantially from the reactivity in aqueous solution, 

both in terms of extent of reaction and reaction 

pathway (Pionke and Chesters, 1973; Karickhoff, 

1984). The transport in the downstream direction of 

the sorbed VOCs will be largely with the suspended 

sediment of the flow. However, there may be some 

downstream movement of the sorbed VOCs with 

the surface bed sediments, depending on the water 

velocity and the shear stresses at the water/streambed 

interface. Under conditions of aggradation, some 

sediment particles with sorbed VOCs may be buried 

for long periods of time, which could threaten human 

and ecosystem health long after the initial source of 

the VOCs has dissipated. Organic compounds buried

Figure 11. Fraction of the concentration of naphthalene remairing 

as a function of distance downstream for volatilization and for 

volatilization and microbial degradation combined at 20 degrees 

Celsius; Henry's law constant is 36.6 pascals cubic meter per 

gram mole.
60 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

with the sediment may appear in the stream water at a 

later time during periods of high flow when sediment 

is being resuspended, or they may gradually desorb 

from the sediment particles and subsequently move 

by molecular diffusion through the pore water into the 

stream water.

FUNDAMENTALS OF THE 

SORPTION PROCESS

Sorption has been defined as any accumulation 

of a dissolved organic compound by solid particles 

(Voice and Weber, 1983). Sorption is generally 

considered to involve the processes of adsorption and 

partitioning. Adsorption is the formation of a chemical 

or physical bond between the organic compound and 

the mineral surface of the solid particle. Partitioning 

involves the "dissolution" of the organic compound 

into the organic matter associated with the solid, much 

like the dissolution of an organic compound into an 

organic solvent.

Adsorption to the mineral surfaces of the solid 

particles can result from physical adsorption due to 

van der Waal's forces or chemisorption resulting from 

electronic interactions between the organic molecule 

and the solid surface (Voice and Weber, 1983). 

The adsorption process is characterized by high heats 

of sorption, nonlinear isotherms, and competition 

between the molecules of the organic compound 

for the adsorption sites on the solid surfaces (Chiou 

and others, 1979; Chiou and others, 1983). The 

partitioning process is characterized by low heats of 

sorption, linear isotherms, noncompetitive sorption, 

and a dependence of the sorption coefficient on the 

water solubility of the organic compound (Chiou and 

others, 1979; Chiou and others, 1983).

Some debate exists in the literature regarding 

the relative importance of the adsorption and 

partitioning processes. Mingelgrin and Gerstl (1983) 

indicated that there may be a continuum of possible 

combinations of these two processes, ranging from 

pure fixed-site adsorption to true partitioning, with the 

contribution of each depending on the characteristics 

of both the organic molecule and the solid. Similarly, 

Voice and Weber (1983) suggested that the actual 

sorption process may involve varying contributions 

from different types of reactions but with one type 

often predominating. Chiou and others (1983) noted 

for nonpolar molecules that solute partitioning 

predominates over adsorption because polar water 

molecules competitively suppress solute adsorption

on polar mineral surfaces. Conversely, adsorption may 

contribute or dominate for highly polar molecules 

or for solids with low organic-matter contents, 

especially when the solids also have a high clay 

content (Karickhoff, 1984).

Studies of the sorption of nonpolar organic 

compounds, including several VOCs, on various soils 

and sediments indicated that the isotherm s were linear 

over wide concentration ranges and that the sorption 

coefficients were dependent on the water solubilities 

of the compounds (Chiou and others, 1979; Chiou 

and others, 1983; Kile and others, 1995). Such 

behavior is consistent with the partitionir^ process 

being dominant for these compounds. Therefore, 

because the target analytes (table 1) are nonionic 

compounds, sorption to solids is expected to be largely 

a partitioning process.

SORPTION COEFFICIENTS

The defining equation for the sorption coefficient is

=Kp Cs (75)

where

S is the mass of organic compound sorbed per

unit mass of solid (g/kg), 

Cs is the concentration of the compound in the

water (g/L), and

Kp is the sorption coefficient (L/kg). 

By convention, concentration units are selected such 

that the mass unit for the solution concentration is 

the same as the mass unit for the sorbed compound 

on the sediment (Karickhoff, 1984). Thus\ typical 

units for Kp are L/kg or rnL/g. Equation 75 indicates 

a linear isotherm with constant partitioning. Values 

of Kp are determined from laboratory measurements 

of the sorbed concentration as a function of the water 

concentration at constant temperature and constant 

solids concentration. Sorption isotherms determined 

in this way usually are linear if the organic-compound 

concentration is less than the smaller of 10 molar or 

about one-half the water solubility (Karickhoff, 1981). 

The Freundlich isotherm also is used to describe 

sorption data. This isotherm has the form

S = Kf Cs

l/n (76)
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

61

where

Kf and n are the constants of the isotherm. 

The Freundlich Kf has been used extensively to 

quantify the extent of sorption and to compare 

different solids, solutes, and experimental conditions 

(Voice and Weber, 1983). Values of n usually range 

from 0.7 to 1.1, although extremes of 0.3 and 1.7 have 

been observed (Lyman, 1990c). This parameter n 

indicates the degree of linearity of the isotherm. When 

n = 1.0, the isotherm is linear, and equation 76 reduces 

to equation 75 and Kf = Kp. Weber and others (1983) 

found that sorption of PCBs on various solids was best 

described by the linear model (eq. 75) over narrow 

ranges of concentrations and by the Freundlich model 

(eq. 76) over larger concentration ranges.

Sorption of nonpolar organic molecules is 

largely partitioning into the organic matter of the 

solids. This organic matter usually is indicated by the 

organic-carbon fraction of the solids. Consequently, 

sorption coefficients commonly are normalized with 

the organic-carbon fraction to give

Koc = ^ (77)

where 

Koc is the normalized sorption coefficient

(L/kg oc), and 

foc is the weight fraction of organic carbon in

the solid.

The weight unit for Koc is on a kg of organic 

carbon (oc) basis rather than a kg of solids basis. 

Such normalization reduced a twentyfold to thirtyfold 

variation in Kp for sorption of phenanthrene and 

pyrene on various soils and sediments to less than 

a twofold variation in Koc values (Karickhoff, 1981). 

Similar behavior was noted for benzene, naphthalene, 

and anthracene. However, in another study (Weber 

and others, 1992), normalization of the Freundlich 

Kf values for three solutes on six soils did not reduce 

the data to a single isotherm. Kile and others (1995) 

determined that Koc values for tetrachloromethane 

and 1,2-dichlorobenzene for large sets of soils and bed 

sediments were relatively constant for each type of 

solid, although the values for the bed sediments were 

about twice those for soils.

Several authors (Mingelgrin and Gerstl, 1983; 

Friesel and others, 1984; Chiou and others, 1985) 

used soil organic-matter fraction rather than soil

organic-carbon fraction. The soil organic-matter 

fraction can be converted to the soil organic-carbor 

fraction by dividing by 1.724 (Kenaga and Goring, 

1980). This procedure is approximate, however, in that 

it assumes the soil or sediment organic-matter consists 

of 58 percent carbon (Smith and others, 1988). If 

precise values are needed, the organic-carbon fraction 

should be measured.

Normalizing the sorption coefficient with the 

fraction of organic carbon removes much, but not all, 

of the variability observed for the sorption of the same 

organic compound on different soils and sediments. 

For example, Garbarini and Lion (1985) determined1 

that the Koc values for methylbenzene on a variety of 

sorbents varied by a factor of 2.5, and the sorbents that 

had the highest and lowest organic-carbon fractions 

had almost the same Koc value. This Koc value also 

was the largest for methylbenzene for all the sorbents 

considered. Sorption coefficients compiled from the 

literature by Mingelgrin and Gerstl (1983) and normalized with the organic-matter fraction also showed 

considerable variability. Factors contributing to this 

remaining variability include inherent differences in 

the sorption characteristics of the organic matter of the 

different soils, differences in experimental procedures, 

and contributions of other soil properties and constituents to the sorption process (Kenaga and Goring, 

1980). Much of this remaining variability is eliminated 

by using a consistent experimental procedure, as was 

demonstrated for a large number of sediments and so'ls 

by Kile and others (1995).

FACTORS AFFECTING 

THE SORPTION PROCESS

In addition to the organic-carbon fraction of 

the solids, the sorption process also may depend on 

other factors such as temperature, pH, size distribution 

and surface area of the solids, salinity, clay mineral 

composition, and cation exchange capacity (Hamaker 

and Thompson, 1972; Karickhoff, 1984; Lyman, 

1990c).

Experimental studies of the effect of temperature are limited and sometimes contradictory. Huang 

(1971) determined that temperature had no significant 

effect on the sorption and desorption of dieldrin 

on a montmorillonite clay. This result is contrary 

to expectations because sorption to a clay should bo 

largely adsorption, which is characterized by a high 

heat of sorption. Chiou and others (1979) determined
62 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

that the sorption of 1,1,1 -triehloroethane on a soil was 

significantly higher at 20°C than at 3.5°C; conversely, 

data for 1,2-dichlorobenzene at these temperatures 

were virtually indistinguishable. This difference in 

behavior for these two compounds was attributed to 

the fact that the solubility of 1,2-dichlorobenzene 

varied little between these temperatures (133 mg/L at 

3.5°C and 148 mg/L at 20°C), whereas the solubility 

of 1,1,1 -trichloroethane decreased from 1,790 mg/L at 

3.5°C to 1,360 mg/L at 20°C (Chiou and others, 1979). 

Weber and others (1983) determined that the sorption 

of a PCB to a river sediment at 12°C was significantly 

less than at 22°C. Oliver (1985) determined the 

temperature dependence of the desorption of 

20 chlorinated nonionic organic compounds from 

three sediments from Lake Ontario. Desorption coefficients for 18 of the 20 compounds decreased as the 

temperature increased from 10°C to 30°C, with 

percentage decreases ranging from -3.0 to -63 percent 

and averaging -43 percent. The increases observed for 

1,2-dichlorobenzene and one of the PCB compounds 

were apparently anomalous.

A review of the effect of temperature on the 

sorption process (ten Hulscher and Cornelissen, 1996) 

also indicated various and sometimes contradictory 

results. The extent of sorption decreased as the 

temperature increased for 1,4-dichlorobenzene and 

1,2,4-trichlorobenzene, whereas three measurements 

for naphthalene indicated a decrease in sorption with 

increasing temperature for two of the isotherms and an 

increase for the third. These sometimes contradictory 

results probably occur because the overall process of 

sorption consists of a combination of adsorption to the 

mineral surface and partitioning into the organic 

matter of the solid, with a continuum of combinations of these two processes possible depending on 

the characteristics of the compound and the solid 

(Mingelgrin and Gerstl, 1983). The adsorption process 

has a high heat of sorption, the partitioning process has 

a low heat of sorption (Chiou and others, 1979), thus 

giving a range of possible temperature effects on the 

sorption process.

Sorption of nonionic organic compounds is 

affected little in the pH range of 5 to 9 considered 

normal for surface waters (Lyman, 1990c). Sorption 

and desorption of dieldrin on a montmorillonite 

clay (Huang, 1971) at pHs of 6.0, 8.0, and 10.0 

indicated that pH had only a slight effect. Hamaker 

and Thompson (1972) indicated that sorption of 

neutral molecules would be expected to show virtually

no effect as a result of pH changes unless the pH 

change resulted in significant change of rome 

important property of the soil. They also noted 

that polar molecules, while not actually forming 

ionic species, could form hydrogen bond'' and, 

therefore, pH could have some effect for these 

molecules.

The size distribution and surface area of the 

solids could affect the sorption coefficient. Richardson 

and Epstein (1971) observed that sorption of three 

insecticides on various size classes of particles from 

two soils was greatest for the <0.08-^im colloidal 

fraction and the 0.08-0.5-^m fraction. Karickhoff and 

others (1979) reported that sorption coefficients were 

very small for the sand sizes, largest for the medium 

and fine silt sizes, and somewhat smaller for the clay 

sizes. Karickhoff (1981) concluded that rorption 

occurred mostly on the fine material and that the sand 

acted primarily as a diluent. The dependence of the 

sorption coefficient on particle size can be determined 

if the organic-carbon fraction of each size fraction is 

known. However, such information normally is not 

available.

Voice and Weber (1983) cited three references 

suggesting an inverse relationship between particle 

size and sorption; however, the correlations were 

generally poor. Weber and others (1983) observed 

increased sorption capacity for sediment particles 

smaller than 75 jam, which was attributed to the 

greater organic-carbon fraction of the smaller 

particles. They concluded that the observed effect 

of particle size on sorption capacity of natural 

sediments probably was the result of the nonhomogeneous physicochemical nature of the sediment 

rather than the particle size itself. Weber and others 

(1983) also observed that surface area had some effect 

because the amount of a PCB compound sorbed on a 

pure montmorillonite clay was greater than on a pure 

kaolinite clay having a substantially smaller specific 

surface area. However, amounts sorbed on these pure 

clays were much smaller than the amount sorbed 

on a natural clay, presumably because of the 

partitioning of the PCB compound into the organic 

matter of the natural clay. This is an experimental 

example of the dominance of the partitioning process 

over the adsorption process for sorption on nonionic 

organic molecules by natural soils and sediments.

Salinity or ionic strength can have a significant 

effect on the sorption of ionic molecules but have 

little effect on the sorption of nonionic o- nonpolar
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

63

molecules. In a situation where adsorption was likely 

dominant, Huang (1971) determined little effect on the 

sorption of dieldrin on montmorillonite at concentrations of sodium chloride of 0.03, 0.3, and 3.0 percent 

by weight. In a situation where partitioning was likely 

to dominate, Karickhoff and others (1979) measured 

about a 15-percent increase in the sorption coefficient 

of pyrene on a pond sediment with an organic-carbon 

fraction of 0.0234 when the salt content was increased 

from zero to 20 mg/mL. Similarly, Garbarini and Lion 

(1985) determined no significant effect of 0.1 molar 

sodium chloride on the sorption of trichloroethene by 

a commercial humic acid compared with sorption in 

distilled water.

SOLIDS CONCENTRATION EFFECT

One of the important and controversial issues 

in the sorption of nonpolar compounds by solids is 

whether the sorption coefficient depends on the solids 

concentration. O'Connor and Connolly (1980) 

determined that the sorption coefficient for four 

pesticides decreased with increasing solids concentration. DDT showed the greatest effect, and this 

compound also had the greatest tendency to partition 

to the solids. Similarly, Voice and others (1983) 

determined that the sorption coefficients of chlorobenzene, naphthalene, and two PCBs on sediments from 

Lake Michigan decreased with increasing solids 

concentration. They observed about an order of 

magnitude increase in the partition coefficient for 

every two orders of magnitude decrease in the solids 

concentration. They concluded that this phenomenon 

was the result of the presence of microparticles from 

the sediments that were in solution and not removed 

by the conventional separation procedures such as 

centrifugation and filtration used to separate the solids 

from the water.

Other investigations also have attributed the 

observed concentration dependence to experimental 

artifacts. Karickhoff (1984) suggested the effect is 

the result of failure to achieve sorption equilibrium, 

which depends significantly on the solids concentration; failure to correct the aqueous phase sorbed 

concentration for the concentration associated with 

the colloid phase; and failure to account for sorption 

to other surfaces, such as the container walls, at 

low solids concentrations. Gschwend and Wu (1985) 

demonstrated that sorption coefficients for seven 

PCBs were constant over a wide range of solids

concentrations if precautions were taken to account 

for sorption to nonsettling or colloidal particles. 

Standard laboratory procedures such as filtration or 

centrifugation do not remove such particles, and 

the compound sorbed to these particles would be 

incorrectly reported as dissolved rather than as sorbed. 

A three-phase model including these nonsettling or 

nonfilterable particles was applied to PCBs sorbed 

on sediments from Lake Superior, and the results 

indicated that the colloid-associated PCBs may be the 

dominant species in surface waters (Baker and others, 

1986).

Other studies also have indicated that the 

observed dependence of the sorption coefficient on 

the solids concentration is an experimental artifact. 

McCarthy and others (1986) argued that even 

relatively small amounts of impurities in the radioactive tracers used for sorption studies could result in 

underestimates of sorption coefficients, the appearance of irreversible sorption, and sorption coefficients 

that appeared to decrease as the solids concentration 

increased. Jepson and others (1995) observed that the 

sorption coefficient for hexachlorobenzene on natural 

sediments decreased as the solids concentration 

increased. This observation was attributed to two 

factors. For experimental times longer than 10 to 

20 days, significant amounts of hexachlorobenzene 

sorbed to the colloidal matter initially present in the 

tap water used for the experiments. For experimental 

times shorter than 10 to 20 days, the slow dissolution 

of the hexachlorobenzene in the water resulted in 

the persistence of undissolved hexachlorobenzene. 

Correction for these two factors resulted in sorption 

coefficients that were independent of the solids 

concentration. Thus, the presence of organic material 

of a molecular size such that the material is in the 

phase normally considered "dissolved" can significantly affect the results of some sorption experiments, 

with the effect dependent on the concentration and 

characteristics of the compound being sorbed.

Sorption of nonpolar organic compounds to 

colloidal or dissolved organic material can decrease 

sorption by solids (Hassett and Anderson, 1982), 

enhance the apparent water solubility (Chiou and 

others, 1986), or decrease the sorption coefficient as 

the solids concentration increases (Voice and others, 

1983). The importance of these processes depends on 

the characteristics of both the organic compound and 

the organic matter. For example, Chiou and others 

(1986) indicated that solubility enhancement by
64 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

moderate concentrations of dissolved organic matter is 

most likely to be important for highly water-insoluble 

compounds. For a more soluble compound such as 

1,2,3-trichlorobenzene with a water solubility of 

18.0 mg/L at 25°C, solubility enhancement will be 

small for dissolved organic-matter concentrations less 

than 100 mg/L (Chiou and others, (1986).

KINETICS OF THE SORPTION PROCESS

Equations 75 and 76 are equilibrium expressions; that is, Kp and Kf relate the sorbed concentrations to the water concentrations at equilibrium. Many 

investigators (for example, Schnoor and others, 1982; 

DiToro and others, 1982; Thomann and DiToro, 

1983; Burns, 1983; Corwin, 1986; and Griffioen and 

van de Meent, 1988) assumed that this linear isotherm 

(eq. 75) applies in environmental systems. This 

assumption requires that equilibrium be rapid and 

reversible so that equilibrium conditions can be 

maintained on a time scale that is short relative to 

the water velocity of the stream.

Experimental evidence, however, indicates that 

the attainment of equilibrium may be slow. Hamaker 

and Thompson (1972) discussed several studies 

showing a slow approach to equilibrium. They also 

noted that the sorption process could contain a very 

slow step such as intraparticle diffusion, which is 

diffusion within the pores of the individual solid 

particles. In this case, an apparent steady-state 

condition would be reached very quickly, which 

would be mistaken for equilibrium. Thus, with 

short equilibration times, a slow drift in this 

apparent equilibrium concentration would be missed. 

Equilibration times in many experimental studies 

were relatively short because of complications that 

become more important as time increases. Such 

complications include microbial degradation, hydrolysis, or oxidation of the test compound. Kenaga and 

Goring (1980) noted that there appeared to be an 

initial rapid sorption, followed by slow sorption over a 

long time period. Karickhoff (1980) determined that 

the sorption of three polycyclic aromatic hydrocarbons 

could be divided into two stages, with the first stage 

completed in minutes and the second stage requiring 

days to weeks for complete equilibration, with 

the time dependent on the characteristics of the 

compound. Approximately one-half of the total 

sorption was completed in the first stage, with this 

result seemingly independent of the compound or the 

sediment. Desorption also showed two stages, with an

initial fast desorption followed by much slower 

desorption. Desorption coefficients were consistently 

smaller than the sorption coefficients; hovever, the 

difference was at most a factor of two anc\ therefore, 

was not considered significant for these t}pes of data. 

Pignatello and Xing (1996) summarized several 

examples in which the apparent sorption coefficient 

increased from 30 percent to tenfold when contact was 

increased from short times ranging from 1 to 7 days to 

long times ranging from 10 to 300 days.

The use of equation 75 and the assumption of 

equilibrium also assumes sorption is reversible; that is, 

desorption is described by the sorption isc therm. Such 

reversibility is usually assumed; however, there are 

experimental studies showing distinct differences 

between sorption and desorption isotherms (Horzempa 

and DiToro, 1983). Rao and Davidson (19P2) rationalized that these nonsingular isotherms could be the 

result of experimental artifacts, failure to achieve 

complete equilibrium during the sorption process, 

and microbiological or chemical changes during the 

experiments. However, consideration of these factors 

(Horzempa and DiToro, 1983) did not explain the 

nonsingular isotherms observed for the sorption 

and desorption of a PCB compound on sediments. 

Desorption coefficients were substantially larger than 

sorption coefficients, and desorption appeared to 

involve two distinct isotherms, with the slower process 

controlled by strongly bound molecules. Pignatello 

and Xing (1996) noted that data showed greater and 

greater resistance to desorption with time, indicating a 

continuum of resistances. They attributed this to the 

intrinsic heterogeneity of solids as indicated by 

polydisperse primary and secondary particles, a 

wide range of pore sizes, and spatial variations of 

mineral and organic components on the rricroscale. 

Also, they noted that the slow desorption step often 

was markedly dependent on the initial concentration, 

resulting from increased intraparticle retardation as the 

pore concentration decreased.

The observation that some of the molecules 

appeared to be strongly bound is consistent with 

the observation (Karickhoff, 1984) that lengthy 

extraction periods of days, multiple solvents, and/or 

abrasive mixing techniques are necessary for quantitative chemical recovery in some studies. Ir other 

studies (Rutherford and others, 1992), a two-solvent 

system with an extraction time of 24—48 1 ours was 

sufficient to obtain greater than 95-percen4: recovery in 

most cases. Karickhoff (1984) noted that the ease of
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

65

extraction of a sorbed compound decreased as the 

equilibration time of the experiment increased. 

The importance of equilibration time on both the 

magnitude of the measured sorption coefficient 

and the recovery of the compound during desorption 

is demonstrated by data for sorption of hexachlorobenzene (Karickhoff, 1984). For equilibration times 

of 5, 32, and 58 days, the measured sorption coefficients normalized for the organic-carbon fraction 

of the sediment (Koc) were 7.6x104, 1.5xl05, and 

1.9xl05 L/kg oc, respectively. Thus, the measured 

sorption coefficient increased as the equilibration time 

increased. When the desorption study was done using 

a purge technique, about 90 percent of the hexachlorobenzene was recovered from the 5-day equilibration 

time sample with 2 days of purging. Conversely, for 

the 32- and 58-day equilibration time samples, 20 days 

of purging was necessary to recover about 90 and 

85 percent, respectively, of the hexachlorobenzene. 

These results clearly demonstrate that the sorptiondesorption process is at some point controlled by 

diffusion of the organic molecules into or out of the 

particle aggregates and the associated organic-carbon 

matter.

The importance of equilibration time and the 

need for long times in laboratory sorption studies 

are indicated by two studies. Jepson and others 

(1995) studied the sorption of hexachlorobenzene 

to sediments from the Detroit River. They concluded 

that a steady-state equilibrium was always attained. 

However, equilibration times of from 2 to 5 months 

were necessary. Brannon and others (1995) measured 

the sorption of two PCB compounds and one polycyclic aromatic hydrocarbon compound to three 

sediments over a 6-month period. They observed that 

the value of the sorption coefficient normalized with 

the organic-carbon fraction increased slowly with time 

throughout the equilibration period. They attributed 

this slow increase to intraparticle diffusion of the 

organic compounds.

Kinetic models of the sorption-desorption 

process have been developed (Karickhoff and Morris, 

1985; DiToro, 1985). On the basis of previous 

observations in the literature, Karickhoff and Morris 

(1985) developed a two-component model that considered a rapid or labile sorption requiring a few hours 

to complete and highly retarded or nonlabile sorption 

requiring days to weeks to complete. Application 

of this model to the sorption of pyrene, pentachlorobenzene, and hexachlorobenzene to various

sediments and soils indicated that the fraction of the 

sorption that was labile ranged from 0.25 to 0.6. 

However, for highly hydrophobic compounds and 

high solids concentrations, the labile fraction was 0.1 

or less. The more highly sorbed compounds sorbed at 

the slowest rates. Desorption studies using a purg? 

technique indicated no irreversible bonding; however, 

in some cases, recovery of the final 10 percent of the 

compound required more purge time than the time 

used for the first 90 percent. DiToro (1985) developed 

a model that assumes the sorbed compound consists 

of a reversible component and a component that 

does not appreciably desorb under the conditions 

of the experiment. The key assumption of the model 

is that the linear isotherm applies both to the sorption 

and desorption of the reversible component. The 

resultant model contained three parameters that were 

determined from a data set of 128 partition coefficients by using a nonlinear least-squares procedure. 

The agreement between calculated and experimental 

partition coefficients was good. This analysis 

indicated that the reversible sorption coefficient 

normalized with the organic-carbon fraction was equal 

to the octanol-water partition coefficient, indicating 

that partitioning into the organic carbon was 

physically the same as partitioning into octanol.

The subject of reversibility was considered 

also by Gschwend and Wu (1985). They conducted 

sorption-desorption experiments by using a PCB 

compound and prewashed sediments from Lake 

Superior. Sorption and desorption isotherms were 

identical. The purpose of the prewashing was to 

remove nonsettling colloidal material. They also 

demonstrated through calculations that the nonsingular 

isotherms observed by DiToro and Horzempa (1982) 

could be explained by the presence of nonsettling 

microparticles or colloidal material. These particles 

and/or colloidal material served as sorbing media but 

were treated in the analytical technique as part of the 

dissolved phase.

SOLIDS, SOILS, AND SEDIMENTS

Most of the previous discussion referred to the 

sorbing media in sorption studies as the "solids" 

phase. "Solids" is a generic term used to define any 

solid particle to which organic molecules can sorb or 

adsorb. This particle may or may not contain organic 

matter. The terms "soils" and "sediments" also have 

been used in a few specific cases in the previous text. 

Soils and sediments are not the same, however.
66 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

Soils are units formed at the Earth's surface and 

have distinctive zones characterized by differences in 

mineralogy, organic content, texture, and structure 

(Leopold and others, 1964). The two fractions having 

the most effect on sorption are the mineral fraction and 

the organic fraction. The mineral fraction consists 

primarily of layer silicates (clays) and metal hydroxides (Tinsley, 1979). The organic fraction is tightly 

bound to the mineral fraction, and is generally considered to consist of humin, humic acids, and fulvic 

acids (Tinsley, 1979). Humin is operationally defined 

as the material that cannot be extracted by alkaline 

reagents; humic acids are the fraction extracted by 

the alkaline reagents that precipitate upon acidification; fulvic acids are the fraction remaining in 

solution.

Karickhoff and others (1979) defined a 

sediment as an eroded soil that has been subjected 

to continuous redispersion and particle-size fractionation. These processes begin with runoff from the 

soil surface and continue with subsequent transport 

in the stream system. Transport within the system in 

controlled primarily by grain size, which can range 

from gravel and boulders to colloids.

One classification of sediment according to 

grain size is given in table 13 (Chow, 1964). Particles 

generally have different organic-carbon fractions and 

different surface areas for the different size classes. 

Consequently, their capacity for sorption of organic 

compounds varies from size class to size class, mostly 

because the organic-matter fraction varies approximately inversely with the particle size. Sediments 

from different drainage basins are likely to have 

different distributions of these types of particles within 

the sediment and, consequently, different sorption 

characteristics. Normalization with the organic-carbon 

fraction of the sediment removes much, but not all, 

of this variation. Karickhoff and others (1979) 

cited an example of the effect of grain size where 

two hydrophobic compounds concentrated in the clay 

fraction and a more soluble compound concentrated 

on the larger particles.

The organic matter associated with soils and 

sediments also undergoes changes with time. These 

changes, referred to as diagenetic alteration (Young 

and Weber, 1995), involve conversion by biotic 

and abiotic processes of biopolymers such as carbohydrates, proteins, and lipids to humic substances such 

as fulvic acids, humic acids, and humin. Over longer 

time periods and under more extreme conditions, these

humic substances are converted to materials such as 

kerogens, coals, or graphite. This diagenetic alteration 

usually results in removal of oxygen-containing 

functional groups, increases in molecular weight, and 

increases in the fraction of carbon present in condensed 

aromatic nuclei compared with the fraction present in 

the aliphatic linkages (Young and Weber, 1995). In 

qualitative terms, Young and Weber (1995) modeled 

soil organic matter as a macromolecule whh a structure 

ranging from one that is completely amorphous (fulvic 

acids) in a relatively young soil to one that is increasingly condensed as diagenetic alteration occurs, 

eventually becoming highly crystalline in an old 

organic-carbon structure such as anthracite coal.

The fact that soils and sediments ard the associated organic matter have different properties results in 

different sorption characteristics. Kile and others 

(1995) determined the sorption of tetrachloromethane 

and 1,2-dichlorobenzene on a number of soils and 

sediments from various parts of the United States and 

China. The mean sorption coefficients normalized 

with the organic-carbon fraction were 1.7 times larger 

for the sediments than for the soils. This difference 

indicates that the erosion process that converts soils 

into sediments results in a change in the characteristics 

of the organic matter. Kile and others (19°5) hypothesized that the erosion process and subsequent water 

transport result in dissolution of the more polar and 

more water-soluble constituents of the organic matter, 

leaving the less polar organic constituents associated 

with the sediments. Analysis of 17 soil and 7 sediment 

samples by 13C nuclear magnetic resonance (NMR) 

spectroscopy indicated median polar organic-carbon 

contents of 51.6 and 42.2 percent for the soils and 

sediments, respectively, supporting this hypothesis 

(Kile and others, 1996).

Table 13. Classification of sediments according to grain size 

(from Chow, 1964)

[mm, millimeter; ^m, micrometer (10~6 meter); >, greater than; 

<, less than]

Class Grair size

Gravel, boulders

Sand

Silt

Clay

Colloid

>2 mm

<2 mm, >64 jim 

<64 urn, >5 ^m 

<5 urn, >0.5 jum 

<0.5 im
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

67

The dividing line between soils and sediments is 

not always definitive, however, because the exact time 

of erosion from the land surface is unknown. Kile and 

others (1995) determined a sorption coefficient for one 

sediment sample from a creek that was significantly 

lower than the values for other sediments but was 

similar to the coefficients for the soils. They attributed 

this result to the fact that this sediment may have been 

a recently eroded soil. Values for two other sediments 

were somewhat lower than expected, possibly 

indicating that the conversion from a soil to a sediment 

was incomplete. Sorption coefficients for sediments 

from large rivers and lakes were comparable, presumably because the sediments were aged and had little 

contribution from recently eroded soils.

Kile and others (1995) limited their soils and 

sediments to those with organic-carbon contents 

greater than 0.1 percent, and they used relatively 

nonpolar solutes to emphasize the partitioning part of 

the adsorption process. For solutes with higher polarities and sorbents with lower organic-carbon contents, 

adsorption to the mineral surfaces may be important. 

Such adsorption will result in higher overall sorption 

coefficients.

Karickhoff (1984) presented some general 

guidelines for estimating the contribution of mineral 

adsorption to total sorption. These are listed in 

table 14. The target analytes (table 1) are largely 

nonpolar, and all have less than 10 carbon atoms. 

Therefore, the results in table 14 indicate that the 

sorption coefficient normalized with the organiccarbon fraction, K^, will be from 50 to 100 times 

the coefficient for sorption to the mineral surfaces, 

Km. Also, the threshold ratio indicates that mineral 

adsorption will become important only for soils 

and sediments with a mineral fraction greater than 

60 times the organic-carbon fraction. These results 

indicate that partitioning of the target analytes into the 

organic matter of sediments is likely to be much more 

important than adsorption to the mineral surfaces for 

most environmental situations.

Other natural sorbents that may contribute 

to the behavior of VOCs in a stream system (fig. 1) 

are colloids, microorganisms, and algae. Colloids 

are of such a size (table 13) that they will not be 

separated from the water phase by common filtration 

and centrirugation techniques. Consequently, colloids 

and any organic compounds sorbed to the colloids will 

be treated as dissolved by normal laboratory separation techniques. This effect has been proposed as an

explanation for both the solids concentration 

dependence of the sorption coefficient and nonsingularities in the sorption-desorption isotherms. 

Also, Hassett and Anderson (1982) noted decreased 

sorption of two hydrophobic organic compounds from 

concentrated solutions of dissolved organic matter 

when compared with sorption from distilled wate- or 

an unconcentrated solution of the organic matter. This 

effect was attributed to sorption of the compounds by 

the dissolved organic matter. Thus, sorption to the 

colloid phase can be a very important process in 

understanding the sorption of organic compounds 

in a stream system. Sorption to microorganisms and 

algae is discussed in the next section.

Table 14. Threshold values of the ratio of the mineral fraction to 

the organic-carbon fraction at which mineral adsorption becomes 

important for three types of organic compounds (from KaricVhoff, 

1984)

[Koc , sorption coefficient normalized with the organic-carbon fraction 

by weight; Km, mineral sorption coefficient; cm, fraction by weight 

of mineral in the sediment; oc, fraction by weight of organic carbon 

in the sediment; <, less than; >, greater than; CIQ denotes molecule 

with 10 carbon atoms]

Type of 

compound

Neutral organics with polar 

functional groups

Small nonpolar organics

Large nonpolar organics

Koc 

Km

10-50

50-100

>100

Threshold

cm 

oc

25-60

>60

insignificant at 

any value

SORPTION TO OTHER SORBENTS

The widespread occurrence of VOCs has 

resulted in the need for information on the sorpticn 

of VOCs to surfaces in addition to the soils and 

sediments of the environment. Dobbs and others 

(1989) measured the sorption of six VOCs and 

seven high-molecular-weight compounds on 

primary, mixed-liquor, and digested solids from 

municipal wastewater-treatment plants. These solMs 

had organic-matter contents ranging from 65 to 

85 percent, which are much higher that the organicmatter contents of natural soils and sediments. When 

the sorption coefficients were normalized with the 

organic-matter content, the coefficients were equivalent for all the solids.
68 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

Steen and Karickhoff (1981) studied the 

sorption of two polycyclic aromatic hydrocarbon 

compounds to bacterial populations isolated from 

14 soils and sediments. The results indicated that 

sorption coefficients normalized with the bacterial 

mass were nearly independent of the source of the 

bacteria. When the sorption coefficients were normalized with the carbon content of the bacteria, the 

resultant sorption coefficients approximated the values 

of the coefficients for sorption of the two compounds 

to soils and sediments.

A study (Whitman and others, 1995) of the 

sorption of naphthalene in a system containing both 

soil and bacteria indicated that sorption to the bacteria 

was negligible when the soil had an organic-carbon 

content of 5.1 percent. However, when a clean sand 

containing no organic carbon was used, as much as 

10 percent of the initial naphthalene was sorbed to the 

bacteria.

Weber and others (1983) measured the sorption 

of a PCB compound to samples of live and dead 

bacteria and algae. The sorption coefficient for the live 

bacteria was about 10 percent larger than the coefficient for the dead bacteria. The sorption coefficient for 

the live algae was more than twice the coefficient for 

dead algae. This result indicates movement of the 

molecules into the viable cells of the algae, allowing 

for more surface sorption. The sorption coefficient for 

the live algae was the largest measured for any of the 

sorbents and was about 11 times larger than the largest 

measured coefficient for natural stream sediments and 

suspended solids.

VOCs in snow also may involve sorption 

processes. Czuczwa and others (1988) noted that snow 

contained the highest concentrations of alkylbenzenes 

in all types of precipitation. They also determined that 

concentrations of all the compounds considered in the 

study were the highest in winter precipitation, possibly 

indicating the contribution of a sorption process. This 

possibility has resulted in fundamental studies of the 

processes involved in the sorption of VOCs by ice 

and snow (Goss, 1993; Hoff and others, 1995). 

Preliminary results indicated that sorption of nonpolar organic molecules to ice surfaces is similar 

in magnitude to that for a water surface, although 

some data indicated ice sorption may be as much 

as a factor of two greater than for sorption to water 

surfaces.

A combined soil-sorption and snow-sorption 

event was observed in the central Canadian Arctic 

when brown snow occurred (Welch and others, 1991). 

Various organic compounds were detected in the 

brown snow, melted snow, and sorbed to the soil 

particles in the snow. Most of these were highmolecular-weight trichlorobenzenes, pesticides, 

and polycyclic aromatic hydrocarbon corrpounds. 

Naphthalene, one of the target analytes (tfble 1), 

was one of the dominant polycyclic aromatic 

hydrocarbons. Air-mass trajectories, clay-mineral 

composition of the soil particles, and visible organic 

detritus such as straw indicated an Asian rource of 

these materials. The source of the organic compounds 

could not be estimated with the same certa; nty because 

of unknown sources and source strengths and also 

possible atmospheric transformations.

EXPERIMENTAL SORPTIOI T 

COEFFICIENTS

Much of the experimental work on aqueous 

sorption to sediments and soils has been directed 

toward pesticides and PCB compounds because 

these compounds have low water solubilities and, 

therefore, are most likely to be involved in sorption 

processes. Sorption coefficients have been measured, 

however, for a number of VOCs on the target analyte 

list (table 1). These experimental sorption coefficients 

are presented in table 15 as K^ values, that is, sorption 

coefficients normalized with the organic-carbon 

fraction, along with information on the type of 

sorbents used in the studies and the organ5 c-carbon 

fractions of these sorbents. Average coefficients 

for each VOC were not computed because of 

expected differences in the sorption characteristics 

of the various sorbents. Although normalization of 

the sorption coefficient with the organic-carbon 

fraction removes much of the variation among 

sorbents, normalization does not remove all the 

variation.

Aqueous sorption coefficients in table 15 

ranged from 0.05 L/kg oc for methylbenzene on 

cellulose (Garbarini and Lion, 1986) to 14,800 L/kg oc 

for 1,2,3-trichlorobenzene on plant cuticle (Boyd 

and others, 1990). Cellulose was chosen for study 

because of its abundance in plants and its use in many 

industrial applications (Rutherford and otvers, 1992).
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

69

Table 15. Experimental sorption coefficients for 27 target analytes

[USGS, U.S. Geological Survey; foc, fraction of organic carbon by weight; Koc, sorption coefficient normalized with the fraction by weight of organic 

carbon; L/kg oc, liter per kilogram of organic carbon; ns, not specified; mg/L, milligram per liter; <, less than; >, greater than]

USGS 

parameter 

code

34423

32106

32102

32103

34506

34511

77651

34541

82625

34501

39180

34475

34704

34699

Compound

dichloromethane

trichloromethane

tetrachloromethane

1 ,2-dichloroethane

1,1,1 -trichloroethane

1 , 1 ,2-trichloroethane

1,2-dibromoethane

1 ,2-dichloropropane

l,2-dibromo-3- 

chloropropane

1 , 1-dichloroethene

trichloroethene

tetrachloroethene

cis- 1 ,3-dichloropropene

trans- 1 ,3-dichloropropene

Sorbent

3 sewage solids

3 sewage solids

3 sewage solids

Extracted peat

Peat

Muck

Cellulose

32 soils

36 sediments

Soil

Soil

32 soils

3 soils

Soils

Soils

Soil

Soils

Soils

3 soils

Soil

Soil

Soils

3 sewage solids

32 soils

3 soils

7 organic compounds

5 soil fractions

Peat

Soil

Aquifer material

32 soils

3 soils

3 sewage solids

Soils

Soils

oc

0.38-0.49

0.38-0.49

0.38-0.49

0.640

0.571

0.531

0.444

0.0016-0.0609

0.001 1-0.0473

0.0093

0.0093

0.0012-0.31

0.002 

0.022 

0.037

<0.087

>0.087

0.0093

ns

0.00290-0.126

0.011 

0.0161 

0.0165

0.0093

0.0093

ns

0.38-0.49

0.0012-0.31

0.002 

0.022 

0.037

0.12-0.65

0.0014-0.833

0.571

0.0093

0.0015

0.0012-0.31

0.002 

0.022 

0.037

0.38-0.49

<0.087

>0.087

<0.087

>0.087

KOC 

(L/kg oc)

198

243

771

115

78

52

4

60

102

33

179

107

60 

64 

108

32

14

62

44

36-160

135 

129 

103

47

129

129

293

101

72 

96 

142

2.0-120

106-460

58

362

373

237

177 

205 

348

1,720

26

27

28

27

Reference

Dobbs and others (1989)

Dobbs and others (1989)

Dobbs and others (1989)

Rutherford and others ( 1 992)

Rutherford and others (1992)

Rutherford and others ( 1 992)

Rutherford and others (1992)

Kile and others (1995)

Kile and others (1995)

Chiou and others (1979)

Chiou and others (1979)

Friesel and others ( 1 984)

Seip and others (1986) 

Seip and others (1986) 

Seip and others (1986)

Hamaker and Thompson (1972)

Hamaker and Thompson (1972)

Chiou and others (1979)

Kenaga and Goring (1980)

Mingelgrin and Gerstl (1983)

Steinberg and others (1987) 

Steinberg and others (1987) 

Steinberg and others (1987)

Chiou and others (1979)

Chiou and others (1979)

Kenaga and Goring (1980)

Dobbs and others (1989)

Friesel and others (1984)

Seip and others (1986) 

Seip and others (1986) 

Seip and others (1986)

Garbarini and Lion (1986)

Garbarini and Lion ( 1 986)

Rutherford and Chiou (1992)

Chiou and others (1979)

Schwarzenbach and Westall (1981)

Friesel and others (1984)

Seip and others (1986) 

Seip and others ( 1986) 

Seip and others (1986)

Dobbs and others (1989)

Hamaker and Thompson (1972)

Hamaker and Thompson (1972)

Hamaker and Thompson (1972)

Hamaker and Thompson (1972)
70 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

Table 15. Experimental sorption coefficients for 27 target analytes—Continued

[USGS, U.S. Geological Survey; foc , fraction of organic carbon by weight; Koc , sorption coefficient normalized with the fraction by weight of organic 

carbon; L/kg oc, liter per kilogram of organic carbon; ns, not specified; mg/L, milligram per liter; <, less than; >, greater than]

USGS 

parameter Compound 

code

34030 benzene

34696 naphthalene

340 1 0 methylbenzene

34371 ethylbenzene

77342 n-butylbenzene

85795 1,3-dimethylbenzene

85795 1 ,4-di methylbenzene

Sorbent

3 sediments

2 soils

17 sediments and soils

Soil

3 soils

2 soils

Corn residues

Plant cuticle

Extracted peat

Peat

Muck

Cellulose

3 sediments

1 7 sediments and soils

2 sediments

(10 mg/L) 

(100 mg/L) 

(300 mg/L)

(10 mg/L) 

(100 mg/L) 

(300 mg/L)

Soil

Aquifer material

2 soils

7 organic compounds

4 soil fractions

3 soils

Soil

2 soils

Corn residues

Plant cuticle

Aquifer material

3 soils

Aquifer material

foc

0.00086-0.0329

0.026 

0.018

0.0011-0.0238

0.0093

0.002 

0.022 

0.037

0.0094 

0.0184

0.427

0.99

0.640

0.571

0.531

0.444

0.00086-0.0329

0.0011-0.0238

0.029

0.038

0.051

0.0015

0.0023

0.0041

0.12-0.65

0.0014-0.0751

0.002 

0.022 

0.037

0.0093

0.0094 

0.0184

0.427

0.99

0.0015

0.002 

0.022 

0.037

0.0015

KOC

(L/kg oc)

83

100 

92

60

31

38 

44

54

31 

38

40

155

32

22

14

1

1,300

870

14,400 

3,860 

2,070

12,900 

4,260 

2,530

804

247

77

191

0.05-151

151-348

56 

94 

134

165

112 

187

132

1,630

2,460

129 

158 

289

333

Reference

Karickhoff and others (1979)

Rogers and others (1980) 

Rogers and others (1980)

Karickhoff (1981)

Chiou and others (1983)

Seip and others (1986) 

Seip and others ( 1 986) 

Seip and others (1986)

Boyd and others (1990) 

Boyd and others (1990)

Boyd and others (1990)

Boyd and others (1990)

Rutherford and others (1992)

Rutherford and others (1992)

Rutherford and others (1992)

Rutherford and others (1992)

Karickhoff and others (1979)

Karickhoff (1981)

Voice and others (1983)

Voice and others (1983)

Whitman and other? (1995)

Schwarzenbach and Westall (1981)

Garbarini and Lion (1985)

Garbarini and Lion (1985)

Garbarini and Lion (1986)

Garbarini and Lion (1986)

Seip and others (19^6) 

Seip and others (19^6) 

Seip and others (19^6)

Chiou and others (1983)

Boyd and others (1990) 

Boyd and others (1990)

Boyd and others (1990)

Boyd and others (1990)

Schwarzenbach anc1 Westall (1981)

Seip and others (19^6) 

Seip and others (19^6) 

Seip and others (19^6)

Schwarzenbach and Westall (1981)
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

71

Table 15. Experimental sorption coefficients for 27 target analytes—Continued

[USGS, U.S. Geological Survey; foc , fraction of organic carbon by weight; Koc , sorption coefficient normalized with the fraction by weight of organic 

carbon; L/kg oc, liter per kilogram of organic carbon; ns, not specified; mg/L, milligram per liter; <, less than; >, greater than]

USGS 

parameter Compound Sorbent 

code

34301 chlorobenzene Aquifer material

Soil

2 sediments 

(10 mg/L) 

(100 mg/L) 

(300 mg/L)

(10 mg/L) 

(100 mg/L) 

(300 mg/L)

3 sewage solids

34536 1 ,2-dichlorobenzene Soil

Soil

Peaty soil

32 soils

36 sediments

34556 1,3-dichlorobenzene Soil

Peaty soil

34571 1 ,4-dichlorobenzene Aquifer material

Soil

Peaty soil

Sediment

77613 1,2,3-trichlorobenzene Aquifer material

2 soils

Corn residues

Plant cuticle

34551 1,2,4-trichlorobenzene Aquifer material

Soil

Peaty soil

Sediment

However, despite the high organic-carbon fraction of

cellulose (table 15), sorption of VOCs on this sorbent

was minimal (Garbarini and Lion, 1986; Rutherford

and others, 1992). Plant cuticle is virtually all organic

matter, and sorption on this sorbent was about

10 times greater than on soils for ethylbenzene and

1,2,3-trichlorobenzene and about 5 times greater for

benzene (table 15). The lesser effect for benzene is

presumably because benzene is more soluble than the

foe

0.0015

0.0093

0.029

0.038

0.38-0.49

0.0093

0.0093

0.17

0.0016-0.0609

0.0011-0.0473

0.0093

0.17

0.0015

0.0093

0.17

0.0816

0.0015

0.0094

0.0184

0.427

0.99

0.0015

0.0093

0.17

0.0816

i^

,. ., oc ^ Reference (L/kg oc)

260 Schwarzenbach and Westall (1981)

83 Chiou and others ( 1 983)

Voice and others (1983) 

5,000 

690 

269

Voice and others (1983) 

6,200 

1,240 

579

753 Dobbs and others (1989)

3 1 0 Chiou and others (1979)

321 Chiou and others (1983)

529 Friesel and others ( 1 984)

290 Kile and others ( 1 995 )

502 Kile and others ( 1 995 )

293 Chiou and others (1983)

478 Friesel and others ( 1 984)

733 Schwarzenbach and Westall (1981)

273 Chiou and others ( 1 983)

429 Friesel and others ( 1 984)

1 ,070 Wu and Gschwend ( 1 986)

2,650 Schwarzenbach and Westall (1981)

1 ,380 Boyd and others (1990)

1,630 Boyd and others (1990)

1,050 Boyd and others (1990)

14,800 Boyd and others (1990)

2,350 Schwarzenbach and Westall (1981)

864 Chiou and others ( 1 983)

1 ,440 Friesel and others ( 1 984)

3,250 Wu and Gschwend ( 1 986)

two substituted benzenes. Sorption of VOCs on the

sewage solids with organic-carbon fractions from C .38

to 0.49 was variable, with the sorption coefficients for

some VOCs such as tetrachloroethene being larger

than for the other sorbents. This brief discussion of the

sorption coefficients for VOCs indicates that normalization with the organic-carbon fraction of the sorbent

is not sufficient to remove all the variability of the

sorption coefficients.
72 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

However, the sorption coefficients presented 

in table 15 are for sorbents with a much wider range 

of organic-carbon fractions than will normally be 

present in streams. Hamaker and Thompson (1972) 

arbitrarily used an organic-carbon fraction of 0.087 as 

the dividing line between "real world" or normal soils 

and sediments and sorbents with high organic-carbon 

fractions. Many of the sorption coefficients in table 15 

are for "real world" soils and sediments. Sorbents with 

organic-carbon fractions larger than 0.087 are atypical 

but are included in table 15 because of the possibility 

that they could be important in some situations. The 

variability in the sorption coefficients in table 15 

indicates that composition of the sorbent also is 

important because the effect of the organic-carbon 

fraction has theoretically been removed through the 

normalization process.

Voice and others (1983) observed that the 

sorption coefficients for naphthalene and chlorobenzene decreased as the sediment concentration 

increased as a result of the "solids concentration" 

effect discussed previously. Consequently, coefficients for these two VOCs are presented in table 15 

at three sediment concentrations interpolated from 

the sorption coefficient-concentration relationships presented by Voice and others (1983). These 

coefficients are included as an indication of the 

possible effect of concentration on the sorption 

coefficient.

PREDICTION EQUATIONS FOR 

SORPTION COEFFICIENTS

The sorption of nonionic organic compounds 

by soils and sediments containing organic matter is 

dominated by the process of partitioning of the organic 

molecules into the organic matter. Partitioning of 

the molecules between the water and organic-matter 

phases is analogous to partitioning between water 

and an organic solvent. Consequently, sorption coefficients frequently are correlated with the octanol-water 

partition coefficient. Water solubility also is important 

because the less soluble a compound, the greater 

the tendency to partition into the octanol and, by 

inference, the greater the tendency to sorb to soils 

and sediments. Thus, sorption coefficients also are 

correlated with the water solubility of the compound.

Most of the equations resulting from correlation of sorption coefficients with the octanol-water 

partition coefficient or the water solubility were

developed for sorption of pesticides by sc ; ls. A few 

were based on data bases that included chlorinated 

aliphatic compounds, aromatic and chlorinated 

aromatic compounds, and polycyclic aromatic 

hydrocarbon compounds. Equations considered most 

appropriate for estimating the sorption coefficients for 

three general classes of VOC compounds are given in 

table 16. Independent variables were either the water 

solubility or the octanol-water partition coefficient. 

Original equations that were normalized with the 

organic-matter content were converted to organiccarbon content by assuming the organic matter was 

58 percent carbon by weight (Smith and others, 1988). 

Also, original equations having solubilities in 

millimoles or micromoles per liter were converted to 

moles per liter, and base e logarithms were converted 

to base 10.

The data base for equation 78 in taHe 16 

consisted of the target analytes 1,2-dichlcroethane, 

1,2-dichloropropane, 1,2-dibromo-3-chlo^opropane, 

1,1,1-trichloroethane, l,2-dichlorobenzen<% tetrachloroethene, 1,2-dibromoethane plus one additional 

VOC, four pesticides, and three PCS corroounds. 

The data base for equation 79 consisted of the target 

analytes 1,1,1-trichloroethane, trichloroethene, and 

tetrachloroethene. The data base for equations 80 

and 81 consisted of the target analytes benzene 

and naphthalene plus six other polycyclic aromatic 

hydrocarbons, one pesticide, and one PCB compound. 

Equation 82 was based on data for benzene, naphthalene, and three other polycyclic aromatic hydrocarbons. Equation 83 was based on data for the target 

analytes methylbenzene, 1,4-dimethylbenzene, 

n-butylbenzene, tetrachloroethene, chlorobenzene, 

1,4-dichlorobenzene, 1,2,4-trichlorobenzene, and 

1,2,3-trichlorobenzene plus three other alkyl 

benzenes and two other chlorinated benzenes. 

The data base for equations 84 and 85 consisted of 

the target analytes benzene, chlorobenzene, ethylbenzene, 1,2-dichlorobenzene, 1,3-dichlorobenzene, 

1,4-dichlorobenzene, and 1,2,4-trichlorobenzene 

plus one other substituted benzene and fcrir PCB 

compounds.

Each of the equations in table 16 was developed 

from data for a somewhat limited range of types 

of compounds. Kenaga and Goring (1980) developed 

equations using all the data available at tl at time 

as one group. The resultant equations were
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

73

Table 16. Equations for predicting sorption coefficients of volatile organic compounds

[VOC, volatile organic compound; log, logarithm to the base 10; Koc, sorption coefficient normalized with the organic carbon content, in liters per kilogram 

of organic carbon; S, water solubility, in moles per liter; Kow, octanol-water partition coefficient]

VOC 

class

Halogenated aliphatic s 

Aromatic

Mixed aromatic s and 

chlorinated aromatics

Equation 

number

(78) 

(79) 

(80)

(81)

(82)

(83)

(84)

(85)

Equation

log Koc = 0.935 -0.557 log S 

log KOC =1.35 -0.337 logS 

logKoc = 1.3 8 -0.54 log S

log Koc = -0.21+ log Kow

log Koc = -0.346 + 0.989 log Kow

log Koc = 0.49 + 0.72 log Kow

log Koc = 0.23 8 -0.729 log S

log Koc = -0.542 + 0.904 log Kow

Reference

Chiou and others (1979) 1 

Friesel and others ( 1 984)2 

Karickhoff and others ( 1 979)3

Karickhoff and others ( 1 979)

Karickhoff (1981)

Schwarzenbach and Westall (1981)

Chiou and others (1983)4

Chiou and others (1983)

'Equation converted from organic-matter normalization to organic-carbon normalization assuming organic matter is 58 percent by weight 

organic carbon; solubility changed from micromoles per liter to moles per liter.

2Equation converted from organic-matter normalization to organic-carbon normalization assuming organic matter is 58 percent by weight 

organic carbon; solubility changed from millimoles per liter to moles per liter; converted from base e logarithms to base 10 logarithms. 

Solubility changed from mole fraction solubility to moles per liter.

4Equation converted from organic-matter normalization to organic-carbon normalization assuming organic matter is 58 percent by weight 

organic carbon.

log 10 Koc = 3.64-0.55 log10 WS (86)

and

l°gioKoc = 1.377 +0.544 log 10 Kc (87)

where

WS is the water solubility (mg/L). 

Equations 86 and 87 were based on 106 and 45 pairs 

of data points, respectively, and the correlation coefficients were -0.84 and +0.86. The correlation coefficients for the equations in table 16 ranged from 0.970 

to 1.00. These coefficients were larger than the coefficients for equations 86 and 87, presumably because of 

the smaller data bases and the greater uniformity of the 

characteristics of the compounds within each of the 

data bases. However, all of these equations are based 

on log-log correlations; therefore, considerable 

variation about the line of agreement should be 

expected, even when the correlation coefficients are 

large.

Molecular topology theory also has been 

applied to the prediction of sorption coefficients 

(Sabljic, 1987; Sabljic and others, 1995). This

procedure correlated the first-order molecular connectivity index, *%, with the sorption coefficient. This 

index was computed from

(88)

k= 1

where each nonhydrogen atom in the molecule was 

described by its atomic 8 value. This 8 value was enual 

to the number of adjacent nonhydrogen atoms, and the 

i and j subscripts corresponded to the pairs of adjacent 

nonhydrogen atoms. Summation was over all bonds 

between nonhydrogen atoms, where n was the number 

of these bonds. The resultant equation for predominantly hydrophobic organic compounds was

log 10 Koc = (0.52) 0.70 (89)

The correlation coefficient for equation 89 was 

0.980 for a sample size of 81 compounds, of which 

33 were target analytes. Equation 89 requires only 

a knowledge of molecular structure to predict the 

sorption coefficient.
74 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

Sorption coefficients for the 28 target analytes 

for which experimental coefficients were not available 

were predicted from the equations in table 16 and 

equations 86, 87, and 89. The water solubilities and 

octanol-water partition coefficients were mostly from 

the compilations of Mackay and others (1992a; 1993). 

Exceptions were as follows. Solubilities for trichlorofluoromethane and dichlorodifluoromethane were 

from Warner and Weiss (1985). The solubility for 

l,l,2-trichloro-l,2,2-trifluoroethane was fromBu and 

Warner (1995). The solubility and the octanol-water 

partition coefficient for bromoethene were estimated 

as discussed previously for the calculation of the 

Henry's law constant for this compound. The 

solubility for ethyl tertiary-butyl ether was from Evan 

and Edlund (1936), and the solubility for tertiary-amyl 

methyl ether was from Stephenson (1992). The 

octanol-water partition coefficients for these two 

compounds were assumed to be the same as the 

measured value of Funasaki and others (1985) for 

diisopropyl ether because these authors showed that 

the log Kow is related to the molar volume, and all 

three of these ethers have the same computed molar 

volume. The solubility for 2-propenenitrile was from 

Lyman (1990b, Table 2-14), and the octanol-water 

partition coefficient was from Verschueren (1983). 

Finally, the solubility for 2-propenal was from 

Verschueren (1983), and the octanol-water partition 

coefficient was from Callahan and others (1979).

The results presented in table 17 indicate 

considerable variation among the coefficients 

predicted with the different equations, as was expected 

considering the log-log nature of the equations. For the 

halogenated alkanes and alkenes, the results indicate 

that the overall equation (eq. 87) based on the octanolwater partition coefficient always predicted the highest 

sorption coefficient, whereas equation 89 based on 

molecular structure (Sabljic, 1987; Sabljic and others, 

1995) generally predicted the lowest coefficients. The 

results for the alkyl benzene target analytes indicate 

that the overall equation (eq. 87) gave the highest 

predicted coefficients and equation 84 (Chiou and 

others, 1983) gave the lowest coefficients, although 

equation 86 (Kenaga and Goring, 1980) gave values 

comparable to equation 84 for most compounds. 

The predicted coefficients for the ether compounds, 

2-propenenitrile, and 2-propenal were generally small, 

presumably because these compounds are the most 

soluble of any of the target analytes. The coefficients 

predicted from equation 89 on the basis of molecular

structure were at or near the lower limit c f the range 

of coefficients predicted by the other equations for 

the halogenated alkanes and alkenes, and within the 

range of predicted coefficients for the one aromatic 

hydrocarbon and the alkylbenzenes. For the ethers, 

2-propenenitrile, and 2-propenal, the coefficients 

predicted with equation 89 were the highest of the 

predicted values.

ESTIMATION OF THE 

IMPORTANCE OF SORPTIC N

The fraction of a target analyte that will be 

sorbed to sediment at equilibrium can be estimated 

following a procedure used by Voice and Weber 

(1983). Combining a mass balance for the analytewater-sediment system at equilibrium with 

equations 75 and 77 gives

f s = (90)

where

fs is the fraction by weight of the z nalyte that is

sorbed to the sediment, 

foc is the weight fraction of organic carbon in

the sediment, and

Csd is the sediment concentration (l"g/L). 

Equation 90 indicates that the fraction sobbed is 

dependent on the sorption coefficient as expected, 

the organic-carbon fraction of the sediment, and 

the sediment concentration.

Equation 90 was applied to the sorption of 

hexachlorobutadiene, 1,2,3-trichlorobenzene, and 

naphthalene for sediments with organic-carbon 

fractions of 0.01, 0.04, and 0.087, where 0.087 is the 

arbitrarily assumed dividing line between "real-world" 

soils and sediments and sorbents with high organiccarbon fractions (Hamaker and Thompsor, 1972). 

Hexachlorobutadiene was selected because it had the 

highest predicted sorption coefficient of any of the 

analytes (table 17). The analyte 1,2,3-tricl lorobenzene 

was selected because it had the highest exoerimental 

sorption coefficient of any of the analytes (table 15), 

exclusive of the plant cuticle values and tH values 

of Voice and others (1983), which will be discussed in 

more detail below. Naphthalene was selected because it 

is a solid at normal environmental temperatures, with 

a melting point of 80.5°C (Mackay and others, 1992b).
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

Table 17. Predicted sorption coefficients for 28 target analytes 

[USGS, U.S. Geological Survey; Koc, sorption coefficient normalized with the fraction by weight of organic carbon; eq., equation]

uses

parameter Compound 

code

Halogenated alkanes

344 1 8 chloromethane

344 1 3 bromomethane

3 2 1 04 tr ibromomethane

32101 bromodichloromethane

32105 chlorodibromomethane

343 1 1 chloroethane

34496 1 , 1 -dichloroethane

34396 hexachloroethane

77443 1,2,3-trichloropropane

34488 trichlorofluoromethane

34668 dichlorodifluoromethane

77652 1,1,2-trichloro- 1,2,2- 

trifluoroethane

Halogenated alkenes

39 1 75 chloroethene

77093 cis- 1 ,2-dichloroethene

34546 *ra«.s-l,2-dichloroethene

50002 bromoethene

39702 hexachlorobutadiene

Aromatic hydrocarbon

77128 styrene

Alkyl benzenes

77224 n-propylbenzene

77223 iso-propylbenzene

77135 1,2-dimethylbenzene

77222 1,2,4-trimethylbenzene

Ethers and other compounds

78032 methyl tertiary-butyl ether

50004 ethyl tertiary-butyl ether

50005 tertiary-amyl methyl ether

81577 dii sop ropyl ether

34210 2-propenal

342 1 5 2-propenenitrile

eq.78

30.0

23.6

99.4

80.4

101

33.1

43.9

1,350

97.2

109

223

220

eq.78

49.5

54.8

39.5

30.7

4,660

eq. 80

605

eq. 79

47.6

41.2

98.3

86.5

99.2

50.6

60.0

476

97.0

104

160

159

eq. 79

64.5

68.6

56.2

48.3

1,010

eq. 81 eq. 82

537 364

KOC

(liter per kilogram 

of organic carbon)

eq. 86

38.8

21.7

52.2

53.7

59.0

37.4

39.0

710

68.7

80.3

174

135

eq. 86

56.6

49.1

35.5

26.3

2,290

eq. 84 eq. 85 eq. 86

135 131 203

eq. 83 eq. 84 eq. 85

1,360 497 597

1,170 440 495

502 156 171

1,630 464 750

eq.87

74.5

106

447

331

394

143

224

3,270

430

567

356

1,250

eq.87

134

245

315

160

9,490

eq.87

947

eq. 86

502

459

224

477

eq. 86

13.6

24.9

27.8

34.2

5.2

8.8

eq. 89

16.6

16.6

39.9

39.9

39.9

27.2

39.9

245

145

55.0

55.0

245

eq. 89

27.2

49.6

49.6

27.2

1050

eq. 89

555

eq.87

2,360

2,110

1,110

2,710

eq.87

77.3

160

160

160

21.3

7.5

75

eq 89

1,010

1,300

477

764

eq 89

108

196

210

212

38.7

49.6
76 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

It also has relatively low water solubility and, therefore, 

is likely to sorb to sediments. For reference purposes, 

the average of six naphthalene water solubilities at 

25 °C determined by the generator column technique 

was 31.5 mg/L (Mackay and others, 1992b). The 

solubility of hexachlorobutadiene was 3.23 mg/L at 

25°C (Banerjee and others, 1980), and the solubility of 

1,2,3-trichlorobenzene was 12.3 mg/L at 25°C (Miller 

and others, 1985).

The results presented in figure 12 indicate that 

very little of these three compounds will be sorbed 

to sediments for sediment concentrations less than 

100 mg/L. For higher concentrations, the fraction 

of the organic compound sorbed increases rapidly. 

Hexachlorobutadiene is sorbed to the greatest extent, 

and naphthalene is sorbed the least, in agreement with 

the previous indication that the extent of sorption is 

inversely proportional to the water solubility. The 

results in figure 12 also indicate that the extent of 

sorption increases as the organic-carbon fraction 

of the sediment increases.

Suspended-sediment concentrations in streams 

vary, depending on flow conditions, characteristics of 

the stream, and the drainage-basin characteristics. As 

one point of reference, suspended-sediment concentrations in two urban stormwater-runoff studies ranged 

from 101 to 1,400 mg/L (von Guerard and Weiss, 

1995) and from 0.5 to 750 mg/L (Pitt and others, 

1995). Consequently, sorption of the target analytes to 

sediments could be a significant process, particularly 

for the less soluble analytes. Each situation should be 

evaluated individually to determine if sorption to 

sediments is a significant process in determining the 

fate of VOCs in streams.

The results of Voice and others (1983) indicated 

an apparent dependence of the sorption coefficient on 

the sediment concentration (table 15). The organiccarbon fractions of the two sediments used in this 

study were 0.029 and 0.038. Combining these fractions 

with the sorption coefficients appropriate for the 

corresponding sediment concentrations gives sorbed 

fractions ranging from 0.0042 to 0.028 for naphthalene. 

These fractions are in qualitative agreement with the 

low fractions for the concentration range from 10 

to 300 mg/L indicated in figure 12. Corresponding

Figure 12. Fraction of the organic compound sorbed to sediment 

as a function of the sediment concentration for three target 

analytes.

NAPHTHALENE 

ORGANIC-CARBON FRACTION

0.01

'1 10 100 1,000 10,000 

SEDIMENT CONCENTRATION, IN MILLIGRAMS PER LITER
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

77

fractions for chlorobenzene ranged from 0.0014 to 

0.0067, indicating little sorption of this compound for 

sediment concentrations between 10 and 300 mg/L.

The development of equation 90 assumes 

that the water and sediment phases are in equilibrium, 

and a rapid and reversible equilibrium frequently 

is assumed in sorption studies. However, the review 

of the kinetics of the sorption process presented 

previously indicated that the assumption of a rapid 

and reversible equilibrium may not be completely 

valid because of strongly bound molecules. As an 

example of the effect of strongly bound molecules, 

Steinberg and others (1987) determined in desorption 

experiments that 1,2-dibromoethane applied to soils 

as long as 19 years earlier was far from equilibrium 

with the water phase. This lack of equilibrium was 

attributed to the fact that the 1,2-dibromoethane 

had diffused into the soil micropores where it was 

inaccessible to soil bacteria and not readily available 

for equilibration with the air and water phases of 

the system. This fraction of the 1,2-dibromoethane 

was estimated to have half-equilibration times for a 

1:2 soil-water mixture of two to three decades at 25°C.

Another example of the importance of intraparticle diffusion is that of Wu and Gschwend (1986), 

who studied the sorption of four chlorobenzene 

compounds on three soils and sediments. They 

reported a slower approach to equilibrium for large 

particles; however, when the particles were disaggregated by sonication, the approach to equilibrium was 

faster. Wu and Gschwend also observed that the more 

hydrophobic (less water soluble) compounds had 

slower sorption rates; consequently, the greater the 

tendency to sorb, the slower the approach to equilibrium. They suggested on the basis of their work that 

the assumption of equilibrium may be valid for contact 

times of days and months. However, for situations 

where the water is continually being replaced with 

"new" water and contact times are much shorter, 

the assumption of equilibrium probably is not valid. 

This suggestion is supported by the work of Rostad 

(1996) who concluded that concentrations of several 

pesticides, herbicides, and PCB compounds in the 

water, colloid, and suspended-sediment phases of the 

Mississippi River were not in equilibrium. Similarly, 

Cheng and others (1995) used calculations to predict 

that the mass of organic compound desorbed from 

sediments in the Buffalo River, New York, during a 

resuspension event would be overpredicted with the 

assumption of equilibrium conditions.

The effect of failing to achieve equilibrium 

will be to overpredict the mass sorbed during a 

sorption process and to underpredict the mass 

desorbed during a desorption process. Consequently, 

the fractions shown in figure 12 are likely to be 

maximum predictions.

RECENT DEVELOPMENTS

The partition model (eq. 75) for the sorption 

of nonionic compounds to solids containing organic 

matter has been used extensively. This model has the 

advantage that sorption isotherms are linear, and the 

equilibrium expressed by equation 75 normally is 

assumed to be established rapidly. Consequently, 

incorporating into a numerical model the effect of 

sorption on the transport and behavior of organic 

compounds in streams is relatively easy. Recent 

developments, however, indicate that some of the 

assumptions of the partition model may not be valid in 

some situations. Specifically, nonlinear isotherms and 

competitive sorption have been determined. These 

results are inconsistent with the partition model, and 

several models have been proposed in attempts to 

explain these phenomena.

Weber and others (1992) and McGinley and 

others (1993) proposed a distributed reactivity me del 

(DRM) to account for isotherm nonlinearities and 

competitive effects determined in their sorption 

studies. This model allows for the fact that soils which 

appear homogeneous on a macroscopic scale can be 

highly heterogeneous on a microscopic scale, witl 

reactivities distributed among a variety of matrix 

components. To test this model, Weber and other? 

(1992) measured isotherms for the sorption of 

tetrachloroethene, 1,4-dichlorobenzene, and 

1,2,4-trichlorobenzene on six soils. The Freundlich n 

(eq. 76) was used as an indicator of nonlinear 

behavior, and n ranged from 0.68 to 1.20, with 16 

of 18 values less than the 1.0-value corresponding 

to linear behavior. The degree of nonlinearity 

increased with increasing sorption capacity of the soils 

and increased as the solute hydrophobicity increased. 

The degree of nonlinearity varied only slightly airong 

solutes for a specific soil and varied considerably 

among soils for a specific solute. McGinley and others 

(1993) determined the sorption of tetrachloroethene in 

the presence of a constant initial concentration of 

1,2,4-trichlorobenzene. The 1,2,4-trichlorobenzeire 

markedly reduced the sorption of trichloroethene on
78 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

one soil, reduced the sorption to somewhat lesser 

extent on a second soil, and had little effect on a third 

soil. The results for the first two soils indicated 

competitive sorption, which was contrary to expectations based on the partition model.

The DRM is somewhat qualitative at this point 

in its development, and from a practical standpoint, 

only one or two nonlinear components can be differentiated in a specific situation. However, model calculations (McGinley and others, 1993) indicated for 

0.1 weight percents of nonlinearly sorbing and 

linearly sorbing organic carbon that only about 

10 percent of the sorption of tetrachloroethene was 

attributed to the linearly sorbing component at a 

solution concentration of 1 p.g/L. On the other hand, 

about 50 percent was attributable to the linearly 

sorbing component at a solution concentration of 

10,000 p.g/L. This calculation indicates that the 

nonlinearly sorbing component will have a much 

greater effect at the low concentrations typical of 

environmental situations.

Young and Weber (1995) expanded the DRM 

concept by suggesting that sorption could lie at any 

point along a spectrum ranging from simple phase 

partitioning to pure surface adsorption. This spectrum 

can be either continuous or discontinuous, with 

different organic-matter fractions responsible for 

different degrees of reactivity. The degree of 

diagenetic alteration that each organic-matter fraction 

has undergone is the principal factor determining 

the reactivity of each fraction and where the 

overall sorption process lies along the spectrum. 

Consequently, they modeled soil organic matter 

as a macromolecule with a structure ranging from 

completely amorphous in a relatively young soil 

such as fulvic acids to a structure that is increasingly 

condensed as diagenesis occurs. The ultimate endpoint 

is a highly crystalline structure of extremely old 

organic carbon such as anthracite coal.

Young and Weber (1995) determined the effect 

of temperature on the sorption of phenanthrene to 

three soils. The changes in the Freundlich isotherm 

parameters with temperature were not statistically 

significant. However, trends were consistent with 

adsorption theory, with a weak trend toward 

increasing linearity and decreasing sorption capacity 

as the temperature increased. They concluded that 

the DRM provides a qualitative basis for explaining 

these temperature effects, as well as the isotherm

nonlinearities and competitive sorption effects 

determined previously by Weber and others (1992) 

and McGinley and others (1993).

Spurlock and Biggar (1994a; 1994b) 

developed a thermodynamic partition model for 

linear and nonlinear organic-carbon-based sorption. 

The Freundlich partition coefficient, Kf, wf related 

to the aqueous-phase activity coefficient and sorbate 

solubility in the humic phase. The Freundlich exponent, 

n, was related to the variation of the sorbate partial 

molar free energy with the sorbed-phase loading. 

Spurlock and Biggar (1994b) determined tH sorption 

of five substituted phenylureas on three sorbents. The 

sorbate residual free energies decreased rapidly as the 

sorbed-phase concentrations decreased, indicating 

strong specific sorbate-sorbent interactions at low 

sorbed-phase loadings. Nonlinearity was related 

to the variations of the sorbate partial molar free energy 

in the sorbed phase.

Xing and others (1996) hypothesizec1 that soil 

organic matter is a dual-mode sorbent, with sorption 

occurring by a combination of partition anc? hole-filling 

mechanisms. The holes are conceptualized as specific 

sites within the matrix where complexatiorj follows 

the Langmuir isotherm. The holes are considered to 

discriminate on the basis of molecular structure, and it 

is likely that more than one type of hole exists because 

of the intrinsic physical/chemical heterogeneity 

of the soil organic matter. Xing and Pignatello (1996) 

determined 1-day, 30-day, and 180-day isotherms 

for the sorption of 1,3-dichlorobenzene, 2,4- 

dichlorophenol, and metolachlor on loam and peat soils. 

The Freundlich partition coefficient, Kf, increased by as 

much as 2.7 fold for the times greater than 1.0 day. The 

Freundlich exponent, n, was less than 1.0 in all cases, 

indicating nonlinear isotherms. The n values for 

the 30-day and 180-day isotherms were significantly 

smaller that the values for the 1-day isotherms, 

indicating that partitioning was much less ideal 

for the longer times. They concluded that the results 

were consistent with the dual-mode sorption model, 

with the adsorption process (hole-filling mechanism) 

being more important for the longer times. This 

indicates that the adsorption sites are internal to 

the soil organic-matter matrix and unevenly distributed 

with respect to access by the molecules being sorbed. 

Xing and Pignatello (1996) also noted that an important 

consequence of the nonlinear isotherm was that, as the 

concentration decreased, sorption became greater than 

would be expected for a linear isotherm.
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

79

Weber and Huang (1996) proposed a threedomain particle for modeling the sorption process. 

Domain I consists of exposed inorganic mineral 

surface where sorption is expected to be largely an 

adsorption process and, therefore, approximately 

linear. Domain II consists of amorphous soil organic 

matter where sorption is expected to be largely a 

partitioning process and, therefore, approximately 

linear. Domain III consists of condensed soil organic 

matter. They rationalized that the processes in 

domains I and II are relatively rapid. Consequently, 

once domains I and II have reached equilibrium, the 

overall sorption process is dominated by domain III. 

Weber and Huang (1996) hypothesized that sorption 

by the condensed soil organic matter of domain III 

will be energetically more favorable and more 

nonlinear than sorption by the amorphous soil organic 

matter in domain II because of the more heterogeneous composition and less polar nature of the 

condensed organic matter. Also, sorption by the 

condensed organic matter appears to be very slow, 

and this was attributed to slow intradomain diffusion.

Weber and Huang (1996) determined the 

sorption of phenanthrene on one soil and three 

sediments. Sorption isotherms were nonlinear, with 

Freundlich n values ranging from 0.726 to 0.890. The 

sorption capacity indicated by the Freundlich partition 

coefficient, Kf, increased rapidly with time, and n 

decreased rapidly with time. The changes in Kf and n 

appeared to occur in three stages: an initiation 

stage, a logarithmic stage, and an apparent equilibrium 

stage. Weber and Huang (1996) concluded that the 

logarithmic stage corresponded to domain III of the 

three-domain model.

Xing and Pignatello (1997) measured the 

sorption of chlorobenzene, 1,2-dichlorobenzene, and 

1,3-dichlorobenzene on a peat soil, extracts of this 

soil, and two humic acids extracted from Canadian 

soils. The isotherms frequently were nonlinear, with 

Freundlich n values ranging from 0.777 to 0.956. The 

nonlinearity occurred over three orders of magnitude 

change in concentration and extended to very small 

concentrations. Competitive sorption also occurred 

between structurally similar molecules. The results 

were consistent with the dual-mode sorption model 

in which soil organic matter is conceptualized as an 

amalgam of rubbery and glassy phases. Each of these 

phases has a dissolution domain, and the glassy phase 

also contains holes that provide complexation sites. 

The holes are limited in size and number, and they

cover a range of steric and electronic characteristics. 

These results conflict with the partition sorption 

model that assumes isotherms are linear and, 

consequently, that sorption coefficients are independent of concentration.

These recent developments have established 

that sorption isotherms frequently are nonlinear. The 

effect of this nonlinearity increases as the concentration decreases, with the result that sorption is most 

likely to be underpredicted in the low-concentration 

range of importance in environmental situations. 

The various versions of the DRM, dual-mode sorHnt, 

and three-domain models can be used to explain 

qualitatively nonlinear isotherms, competitive 

sorption, and thermal effects inconsistent with the 

partition model of sorption. However, these models 

are not sufficiently developed to allow prediction 

of sorption coefficients and their concentration 

dependence in nonlinear systems. Consequently, the 

sorption coefficients presented in tables 15 and 17 

must be used to estimate the effect of sorption on 

the concentrations of the target analytes in streams. 

This estimation should be done, however, with the 

understanding that sorption could be underestimated 

because of nonlinear effects.

HYDROLYSIS

Hydrolysis is the reaction of an organic 

compound with water. This reaction can involve 

the hydrogen ion or the hydroxyl ion of water, or 

the reaction can be neutral. An example of a neutral 

hydrolysis reaction is

RX + HOH = ROH + HX (91)

where

RX is the organic compound with the functional group X that is capable of being 

hydrolyzed.

Examples of hydrolyzable functional groups include 

halogenated aliphatics, epoxides, organophosphorus 

esters, carboxylic acid esters, anhydrides, amides, 

carbamates, and ureas (Larson and Weber, 1994). 

The rate of hydrolysis can be described by a 

general first-order expression of the form

—- 

d[RX] _~ K (92)
80 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

where 

[RX] is the concentration of the organic

compound, 

t is time, and

Kh is the rate coefficient for hydrolysis. 

However, because the hydrolysis process is known to 

be dependent on pH, this simplified equation must be 

modified to account for pH effects. This can be done 

by using the relationship (Washington, 1995)

d[RX] = K- a[H+ ][RX]

dt

+ K'b [OH-][RX] (93)

where

K'a and K'b are the second-order rate coefficients 

for the acid and base hydrolysis 

reactions, 

Kn is the first-order rate coefficient for the

neutral hydrolysis reaction, 

[H+] is the hydrogen-ion concentration, and 

[OtT] is the hydroxyl-ion concentration. 

Comparing equations 92 and 93 indicates that

Kh [RX] = K'a [H+ ][RX] + Kn [RX] 

+ K'b [OH-][RX].

(94)

When the concentration of the organic 

compound is dilute, such as is common in environmental situations, concentration effects on the rate of 

hydrolysis are minimal, and the concentration can be 

canceled in equation 94. Also, the [OH~] ion concentration can be eliminated from equation 94 by substituting the kinetic expression for the dissociation of 

water. With these changes, equation 94 becomes

Kh = K'a [H+ ] (95)

where

Kw is the rate coefficient for the dissociation of

water, which is defined as Kw = [OH~][H+]. 

Equation 95 indicates that the hydrolysis rate coefficient depends on the relative magnitudes of the rate 

coefficients K'a, Kn, and K'b. In general, the first term 

on the right side of equation 95 will dominate for 

acidic conditions, the third term will dominate for 

basic conditions, and the middle term will dominate

for neutral conditions. Washington (1995) indicated 

that neutral hydrolysis reactions are most likely to be 

important for VOCs in water.

Because of the assumptions involved in 

equations 92 and 93, the overall rate coefficient 

for hydrolysis is sometimes termed a pseu'io firstorder rate coefficient (Mabey and Mill, 1978; 

Washington, 1995). Consequently, integration of 

equation 92 gives

[RX]t = [RX] 0 exp(-Kh t)

where

[RX]t is the concentration at time t, and 

[RX]0 is the concentration at time zero.

The corresponding half-life for hydrolysis is

(96)

0.693

L 0.5h (97)

where 

to 5h is the time required for the concentration to

be reduced to one-half by the process of

hydrolysis.

Because rates of hydrolysis at environmental 

temperatures are slow for most compounds, laboratory 

hydrolysis experiments usually are done at high 

temperatures to increase reaction rates. Th^se experiments are done in sealed glass ampules, with corrections applied for the amount of the compOMnd that 

volatilizes into the head space. Results fro*n these 

high-temperature conditions are extrapolated to 

lower temperatures characteristic of environmental 

conditions. Consequently, the temperature dependence 

of the rate coefficient for hydrolysis is important. This 

temperature dependence usually is expressed by the 

Arrhenius equation (Washington, 1995), which has the 

form

Kh = Ah exp[- — (98)

where 

Ah

R

T

is a constant (day J );

is the activation energy for the hydrolysis

process (kJ/g mol);

is the ideal gas constant (kJ/g mol/K); and 

is the temperature, in kelvins.
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

81

The average activation energy for a number of VOCs 

was 110.1 kJ/g mol (Washington, 1995), which is 

larger than the value of 65 kJ/g mol used for a group 

of many different types of organic compounds (Mabey 

and Mill, 1978). This difference indicates that the 

hydrolysis rate coefficients for VOCs are more 

sensitive to temperature than the coefficients 

for other types of organic compounds.

The concentration also is important. Laboratory 

experiments usually are done at concentrations much 

higher than concentrations present in typical environmental situations. The extrapolation of results for high 

concentrations to low concentrations is rationalized on 

the basis that rate processes that are first order at high 

concentrations remain first order at low concentrations; that is, the rate coefficient is independent of 

concentration (Mabey and Mill, 1978). Of the VOCs, 

only tetrachloromethane has a rate coefficient for 

hydrolysis that is dependent on concentration (Mabey 

and Mill, 1978). More recent experimental work 

coupled with a reanalysis of the older tetrachloromethane data indicated that hydrolysis was first 

order with a half-life of about 40 years (Jeffers and 

others, 1996). Consequently, rate coefficients should 

be independent of concentration, and extrapolation 

of high concentration results to low concentrations 

should be valid for all the target analytes. Similarly, 

semilogarithmic Arrhenius plots based on equation 98 

resulted in linear plots, thus supporting the procedure 

of extrapolating high temperature results to lower 

temperatures characteristic of environmental 

conditions.

The ionic strength of the solution also affects 

the rate of hydrolysis, but the effect is difficult to 

predict. Mabey and Mill (1978) determined that ionicstrength effects can either increase or decrease the 

hydrolysis rates, depending on the organic compound, 

the specific salts causing the change in ionic strength, 

and their concentrations. However, the ionic strength 

of most natural waters usually is less than 0.01 molar 

in total cation and anion concentrations (Mabey and 

Mill, 1978); consequently, they recommended that 

buffer concentrations of 0.01 molar be used in 

determinations of hydrolysis rate constants to 

minimize any possible effect of ionic strength.

The preceding paragraphs give only a brief 

summary of the principles of hydrolysis of organic 

compounds in water. A more thorough discussion is 

given by Larson and Weber (1994).

HYDROLYSIS HALF-LIVES FOR 

THE TARGET ANALYTES

Hydrolysis half-lives and estimates from 

the literature of the potential for hydrolysis for 

49 target analytes are presented in table 18. Most 

of these values are from literature reviews such as 

Radding and others (1977), Mabey and Mill (1978), 

Callahan and others (1979), Mabey and others (1982), 

Howard and others (1991), Mackay and others (1992a; 

1993), Kollig (1993; 1995), and Washington (1995). 

For some of the VOCs, values in the literature reviews 

differed slightly from values in the original references, 

presumably because of differences in the procedures 

used for extrapolation to environmental temperatures; 

in these cases, both values are presented in table 18. 

Most values are for a pH of 7 and a temperature of 

25°C, although Washington (1995) gives values for a 

pH of 5.6 and a temperature of 10°C. The pH of 5.6 

was chosen because this is the approximate pH of 

water in equilibrium with the atmospheric concentration of carbon dioxide. The temperature of 10°C 

was chosen as more representative of ground-water 

temperatures than the 25°C commonly used for 

environmental studies.

Numerical values of the hydrolysis half-lives 

were available for 30 target analytes, and hydrolysis 

was judged to be insignificant for 19 other target 

analytes, in many cases because the molecule 

contained no hydrolyzable groups. Information 

was not available for iso-propylbenzene, 1,2,3- 

trichlorobenzene, and the four ether compounds 

(table 1).

Estimation of the importance of hydrolysis 

for iso-propylbenzene and 1,2,3-trichlorobenzene 

can be made on the basis of available information 

for similar compounds. The alkyl benzene compounds 

methylbenzene, ethylbenzene, n-propylbenzene, and 

n-butylbenzene were judged to have no hydrolyzable 

groups (Mabey and others, 1982; Kollig, 1993; 

1995), and methylbenzene and ethylbenzene were 

considered to be recalcitrant with respect to hydrolysis 

(Washington, 1995). A similar conclusion seems 

reasonable for iso-propylbenzene. Hydrolysis for 

1,2,4-trichlorobenzene was judged to be not significant (Mabey and others, 1982; Kollig, 1993) or 

to have a half-life of 3.4 years at pH 7 and 25°C 

(Mackay and others, 1992a) as shown in table 18.
82 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

Table 18. Literature values of half-lives for hydrolysis of 30 target analytes and estimates of the potential for hydrolysis of 

19 target analytes

[USGS, U.S. Geological Survey; °C, degrees Celsius; to 5h, half-life for hydrolysis, in years; A, acidic conditions, pH not specified; NS, rot specified; 

NACM, no acid catalyzed mechanism known; B, basic conditions, pH not specified; HPHI, hydrolysis pH independent; HNES, hydrolysfs not environmentally significant; RC, recalcitrant toward hydrolysis; PNS, probably not significant; N, neutral pH condition; NLFG, no labile functional group; 

iNS, isomernot specified; NHFG, no hydrolyzable functional group]

USGS 

parameter Compound 

code

344 1 8 chloromethane

34423 dichloromethane

32106 trichloromethane

32102 tetrachloromethane

pH

7

7

7

7

7

A

5.6

5.6

B

NS

7

7

7

7

A

5.6

5.6

B

NS

7

7

7

7

7

7

A

5.6

5.6

NS

7

7

7

7

7

7

7

A

5.6

5.6

B

Temperature

(°C)

25

25

25

25

10

NS

25

10

NS

20

25

25

25

25

NS

25

10

NS

20

25

25

25

25

25

10

NS

25

10

25

25

25

25

25

25

25

10

NS

25

10

NS

tfl-Sh 

(years)

1.14x10°

9.3x1 0' 1

1.16x10°

3.58xl04

3.16xl05

NACM

9.00x1 05

7.93xl05

HPHI

2.5x10°

7.04x1 02

6.87xl02

6.93xl02

1.05xl07

NACM

2.65xl08

2.77xl09

HPHI

=1.5x10°

3.5xl03

3.16xl04

1.85xl03

6.93xl03

1.76xl03

l.SOxlO4

NACM

5.93xl03

7.85xl04

=1.25x10°

7.0xl03

(1 mg/L)

7.0x10°

(103 mg/L)

4.05X10 1

4.08X10 1

3.81X10 1

4.0x10*

4.41xl02

NACM

3-SlxlO 1

4.41xl02

HPHI

Reference

Radding and others (1977)

Mabey and Mill (1978)

Mabey and others (1982)

Washington (195 5)

Washington (195 5)

Mabey and others (1982)

Washington (195 5)

Washington (195 5)

Mabey and others (1982)

Callahan and others (1979)

Radding and others (1977)

Mabey and Mill (1978)

Kollig(1993)

Washington (1995)

Mabey and other? (1982)

Washington (1995)

Washington (1995)

Mabey and other? (1982)

Dilling and others (1975)

Radding and others (1977)

Mabey and other? (1982)

Jeffers and other? (1989)

Kollig(1993)

Washington (1 99 5)

Washington (1 99 5)

Mabey and other? (1982)

Washington (1 99 5)

Washington (1 99 5)

Dilling and other? (1975)

Radding and others (1977)

Radding and others (1977)

Jeffers and others (1989)

Kollig(1993)

Washington (1995)

Jeffers and others (1996)

Washington (1995)

Mabey and others (1982)

Washington (1995)

Washington (1995)

Mabey and others (1982)
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

83

Table 18. Literature values of half-lives for hydrolysis of 30 target analytes and estimates of the potential for hydrolysis of 

19 target analytes—Continued

[USGS, U.S. Geological Survey; °C, degrees Celsius; t0 5h, half-life for hydrolysis, in years; A, acidic conditions, pH not specified; NS, not specifiedNACM, no acid catalyzed mechanism known; B, basic conditions, pH not specified; HPHI, hydrolysis pH independent; HNES, hydrolysis not enviro" 

mentally significant; RC, recalcitrant toward hydrolysis; PNS, probably not significant; N, neutral pH condition; NLFG, no labile functional group; 

iNS, isomer not specified; NHFG, no hydrolyzable functional group]

USGS 

parameter Compound 

code

344 1 3 bromomethane

32104 tribromomethane

32101 bromodichloromethane

32105 chlorodibromomethane

34311 chloroethane

34496 1,1-dichloroethane

pH

7

7

7

A

B

NS

7

7

7

A

5.6

5.6

7

7

7

A

5.6

5.6

7

7

7

A

5.6

5.6

7

7

7

7

7

A

5.6

5.6

B

7

7

7

7

A

5.6

5.6

Temperature

(°C)

25

25

20

NS

NS

25

25

25

10

NS

25

10

25

25

10

NS

25

10

25

25

10

NS

25

10

25

25

25

25

10

NS

25

10

NS

25

25

25

10

NS

25

10

tfl-Sh 

(years)

5.5xlO~2

5.4xlO"2

1.04x10-'

NACM

HPHI

7.3x1 0~2

6.86xl02

6.90xl02

6.85xl03

HNES

1.73xl04

1.72xl05

1.37xl02

1.28xl02

1.26xl03

NACM

3.23xl03

3.18xl04

2.74xl02

2.77xl02

2.97xl03

NACM

6.96xl03

7.46x1 04

LlOxlO' 1

1.04X10" 1

2.54x10°

2.6x10°

2.72X10 1

NACM

2.54x10°

2.72xlO !

HPHI

6.88xl02

6.13X10 1

5.82X10 1

6.06x1 02

NACM

5.82X10 1

6.06xl02

Reference

Radding and others ( 1 977)

Howard and others ( 1 99 1 )

Howard and others ( 1 99 1 )

Mabey and others (1982)

Mabey and others (1982)

Mackay and others (1 993)

Radding and others (1977)

Washington (1995)

Washington (1995)

Mabey and others (1982)

Washington (1995)

Washington (1995)

Mabey and Mill (1978)

Washington (1995)

Washington (1995)

Mabey and others (1982)

Washington (1995)

Washington (1995)

Mabey and Mill (1978)

Washington (1995)

Washington (1995)

Mabey and others (1982)

Washington (1995)

Washington (1995)

Radding and others (1977)

Mabey and Mill (1978)

Washington (1995)

Jeffers and Wolfe ( 1 996)

Washington (1995)

Mabey and others (1982)

Washington (1995)

Washington (1995)

Mabey and others (1982)

Mabey and others (1982)

Jeffers and others (1989)

Washington (1995)

Washington (1995)

Mabey and others ( 1 982)

Washington (1995)

Washington (1995)
84 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

Table 18. Literature values of half-lives for hydrolysis of 30 target analytes and estimates of the potential for hydrolysis- of 

19 target analytes—Continued

[USGS, U.S. Geological Survey; °C, degrees Celsius; t0 5h, half-life for hydrolysis, in years; A, acidic conditions, pH not specified; NS, not specified; 

NACM, no acid catalyzed mechanism known; B, basic conditions, pH not specified; HPHI, hydrolysis pH independent; HNES, hydrolysis not environmentally significant; RC, recalcitrant toward hydrolysis; PNS, probably not significant; N, neutral pH condition; NLFG, no labile functional group; 

iNS, isomer not specified; NHFG, no hydrolyzable functional group]

USGS 

parameter Compound 

code

32103 1 ,2-dichloroethane

34506 1,1,1 -trichloroethane

345 1 1 1 ,1 ,2-trichloroethane

34396 hexachloroethane

77651 1 ,2-dibromoethane

PH

7

7

7

7

7

7

A

5.6

5.6

7

7

7

7

7

A

5.6

5.6

NS

NS

7

7

7

7

7

A

5.6

5.6

9

7

7

A

B

7

7

7

7

7

7

7.5

NS

Temperature

(°C)

25

25

25

25

25

10

NS

25

10

25

25

25

25

10

NS

25

10

25

25

25

25

25

25

10

NS

25

10

25

NS

25

NS

NS

25

25

25

25

25

25

25

25

*0.5h 

(years)

S.OxlO4

4.4x1 04

7.2xlO !

7-OOxlO 1

7.15x10'

6.44x1 02

NACM

7-OlxlO1

6.44x1 02

1.1x10°

7.3xlO~!

1.08x10°

8.26x1 0-1

1.03X10 1

NACM

8.26x1 0"1

1.03X10 1

S.OxlO"1

9.6x1 0~!

1.39xl02

2.54xl04

3.7X10 1

1.37xl02

9.04xl02

NACM

3.02xl03

2.13xl04

8.94x10°

HNES

RC

NACM

HNES

S.OxlO3

1.1x10°

1.4to2.7xlO"2

8.0x10°

6.4x10°

2.2x10°

2.5x10°

4.1x10°

Reference

Radding and ott ers (1977)

Mabey and others (1982)

Jeffers and othe-s (1989)

Washington (19?5)

Jeffers and Wol p<?( 1996)

Washington (19?5)

Mabey and others (1982)

Washington (19?5)

Washington (19°5)

Jeffers and othe-s (1989)

Mackay and others (1993)

Kollig(1993)

Washington (19°5)

Washington (19°>5)

Mabey and others (1982)

Washington (1995)

Washington (19°5)

Dilling and others (1975)

Haag and Mill (1988)

Jeffers and othe-s (1989)

Kollig(1993)

Mackay and others (1993)

Washington (1995)

Washington (1995)

Mabey and others (1982)

Washington (1995)

Washington (15 95)

Mackay and others (1993)

Mabey and others (1982)

Jeffers and Wolfe( 1996)

Mabey and others (1982)

Mabey and others (1982)

Radding and others (1977)

Kollig(1993)

Mackay and otters (1993)

Mackay and otHrs (1993)

Jeffers and Wolfe ( 1996)

Mackay and otHrs (1993)

Vogel and Rein hard (1986)

Haag and Mill (1988)
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

85

Table 18. Literature values of half-lives for hydrolysis of 30 target analytes and estimates of the potential for hydrolysis of 

19 target analytes—Continued

[USGS, U.S. Geological Survey; °C, degrees Celsius; to 5h, half-life for hydrolysis, in years; A, acidic conditions, pH not specified; NS, not specified; 

NACM, no acid catalyzed mechanism known; B, basic conditions, pH not specified; HPHI, hydrolysis pH independent; HNES, hydrolysis not environmentally significant; RC, recalcitrant toward hydrolysis; PNS, probably not significant; N, neutral pH condition; NLFG, no labile functional group; 

iNS, isomer not specified; NHFG, no hydrolyzable functional group]

USGS 

parameter 

code

34541

77443

82625

34488

34668

77652

39175

34501

77093

34546

Compound

1 ,2-dichloropropane

1 ,2,3-trichloropropane

1 ,2-dibromo-3-chloropropane

trichlorofluoromethane

dichlorodifluoromethane

1,1,2-trichloro- 1,2,2-

trifluoroethane

chloroethene

1,1-dichloroethene

cis- 1 ,2-dichloroethene

trans- 1 ,2-dichloroethene

pH

7

7

7-9

7

A

5.6

5.6

7

7-9

7

7

7

7

9

7

5.6

NS

N,A,B

7

7

N, A, B

NS

5.6

5.6

4.3-9.4

7

7

7

N, A, B

5.6

5.6

7

7

7

5.6

5.6

N,A, B

7

7

N, A,B

5.6

5.6

N,A,B

Temperature

(°C)

25

25

25

10

NS

25

10

25

25

25

25

25

15

25

25,10

25,10

NS

25

25

10

NS

25

25

10

NS

25

25

10

NS

25

10

25

25

10

25

10

25

25

10

NS

25

10

25

tfl-Sh 

(years)

l.SlxlO 1

1.66xlO ]

l.SSxlO 1

1.75xl02

NACM

1.66X10 1

1.75xl02

4.1X10 1

4.4x10'

3.8x10*

3.84xlO ]

1.73xl02

1.41xl02

3.8X10" 1

RC

RC

PNS

NLFG

>9.91xlO°

>1.07xl02

HNES

<1.0xl0 1

>9.91xlO°

>1.07xl02

PNS

1.2xl08

1.19xl08

1.93xl09

HNES

2.99xl09

4.85xl0 10

2.1xl0 10

2.06xl0 10

3.34xlO n

2.06xl0 12

8.40xl0 12

NLFG

2.06xl0 10

3.34xlOn

HNES

5.18xlOn

8.40xl0 12

NLFG

Reference

Kollig(1993)

Washington (1995)

Mackay and others ( 1 993 )

Washington (1995)

Mabey and others (1982)

Washington (1995)

Washington (1995)

Kollig(1993)

Mackay and others (1993)

Burlinson and others (1982)

Howard and others ( 1 99 1 )

Kollig(1993)

Burlinson and others (1982)

Howard and others ( 1 99 1 )

Washington (1995)

Washington (1995)

Callahan and others (1979)

Kollig(1993)

Washington (1995)

Washington (199 5)

Mabey and others (1982)

Hill and others (1976)

Washington (1995)

Washington (1995)

Callahan and others (1979)

Jeffers and others (1989)

Washington (1995)

Washington (1995)

Mabey and others (1982)

Washington (1995)

Washington (1995)

Jeffers and others (1989)

Washington (1995)

Washington (1995)

Washington (1995)

Washington (1995)

Kollig(1993)

Washington (1995)

Washington (1995)

Mabey and others (1982)

Washington (1995)

Washington (1995)

Kollig(1993)
86 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

Table 18. Literature values of half-lives for hydrolysis of 30 target analytes and estimates of the potential for hydrolysis of 

19 target analytes—Continued

[USGS, U.S. Geological Survey; °C, degrees Celsius; IQ 5h , half-life for hydrolysis, in years; A, acidic conditions, pH not specified; NS, not specified; 

NACM, no acid catalyzed mechanism known; B, basic conditions, pH not specified; HPHI, hydrolysis pH independent; HNES, hydrolysis not environmentally significant; RC, recalcitrant toward hydrolysis; PNS, probably not significant; N, neutral pH condition; NLFG, no labile functional group; 

iNS, isomer not specified; NHFG, no hydrolyzable functional group]

USGS 

parameter 

code

39180

34475

50002

iNS

34704

34699

39702

34030

77128

34696

34010

Compound

trichloroethene

tetrachloroethene

bromoethene

1 ,3-dichloropropene

cis-\ ,3-dichloropropene

trans- 1 ,3-dichloropropene

hexachlorobutadiene

benzene

styrene

naphthalene

methylbenzene

pH

7

7

7

7

N,A, B

5.6

5.6

NS

N, A, B

7

7

7

N,A, B

5.6

5.6

NS

N,A,B

NS

7

7

7

7

A

5.6

5.6

B

7

7

N,A,B

N,A, B

7

5.6

N,A,B

N,A,B

N,A,B

N,A,B

7

N,A, B

5.6

N,A, B

Temperature

25

25

25

10

NS

25

10

25

25

25

25

10

NS

25

10

25

25

NS

25

25

25

10

NS

25

10

NS

25

25

NS

25

25,10

25, 10

25

25

NS

25

25,10

NS

25,10

25

(years)

1.3xl06

1.66xl05

^IxlO5

2.08xl06

HNES

2.78xl05

3.61xl06

8.9x10''

NLFG

9.9xl08

9.56xl08

1.33xl010

HNES

2.40x1 0 10

3.34xlOn

7.3x10-'

NLFG

NHFG

1. SSxlO- 1

l.StoS.lxlO" 1

1.62xlO"2

1.38x10"'

NACM

1.62xlQ-2

1.38X10" 1

HPHI

1.73xlO-2

1.73xlO~2

HNES

NLFG

RC

RC

NHFG

NHFG

NHFG

NHFG

RC

NHFG

RC

NHFG

Reference

Jeffers and othe-s (1989)

Washington (1995)

Jeffers and Wol Ml 996)

Washington (1995)

Mabey and others (1982)

Washington (1995)

Washington (1995)

Dilling and others (1975)

Kollig(1993)

Jeffers and othe-s (1989)

Washington (IS 95)

Washington (IS 95)

Mabey and others (1982)

Washington (IS 95)

Washington (1995)

Dilling and others (1975)

Kollig(1993)

Howard and oil ers ( 1 99 1 )

Mabey and others (1982)

Mackay and oil ers (1993)

Washington (IS 95)

Washington (1S95)

Mabey and others (1982)

Washington (1995)

Washington (1995)

Mabey and others (1982)

Kollig(1993)

Kollig(1993)

Mabey and others ( 1 982)

Kollig(1993)

Washington (1995)

Washington (1995)

Kollig(1993)

Kollig(1993)

Mabey and others (1982)

Kollig(1993)

Washington (1995)

Mabey and others (1982)

Washington (1995)

Kollig(1993)
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

87

Table 18. Literature values of half-lives for hydrolysis of 30 target analytes and estimates of the potential for hydrolysis of 

19 target analytes—Continued

[USGS, U.S. Geological Survey; °C, degrees Celsius; t0 5h, half-life for hydrolysis, in years; A, acidic conditions, pH not specified; NS, not specified: 

NACM, no acid catalyzed mechanism known; B, basic conditions, pH not specified; HPHI, hydrolysis pH independent; HNES, hydrolysis not environmentally significant; RC, recalcitrant toward hydrolysis; PNS, probably not significant; N, neutral pH condition; NLFG, no labile functional group; 

iNS, isomer not specified; NHFG, no hydrolyzable functional group]

USGS 

parameter 

code

34371

77224

77342

77135

85795

85795

34301

34536

34566

34571

34551

34210

34215

Compound

ethylbenzene

n-pr opy Ib enzene

n-butylbenzene

1 ,2-dimethylbenzene

1 ,3-dimethylbenzene

1 ,4-dimethylbenzene

chlorobenzene

1 ,2-dichlorobenzene

1 ,3-dichlorobenzene

1 ,4-dichlorobenzene

1 ,2,4-trichlorobenzene

2-propenal

2-propenenitrile

PH

7

N, A, B

5.6

N, A,B

NS

NS

7

5.6

N, A,B

7

5.6

N, A, B

7

5.6

N, A,B

7

N,A, B

5.6

N, A,B

7

N, A,B

5.6

N, A,B

7

7

N, A,B

5.6

7

7

N, A,B

5.6

N, A,B

7

N, A,B

N, A,B

NS

N, A,B

7

5

9

N, A,B

Temperature

(°C)

25, 10

NS

25, 10

25

25

25

25, 10

25, 10

25

25, 10

25, 10

25

25, 10

25, 10

25

25, 10

25

25, 10

25

25, 10

25

25,10

25

25

25,10

25

25, 10

25

25,10

25

25, 10

25

25

NS

25

NS

NS

NS

NS

NS

NS

to.5h 

(years)

RC

NHFG

RC

NHFG

NHFG

NHFG

RC

RC

NHFG

RC

RC

NHFG

RC

RC

NHFG

RC

HNES

RC

NLFG

RC

HNES

RC

NLFG

>8.79xl02

RC

HNES

RC

>8.79xl02

RC

HNES

RC

NLFG

3.4x10°

HNES

NLFG

NHFG

NHFG

1.21xl03

1.88xl02

1.3xlO l

HNES

Reference

Washington (1995)

Mabey and others (1982)

Washington (1995)

Kollig(1993)

Kollig(1995)

Kollig(1995)

Washington (1995)

Washington (1995)

Kollig(1993)

Washington (1995)

Washington (1995)

Kollig(1993)

Washington (1995)

Washington (1995)

Kollig(1993)

Washington (199 5)

Mabey and others (1982)

Washington (199 5)

Kollig(1993)

Washington (1995)

Mabey and others (1982)

Washington (1995)

Kollig(1993)

Mackay and others (199 2 a)

Washington (199 5)

Mabey and others (1982)

Washington (1995)

Mackay and others (1992a)

Washington (1995)

Mabey and others (1982)

Washington (1995)

Kollig(1993)

Mackay and others (1992a)

Mabey and others (1982)

Kollig(1993)

Callahan and others (1979)

Mabey and others (1982)

Howard and others ( 1 99 1 )

Howard and others ( 1 99 1 )

Howard and others (1991)

Mabey and others (1982)
88 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

A similar conclusion seems reasonable for 1,2,3- 

trichlorobenzene. Finally, a literature review (Mackay 

and others, 1993) has a number of aliphatic ethers 

listed in its compilations but no specific information 

on the hydrolysis characteristics of any of these ether 

compounds. Consequently, it seems reasonable to 

assume that hydrolysis will not be important for the 

four ether target analytes. The fact that hydrolysis of 

ether compounds has not been studied also indicates 

that hydrolysis probably is not significant.

Most of the hydrolysis half-lives in table 18 are 

large. However, nine target analytes have half-lives of 

548 days or less, and hydrolysis of these compounds 

could be important in determining the fate of these 

compounds in some streams. Half-lives for some of 

these compounds vary by orders of magnitude from 

study to study (for example, chloromethane, dichloromethane, 1,2-dichloroethane, 1,1 -dichloroethene, 

and tetrachloroethene). Therefore, values for these 

compounds should be used with caution.

Hydrolysis half-lives in table 18 increased as 

expected as the temperature increased for every 

compound for which temperature data were available. 

Half-lives increased as the pH decreased for about 

60 percent of the compounds for which data were 

available; half-lives did not change for the other 

40 percent.

Table 18 indicates that 2-propenal is not subject 

to hydrolysis. However, opinions differ about this 

compound. Kollig (1993) stated that 2-propenal 

undergoes rapid hydrolysis through the addition of 

water across the double bond. Half-life for this neutral 

hydrolysis reaction was l.OxlO"9 years. Callahan and 

others (1979) stated, however, that this water addition 

reaction is not a hydrolysis process because the 

reaction is reversible, and they suggested that 

2-propenal contains no hydrolyzable groups suitable 

for environmental transformations, as indicated in 

table 18. Nordone and others (1996) determined 

overall half-lives of 1.16xlO~3 and 2.83xlO~3 years 

for the disappearance of 2-propenal from weedy and 

nonweedy canals by all processes. These results 

indicate that 2-propenal is not undergoing the rapid 

hydrolysis indicated by the short half-life reported by 

Kollig (1993).

There also is some disagreement regarding the 

effect of the hydrolysis process on 2-propenenitrile. 

Mabey and Mill (1978) indicated that nitriles should 

be considered hydrolytically inert under ordinary

conditions, whereas Schnoor and others (1987) 

indicated that nitriles are generally susceptible to 

hydrolysis. Half-lives in table 18 range from 13 years 

for basic hydrolysis to 1,210 years for neutral hydrolysis, indicating some hydrolysis but not complete 

recalcitrance.

ESTIMATION OF THE 

IMPORTANCE OF HYDROLYSIS

The importance of hydrolysis of the target 

analytes with respect to volatilization wa? estimated 

using a procedure similar to that used for estimating 

the importance of microbial degradation, with one 

difference. Equation 74 used for microbial degradation 

gives the fraction of the concentration remaining at 

distance L downstream. To demonstrate an alternative 

procedure, the fraction of the compound removed at 

distance L downstream was used for comparing the 

relative effects of hydrolysis and volatilization. The 

fraction-removed procedure is likely to be of most use 

to those interested in estimating the contributions of 

various processes determining the fate of VOCs in 

streams. The fraction-remaining procedure is likely to 

be of most use to individuals such as water managers 

who are interested in estimating the concentrations of 

VOCs remaining at various points downstream. In 

either case, the fractions are easily converged from one 

to the other.

Substituting the hydrolysis rate coefficient for 

the microbial degradation coefficient in equation 74, 

it can be shown that the fraction of the total concentration, Tf, removed by hydrolysis and volat; lization at 

distance L, is given by.

Tf = l-exp[-i(Kh (99)

Similarly, the fractions removed by hydrolysis, Hf, 

and volatilization, Vf, are given by

Hf = l-exp[-t(Kh )] (100)

and

Vf = l-exp[-t(Kv )] . (101)
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

89

These fractions are not additive, however, because 

of the exponential nature of equation 99. It can be 

shown that

Tf = Hf+Vf-(Hf)(Vf) . (102)

The importance of hydrolysis relative to volatilization was estimated for the nine target analytes 

with hydrolysis half-lives of 548 days or less for a 

temperature of 25°C and a pH of 7. First-order rate 

coefficients for hydrolysis were computed from the 

half-lives in table 18 and equation 97. The compound 

l,2-dibromo-3-chloropropane has a half-life of 

0.38 year at pH 9; however, this compound was not 

included because the half-lives at neutral pH were 

much longer than 548 days. The same three sets of 

hydraulic conditions used previously for estimating 

the importance of microbial degradation were used to 

obtain a range of volatilization coefficients.

Distances downstream of 100,000, 20,000, 

and 170 m were used for the hydraulic conditions 

corresponding to the O'Connor-Dobbins (1958), 

Churchill and others (1962), and Owens and others 

(1964) equations. This distance was varied so as to 

obtain approximately the same fractional loss by 

volatilization for each of the three sets of hydraulic 

conditions. The short distance required for the Owens 

and others equation and the long distance required for 

the O'Connor-Dobbins equation again indicates the 

importance of flow depth in determining the volatilization coefficient. Minimum hydrolysis half-lives 

were used for those compounds with the widely 

varying experimental coefficients discussed 

previously so as to maximize the effect of hydrolysis.

The results are presented in table 19. These 

results indicate that hydrolysis was significant for 

the deep, low-velocity stream for several of the 

target analytes. The fraction removed by hydrolysis 

exceeded the fraction removed by volatilization 

for bromomethane, 1,2-dibromoethane, and 1,3- 

dichloropropene. Hydrolysis also was significant 

for chloroethane. For the other compounds, fractions 

removed by hydrolysis were much smaller, ranging 

from 0.0207 for dichloromethane to 0.0610 for 1,1,1- 

trichlorethane and 1,1-dichloroethene, with the 

corresponding fractions removed by volatilization 

ranging from 0.346 for dichloromethane to 0.310 for 

1,1,1-triehloroethane. Consequently, hydrolysis has a 

measurable effect in a deep, low-velocity stream, even 

for a hydrolysis half-life as long as 548 days.

The results in table 19 indicate that hydrolysis 

was not significant for the shallow, medium-velocity 

stream. Similarly, hydrolysis was not important for 

the intermediate-depth, high-velocity stream, with 

the possible exception of 1,2-dibromoethane and 

1,3-dichloropropene, for which hydrolysis has some 

significance relative to volatilization.

This analysis indicates that hydrolysis of VOCs 

is most likely to be important in deep, low-velocity 

streams, where the first-order rate coefficient for 

hydrolysis is comparable to or larger than the volatilization coefficient. As the water velocity increases 

and/or the flow depth decreases, the volatilization 

coefficient increases, resulting in less effect of hydrolysis in the determination of the overall fate of VOCs in 

streams.

This analysis of the importance of hydrolysis 

of the target analytes should be considered with th^ 

following qualifications. Experimental values for the 

half-lives for most of the analytes for which hydrolysis 

was important vary considerably, in some cases, by 

orders of magnitude (table 18). Minimum half-lives 

were used in the analysis to maximize the contribution 

of hydrolysis. Consequently, these estimates should be 

considered as qualitative only, and each situation 

should be evaluated specifically to determine if 

hydrolysis is contributing to the determination of the 

fate of VOCs in streams.

AQUATIC PHOTOLYSIS

Photolysis of organic compounds in stream? is 

the reaction of the molecules of the compounds with 

sunlight. The molecules absorb light and are raised to 

an excited state. These excited molecules return to 

their ground states by photochemical degradation 

(photolysis) or by photophysical transitions such as 

fluorescence, quenching, radiationless transfer, and 

phosphorescence (Harris, 1990). The photolysis 

process must be rapid to compete favorably with 

photophysical processes, and it is estimated that 

more than 95 percent of the light absorbed results 

in photophysical deactivation rather than photolys;s 

(Harris, 1990). Photolysis generally results in conversion to simpler compounds rather than complete 

conversion to inorganic products such as carbon 

dioxide, water, salts, and oxides (Bedding and others, 

1983). In many cases, products of photolysis are 

the same as metabolites resulting from microbial 

degradation.
90 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

Table 19. Fractions removed by volatilization and hydrolysis and total fraction removed for nine target analytes with hydrolysis halflives of 548 days or less

[USGS, U.S. Geological Survey; Hyd, hydrolysis; V fract, fraction of the concentration removed by volatilization; H fract, fraction of the concentration 

removed by hydrolysis; T fract, total fraction of the concentration removed by both processes; U, water velocity; m/s, meters per second; Y, flow depth; 

m, meters; L, distance downstream; iNS, isomer not specified]

USGS 

parameter 

code

34418

34423

34413

34311

34506

77651

34501

34475

iNS

34418

34423

34413

34311

34506

77651

34501

34475

iNS

Hyd half-life Compound ,

(days)

O'Connor-Dobbins (1958);

chloromethane

dichloromethane

bromomethane

chloroethane

1,1,1 -trichloroethane

1 ,2-dibromoethane

1,1-dichloroethene

tetrachloroethene

1 ,3-dichloropropene

Churchill and others (1962)

chloromethane

dichloromethane

bromomethane

chloroethane

1,1,1 -trichloroethane

1 ,2-dibromoethane

1 , 1 -dichloroethene

tetrachloroethene

1 ,3-dichloropropene

U = 0.070 m/s;

339

548

20

38

182

5.1

182

266

5.5

; U = 1.52 m/s;

339

548

20

38

182

5.1

182

266

5.5

Owens and others (1964); U = 0.56 m/s;

34418

34423

34413

34311

34506

77651

34501

34475

iNS

chloromethane

dichloromethane

bromomethane

chloroethane

1,1,1 -trichloroethane

1 ,2-dibromoethane

1 , 1 -dichloroethene

tetrachloroethene

1 ,3-dichloropropene

339

548

20

38

182

5.1

182

266

5.5

V fract

Y = 10 m;L = 100,000m

0.377

.346

.373

.346

.310

.314

.332

.301

.311

Y = 1.5 m; L = 20,000 m

.390

.343

.380

.360

.323

.243

.348

.314

.282

Y = 0.12 m;L = 170m

.405

.318

.379

.378

.342

.148

.373

.332

.219

H fract

0.0332

.0207

.436

.260

.0610

.894

.0610

.0422

.876

.0003

.0002

.0053

.0028

.0006

.0205

.0006

.0004

.0190

.0000

.0000

.0001

.0001

.0000

.0005

.0000

.0000

.0004

T fract

0.398

.360

.646

.516

.352

.928

.373

.330

.914

.391

.343

.383

.362

.323

.259

.348

.314

.296

.405

.318

.379

.378

.342

.148

.373

.332

.219

Two types of photolysis can occur. Direct 

photolysis involves absorption of light by the 

molecules of the compound followed by chemical 

reaction (Zepp and Cline, 1977). Indirect photolysis 

or sensitized photolysis occurs when energy is 

transferred from energized molecules of one chemical 

species in the aquatic system to molecules of another 

species, followed by chemical reaction (Harris, 1990). 

This first species is called a photosensitizer (Bedding 

and others, 1983); compounds such as humic

substances, transition-metal ions, free radicals of 

organic molecules, peroxides, singlet oxygen, and 

ozone can serve as photosensitizers (Smith and others, 

1988). Chlorophyll also has been considered a 

photosensitizer (Bedding and others, 1983). Indirect 

photolysis has not been defined as completely as direct 

photolysis (Callahan and others, 1979; Baughman and 

Burns, 1980; Smith and others, 1988); consequently, it 

is difficult to separate the effect of these t^ro processes 

in natural aquatic systems.
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

91

Photolysis depends on the properties of the 

organic molecule of interest, the characteristics of the 

stream, and the input of sunlight (Smith and others, 

1988). Molecules most likely to undergo photolysis 

are those with an extended conjugated hydrocarbon 

system, such as a polycyclic aromatic hydrocarbon, 

and those with a carbonyl, azo, or nitro group (Harris, 

1990). Characteristics of the stream that are important 

include the presence of the photosensitizers, and the 

presence of quenching agents such as dissolved 

molecular oxygen and inorganic ions, which could 

change the yields of the photolysis reaction. 

Suspended-sediment concentrations and water depth 

are important considerations in determining light input 

to the water column, and these characteristics are 

discussed further below. The sunlight input is complex 

because it varies diurnally, seasonally, and with 

geographic location. This variation is discussed in 

more detail in the next section.

KINETICS OF THE AQUATIC 

PHOTOLYSIS PROCESS

f d[Cph] l 

The average rate of photolysis, } ~~^— r at

I J A,

wavelength X in a completely mixed body of water, 

is directly proportional to the rate of absorption of 

light by the molecules of the compound of interest 

(Zepp and Cline, 1977). The amount of light of 

wavelength X absorbed per unit time, 1^, by the water 

alone can be computed from the Beer-Lambert law in 

the form

I x = (103)

where

I0^ is the incident light intensity (photon/cm2/s)

on the water surface, 

a^ is the absorption coefficient of the water

(cm" 1 ), and

z is the pathlength of the light (cm). 

Light incident on the water surface consists 

of direct radiation and sky radiation, where sky 

radiation is that resulting from molecular and aerosol 

scattering of direct radiation (Smith and others, 1988). 

Consequently, the average rate of absorption of 

light of wavelength X per unit volume of water, Ia^ 

(photon/cm2/s), is given by (Zepp and Cline, 1977)

Y

(104)

P h

where 

I and are the incident light intensities 

(photon/cm2/s) at the water surface 

resulting from direct and scattered 

radiation, 

zd and zs are the corresponding pathlengths

(cm), and

Yph is the water depth (cm).

Zepp and Cline (1977) determined that the pathlergth 

for direct radiation, zd, is Yphsecant 0 where 0 is the 

angle of refraction of the light in the water. The angle 

of refraction is dependent on the angle of incidence, 

which for sunlight is the solar zenith angle. As this 

angle increases during the solar day, the degree of 

refraction increases. Thus, for light coming from t>e 

horizon where the solar zenith angle is greater than 

85°, the light is strongly refracted to an angle of 48° 

(Zepp and Cline, 1977). Because the secant of 48° is 

1.49, the pathlength for direct radiation, zd, will be less 

than or equal to 1 .49 Yph during the solar day. The 

pathlength for scattered radiation, zs, was shown to be 

equal to 1.20 Yph, assuming the index of refraction of 

water is 1.34 (Zepp and Cline, 1977).

The presence of the molecules of the organic 

compound of interest in the water changes the 

absorption coefficient from o^ for the water alone to 

ax, + 8^ [Cph] where e^ is the molar extinction coefficient of the organic compound (L/g mol/cm) and [Cpt,] 

is the molar concentration (g mol/L). The fraction 

of the light absorbed by the molecules of the organic 

compound of interest is e^ [Cph]/(a^ + e^ [Cph]). 

Because concentrations in environmental situations 

are generally low, it follows that the ejJCph] term is 

negligible with respect to o^, and the fraction can 

be approximated by e^ [Cph]/a^. Consequently, the 

average rate of light absorption of wavelength X by *)

the organic compound, IaX' (photon/cm /s), is

(105)

where

90 j is a conversion factor equaling 6.02x 10 

when the intensities are expressed in

0

photon/cm /s and concentration is in 

g mol/L.
92 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

Equation 105 can be written in the form

(106)

where

a pseudo first-order rate coefficient for 

photolysis given by

(107)

Equation 104 simplifies for two limiting cases. 

For deep flows where essentially all the incident 

light is absorbed and a^zj and a^zs are both 

greater than 2, equation 104 reduces to (Zepp and 

Cline, 1977)

(108) YP h

and equation 107 becomes

(109)

Equation 109 indicates that the photolysis rate coefficient is inversely proportional to the water depth for 

depths greater than the photic zone, where the photic 

zone is the depth to which the light penetrates. 

2. For shallow flows where a^zd and a^zs are 

both less than 0.02, the terms (1 minus 10) 

approximate as 2.303a^z, and equation 104 

simplifies to

2.303

Y

(110)

P h

and equation 107 becomes

2.303

JY. (Ill)

P h

The efficiency with which the absorbed light is 

used to produce products by the photolysis reaction is 

the reaction quantum yield, fa (Mill and others, 1982), 

where the quantum yield is in gram moles reacted per 

photon of light absorbed. Consequently,

^phX (112)

Equation 112 applies to a specific wavelength 

of light, X. Sunlight, however, is not of a constant 

wavelength but has contributions from the ultraviolet, 

visible, and infrared parts of the spectrum. About 

10 percent of the incident light energy is in the ultraviolet range and 45 percent is in each of tH visible 

and infrared ranges (Harris, 1990). A comparison 

of energy, wavelength, and frequency ranges with 

typical chemical bond dissociation energies indicates 

that only the electronic transitions corresponding 

to ultraviolet/visible absorption are energetic 

enough to result in chemical reactions (Harris, 

1990). Consequently, the wavelength range of 

importance for photolysis reactions is from 110 to 

750 nm. Within this range, wavelengths loss than 

300 nm are filtered out by the ozone layer, leaving an 

effective wavelength range from about 300 to 750 nm.

Because of this variation in wavelength, 

equation 112 must be integrated to obtain the overall 

photolysis rate, [ dt cph^ .Thus, 

Idt

d[Cph ] 

dt

where integration usually is over the wave'ength range 

from about 300 to 750 nm. Zepp and Cline (1977) 

indicated that the quantum yield, fa, is independent of 

wavelength, and Harris (1990) argued that the solar 

spectrum is sufficiently constant in its wavelength 

distribution so that the assumption of a constant 

quantum yield is reasonable for comparison of similar 

solar irradiation situations. However, Harris (1990) 

also noted that quantum yields determined in the 

laboratory at 254 nm are not equal to quartum yields 

at 300 nm, and in some cases, processes and products 

at these two wavelengths are different. Thus, there 

are arguments both for and against assuming that 

the quantum yield is independent of wave^ngth.

For the purpose of this discussion, the quantum 

yield was assumed independent of wavelength as was 

done by Zepp and Cline (1977). With this assumption 

and approximating the integral as a sum, equation 113 

becomes
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

93

df~ L-^CP^ZV • (114)

Equation 114 requires kpWl, and application of 

equations 109 and 111 for kph^ is complicated 

by the fact that the sunlight intensity at a specific 

wavelength, 1^, is a function of the time of day, 

the day of the year, and the geographic location as 

specified by the latitude. The sunlight decreases from 

midday to sunset, from summer to winter, and from 

points on the equator to higher and lower latitudes. To 

simplify calculations, Zepp and Cline (1977) provided 

tabular values of the light-intensity functions in 

equations 109 and 111.

The rate coefficients for the two limiting cases 

(eqs. 109 and 111) discussed previously are written 

(Zepp and Cline, 1977)

(115)

where

and

2.303

"phX

j

(116)

where

Z/L = Idx secant 9 + 1.21^. 

Values of \V\ and Z^ are presented by Zepp and 

Cline (1977) as a function of wavelength at 40° north 

latitude for midday and midseason of each of the four 

seasons. Equation 115 applies for deep depths where 

essentially all the incident light is absorbed and 

equation 116 applies to shallow depths in natural 

waters and up to 0.5 m in distilled water (Zepp and 

Cline, 1977). The actual depth to which equation 116 

applies depends on the absorption coefficient of the 

natural water and the degree of error that can be 

tolerated.

An analysis (Harris, 1990) of the W^ and 

the Z^ factors indicated that the summer and winter 

values differed by no more than a factor of four for 

wavelengths longer than 320 nm. Conversely, for

wavelengths in the 295- to 320-nm range where 

organic molecules are most likely to absorb light, 

the summer value was as much as 36 times larger 

than the winter value. This variation indicates that the 

assumption that the quantum yield is independent of 

wavelength used in deriving equation 114 may be less 

valid for seasonal comparisons.

Values of W^ and Z^ presented by Zepp and 

Cline (1977) are for clear-sky conditions and average 

ozone concentrations. Ozone concentrations, however, 

increase with latitude in the northern hemisphere and 

from fall to spring. Cloud cover can decrease the 

sunlight intensity in the 280- to 320-nm range by as 

much as 50 percent, and intensity increases 15 to 

20 percent for each kilometer increase in elevatior 

above sea level (Zepp and Cline, 1977). In the case of 

streams, shielding by the banks and vegetation could 

result in an additional restriction of the light reaching 

the water surface.

Absorption of the incident sunlight within the 

water column, indicated by the absorption coefficient 

a^, can occur as a result of the water itself or constituents within the water such as phytoplankton and 

organic degradation products. The transmission and 

attenuation of the light also depends on the sediment 

concentration. Suspended sediments in the water 

column increase the degree to which the incident light 

is attenuated and, therefore, decreases the depth to 

which the light penetrates. This effect is likely to 

be one of the most important factors in determining 

rates of photolysis of organic compounds in streams 

(Bedding and others, 1983). Reported results of 

the effect of suspended sediment are contradictory, 

however. One study reported that photolysis rates 

generally were more rapid in turbid water than in clear 

water, the explanation being that the diffusiveness of 

the light caused by the scattering increased the rates 

(Bedding and others, 1983). Conversely, a second 

study resulted in decreased photolysis rates, the 

explanation being that the sediment particles shielded 

the organic molecules from the light. Suspended 

sediments also directly affected photolysis in that 

higher rates of photolysis were determined for 

pesticides and polycyclic aromatic hydrocarbons 

sorbed on sediments compared with rates for the 

dissolved compounds (Bedding and others, 1983). In 

streams, a dynamic equilibrium will exist as sediments 

containing these sorbed compounds move into and out 

of the photic zone by the processes of deposition and 

resuspension.
94 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

The overall photolysis process can be 

described by

dt J T

(117)

where 

Kph is the first-order rate coefficient (day" 1 )

for photolysis. 

Comparing equations 117 and 114 indicates that

LP ia (118)

where

kph^ is given by equation 116 for shallow

flows and by equation 115 for deep flows. 

The corresponding half-life for photolysis, t0 5ph, is 

given by

o.sph = 0-693/Kph (119)

PHOTOLYSIS HALF-LIVES FOR 

THE TARGET ANALYTES

Very little information was found on the halflives for photolysis of the target analytes in aquatic 

systems. References such as Callahan and others

(1979), Mabey and others (1982), Harris (1990), 

Howard and others (1991), and Mackay and others 

(1992a; 1992b; 1993) contained numerical values of 

the half-life for only four target analytes. This lack of 

information is consistent with the previously discussed 

statement (Harris, 1990) that organic molecules 

having moderate-to-strong absorption in the greater 

than 290-nm wavelength range characteristic of 

sunlight generally have either an extended conjugated 

hydrocarbon structure such as the fused ring of 

polycyclic aromatic hydrocarbons or a functional 

group having an unsaturated bond such as a carbonyl, 

azo, or nitro group. Because few of the 55 target 

analytes (table 1) have such structures, photolysis in 

the water phase of the environment likely will not be 

important for most.

Values of experimental half-lives for aquatic 

photolysis of benzene, naphthalene, chlorobenzene, 

and 1,2,4-trichlorobenzene are presented in table 20. 

Also presented in table 20 are negative results from 

photolysis studies of four target analytes. nhese results 

are presented as time periods for which concentration 

changes as a result of photolysis were not observed; 

consequently, if photolysis of these compounds 

occurred, the half-lives must be longer than the 

specified times. Hill and others (1976) observed no 

direct photolysis of chloroethene, but they did observe 

rapid photolysis in the presence of photos^nsitizers 

such as acetone and hydrogen peroxide. TTiey noted 

that only high-energy sensitizers such as f:etone are

Table 20. Photolysis half-lives for nine target analytes; all values for the surface layer of water except the 5-meter value for naphthalene 

[USGS, U.S. Geological Survey; to 5ph, half-life for photolysis; oo, infinite half-life; >, greater than; m, meters]

USGS 

parameter 

code

34423

32106

34506

77652

39175

34030

34696

34301

34551

Compound

dichloromethane

trichloromethane

1,1,1 -trichloroethane

1,1 ,2-trichloro-l ,2,2-trifluoroethane

chloroethene

benzene

naphthalene

chlorobenzene

1 ,2,4-trichlorobenzene

to.5ph

(days)

>365

>365

>365

00

>3.8

117-673

3* (surface)

550* (5 m)

6.2x1 04**

1.6xl05 **

Reference

Dilling and others (1975)

Dilling and others (1975)

Dilling and others (1975)

Howard and others (1991)

Hill and others (1976)

Mackay and others (1992a)

Mackay and others (1992a)

Dullin and others (19F6)

Dullin and others (19F6)

*Midday summer values at 40 degrees north latitude. 

* Twenty-four-hour summer values at 40 degrees north latitude.
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

95

effective with chloroethene because of the high 

triplet-state energy of this compound. Because little 

is known about the concentration and distribution 

of high-energy sensitizers in the environment, the 

importance of this indirect photolysis process can 

not be estimated. Another complicating factor is 

the presence in the environment of competing energy 

acceptors such as dissolved oxygen. Photolysis of 

1,1,2-trichloro-1,2,2-trifluoroethane was not observed; 

thus, the half-life is considered to be infinite for this 

compound (Howard and others, 1991).

Finite values of the half-lives for benzene 

and chlorobenzene (table 20) contrast somewhat 

with the conclusion (Mabey and others, 1982) that 

aquatic photolysis was not significant for these 

compounds. One might argue, however, that a half-life 

of 6.2x104 days for chlorobenzene is insignificant for 

practical purposes. Half-lives for benzene were 

much shorter, and photolysis might be important 

for this compound. Photolysis of benzene is considered in more detail in the discussion of the relevant 

importance of photolysis and volatilization of 

naphthalene and benzene.

Photolysis was considered to be unimportant 

for 32 target analytes and an unspecified isomer 

of 1,3-dichloropropene. This assessment was based 

on the chemical structure and comparisons with 

similar types of molecular structures and the fact that 

some of the molecules contained no chromophores 

absorbing in the spectral range of sunlight (Callahan 

and others, 1979; Mabey and others, 1982; Howard 

and others, 1991). These target analytes are listed in 

table 21. Three of these are compounds listed in 

table 20 as having no observable change in concentration as a result of photolysis over a 365-day 

period (dichloromethane, trichloromethane, and 

1,1,1-trichloroethane) and one (chloroethene) over 

a period of 3.8 days. Hexachlorobutadiene has an 

absorption spectrum that extends into the sunlight 

spectrum; however, the absorption coefficient may not 

be large enough for direct photolysis to be important 

(Callahan and others, 1979).

No information was available on the photolysis characteristics in aquatic systems of 18 target 

analytes. These analytes are listed in table 22. 

Two ether compounds (ethyl tertiary-butyl ether 

and tertiary-amyl methyl ether) were not listed in 

any of the references cited previously. However, 

Harris (1990) pointed out that ethers do not absorb

Table 21 . Thirty-three target analytes for which aquatic 

photolysis was not considered significant

[USGS, U.S. Geological Survey; iNS, isomer not specified]

USGS

parameter

code

Compound Reference

34418 chloromethane

34423 dichloromethane

32106 trichloromethane

32102 tetrachloromethane

34413 bromomethane

32104 tribromomethane

32101 bromodichloromethane

32105 chlorodibromomethane

34311 chloroethane

34496 1,1 -dichloroethane

32103 1,2-dichloroethane

34506 1,1,1-trichloroethane

34511 1,1,2-trichloroethane

34396 hexachloroethane

34541 1,2-dichloropropane

34488 trichlorofluoromethane

34668 dichlorofluoromethane

39175 chloroethene

34501 1,1-dichloroethene

34546 /r<ms-l,2-dichloroethene

39180 trichloroethene

34475 tetrachloroethene

iNS 1,3-dichloropropene

39702 hexachlorobutadiene

34030 benzene

34010 methylbenzene

34371 ethylbenzene

34301 chlorobenzene

34536 1,2-dichlorobenzene

34566 1,3-dichlorobenzene

34571 1,4-dichlorobenzene

34210 2-propenal

34215 2-propenenitrile

Mabey and others (1982)

Callahan and others (1979) 

Mabey and others (1982)

Callahan and others (1979) 

Mabey and others (1982)

Mabey and others (1982) 

Howard and others (1991)

Mabey and others (1982) 

Mabey and others (1982) 

Mabey and others (1982) 

Mabey and others (1982) 

Mabey and others (1982) 

Mabey and others (1982) 

Mabey and others (1982)

Callahan and others (1979) 

Mabey and others (1982)

Callahan and others (1979) 

Mabey and others (1982)

Mabey and others (1982)

Callahan and others (1979) 

Mabey and others (1982)

Mabey and others (1982) 

Mabey and others (1982) 

Callahan and others (1979)

Callahan and others (1979) 

Mabey and others (1982)

Callahan and others (1979) 

Mabey and others (1982)

Callahan and others (1979) 

Mabey and others (1982)

Callahan and others (1979) 

Mabey and others (1982)

Mabey and others (1982)

Callahan and others (1979) 

Mabey and others (1982)

Mabey and others (1982) 

Mabey and others (1982) 

Mabey and others (1982) 

Mabey and others (1982) 

Mabey and others (1982) 

Mabey and others (1982) 

Mabey and others (1982) 

Mabey and others (1982) 

Mabey and others (1982)
96 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

light in the sunlight spectral range; therefore, photolysis in aquatic systems is not likely to be important 

for these compounds. It also seems reasonable to 

assume on the basis of information in tables 20 

and 21 that photolysis in aquatic systems will not 

be important for the halogenated alkanes and alkenes 

in table 22. Conversely, photolysis could be significant for the aromatic hydrocarbons, the alkyl 

benzenes, and the halogenated aromatics because 

these are the types of compounds for which photolysis 

occurs to some extent (table 20). Finally, Callahan 

and others (1979) noted that 2-propenenitrile does 

not absorb significantly in the sunlight spectral 

range. However, because of the conjugate double 

bonds of this molecule, they rationalized that 

indirect photolysis might be important in polluted 

streams where the appropriate photosensitizers are 

present.

Table 22. Eighteen target analytes for which no information 

was available on aquatic photolysis

[USGS, U.S. Geological Survey]

USGS 

parameter 

code

77651

77443

82625

77093

50002

34704

34699

77128

77224

77223

77342

77135

85795

85795

77222

77613

78032

81577

Compound

1 ,2-dibromoethane

1 ,2,3-trichloropropane

l,2-dibromo-3- 

chloropropane

cis- 1 ,2-dichloroethene

bromoethene

cis-\ ,3-dichloropropene

trans- 1 ,3-dichloropropene

styrene

n-propylbenzene

iso-propylbenzene

n-butylbenzene

1 ,2-dimethylbenzene

1 ,3-dimethylbenzene

1 ,4-dimethylbenzene

1 ,2,4-trimethylbenzene

1 ,2,3-trichlorobenzene

methyl tertiarybutyl ether

diisopropyl ether

Reference

Howard and others (1991)

Howard and others (1991)

Howard and others (1991)

Howard and others (1991)

Howard and others (1991)

Callahan and others (1979)

Callahan and others (1979)

Howard and others (1991)

Mackay and others (1992a)

Howard and others (1991)

Mackay and others (1992a)

Howard and others (1991)

Howard and others (1991)

Howard and others ( 1 991 )

Howard and others ( 1 99 1 )

Mackay and others (1992a)

Howard and others (1991)

Mackay and others (1993)

ESTIMATION OF THE IMPORTANCE 

OF AQUATIC PHOTOLYSIS

The results in table 20 indicate that photolysis 

in the aquatic phase is most likely to be important for 

naphthalene. The results in table 20 also indicate water 

depth has a strong influence on the importance of 

photolysis of naphthalene, with the effect of photolysis 

decreasing rapidly as the depth increases. The photolysis half-life for a 5-m depth was 183 times longer 

than the half-life for photolysis in the surface layer of 

the water. This increase in the half-life wes the result 

of light attenuation with increasing depth.

The comparison of the relative effects of 

photolysis and volatilization on the concentrations of 

naphthalene in streams must be done over a 24-hour 

period, or a multiple thereof, because photolysis 

ceases during the nighrtime. Volatilization, however, 

continues at virtually the same rate, except possibly 

for a small decrease as a result of lower water temperatures during the nighttime. The half-lives in table 20 

for naphthalene are midday summer values; hence, 

they correspond to the maximum photolysis rate. 

These values are not, however, applicable over a 

24-hour period because of the diurnal variation of 

the sunlight intensity.

To obtain a mean rate coefficient fcr photolysis over a 24-hour period, the half cycle sine wave 

approximation of the production of oxygen by 

photosynthesis (O'Connor and DiToro, 1970) was 

used to indicate the variation of the sunlight intensity 

with time over 24 hours. Adapted for photolysis, this 

equation has the form

where

is the rate coefficient for photolysis 

(day" 1 ) at time t (fractional day), 

kphm is the maximum photolysis rate coefficient

(day- 1 ),

p is the fraction of the 24-hour t;me period 

over which sunlight reaches the water 

surface, and 

ts is the time (fractional day) at which the

sunlight first reaches the water surface. 

A solar day from 0500 to 2100 hours was assumed, 

resulting in values of 0.208 for ts, 0.667 for p, and 0.875 

for the end of the solar day. The deep-flow half-life for
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

97

naphthalene in table 20 (550 days) corresponds to a 

photolysis rate coefficient of 0.00126 day" 1 , and this 

was used for kphm in equation 120. This maximum rate 

according to equation 120 occurred just past midday at 

0.542 fractional days.

Photolysis rate coefficients were computed from 

equation 120 as a function of time over the solar day, 

and an integral mean coefficient was computed from 

these values. The result for a depth of 500 cm was 

0.000798 day"1 , which is 37 percent of the midday 

value. For deep flows, equation 115 indicates that the 

photolysis rate coefficient is inversely proportional to 

the water depth, Yph- Consequently, it follows that

0.399 (121)

which gives a rate coefficient of 0.000798 day" 1 for a 

depth of 500 cm.

Equations 99 and 100 were modified by 

replacing the first-order rate coefficient for hydrolysis 

with the first-order rate coefficient for aquatic photolysis. These modified equations were then used with 

equation 101 and the O'Connor and Dobbins equation 

(eq. 55) to estimate the relative contributions of 

photolysis and volatilization to the fate of naphthalene 

in streams. Volatilization coefficients were computed 

as a function of depth from equations 55, 27, 28, 16, 

and 13 by using procedures discussed previously. An 

average air-film mass-transfer coefficient at 20°C of 

756 m/day and depths covering the range (0.27 to 

11.3 m) appropriate for the O'Connor-Dobbins 

equation (fig. 6) were used. The results indicated that 

the fraction of naphthalene removed by photolysis for 

all depths between 0.27 and 11.3 m was less than 

2 percent of the total removed by both processes.

The shallow approximation (eq. 110) requires 

that the factors a^zd and a^zs be less than 0.02. Zepp 

and Cline (1977) indicated that an a value of the order 

of 0.11 cm at a wavelength of 313 nm is appropriate 

as an average value for 10 streams hi the southeastern 

United States, and a value for pure water at this 

wavelength is 0.0020 cm" 1 . The corresponding 

maximum water depths for the z^ and zs parameters 

meeting the criterion are 0.15 and 0.12 cm for natural 

waters and 8.3 and 6.7 cm for pure water. Gas-transfer 

coefficients for such shallow flows in natural streams 

have not been measured, but large coefficients would 

be expected.

For comparison, the minimum flow depth of 

11.9 cm from the work of Owens and others (1964) 

was used as an example of a shallow flow. Again, the 

minimum water velocity of their data set (0.04 m/s) 

was used to minimize the contribution of volatilization. Estimating the volatilization coefficient for 

naphthalene for these conditions by using equation 53 

and procedures discussed previously resulted in a 

coefficient of 11.4 day" 1 . The integral mean coefficient for photolysis for the solar day corresponding to 

the midday shallow-flow half-life of 3 days (table 20) 

was 0.146 day" 1 . Using a 24-hour time period ftr 

volatilization and a 16-hour time period for photc lysis 

results in estimated fractions removed of 0.999999 for 

volatilization and 0.0928 for photolysis, assuming 

each process acting alone. The contribution of photolysis is minimal relative to volatilization, and it is 

expected that the volatilization coefficient would be 

even larger for the shallower flows for which the 

shallow approximation applies.

The previous analysis of the importance of 

photolysis of naphthalene was based on deep-flow 

and shallow-flow estimates of the photolysis halflives. The deep value is for a 5-m quiescent water 

column. Water in streams, however, is not quiescent 

but is subject to continual mixing as a result of shear 

forces at the bottom and banks of the stream. It was 

previously noted that concentration gradients of 

dissolved constituents over the vertical dimension 

(depth) in streams generally are not observed because 

of this mixing process. Consequently, water parcels 

are moving continuously to the surface where 

photolysis can occur hi the surface layers. It can be 

argued then that a more appropriate evaluation of the 

relative contributions of photolysis and volatilization 

would be to compare the shallow-flow photolysis 

process with the deep-flow volatilization process.

The results of this comparison are presented 

in figure 13 for naphthalene and benzene. A flowdepth range of 0.27 to 11.3 m appropriate for the 

O'Connor-Dobbins equation was used as in the 

previous example. The results are for a downstream 

distance of 5,010 m, which is the distance traveled in 

24 hours at a water velocity of 0.058 m/s. Figure 13 

indicates that the fraction removed by volatilization 

decreases rapidly as the depth increases. Conversely, 

with the assumption that photolysis occurs only in 

surface layers continually replaced with water frcm 

below, the fraction removed by photolysis is constant. 

Figure 13 indicates that the fractions removed by the
98 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

two processes are approximately equal for a flow 

depth of about 3 m for naphthalene and 5.7 m for 

benzene. For deeper flows, the fractions removed by 

photolysis exceed the fractions removed by volatilization, indicating that photolysis is significant for these 

two compounds for deep, low-velocity streams.

This analysis is based on the assumption that 

photolysis in the surface layers is not rate-limited by 

the movement of water parcels to the surface from the 

lower parts of the water depth and, thus, the results

1.0

0.8

Q 

LU

o

LU 

DC

0.6

g

Q0.4

CC

LL

0.2

0 

1.0

0.8

Q 

LU

O 

LU

cc

0.6

O 

CJ0.4

CC 

LL

0.2

NAPHTHALENE 

— VOLATILIZATION 

PHOTOLYSIS

BENZENE 

VOLATILIZATION 

PHOTOLYSIS

468

FLOW DEPTH, IN METERS

10 12

Figure 13. Fraction of naphthalene and benzene removed 

by volatilization and photolysis as a function of flow depth 

at 20 degrees Celsius; Henry's law constants are 36.6 and 

507 pascals cubic meter per gram mole for naphthalene and 

benzene, respectively.

should be considered as an upper bound on the effect 

of photolysis. The fact that vertical concentration 

gradients of dissolved constituents do not occur in 

streams would seem to argue in favor of no rate limitation. On the other hand, volatilization decreases as 

the flow depth increases, presumably because the 

thickness of the water film increases as th^ depth 

increases. However, the effect of this increased film 

thickness on photolysis in the surface layer and the 

rate of replacement of the photolyzed compound 

with compound from the deeper parts of the flow is 

unknown. Experimental verification of the effect of 

vertical mixing on photolysis is needed.

Bedding and others (1983) discussed the 

possibility that organic compounds might accumulate 

in the surface films of natural waters whe^e they 

would be subject to the maximum sunlight intensity. 

No experimental data were available to support this 

hypothesis, and they concluded that the mass of 

compounds present in these thin films is insignificant relative to the mass in the water column, thus 

resulting in an insignificant effect on the overall rate 

of photolysis. This conclusion, however, neglects 

the continual replacement in streams of the surface 

concentrations with concentrations of the compound 

from below.

The analysis of the photolysis characteristics of 

naphthalene was limited to a wavelength of 313 nm 

because this is the wavelength at which naphthalene 

molecules most strongly absorb (Harris, 1990). 

Actually, the contributions of the various sunlight 

wavelengths should be summed as indicated by 

equation 114, and such summation would probably 

result in lower rate coefficients because of lower 

contributions at the longer wavelengths. Also, the 

photolysis half-lives were for summer conditions, and 

the half-lives for the other three seasons of the year are 

likely to be longer. Consequently, the importance of 

photolysis relative to volatilization would be less than 

that indicated by the previous analyses. Finally, the 

summation indicated by equations 114, 115, and 116 

requires molar absorption coefficients as a function of 

wavelength, and such information is available for only 

a few polycyclic aromatic hydrocarbons and pesticides 

(Harris, 1990).

These estimates of the importance of photolysis 

in streams were based on a consideration of direct 

photolysis only. Zepp and others (1975), however, 

noted that the overall photolysis rate of an ester in a 

stream water was twice the rate observed in distilled
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

99

water, presumably because of the presence of photosensitizers in the natural water. Therefore, this 

possibility should always be evaluated when considering the transport and fate of the target analytes in 

streams, especially those analytes with an aromatic 

ring structure.

OXIDATION AND 

CHEMICAL REACTION

Oxidation is the reaction of an organic molecule 

with an oxidant that is a reactive electron-deficient 

species. The most common oxidants that might be 

significant in oxidation processes in environmental 

waters are the peroxy radical (RO2 »), photoexcited 

diradicals (RO»), the hydroxy radical (HO»), singlet 

oxygen (^2), and ozone (O3 ) (Mill, 1980). These 

oxidants apparently result from photolysis of natural 

humic materials present in the water (Mill, 1980). 

Molecular oxygen (©2) is rarely involved directly 

in reactions but is the ultimate electron sink 

in most environmental oxidation reactions (Mill 

and others, 1982).

Oxidation can be described by

= kox [C][OX] (122)

where 

[C]

t

is the molar concentration of the organic

compound of interest, 

is time, 

is the second-order rate coefficient for

oxidation, and 

is the molar concentration of the oxidant

(Mill and others, 1982).

The units of kox are usually M" 1 s" 1 , where M is 

the molar concentration. The total rate of oxidation 

is the sum of the rates of reaction for each of the 

oxidants, or

[OX]

d[C] = 

dt (123)

i= 1

If some type of average effective oxidant concentration can be determined for the time of interest, then 

[OX]j can be considered constant and equation 123 

can be integrated to give

[C]t = [C] 0 exp ,: t (124)

where

is the molar concentration at time t,

is the molar concentration at time zero, and

is the pseudo first-order rate coefficient for

oxidation equivalent to koxi [OX]j. 

The corresponding half-life for oxidation is

[C]0 k'

0X1

0.693

O.Soxi (125)

Use of an average oxidant concentration might 

be questioned because the oxidants apparently result 

from photolysis of natural humic materials (Mill, 

1980), and the sunlight intensity that drives photolysis 

varies diurnally. However, little is known about tH 

variation of the concentrations of these oxidants with 

time. Mabey and others (1984) indicated that concentrations of these oxidants in water averaged over 

several diurnal cycles for the spring and summer 

seasons are known only within a factor of 10.

Mabey and others (1984) estimated that the 

peroxy radical (RO2») has a concentration of 10~9 

to 10~10 M in natural waters. Consequently, only 

compounds with second-order rate coefficients larger 

than about 103 M" 1 s will oxidize in water at significant rates. Combining this rate coefficient with a 

peroxy radical concentration of 10~9 M gives a halflife of 8 days, and a pseudo first-order oxidation 

rate-coefficient of 10"6 s" 1 . Mabey and others (1984) 

concluded that aromatic amines and phenols are the two 

major classes of organic compounds reactive enough 

toward the peroxy radical for oxidation to be significant 

in natural waters. None of the target analytes (table 1), 

however, are these types of compounds. Similarly, Mill 

(1980), using rate coefficients from the literature for 

reaction with the peroxy radical, computed oxidation 

half-lives of 9x104 days for olefins and 8 days for 

polycyclic aromatic hydrocarbons. Because naphthalene (table 1) is a polycyclic aromatic hydrocarbon, 

oxidation of this target analyte might be important in 

some types of streams.

Estimated half-lives for oxidation by the 

photoexcited diradical (RO») using a radical concentration of 10~14 M were 73 years and 334 days
100 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

for alkanes and olefins, respectively (Mill, 1980). 

Oxidation by the hydroxy radical (HO«) generally 

is assumed to be negligible because the concentration 

of this radical is so low in natural waters (Mill, 1980). 

Conversely, in the atmosphere, the hydroxy radical 

is the only radical oxidant of importance. However, 

discussion of atmospheric processes is beyond the 

scope of this report.

Singlet oxygen (] O2) is an electronically excited 

form of oxygen that undergoes a wide variety of 

reactions with organic molecules; most of these 

reactions are not observed with triplet oxygen. Using 

a reported aquatic concentration of about 10~ 12 M 

and rate coefficients from the literature, estimated 

half-lives for oxidation by singlet oxygen are 

7.3 years, 8 days, and 19 hours for olefins, substituted olefins, and dienes, respectively (Mill, 1980). 

Consequently, oxidation by singlet oxygen may be 

significant for target analytes- (table 1) within these 

classifications, such as the halogenated alkenes and 

hexachlorobutadiene.

Significant concentrations of ozone have not 

been observed in natural waters. However, ozone 

is used to disinfect drinking water and wastewaters 

before discharge, and such use could result in direct 

oxidation of some organic compounds. Also, because 

ozone in water can form the hydroxy radical, oxidation 

of compounds resistant to ozone itself may occur 

(Mill, 1980) through oxidation by the hydroxy radical.

OXIDATION HALF-LIVES FOR 

THE TARGET ANALYTES

Experimental values of the half-lives for 

oxidation of 12 target analytes are presented in 

table 23 for low (5xlO~19 M) and high (2xlO" 17 M) 

concentrations of the hydroxy radical likely to be 

present in relatively oligotrophic and eutrophic natural 

waters. Half-lives from Haag and Yao (1992) were 

calculated from their experimental second-order rate 

coefficients for oxidation by the hydroxy radical and 

these expected low and high concentrations of the 

hydroxy radical by using equation 125.

It was previously suggested (Mill, 1980) that 

oxidation by the hydroxy radical is generally assumed 

to be negligible in natural waters because the concentration of the hydroxy radical is so low. The results 

listed in table 23 indicate that this is a valid assumption for the halogenated alkanes but that this assumption may not be valid for the aromatic hydrocarbons, 

the alkyl benzenes, and the halogenated aromatics

where many of the half-lives were less than 1 year for 

the possible high concentrations of hydrory radical in 

natural waters.

Calculated values of half-lives for the oxidation 

of 39 target analytes by the peroxy radical, singlet 

oxygen, and the hydroxy radical are presented in 

table 24. Concentrations of 10~9 M for the peroxy 1 *?

radical and 10 M for singlet oxygen were used as 

appropriate for surface waters under natural sunlight 

conditions (Mabey and others, 1982). Exceptions 

were the Howard and others (1991) results for 1,2-, 

1,3-, and 1,4-dimethylbenzene for which peroxy 

radical concentrations of 5xlO~10 and 10~u M were 

used. The lower and higher concentration;' of the 

hydroxy radical presented previously wer? used for 

this oxidant. Half-lives were computed from these 

concentrations, equation 125, and calculated values 

of the second-order rate coefficients for oxidation by 

the different oxidants.

Mabey and others (1982) used two procedures 

to calculate second-order rate coefficients for 

oxidation by the peroxy radical. The first consisted 

of using a measured rate coefficient available from the 

literature for a similar chemical structure. The second 

was based on structure-activity relationships for 

hydrogen-atom transfer and addition to double bonds. 

The latter procedure was used most frequently. 

Only a few target analytes are reactive toward singlet 

oxygen. When no reactive centers were noted, the 

compounds were assigned a rate coefficient of less 

than 360 M" 1 hour" 1 , which corresponds to a 

half-life of greater than 2.2x105 years. Reactive 

compounds were given rate coefficients on the 

basis of similar structures with known rate coefficients 

for reaction with singlet oxygen (Mabey rnd others, 

1982).

Values for the three dimethylbenzenes (table 24) 

were from scientific judgment based on estimated 

rate coefficients for alkyl peroxy radicals in aqueous 

solution (Howard and others, 1991). The calculated 

values of Haag and Yao (1992) were obtained using 

two procedures. Values for the halogenated alkanes 

were estimated from a correlation of experimental 

rate coefficients for oxidation by hydroxy radical in 

water as a function of the experimental rate coefficient 

for oxidation in the atmosphere. Chlorinated aromatics 

did not fit the same correlation; consequently, halflives for these compounds were estimated using the 

Hammett a constant (Haag and Yao, 1992).
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

Table 23. Experimental values of the half-lives for oxidation of 12 target analytes in water by the hydroxy radical 

[USGS, U.S. Geological Survey; t0 5ox, half-life for the oxidation process]

101

USGS

parameter

code

34423

32106

32104

34511

34541

82625

34030

34010

77223

77222

34301

34566

Oxidant

Compound concentration , °-5ox /yparsi Reference

(moles per liter)

dichloromethane

trichloromethane

tri bromomethane

1 , 1 ,2-trichloroethane

1 ,2-dichloropropane

1 ,2-dibromo-3-chloropropane

benzene

methylbenzene

iso-propylbenzene

1 ,2,4-trimethylbenzene

chlorobenzene

1 ,3-dichlorobenzene

The calculated values of the half-lives in table 24

are greater than 2 years, with the following exceptions.

The half-lives for oxidation by singlet oxygen of

chloroethene, 1, 1 -dichloroethene, 1 ,3-dichloropropene

(isomer not specified), and 2-propenenitrile were

estimated as greater than 0.79 year. The "greater

than" value results from the fact that the estimated rate

01 i

coefficients were listed as less than 10 M hour" by

2xlO" 17 5.0x10'

5xlO~ 19 2.0xl03

2xlO" 17 1.9x10'

5xlO~ 19 7.6xl02

2xlO~ 17 1.2x10'

5xlO~ 19 4.9xl02

2xlO~ 17 7.8x10'

5xlO~ 19 3.1xl03

2xlO~ 17 l.OxlO2

5xlO~19 4.0xl03

2xlO~ 17 2-OxlO 1

5xlO~ 19 8.1xl02

2xlO~ 17 8.4x10°

5xlO~ 19 3.4xl02

2xlO~ 17 l.OxlO 1

5xlO~ 19 4.0xl02

2xlO~17 8.4x10°

5xlO~ 19 3.4xl02

2xlO~ 17 2.9x10°

5xlO~ 19 1.2xl02

2xlO~ 17 3.4x10°

5xlO~ 19 1.4xl02

2xlO~ 17 2.6x10°

5xlO" 19 l.OxlO2

2xlO"17 9.2x10-'

5xlO" 19 3.7x10'

2xlO~17 3.6xlO~2

5xlO~ 19 1.5xlO~'

2xlO~ 17 3.7X10" 1

5xlO~ 19 1.5x10'

2xlO~ 17 1.2xlO~'

5xlO~ 19 4.9x10°

2xlO~ 17 l.SxlO" 1

5xlO~ 19 7.1x10°

2xlO~ 17 2.2xlO~'

5xlO~ 19 8.8x10°

Mabey and others (1982). The

H aag and Yao ( 1992)

Haag and Yao ( 1992)

H aag and Yao (1992)

Howard and others ( 1 99 1 )

Haag and Yao (1992)

Haag and Yao (1992)

Haag and Yao (1992)

Haag and Yao (1992)

Haag and Yao (1992)

Haag and Yao (1992)

Haag and Yao (1992)

Haag and Yao (1992)

Howard and others ( 1 99 1 )

Howard and others (1991 )

Howard and others (1991 )

Howard and others ( 1 99 1 )

Howard and others ( 1 991 )

Haag and Yao (1992)

half-lives for oxidaton

by the hydroxy radical at the maximum concentration

(2xlO~17 M) likely to be present in natural waters were

0.28 year for 1,2- and 1 ,4-dichlorobenzene and 1,23-

and 1,2,4-trichlorobenzene. Calculated half-lives for

naphthalene were greater than 10 years, in contrast to

the 8 days estimated previously (Mill, 1980) for polycyclic aromatic hydrocarbons.
102 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

Table 24. Calculated values of the half-lives for oxidation of 39 target analytes in water

[USGS, U.S. Geological Survey; t0 5ox , half-life for the oxidation process; 'O^, singlet oxygen at a concentration of 10 l2 moles per liter; RO^*, peroxy 

radical at a concentration of 10~9 moles per liter, except for the data of Howard and others (1991), which used high and low concentrations of 5x 10~ l ° and 

10 moles per liter; OH«, hydroxy radical at concentrations of 2x10 and 5x10 moles per liter; >, greater than; 00, infinite half-life; =, approximately 

equal to]

USGS 

parameter 

code

34418

34423

32106

32102

34413

32104

32101

32105

34311

34496

32103

34506

34511

34396

34541

82625

34488

Compound

chloromethane

dichloromethane

trichloromethane

tetrac hlorom ethane

bromomethane

tribromomethane

bromodichloromethane

chlorodibromomethane

chloroethane

1 , 1 -dichloroethane

1 ,2-dichloroethane

1,1,1 -trichloroethane

1 , 1 ,2-trichloroethane

hexachloroethane

1 ,2-dichloropropane

1 ,2-dibromo-3-chloropropane

tr ichlorofl uoromethane

Oxidant

'02

RO2 -

'of

RO2-

OHOH-

'02

ROT

'of

RO2 -

OHOHIO,

ROT

'of

ROT

'of

ROT

'of

ROT

'of

ROT

'of

ROT

'of

ROT

'of

RO2-

'02

RO2-

OHOH-

'02

ROT

'02

RO2 -

OHOH-

'02

RO2-

(years)

>2.2xl05

1.6xl06

>2.2xl05

4.0x 1 O5

9.2x10°

3.7xl02

>2.2xl05

l.lxlO5

>2.2xl05

>7.9xl04

>5.5xl02

>2.2xl04

>2.2xl05

7.9xl05

>2.2xl05

1.6xl05

>2.2xl05

4.0x1 05

>2.2xl05

1.6xl05

>2.2xl05

>7.9xl04

>2.2xl05

7.9x1 04

>2.2xl05

>7.9xl04

>2.2xl05

7.9x1 04

>2.2xl05

2.6x1 04

5.5x10°

2.2xl02

oo

oo

>2.2xl05

=7.9x1 04

3.8x10°

1.5xl02

oo

oo

Reference

Mabey and others (1982)

Mabey and others (1982)

Haagand Yao(19">2)

Mabey and others (1982)

Mabey and others (1982)

HaagandYao(19">2)

Mabey and others (1982)

Mabey and others (1982)

Mabey and others (1982)

Mabey and others (1982)

Mabey and others (1982)

Mabey and others (1982)

Mabey and others (1982)

Mabey and others (1982)

Mabey and others (1982)

HaagandYao(19?2)

Mabey and others (1982)

Mabey and others (1982)

Haag and Yao ( 19^2)

Mabey and others (1982)
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

103

Table 24. Calculated values of the half-lives for oxidation of 39 target analytes in water—Continued

[USGS, U.S. Geological Survey; t0 5ox, half-life for the oxidation process; 'O2 , singlet oxygen at a concentration of 10~ 12 moles per liter; RO2 v pero?-y 

radical at a concentration of I (T9 moles per liter, except for the data of Howard and others (1991), which used high and low concentrations of 5x 10~' ° and 

KT H moles per liter; OH«, hydroxy radical at concentrations of2xl(Tl7 and 5x10 19 moles per liter; >, greater than; oo, infinite half-life; s, approximately 

equal to]

USGS 

parameter 

code

34668

39175

34501

34546

39180

34475

iNS

34030

34696

34010

34371

77135

85795

85795

34301

34536

34566

34571

Compound

dichlorodifluoromethane

chloroethene

1,1-dichloroethene

trans- 1 ,2-dichloroethene

trichloroethene

tetrachloroethene

1 ,3-dichloropropene

benzene

naphthalene

methylbenzene

ethylbenzene

1 ,2-dimethylbenzene

1 ,3-dimethylbenzene

1 ,4-dim ethylbenzene

chlorobenzene

1 ,2-dichlorobenzene

1 ,3-dichlorobenzene

1 ,4-dichlorobenzene

Oxidant

'02

ROT

'02

R02 -

] 02

ROT

! 02

R02 -

'°2

ROT

] °2

ROT

"02

ROT

'°2

ROT

W

ROT

] 02

R02 -

'02

ROT

ROT

ROT

R02 .

RO2 -

R0r

ROT

! 02

ROr

'02

RO2 -

OHOHlQ2

ROT

W

RO2 -

OHOH-

(years)

00

00

>7.9xlO-'

2.6x1 04

>7.9xlO"'

2.6x1 04

>7.9xl02

1.3xl04

>7.9xl04

1.3xl04

>7.9xl05

1.3xl04

>7.9xlO~1

1.8xl03

>2.2xl05

>7.9xl04

>2.2xl05

>7.9xl04

>2.2xl05

5.5xl02

>2.2xl05

l.lxlO2

4.3xlOJ

3.1xl04

5.5xl02

2.7x1 04

3.1xl02

1.6xl04

>2.2xl05

>7.9xl04

>2.2xl05

>7.9xl04

2.8X10- 1

l.lxlO1

>2.2xl05

>7.9xl04

>2.2xl05

>7.9xl04

2.8x1 0" 1

l.lxlO1

Reference

Mabey and others (1982)

Mabey and others (1982)

Mabey and others (1982)

Mabey and others (1982)

Mabey and others (1982)

Mabey and others (1982)

Mabey and others (1982)

Mabey and others (1982)

Mabey and others (1982)

Mabey and others (1982)

Mabey and others (1982)

Howard and others (1991)

Howard and others ( 1 99 1 )

Howard and others (1991)

Mabey and others (1982)

Mabey and others (1982)

Haag and Yao (1992)

Mabey and others (1982)

Mabey and others (1982)

Haag and Yao (1992)
104 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

Table 24. Calculated values of the half-lives for oxidation of 39 target analytes in water—Continued

[USGS, U.S. Geological Survey; t0 5ox, half-life for the oxidation process; { O2 , singlet oxygen at a concentration of 10~12 moles per liter; FO2 «, peroxy 

radical at a concentration of 10~9 moles per liter, except for the data of Howard and others (1991), which used high and low concentrations of 5xlO~10 and 

10~'' moles per liter; OH«, hydroxy radical at concentrations of 2xlO~ 17 and 5xlO~ 19 moles per liter; >, greater than; °o, infinite half-life; =, approximately 

equal to]

USGS 

parameter 

code

77613

34551

34210

34215

Compound

1 ,2,3-trichlorobenzene

1 ,2,4-trichlorobenzene

2-propenal

2-propenenitrile

Oxidant

OHOH-

'02

RO2'

OH'

OHIo2

ROT

w

ROT

to.Sox 

(years)

2.8X10" 1

l.lxlO1

>2.2xl05

>7.9xl04

2.8x10-'

l.lxlO1

7.9x10°

2.3x10'

>7.9xlQ-'

2.2xl03

Reference

Haag and Yao ( 1992)

Mabey and others (1992)

Haag and Yao (1992)

Mabey and others (1982)

Mabey and others (1982)

ESTIMATION OF THE 

IMPORTANCE OF OXIDATION

Fourteen target analytes have experimental or 

calculated half-lives for oxidation of less than 1 year 

(tables 23 and 24). Four of these are calculated values 

of greater than 0.79 year for reaction with singlet 

oxygen (table 24). These values are included because 

present knowledge does not indicate how much larger 

than 0.79 year these half-lives might be. The halflives for these 14 analytes range from 0.036 year for 

methylbenzene to 0.92 year for benzene, both for the 

hydroxy radical.

The oxidants causing oxidation of the target 

analytes presumably result from photolysis of natural 

humic materials in the water. Consequently, the 

concentrations of the oxidants are dependent on the 

diurnal variation of the sunlight intensity. Also, it has 

been estimated (Mill, 1980) that these oxidants persist 

in natural waters for only a few milliseconds.

Estimation of the importance of oxidation of the 

target analytes in streams is similar to estimation of the 

importance of aquatic photolysis. The question is one of 

deep flows compared to shallow flows and whether or 

not vertical mixing in streams is rapid enough that the 

shallow-flow approximation applies to deep flows. For 

the purpose of this report, it was assumed that vertical 

mixing is rapid enough that penetration of the sunlight 

to the deeper depths is not limiting the extent of

oxidation, with oxidation occurring primarily in the 

surface layers that are continually replaced with water 

from the deeper flows by the vertical mixing processes. 

This approach will maximize the contribution of 

oxidation relative to volatilization.

The fraction of the concentration removed by 

oxidation was compared with the total fraction removed 

by volatilization and oxidation combined, using the 

same procedure as was used for estimating the 

importance of aquatic photolysis. Figure 14 shows the 

ratio of the fraction removed by oxidation to the total 

removed by both processes for four target analytes. 

The flow depth was varied from 0.27 to 11.3 m 

for a downstream distance of 5,010 m. This is the 

distance traveled by a water parcel in 24 hcurs at a 

water velocity of 0.058 m/s. Integral mean first-order 

rate coefficients for oxidation were computed from 

equation 125, using the half-lives from tabHs 23 

and 24.

Figure 14 indicates that the importance of 

oxidation relative to volatilization increases as the 

flow depth increases because of reduced volatilization 

at the increased depths, as was observed for other 

comparisons. The effect was greatest for methylbenzene, least for benzene, with intermediate effects for 

1,2,4-trimethylbenzene and 1,3-dichlorobenzene. 

These four compounds were selected as representative 

of the 14 target analytes with oxidation half-lives 

of less than 1 year, and the results for the other
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

105

10 analytes fall within the range of results shown 

in figure 14. The other 10 analytes were isopropylbenzene, chlorobenzene, 1,2- and 1,4- 

dichlorobenzene, 1,2,3- and 1,2,4-trichlorobenzene, 

chloroethene, 1,1-dichloroethene, 1,3-dichloropropene 

(isomer not specified), and 2-propenenitrile. The latter 

four analytes had fractions removed comparable to 

those for benzene shown in figure 14. The other six 

analytes had fractions removed comparable to those 

for 1,3-dichlorobenzene shown in figure 14, as would 

be expected.

These results should be considered as upper 

bounds on the extent of oxidation because no depth 

effect on the rate of replacement of the surface layer 

of water in which oxidation occurs was assumed. The 

fact that vertical concentration gradients of dissolved 

constituents do not occur in streams would seem to 

argue in favor of no rate limitation. On the other hand, 

volatilization decreases as the flow depth increases, 

presumably because the thickness of the water film 

increases as the depth increases. However, the effect 

of this increased film thickness on oxidation in the 

surface layer and the rate of replacement of the 

oxidized compound with compound from the deeper 

parts of the flow is unknown. Experimental verification of the effect of vertical mixing on oxidation is 

needed.

a 0.7 m

m 

a: o.e

ir 0.5

t: 0.4 

O

I

X0.3 

O

a

£0.2

O

———— METHYLBENZENE

.......... 1,2,4-TRIMETHYLBENZENE

_. __. 1,3-DICHLOROBENZENE

——— BENZENE

468 

FLOW DEPTH, IN METERS

Figure 14. Ratio of the fraction of four target analytes removed 

by oxidation to the fraction removed by oxidation and volatilization as a function of flow depth.

The results shown in figure 14 indicate that 

oxidation is significant relative to volatilization for 

methylbenzene, particularly for the deeper flows, 

whereas the effect of oxidation on benzene is minimal. 

The fractions removed for the other alkyl benzenes 

and halogenated aromatics are intermediate between 

these extremes. Thus, oxidation is negligible for most 

of the target analytes. However, in any assessment 

of the processes contributing to the determination cf 

the fate of VOCs in streams, all processes should be 

considered as possible during the initial stages of tie 

evaluation, with processes being eliminated only as it 

is determined that they are not significant for the VOC 

of interest.

CHEMICAL REACTION

The possibility of chemical reaction of the target 

analytes with other chemical compounds in streams' 

is remote, other than the hydrolysis, aquatic photolysis, and oxidation reactions previously discussed. 

The target analytes may undergo various chemical 

reactions, but these most likely would occur at temperatures, pressures, and in the presence of catalysts not 

normally observed in natural water bodies. There also 

is the possibility of chemical reaction occurring in the 

sorbed phase. As indicated in the sorption section, 

however, most of the target analytes have little 

tendency to sorb to sediments. Hence, contributionr 

of chemical reactions other than those previously 

discussed are considered negligible in determining the 

fate of VOCs in streams.

BIOCONCENTRATION AND 

AQUATIC TOXICOLOGY

Bioconcentration is the uptake of organic 

compounds from water by organisms within an 

aquatic system. The tendency of an organic compound 

to accumulate in an aquatic organism is indicated by 

the bioconcentration factor, BCF. This factor is 

defined (Bysshe, 1990) as follows:

„„„ riLr _ — —————————:——:———————concentration in organism

concentration in water

(126)

Units for the concentration in the organism are usually 

ng/kg wet weight, and water concentrations are 

usually in ng/L. BCFs are usually for fish, and with the 

assumption that 1 gram offish is equivalent to 1 mL of
106 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

water (Oliver and Niimi, 1985), the BCF is dimensionless. Values of the BCF can range from about 1 to 106 

(Bysshe, 1990). BCFs commonly are expressed on a 

logarithm base 10 basis. Consequently, log BCF varies 

from about 0 to 6.

Absorption of an organic compound from water 

by an organism involves a number of steps. These 

include movement of the chemical to the absorbing 

membrane (for example, the gill or stomach wall of 

a fish), diffusion through these membranes into the 

bloodstream, and transport by the blood to the lipid 

reservoir (Barren, 1990). Each of the steps can be ratelimiting and, thus, can control the uptake rate and the 

extent of accumulation.

The BCF depends on the species of organism, 

the body size, environmental factors, and the chemical 

properties of the organic compound (Barren, 1990). 

The most important chemical property is the lipophilicity of the compound because bioconcentration is a 

thermodynamically driven partitioning between the 

lipid phase of the organism and the surrounding water, 

with the final state that of equilibrium between the two 

phases.

Equation 126 is an equilibrium expression 

describing this partitioning of the organic compound 

between the lipid reservoir of the organism and the 

water of the aquatic system. However, unlike a watersolvent system where equilibrium is reached rapidly, 

equilibrium may be approached slowly because of the 

necessity for molecular diffusion of the compound 

into the lipid phase of the organism (Smith and others, 

1988). The time to reach equilibrium generally 

increases as the water solubility of the compound 

decreases and as the size of the organism increases. 

Failure to achieve equilibrium results in underestimated values of the BCFs for organisms initially 

exposed to high water concentrations of organic 

compounds. Conversely, failure to achieve equilibrium results in overestimated values of the BCFs for 

organisms previously exposed to high water concentrations and then moved to uncontaminated water.

Concentrations of the organic compound in the 

fish (eq. 126) are usually on a wet-weight basis of the 

total fish. Chiou (1985) and Thomann (1989) argued 

that a lipid-based concentration would eliminate a 

potential major source of variability in the wholebody concentrations because whole-body concentrations may vary with lipid content. Thomann (1989) 

noted, however, that not all the variability is eliminated 

by using lipid-based concentrations. Chiou (1985)

concluded that lipid-based BCFs for guppies and 

rainbow trout were basically identical when the 

uncertainties associated with the measurement 

of the BCFs were considered. Conversely, Sijm and 

van der Linde (1995) determined that lipid-normalized 

BCFs for six size classes of fathead minnow varied 

much more than the BCFs based on total f sh weight. 

Also, Barren (1990) cited several models showing 

only a weak relationship between organic-compound 

accumulation and lipid content of the organisms, and 

he concluded that the lipid content may not be as 

important as previously thought. C.T. Chiou 

(U.S. Geological Survey, written commun., 1996) 

indicated that the accumulation of relative1 y nonpolar 

organic compounds would be more closely related to 

the lipid content of the organism than the accumulation 

of more polar compounds. On the other hand, the more 

polar organic compounds would tend to accumulate in 

the more polar nonlipid materials of the organism such 

as the proteins (C.T. Chiou, U.S. Geological Survey, 

written commun., 1996). Both lipid-based and weightbased BCFs have been reported in the literature, with 

the lipid-based factors larger because the lipid content 

is only a small fraction of the total weight.

There are numerous factors that contribute to 

variability in measured values of BCFs. In laboratory 

studies, these factors include the time to reach 

equilibrium, which can vary substantially for different 

compounds and different species of organisms. Lipid 

content not only varies between species but can vary 

within a specific species, depending on growth stage 

and position in the reproductive cycle (Bysshe, 1990). 

Lipid content varies with the particular fish taken for 

analysis. A PCB concentration in fish was largest in the 

liver, with decreasing concentrations in the gills, whole 

fish, heart, brain, and muscle (Chiou, 1981). Water 

temperature and dissolved-oxygen concentration could 

affect the time required to reach equilibrium (Bysshe, 

1990). The water concentration also may contribute 

to variability. Oliver and Niimi (1983) found that 

the BCFs at a high-concentration level were about 

2.2 times larger than BCFs at a low-concentration level. 

Conversely, Smith and others (1988) cited two studies 

that observed no dependence on the water concentration. Barren (1990) indicated that the observation of 

Oliver and Niimi (1983) may have been trn result of 

saturation of the processes controlling uptrke, distribution, and excretion, causing deviations at t^e high 

concentrations from the first-order processes usually 

observed at low concentrations.
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

107

Three experimental procedures have been 

used to measure BCFs in the laboratory (Kenaga 

and Goring, 1980): (1) a flow-through system in 

which fish are exposed to a continuous flow of 

water containing the organic compound of interest, 

(2) a model ecosystem containing various organisms 

in quiescent water that contains the organic compound 

of interest, and (3) a terrestrial/aquatic model 

ecosystem containing various organisms in a quiescent 

soil/water combination. Data from flow-through 

systems are preferred because model ecosystem tests 

do not always represent the maximum bioconcentration potential (Bysshe, 1990). The concentration of the 

organic compound is likely to be much more variable 

in ecosystems than in flow-through systems because 

of losses from other processes such as volatilization, 

sorption to surfaces, biotransformation, or microbial 

degradation. Thus, while ecosystems were designed to 

mimic the real environment, results from flow-through 

systems are probably more reliable.

Many of the parameters causing variability 

in laboratory BCFs also result in variability in field 

BCFs, with the variability in the field values likely 

to be greater because of greater ranges in ambient 

conditions (Bysshe, 1990). Other factors, however, 

are of importance in the environment. Fish may avoid 

areas of contamination by moving to other parts of the 

aquatic system, while such avoidance is not possible in 

laboratory systems. Exposure depends also on whether 

the fish are surface or bottom dwelling, migratory, or 

remaining in a relatively small volume of the system 

(Bysshe, 1990).

Seasonal as well as daily variations in factors 

such as temperature, ionic strength, pH, dissolved 

oxygen, and other water-quality parameters could 

affect the bioconcentration process. Temperature 

effects have not been studied extensively. However, 

temperature affects the physiology and the biochemistry of aquatic organisms (Barren, 1990). Generally, 

increasing temperature increases the rates of uptake 

and excretion, with the effect on bioconcentration 

dependent on the relative effects of temperature 

on each of these two processes. Ionic strength also 

has not been studied extensively, and limited data 

are conflicting. One study reported no differences 

in bioconcentration between marine and freshwater 

fish, whereas a second study reported much higher 

BCFs for freshwater fish (Barren, 1990). The 

pH of the water appears to be important for ionizable 

compounds because it affects the concentration of the

nonionized form, which is absorbed faster that the 

ionized form (Barren, 1990). Sorption to dissolved 

organic matter may reduce the bioavailability of seme 

compounds; the effect is most likely to be important 

for the very hydrophobic compounds. Barron (1991) 

concluded that the effects of environmental conditions 

on bioconcentration are largely unpredictable. 

Finally, the type of source of the organic 

compound is important. The BCF (eq. 126) is an 

equilibrium parameter, and accurate determination 

of the parameter requires long-term exposure of th^ 

organism to a constant water concentration of the 

organic compound. With present-day controls on 

waste discharges, long-term constant concentrations 

of organic compounds in streams may be the 

exception rather than the general rule. Connolly 

and Pedersen (1988) noted that it may be invalid 

to compute a BCF on the basis of the present water 

concentration and present body burden of an 

organism when the organism has been exposed 

to a fluctuating concentration of the organic 

compound.

KINETICS OF THE 

BIOCONCENTRATION PROCESS

The uptake and excretion of an organic 

compound by a fish can be described by the equation 

(Oliver and Niimi, 1985)

dC (127)

where 

C and C are the concentrations of the organic

compound in the fish and in the water, 

ku is the rate coefficient for the uptake of the

compound by the fish, 

ke is the rate coefficient for excretion of the

compound by the fish, and 

t is time. 

At equilibrium, dCf/dt = 0, and it follows that

C

-

V_^,,

(128)

Equation 128 can be used to compute the BCF from 

a measurement of the concentration of the organic 

compound in the fish after exposure to water
108 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

containing a concentration of Cw of the target 

compound. The exposure time must be long enough 

to ensure equilibrium between the fish and the water.

A kinetic approach (Oliver and Niimi, 1985) 

also can be used. If a flow-through system is used such 

that the concentration of the organic compound in the 

water is constant, equation 127 can be integrated to 

give

Cf = (ku ) -r= [l-exp(-ke t)] (129)

The rate coefficient ke is determined by exposing the 

fish to contaminated water, transferring the fish to 

clean water, and measuring the excretion rate of the 

organic compound from the fish. The excretion rate 

coefficient is given by

0.693

L 0.5e

(130)

where 

t0 5e is the half-life for excretion of the organic

compound from the fish.

In a separate experiment, the uptake of the organic 

compound by the fish from the water is determined. 

According to equation 129, a plot of the fish concentration as a function of [1 minus exp(-ket)] should be 

linear with a slope of kuCw/ke . Because Cw and ke are 

known, ku can be computed and the BCF can be 

determined from equation 129 (Oliver and Niimi, 

1985).

The one compartment model defined by 

equation 127 is based on several assumptions 

(Barron, 1990). First, it is assumed that the uptake 

and excretion processes are first order such that the 

BCF is independent of the concentration of the organic 

compound. Second, it is assumed that uptake is limited 

only by diffusion processes in the various membranes 

and fluids, and the BCF is determined only by the 

lipophilicity of the organic compound and the lipid 

content of the organism. Finally, the model assumes 

the metabolism of the compound by the organism is 

negligible.

This latter assumption has been questioned by 

some who believe that organisms higher in the food 

chain could consume organisms from the bottom of 

the food chain that have bioconcenrrated some organic 

compound. This higher organism could then be

consumed by another organism still higher in the food 

chain, resulting in movement of the organic compound 

up the food chain. This process is called biomagnification. However, there has been considerable debate 

over whether or not biomagnification actually occurs 

and is significant. Smith and others (1988) cited 

several examples that indicated that biomrgnification 

is not well substantiated by existing data. Similarly, 

Bysshe (1990) cited several laboratory studies that 

indicated that absorption of organic compounds from 

water is far more significant than food ing^stion for 

bioconcentration by fish. However, in natural systems, 

this observation may not necessarily be tme because 

the ambient water concentrations are so low. Also, 

Connolly and Pedersen (1988) concluded that foodchain transfer can result in an increase in concentration with trophic level, with this effect significant for 

compounds with log Kow values greater than about 4. 

Thomann (1989) cited other examples supporting the 

contribution of food to the overall accumulation of 

organic compounds and presented a mode1 to account 

for this contribution.

The Thomann (1989) model is based on 

a four-level food chain involving phytoplankton, 

zooplankton, small fish, and the top predator. Foo.dchain accumulation is defined as the increase in the 

concentration of an organic compound in an organism 

compared with the concentration that would be 

expected from exposure to the water only. Primary 

parameters in the model are the rate coefficients for 

uptake and excretion and the assimilation efficiency 

factor for the organic compound. The modd also 

incorporates a growth factor that accounts for the 

dilution of the concentration of the compound in 

the lipid reservoir as a result of organism growth.

The rate coefficient for uptake indicates how 

fast the organic compound is taken up by the fish, and 

this information is important for short-term exposures 

such as might occur during a spill when eauilibrium is 

not attained (Sijm and van der Linde, 1995). The 

uptake coefficient also is important for long-term 

exposures to organic compounds with low water 

solubilities that approach equilibrium slowly. The 

uptake coefficient depends on properties of the 

organic compound such as the lipid partition coefficient, steric properties, and molecular weight 

(Thomann, 1989). For compounds with low octanolwater partition coefficients, transfer across the gill 

membranes of a fish is rapid through the a queous 

diffusion layer but is much slower through the lipid
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

109

membrane because of low solubility in the lipid. For 

compounds with larger octanol-water partition coefficients, the compound is more soluble in the lipid, 

resistance is reduced, and rate of movement is directly 

proportional to the octanol-water partition coefficient. 

For compounds with very large octanol-water partition 

coefficients, the water solubility of the compound 

limits rate of movement in the water, and chemical 

uptake is reduced. Thomann (1989) presented a plot of 

the efficiency of uptake as a function of the logarithm 

of the octanol-water partition coefficient (log Kow) for 

three fish weights. The efficiency increases with 

increasing log Kow up to a plateau, is constant for 

a range of log Kow values, and then decreases for 

log Kow values larger than about 6. The range of the 

plateaus varied, depending on fish weight. Also, the 

efficiency appears to be larger for the smaller fish with 

log Kow between 4 and 6.

The excretion coefficient indicates how fast 

the organic compound is depurated from the fish and, 

thus, is a measure of the biological half-life of the 

compound in the fish (Sijm and van der Linde, 1995). 

This half-life is defined by equation 130. Thomann 

(1989) assumed that the lipid-normalized BCF is equal 

to the octanol-water partition coefficient for equilibrium conditions with zero growth, where the lipidnormalized BCF is computed from equation 126 by 

using the weight of the lipid tissue rather than the total 

weight of the organism. With these assumptions, it 

follows from equation 127 that

(131) K,

A log-log plot of literature data for laboratory results shows reasonable agreement with 

equation 131, with an increase in slope occurring at a 

log Kow value of about 6. Thomann (1989) attributed 

this change in slope to a varying efficiency of release 

across the membranes of the organisms. Lower release 

rates were observed for the larger fish because of 

decreased respiration and a generally larger ratio 

of volume-to-membrane area. This difference was 

expected to cause about an order of magnitude 

decrease in the excretion coefficient for compounds 

with log Kow values in the 4 to 6 range. Under field 

conditions, the relationship is different because 

of the effect of growth. A plateau is reached for 

compounds with log Kow values larger than about 

6 because of the additional apparent loss resulting

from growth and dilution of the lipid reservoir. Weight 

of the fish was important for both the laboratory and 

field results.

Assimilation of the organic compound through 

ingestion of food involves transport across the 

membrane between the gastrointestinal tract and tH 

bloodstream. Thomann (1989) evaluated data from 

laboratory experiments and found an approximately 

constant assimilation efficiency for some compounds. 

For other compounds with log Kow values greater 

than about 6, the assimilation capacity decreased as 

log Kow increased.

Calibration of the model (Thomann, 1989) and 

application to various organic compounds with fish 

and squid as the organisms resulted in several general 

conclusions. A maximum field log BCF (lipid-based) 

of about 5.5 is expected for a compound with a 

log Kow of about 6. Larger values of the log Kow do 

not result in proportional increases in the log BCF 

because of growth and transfer efficiency effects. 

Measured concentrations in the upper trophic levels 

were significantly higher than those calculated using 

field BCF values for compounds with log Kow values 

in the range of 5 to 7. These higher concentrations 

were attributed to food chain magnification. Such 

magnification does not occur for compounds 

with log Kow values less than 5 because of decreased 

uptake and increased excretion.

The model of Thomann (1989) was discussed 

in some detail as an example of one type of model 

available for bioconcentration and biomagnification 

of organic compounds in organisms. Other models 

have been developed that consider different effects. 

Clark and others (1990) included size- and speciesdependent parameters in their model and also considered bioavailability. The presence of large concentrations of sorbents such as suspended sediments and 

humic materials reduces the bioavailability of the 

organic compounds. For organic compounds with 

log Kow values between 4 and 5, Clark and others 

(1990) estimated that most of the compound will b^ 

in the water. For a log Kow of 6, about one-half of the 

compound will be sorbed, and for a log Kow of 8, most 

of the compound will be sorbed. The concentration 

of the sorbent material is also important, as previously 

discussed in the sorption section of this report. The 

model of Sijm and others (1992) included biotransformation of the organic molecules as well as life stage, 

sex, reproduction, and growth characteristics of 

the fish. Main parameters in the model of Sijm and
110 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

van der Linde (1995) were gill exchange surface area 

and lipid content of the fish and the molecular weight 

and log Kow of the organic compound.

OCTANOL-WATER PARTITION 

COEFFICIENTS FOR THE TARGET ANALYTES

The previous discussion pointed out that the 

logarithm of the octanol-water partition coefficient 

is a good indicator in determining the possibility of 

bioconcentration. For this reason, the logarithms 

of the octanol-water partition coefficients of the target 

analytes are presented in tables 25 and 26. Table 25 

contains experimental values for 40 analytes from the 

compilations of Mackay and others (1992a; 1992b; 

1993). For those analytes for which more than one 

experimental value was available, an arithmetic 

average log Kow value was computed. This is equivalent to computing a geometric mean of the Kow values, 

as suggested by Lyman (1990b). Calculated values 

of the octanol-water partition coefficients for the 

remaining 15 target analytes are presented in table 26. 

These coefficients were calculated using structureactivity relationships and equations relating water 

solubility and the octanol-water partition coefficient. 

The octanol-water partition coefficient for bromoethene was estimated from its structure, as discussed 

previously in the section on the Henry's law constant. 

The coefficients for ethyl tertiary-butyl ether and 

tertiary-amyl methyl ether were estimated from a 

molar volume-log Kow relationship, as discussed 

previously in the section on sorption.

The logarithms of the octanol-water partition 

coefficients of the target analytes range from -0.92 for 

2-propenenitrile to 4.78 for hexachlorobutadiene, with 

a median of 2.34 for tribromomethane. The frequency 

distribution of the log Kow values is presented in 

figure 15.

The results in tables 25 and 26 and figure 15 

indicate that the log Kow values for the target analytes 

are generally small in comparison with values noted in 

the previous discussion of bioconcentration models. 

Only four analytes have values larger than 4. As 

indicated previously, Clark and others (1990) 

estimated that compounds with log Kow values 

between 4 and 5 will be mostly in the water, with little 

bioconcentration. However, while the bulk of the 

organic compound may remain in the water phase, 

organisms may still take up small amounts of organic 

compounds, and BCFs have been measured for a 

number of target analytes.

Table 25. Experimental values of the octanol-water partition 

coefficient for 40 target analytes (from Mackay and others, 1992a; 

1992b; 1993)

[USGS, U.S. Geological Survey; log Kow, logarithm base 10 of the octanolwater partition coefficient]

USGS

parameter 

code

34418

34423

32106

32102

34413

32101

32105

34311

34496

32103

34506

34511

34396

34541

34488

34668

77652

39180

34475

39702

34030

77128

34696

34010

34371

77224

77223

77342

77135

85795

85795

77222

34301

34536

34566

34571

77613

34551

78032

81577

Compound

chloromethane

dichloromethane

trichloromethane

tetrachloromethane

bromomethane

bromodichloromethane

chlorodibromomethane

chloroethane

1 , 1 -dichloroethane

1,2-dichloroethane

1,1,1 -trichloroethane

1 , 1 ,2-trichloroethane

hexachloroethane

1 ,2-dichloropropane

trichloro'fluoromethane

dichlorodifluoromethane

1 , 1 ,2-trichloro- 1 ,2,2-trifluoroethane

trichloroethene

tetrachloroethene

hexachlorobutadiene

benzene

styrene

naphthalene

methylbenzene

ethylbenzene

n-propylbenzene

iso-propylbenzene

n-butylbenzene

1 ,2-dimethylbenzene

1 ,3-dimethylbenzene

1 ,4-dimethylbenzene

1 ,2,4-trimethylbenzene

chlorobenzene

1 ,2-dichlorobenzene

1 ,3-dichlorobenzene

1 ,4-dichlorobenzene

1 ,2,3-trichlorobenzene

1 ,2,4-trichlorobenzene

methyl tertiary-butyl ether

diisopropyl ether

log Kow

0.91

1.25

2.00

2.71

1.19

2.10

2.24

1.45

1.86

1.47

2.48

2.38

3.93

2.00

2.53

2.16

3.16

2.43

2.64

4.78

2.17

2.94

3.35

2.71

3.14

3.66

3.58

4.29

3.15

3.23

3.20

3.78

2.84

3.46

3.55

3.46

4.07

4.04

.94

1.52
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

111

Table 26. Calculated values of the octanol-water partition coefficient for 15 target analytes 

[USGS, U.S. Geological Survey; log Kow, logarithm base 10 of the octanol-water partition coefficient]

uses

parameter 

code

Compound log Kow Reference

32104

77651

77443

82625

39175

34501

77093

34546

50002

34704

34699

50004

50005

34210

34215

tribromomethane

1,2-dibromoethane

1,2,3-trichloropropane

1,2-dibromo-3-chloropropane

chloroethene

1,1-dichloroethene

cis- 1,2-dichloroethene

trans-\ ,2-dichloroethene

bromoethene

cis- 1,3-dichloropropene

tram- 1,3-dichloropropene

ethyl tertiary-butyl ether

tertiary-amyl methyl ether

2-propenal

2-propenenitrile

2.34

1.60

2.36

2.26

1.33

2.00

1.68

1.84

1.52

2.20

2.20

1.52

1.52

-0.09

-0.92

Mackay and others (1993)

Radding and others (1977)

Mackay and others (1993)

Howard (1991)

Mackay and others (1993)

Mackay and others (1993)

Mackay and others (1993)

Mackay and others (1993)

This study

Kollig(1993)

Kollig(1993)

This study

This study

Callahan and others (1979)

Verschueren(1983)

16

CO 12 

LLJ

LU

s

ffi

CO

10

-1.0 -0.5 0.5 1.0 1.5 2.0 2.5 3.0 3.5 4.0 4.5

LOGARITHM BASE 10 OF THE OCTANOL-WATER PARTITION COEFFICIENT

5.0 5.5

Figure 15. Frequency distribution of the logarithm base 10 of the octanol-water partition coefficient for the target analytes.
112 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

BIOCONCENTRATION FACTORS 

FOR THE TARGET ANALYTES

Experimental values of the bioconcentration 

factor on a whole body basis are presented in table 27 

for 26 target analytes for several types of organisms. 

Log BCF values ranged from 0.20 for trichloromethane 

in bluegill sunfish to 4.23 for hexachlorobutadiene 

in rainbow trout. The large value for hexachlorobutadiene is consistent with this compound having the 

largest log Kow of any of the target analytes (tables 25 

and 26).

Most of the BCFs are for fish. However, 

several pure strains of alga and mixtures of algae 

were included, and these organisms generally had high 

BCFs. Algae have high BCFs because of the large 

surface-to-volume ratio relative to other organisms 

(Kenaga and Goring, 1980), which facilitates absorption of the organic compounds. Many algre also have 

high lipid contents which contribute to the high BCFs. 

This high lipid content may be more important than 

the surface-to-volume ratio at equilibrium conditions 

(C.T. Chiou, U.S. Geological Survey, written 

commun., 1996).

Table 27. Experimental values of the bioconcentration factor based on the whole body weight for various organisms for 2( target 

analytes

[USGS, U.S. Geological Survey; log BCF, logarithm base 10 of the bioconcentration factor]

USGS 

parameter Compound 

code

32106 trichloromethane

32 1 02 tetrachloromethane

32103 1,2-dichloroethane

34506 1,1,1-trichloroethane

34396 hexachloroethane

39175 chloroethene

39180 trichloroethene

34475 tetrachloroethene

Common 

name of 

organism

bluegill sunfish

bluegill sunfish

rainbow trout

largemouth bass

catfish

bluegill sunfish

green alga

bluegill sunfish

rainbow trout

bluegill sunfish

rainbow trout

green algae

fish

bluegill sunfish

bluegill sunfish

bluegill sunfish

bluegill sunfish

bluegill sunfish

rainbow trout

bluegill sunfish

fathead minnow

green alga

bluegill sunfish

rainbow trout

green alga

bluegill sunfish

rainbow trout

rainbow trout

rainbow trout

log BCF

0.78

0.20-0.40

0.52-1.01

0.46-0.49

0.52-0.57

0.78

2.84

1.48

1.24

1.48

1.72

2.48

1.24

0.30

0.30

0.95

0.95

2.14

2.71,3.08

2.14

2.85

1.6

1.23

1.59

3.06

1.69

1.59

2.06

1.79

Reference

Veith and others (1980)

Howard (1990)

Howard (1990)

Howard (1990)

Howard (1990)

Mackay and others (1993)

Mackay and others (1993)

Veith and others (1980)

Howard (1990)

Mackay and othe-s (1993)

Mackay and othe-s (1993)

Mackay and othe-s (1993)

Mackay and othe-s ( 1 993)

Veith and others (1980)

Howard (1990)

Veith and others (1980)

Howard (1990)

Veith and others (1980)

Oliver and Niimi ( 1983)

Mackay and othe-s (1993)

Mackay and othe-s (1993)

Mackay and othe-s (1993)

Veith and others (1980)

Mackay and othe-s (1993)

Mackay and othe-s (1993)

Veith and others (1980)

Mackay and othe-s (1993)

Mackay and othe-s (1993)

Mackay and othe-s (1993)
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

113

Table 27. Experimental values of the bioconcentration factor based on the whole body weight for various organisms for 26 target 

analytes—Continued

[USGS, U.S. Geological Survey; log BCF, logarithm base 10 of the bioconcentration factor]

USGS 

parameter 

code

39702

34030

77128

34696

34010

34371

77223

77135

85795

85795

34301

34536

Compound

hexachlorobutadiene

benzene

styrene

naphthalene

methylbenzene

ethylbenzene

iso-propylbenzene

1 ,2-dimethylbenzene

1 ,3-dimethylbenzene

1 ,4-dimethylbenzene

chlorobenzene

1 ,2-dichlorobenzene

Common 

name of 

organism

rainbow trout

Atlantic croaker

blue crab

spotted sea trout

blue catfish

Pacific herring

eels

green algae

goldfish

green alga

goldfish

mussel

mussel

water flea

green alga

bluegill sunfish

eels

Manila clam

mussel

goldfish

green alga

clams

goldfish

green alga

goldfish

eels

clams

goldfish

green alga

eels

clams

goldfish

green alga

eels

goldfish

green alga

Atlantic croaker

blue crab

spotted sea trout

blue catfish

fathead minnow

green algae

fish

bluegill sunfish

rainbow trout

log BCF

3.76, 4.23

2.84

1.66

2.42

3.07

0.64

0.54

1.48

0.63

1.63

1.13

1.64

1.49

2.12

2.11

2.50

1.12

0.22

0.62

0.92

1.99

0.67

1.19

2.31

1.55

1.33

0.79

1.15

2.34

1.37

0.78

1.17

2.40

1.37

1.17

2.41

2.09

2.06

2.36

1.96

2.65

1.70

1.88

1.95

2.43, 2.75

Reference

Oliver and Niimi (1983)

Pereira and others (1988)

Pereira and others ( 1 988)

Pereira and others (1988)

Pereira and others (1988)

Mackay and others ( 1992a)

Mackay and others (1992a)

Mackay and others ( 1992a)

Mackay and others ( 1 992a)

Mackay and others (1992a)

Mackay and others (1993)

Mackay and others ( 1 992V)

Mackay and others (1992b)

Mackay and others (1992V)

Mackay and others ( 1992V )

Mackay and others ( 1992V)

Mackay and others (1992a)

Mackay and others (1992a)

Mackay and others (1992a)

Mackay and others (1992a)

Mackay and others ( 1992a)

Mackay and others (1992a)

Mackay and others (1992a)

Mackay and others (1992a)

Mackay and others (1992 a)

Mackay and others (1992a)

Mackay and others (1992a)

Mackay and others (1992a)

Mackay and others (1992a)

Mackay and others (1992a)

Mackay and others (1992a)

Mackay and others (1992a)

Mackay and others (1992a)

Mackay and others (1992a)

Mackay and others (1992a)

Mackay and others (1992a)

Pereira and others (1988)

Pereira and others (1988)

Pereira and others (1988)

Pereira and others (1988)

Mackay and others (1992a)

Mackay and others ( 1992a)

Mackay and others (1992a)

Veith and others (1980)

Oliver and Niimi (1983)
114 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

Table 27. Experimental values of the bioconcentration factor based on the whole body weight for various organisms for 26 target 

anal ytes—Conti nued

[USGS, U.S. Geological Survey; log BCF, logarithm base 10 of the bioconcentration factor]

USGS 

parameter Compound 

code

34536 1 ,2-dichlorobenzene (Continued)

34566 1 ,3-dichlorobenzene

3457 1 1 ,4-dichlorobenzene

77613 1,2,3-trichlorobenzene

34551 1,2,4-trichlorobenzene

34210 2-propenal

34215 2-propenenitrile

Common 

name of 

organism

Atlantic croaker

blue crab

spotted sea trout

blue catfish

bluegill sunfish

rainbow trout

Atlantic croaker

blue crab

spotted sea trout

blue catfish

fathead minnow

bluegill sunfish

rainbow trout

rainbow trout

Atlantic croaker

blue crab

spotted sea trout

blue catfish

green algae

rainbow trout

fathead minnow

flagfish

rainbow trout

Atlantic croaker

blue crab

spotted sea trout

blue catfish

guppy

fathead minnow

rainbow trout

rainbow trout

Atlantic croaker

blue crab

spotted sea trout

blue catfish

fathead minnow

green sunfish

rainbow trout

bluegill sunfish

green algae

fish

fathead minnow

flagfish

bluegill sunfish

bluegill sunfish

log BCF

2.29

2.16

2.19

2.34

1.82

2.62, 2.87

1.94

1.56

1.63

1.92

1.99

1.78

2.57,2.86

2.71,2.95

2.25

2.23

2.45

2.03

2.00

2.33

2.04

2.47

3.08, 3.42

2.89

2.47

1.58

3.01

3.28

3.00-3.75

3.11,3.51

3.36, 3.57

3.10

2.60

1.91

3.19

3.32, 3.45

3.37

2.95

2.26

2.40

2.69

2.61

3.31

2.54

1.68

Referen TMe

Pereira and others (1988)

Pereira and others ( 1 988)

Pereira and others (1988)

Pereira and others ( 1 988)

Veith and others (1980)

Oliver and Niimi ( 1983)

Pereira and others (1988)

Pereira and others ( 1 988)

Pereira and others (1988)

Pereira and others (1988)

Mackay and others (1992a)

Veith and others (1980)

Oliver and Niimi (1983)

Oliver and Niimi (1985)

Pereira and others (1988)

Pereira and others ( 1 988 )

Pereira and others (1988)

Pereira and others (1988)

Mackay and others (1992a)

Mackay and others (1992a)

Mackay and others (1992a)

Mackay and others (1992a)

Oliver and Niimi (1983)

Pereira and others (1988)

Pereira and others (1988)

Pereira and others (1988)

Pereira and other? (1988)

Mackay and others ( 1 992a)

Sijm and van der Linde (1995)

Oliver and Niimi (1983)

Oliver and Niimi (1985)

Pereira and others (1988)

Pereira and others (1988)

Pereira and others (1988)

Pereira and other? (1988)

Mackay and others (1992a)

Mackay and others (1992a)

Mackay and others ( 1 992a)

Mackay and othe-s (1992a)

Mackay and othe-s ( 1 992a)

Mackay and othe-s (1992a)

Mackay and othe-s ( 1 992a)

Mackay and othe-s (1992a)

Veith and others (1980)

Veith and others (1980)
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

115

Experimental values of the bioconcentration 

factor on a lipid basis are presented in table 28 for 

seven target analytes for several types of organisms. 

Values in table 28 are higher than the corresponding 

values in table 27 because the lipid weight is only a 

small fraction of the whole body weight. Lipid 

fractions were from 0.071 to 0.112 for the rainbow

trout used in the study of Oliver and Niimi (1983), 

0.050 to 0.080 for the rainbow trout used in the study 

of Oliver and Niimi (1985), 0.005 for the blue crab to 

0.033 for the blue catfish used in the study of Pereira 

and others (1988), and from 0.0147 to 0.238 for the six 

sizes of fathead minnows used in the study of Sijm and 

van derLinde( 1995).

Table 28. Experimental values of the bioconcentration factor based on the lipid weight for various organisms for seven target analytes 

[USGS, U.S. Geological Survey; log BCF, logarithm base 10 of the bioconcentration factor]

USGS 

parameter Compound 

code

34396 hexachloroethane

39702 hexachlorobutadiene

34536 1,2-dichlorobenzene

34566 1,3-dichlorobenzene

34571 1 ,4-dichlorobenzene

77613 1,2,3-trichlorobenzene

34551 1,2,4-trichlorobenzene

Common 

name of 

organism

rainbow trout

rainbow trout

rainbow trout

Atlantic croaker

blue crab

spotted sea trout

blue catfish

rainbow trout

Atlantic croaker

blue crab

spotted sea trout

blue catfish

rainbow trout

rainbow trout

Atlantic croaker

blue crab

spotted sea trout

blue catfish

guppy

flagfish

rainbow trout

Atlantic croaker

blue crab

spotted sea trout

blue catfish

guppy

fathead minnow

rainbow trout

rainbow trout

Atlantic croaker

blue crab

spotted sea trout

blue catfish

flagfish

log BCF

3.77, 4.09

4.84, 5.30

3.50, 3.81

3.74

4.36

4.00

3.44

3.70, 3.93

3.60

3.86

3.27

3.40

3.65, 3.92

3.88,4.11

3.91

4.53

4.08

3.51

3.26

3.56

4.17,4.48

4.54

4.77

3.22

4.49

4.11

3.94-5.39

4.20, 4.57

4.54, 4.71

4.76

4.90

3.54

4.68

4.25

Reference

Oliver and Niimi (1983)

Oliver and Niimi (1983)

Oliver and Niimi (1983)

Pereira and others (1988)

Pereira and others (1988)

Pereira and others (1988)

Pereira and others (1988)

Oliver and Niimi (1983)

Pereira and others ( 1 988)

Pereira and others (1988)

Pereira and others (1988)

Pereira and others (1988)

Oliver and Niimi (1983)

Oliver and Niimi (1985)

Pereira and others (1988)

Pereira and others ( 1 988)

Pereira and others (1988)

Pereira and others (1988)

Mackay and others (1992a)

Mackay and others (1992a)

Oliver and Niimi (1983)

Pereira and others (1988)

Pereira and others (1988)

Pereira and others (1988)

Pereira and others (1988)

Mackay and others (1992a)

Sijm and van der Linde (1 995)

Oliver and Niimi (1983)

Oliver and Niimi (1985)

Pereira and others (1988)

Pereira and others (1988)

Pereira and others (1988)

Pereira and others (1988)

Mackay and others (1992a)
116 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

The objective of using the lipid weight rather 

than the whole body weight to compute BCFs is to 

reduce the variability in the computed BCFs. Results 

of a comparison of the data from the four studies cited 

above are presented in table 29. In three of the four 

studies, the relative standard deviation used as an 

indicator of variability was lower for the lipidbased BCFs, whereas in the fourth study, it was 

higher.

Table 29. Comparison of whole-weight and lipid-weight 

bioconcentration factors

[n, number of values; Cv , relative standard deviation, in percent]

Basis

Whole

Lipid

Whole

Lipid

Whole

Lipid

Whole

Lipid

n

14

14

4

4

28

28

6

6

Cv 

(percent)

±16.0

±11.7

±11.0

±7.73

±20.2

±12.0

±7.26

±12.2

Reference

Oliver and Niimi ( 1983)

Oliver and Niimi ( 1983)

Oliver and Niimi (1985)

Oliver and Niimi (1985)

Pereira and others (1988)

Pereira and others (1988)

Sijm and van der Linde ( 1 995)

Sijm and van der Linde (1995)

Comparisons in table 29 were for different types 

of studies. The studies of Oliver and Niimi (1983; 

1985) were for seven target analytes and two target 

analytes, respectively, at two levels of water concentration (tables 27 and 28). Results for the two concentrations were comparable. The study of Pereira and others 

(1988) was for seven target analytes and four organisms 

(tables 27 and 28). The average relative standard 

deviation across the organisms was larger than the 

average across the analytes for the weight-based BCFs 

(±17.8 percent compared to ±12.8 percent) and was 

approximately the same for the lipid-based BCFs 

(±10.1 percent compared to ±9.02 percent). The 

study of Sijm and van der Linde (1995) involved one 

target analyte and six size classes of fathead minnow 

(tables 27 and 28), with the lipid content generally 

decreasing as the age and size of the minnows 

increased. Results in table 29 indicate that the 

variability of the BCFs for the target analytes is 

reduced somewhat when based on the lipid content; 

however, considerable variability remained.

PREDICTION EQUATIONS FOR 

THE BIOCONCENTRATION FAC TOR

Various equations are available in tin literature 

for predicting BCFs for those target analytes for which 

experimental values are not available. BCFs have been 

related to the water solubility, the soil sorp+ion coefficient, and the octanol-water partition coefficient 

(Kenaga and Goring, 1980). The octanol-water partition 

coefficient is the most logical choice for a correlating 

parameter because it describes the partitioning of an 

organic compound between the octanol and water 

phases, which closely mimics the partitioning of the 

same organic compound between the water and lipid 

phase of an aquatic organism. Relationship^1 between 

the BCF and water solubility and the soil sorption 

coefficient follow indirectly because of the relationships of these two parameters to the octanol-water 

partition coefficient.

Bysshe (1990) summarized nine equations for 

estimating the BCF on a wet-weight basis. Six of these 

equations have the octanol-water partition coefficient 

as the independent variable, two have the water 

solubility, and one has the soil-sorption cc efficient. 

The equation best suited for the target analytes is

log BCF = 0.76 log Kow -0.23 (132)

which was based on 84 organic compounds of which 

15 were target analytes. Equation 132 was based on 

data for four species offish and had a correlation coefficient of 0.823. Some of the other equations had larger 

correlation coefficients (Bysshe, 1990). However, 

these equations were based on much smaller numbers 

of compounds than the 84 compounds on which 

equation 132 was based. The numbers of target analytes 

also were much smaller than the 15 target analytes 

included in the data base used for equation 132. 

Equation 132 can be used with the octanol-water 

partition coefficients from tables 25 and 26 to predict 

the BCF for any of the target analytes. If other VOCs 

are important in some situations, then BCFs can be 

estimated from equation 132 and octanol-water 

partition coefficients from the literature.

ESTIMATION OF THE 

IMPORTANCE OF BIOCONCENTRATION

The previous discussion of bioconcentration 

models illustrated the complexity of the process. This 

discussion also indicated that bioconcentration is most
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

117

likely to be important for high-molecular-weight, lowsolubility compounds that persist for long time periods 

in surface waters. Because of the volatile nature of the 

target analytes of the present study, bioconcentration 

is less likely to be significant for these compounds.

The importance of bioconcentration in 

determining the fate of the target analytes in streams 

was estimated by combining a mass balance of an 

organic compound in a compartment of water with 

the defining equation (eq. 126) for the BCF. Assuming 

that bioconcentration is the only process affecting 

the concentration of the organic compound in the 

compartment, and also assuming that the water and 

organism phases are in equilibrium, it follows that

= 0

(BCF)(Wf) 

+ (BCF)(Wf)] (133)

where

BCF is on a wet-weight basis,

f0 is the fraction by weight of the organic

compound in the organism, and 

Wf is the fraction by weight (gram of 

organism/gram of water) of the 

organism in the water.

Most BCF studies have been for fish (tables 27 and 

28); consequently, estimation of the importance of 

bioconcentration will be for fish.

Application of equation 133 requires a value of 

Wf, which is basically the density offish in the water. 

This density can vary widely, depending on conditions 

for productivity in the water. Neely and Mackay (1982) 

reported an areal density offish ranging from 5.6 to 

17 g/m2 for natural lakes of glacial origin. Using a value 

of 10, which is near the middle of this range, gives

x-5

(134)

where

Y is the water depth (m). 

For a water depth of 10 m as used by Neely and 

Mackay (1982), equation 134 indicates a fish density 

of 10 , and the same value was used later (Mackay 

and others, 1992a; 1992b; 1993) in their hypothetical 

terrestrial/aquatic ecosystem model. A value of 10~6 

also was used by McCall and others (1983). Because it 

is customary to assume that 1 g offish is equal to 1 mL 

offish (Oliver and Niimi, 1985), this weight fraction 

is identical to the volume fraction for the fish in the

water. Fractions of the order of 10 6 also have been 

used in experimental studies. Boyle and others (1985) 

used fractions of 1.4x 10~6 and 1.1 x 10~5 for bass 

and bluegill in a pond study of bioconcentration 

of a polycyclic aromatic compound.

The fish densities cited previously were areal 

densities and, thus, were per unit of surface area. 

Consequently, it is not known how the fish density 

varies with depth. An areal definition implies that the 

factor varies inversely with depth, as indicated by 

equation 134, although it also could imply a density 

independent of depth. To encompass a range of 

conditions, the importance of bioconcentration wa^ 

estimated for both a fish density that decreased with 

depth, as indicated by equation 134, and a fish density 

independent of depth. Several values of the base fish 

density also were evaluated.

Hexachlorobutadiene was chosen as the target 

analyte for evaluation because it has the largest 

octanol-water partition coefficient (log Kow = 4.78) 

of any of the analytes (tables 25 and 26) and, 

therefore, should exhibit the greatest tendency to 

bioconcentrate. The fraction removed by bioconcentration was computed from equation 133 for constant 

fish fractions (Wf) of 10~5 , 10~6 , and 10~7 and for fish 

fractions varying from these nominal values with the 

flow depth as indicated by equation 134. Flow depths 

ranged from 0.27 to 11 m and were for a constant 

water velocity of 0.058 m/s. Volatilization coefficients 

were computed from the O'Connor and Dobbins 

equation (eq. 55) and equations 27, 28, 16, and 13 

using procedures described in the volatilization 

section of this report. An average air-film masstransfer coefficient at 20°C of 756 m/day and a 

downstream distance of 5,010 m were used. 

This is the distance traveled during a 1.0-day 

period at a velocity of 0.058 m/s.

The results in figure 16 indicate for a fish 

density decreasing with depth that the fraction of 

hexachlorobutadiene removed by bioconcentratior is 

comparable to that removed by volatilization for tH 

fish fraction of 10"5 . For the 10"6 and 10"7 fish 

fractions, the fractions removed by bioconcentration 

are much smaller than the fraction removed by volatilization for shallow flows. For deep flows, all the 

fractions removed approach zero. The results in 

figure 16 indicate for a constant fish fraction of 10~5 

that a constant 37.6 percent of the hexachlorobutadiene is removed by bioconcentration, which exceeds 

the fraction removed by volatilization over most of the
118 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

flow-depth range. For a constant fish fraction of 10~6, 

the fraction removed by volatilization exceeds the 

fraction removed by bioconcentration to a flow depth 

of about 4 m. For deeper flows, the fraction removed 

by bioconcentration exceeded that removed by volatilization. For a constant fish fraction of 10 , the 

fraction removed by bioconcentration is negligible 

for all depths. The fraction removed by volatilization 

shows the characteristic rapid decrease with increasing 

depth, which was discussed previously in the volatilization section of this report.

1.0

0.8

0.6

< 

CC

UJ 0.4 

O

O 

O 

O

00 0.2

< 

N

=! 1.0

HEXACHLOROBUTADIENE 

VOLATILIZATION

BIOCONCENTRATION WITH A 

CONSTANT FISH FRACTION OF

HEXACHLOROBUTADIENE 

VOLATILIZATION

BIOCONCENTRATION WITH A FISH FRACTION 

DECREASING WITH DEPTH FROM AN INITIAL 

FISH FRACTION OF

10EXP(-5) 

10EXP(-6) 

10EXP(-7)

4 6 8 10

FLOW DEPTH, IN METERS

Figure 16. Fraction of hexachlorobutadiene removed by volatilization and bioconcentration as a function of flow depth.

Deep flows and a low water velocity were used in 

estimating the importance of bioconcentration so as to 

minimize the contribution of volatilization. With these 

conditions and a fish fraction of 10~5 , bioconcentration 

was significant for hexachlorobutadiene. However, a 

fish fraction of 10~5 is an order of magnitude larger 

than the value of 10~6 commonly used, and predicted 

bioconcentration fractions for this fraction were less 

than 0.06 for all except the very shallow flows. This 

analysis also assumes that equilibrium is established 

between the water and lipid phases, and such an 

equilibrium is slow to be established. Failure to 

achieve equilibrium results in lower fractions than 

those predicted in figure 16. Finally, Kenaga and 

Goring (1980) noted that when a hydrophoMc 

compound is easily metabolized, the tendency to 

bioconcentrate is offset by the rapid rate of metabolism, 

and significant levels of the compound will not develop 

in the organism. The target analytes of the present study 

are all volatile organic compounds, and significant 

concentrations are not likely to persist for long times in 

streams because of volatilization losses. Consequently, 

it would seem that the same conclusion wonld be 

appropriate for these volatile organic compounds, with 

bioconcentration not being a significant mechanism for 

loss from streams relative to volatilization.

As with every process that might af "ect the 

concentration of VOCs in streams, each situation must 

be specifically evaluated because exceptions to these 

generalizations may occur. For example, previous 

discussion indicated that algae have highe^ BCFs than 

fish. If a stream has a reach containing a high concentration of algae or aquatic vegetation, significant 

bioconcentration may occur in this section. Such a 

section would be analogous to the use of wetlands to 

treat wastewaters. In general, however, densities of 

aquatic organisms in streams will be small enough that 

the effect of bioconcentration on the concentrations 

of VOCs will be negligible. This does not however, 

necessarily mean that the small amounts removed will 

have negligible effect on the organism in which the 

VOC is concentrated.

AQUATIC TOXICOLOGY

Measuring the effects of organic compounds on 

aquatic organisms is difficult. The general procedure 

is to do a laboratory toxicity bioassay to determine a 

toxicity endpoint for a specific organic compound, 

determine the concentration of this compound in the 

aquatic system, and compare the concentration to the
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

119

endpoint using the quotient method (Bascietto and 

others, 1990). The quotient method is the ratio of the 

concentration to the toxicity endpoint. An example is 

the water concentration divided by the LC50 value, 

where the LC50 value is the concentration resulting 

in death to 50 percent of the organisms. The closer 

the quotient is to unity or larger, the higher the 

probability that an adverse effect will occur. The 

greatest uncertainty in this procedure is interpreting 

the adverse effect because results in the natural system 

are not likely to be the same as in the laboratory. This 

uncertainty is compensated for by adding a safety 

factor to the no-observed-effect level to give a concentration below which exposures are assumed to be safe 

(Bascietto and others, 1990).

The U.S. Environmental Protection Agency 

(1996) assembled from the literature in a computer 

file designated AQUIRE all LC50 and EC50 values 

for organic compounds for six categories of 

organisms. These categories were amphibian, crustacean, fish, insect, invertebrate, and plant. The EC50 

value is the concentration at which 50 percent of the 

organisms are affected, with an endpoint other than 

death. The U.S. Geological Survey (Rowe and others, 

1997) abstracted from the AQUIRE file entries for the 

target analytes that were for freshwater studies and 

were of documentation codes C and M. A documentation code C is a study using published or welldocumented procedures. A documentation code M is 

a study where procedures were generally satisfactory 

but some information was lacking. When more than 

two concentrations were reported for a specific target 

analyte for a specific organism with comparable 

experimental conditions, only the two lowest concentrations were abstracted from the data base.

Selected LC50 and EC50 values from this abstraction are presented in tables 30 and 31 for the target 

analytes. When several values were available for a 

category, maximum and minimum values were 

presented. Consequently, these maximum and 

minimum concentrations of the target analytes 

represent the range of minimum concentrations that 

might result in the indicated effect for the specific 

organism. Concentrations were not available for all six 

categories of organisms for most target analytes. For 

example, there were no LC50 values for the plant 

category for any of the target analytes listed in table 30.

LC50 values varied widely, ranging from 

0.8 |ug/L for bromomethane and guppies to four values 

of 10 ng/L or larger. These values were l.lxlO6 |ag/L 

for 1,2-dichloroethane and medaka; greater than

1 .Ox 106 |ag/L for 1,1,1 -trichloroethane and medaka; 

greater than l.OxlO6 |ag/L for benzene and rotifers; and 

1.3xlO6 |ag/L for methylbenzene and mosquito fish. 

However, values in general are large. Of the 170 values 

in table 30, only 19 are less than 103 ng/L, only 5 are 

less than 102 ng/L, and only 2 are less than 10 ng/L. 

EC5o values varied widely also, ranging from 0.6 |ug/L 

for bromomethane and guppies to 2.0xl06 ng/L for 

dichloromethane and duckweed. Of the 120 values in 

table 31, only 23 are less than 103 ng/L, only 3 are less 

than 102 |ug/L, and one is less than 10 ng/L. The small 

LC50 and EC50 values are concentrated in several 

compounds, with bromomethane, hexachlorobutadiene, 1,2,3-trichlorobenzene, and 2-propenal each 

having three or more values less than 103 ng/L. The 

large number of small values and the small concentrations for 2-propenal were not surprising because this 

compound has been used in dilute solutions to kill 

undesirable vegetation in irrigation canals (National 

Academy of Sciences, 1977).

Most of the LC50 and EC50 values in tables 30 

and 31 are larger than 103 ng/L, and such large 

concentrations of VOCs are not likely to exist and/or 

to persist in streams for long periods of time because 

of the volatile nature of these compounds. On the o*her 

hand, streams in some parts of the Nation freeze 

during the winter, resulting in ice cover that prevents 

VOCs from volatilizing to the atmosphere. As an 

example, concentrations of VOCs in the upper 

Mississippi River were higher under ice cover than 

during ice-free conditions, although concentrations 

generally were in the low microgram-per-liter range 

(Barber and others, 1995). In the event of a spill, 

concentrations might be large for a short period, bit 

exposure should be of short duration, minimizing 

acute exposure of the organisms to the VOCs. This 

variety of conditions requires that each situation be 

evaluated individually to determine possible adverse 

effects of acute exposure of organisms to VOCs.

The effect of long-term chronic exposure of 

aquatic organisms to low concentrations of VOCs 

has not been studied extensively, and results are lil ely 

to be different from short-term acute exposures. 

However, long-term exposures to low concentrations 

of VOCs in streams are unlikely because of the 

volatile, transient nature of VOCs and also because of 

the continuing effort to reduce and eliminate waste 

discharges to the streams of the nation. An exception 

might be those streams discussed above that are 

covered with ice for several months; a fairly long-term 

exposure could occur under these conditions.
120 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

Table 30. Minimum concentrations of 40 target analytes resulting in 50-percent mortality of various aquatic organisms (f"om Rowe and 

others, 1997); when two or more concentrations were available for an organism category, maximum and minimum values are presented; 

when two concentrations for an organism category were approximately the same, the minimum concentration is presented

[USGS, U.S. Geological Survey; LC50 , lethal concentration at 50-percent mortality; ng/L, microgram per liter; >, greater than]

USGS 

parameter 

code

34418

34423

32106

32102

34413

32104

32105

32103

34506

34511

34396

Compound

chloromethane

dichloromethane

trichloromethane

tetrachloromethane

bromomethane

tribromomethane

chlorodibromomethane

1 ,2-dichloroethane

1,1,1 -trichloroethane

1 , 1 ,2 -trichloroethane

hexachloroethane

Category of 

organism

fish

fish

insect

fish

insect

invertebrate

fish

insect

invertebrate

fish

insect

fish

insect

fish

amphibian

crustacean

fish

insect

fish

insect

fish

insect

invertebrate

amphibian

crustacean

Common 

name of 

organism

bluegill

medaka

fathead minnow

water flea

water flea

medaka

rainbow trout

water flea

water flea

rotifer

medaka

medaka

water flea

water flea

flatworm

carp

guppy

water flea

carp

bluegill

water flea

water flea

carp

carp

salamander

salamander

scud

medaka

rainbow trout

water flea

stonefly

medaka

bluegill

water flea

water flea

fathead minnow

guppy

midge

water flea

zebra mussel

snail

bullfrog

bullfrog

crayfish

LC50

(W/L)

5.5xl05

8.4x1 05

1.9xl05

3.1xl05

2.2xl05

5.0-105

1.2xl03

2.9 -105

2.9 -104

2.0 -103

6.7xl05

2.0 -103

3.5xl04

2.3xl03

2.0 -102

1.4xl04

8.0x1 0" 1

2.2xl03

8.0 -104

2.9 -104

5.6xl04

4.4x1 04

5.2xl04

3.4x1 04

6.5xl03

2.5xl03

>1.C<105

l.lxlO6

3.4x10*

2.5xl05

>1.0xl05

>1.0xl06

4.0 -104

>5.3xl05

5.4x1 03

8.2x1 04

4.0 -104

1.5xl05

l.SxlO4

1.9 -105

5.8xl04

2.8xl03

2.4xl03

2.7xl03
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

121

Table 30. Minimum concentrations of 40 target analytes resulting in 50-percent mortality of various aquatic organisms (from Rowe and 

others, 1997); when two or more concentrations were available for an organism category, maximum and minimum values are presented; 

when two concentrations for an organism category were approximately the same, the minimum concentration is presented—Continued

[USGS, U.S. Geological Survey; LC50, lethal concentration at 50-percent mortality; j.ig/L, microgram per liter; >, greater than]

USGS 

parameter 

code

34396

77651

34541

77443

82625

34501

34546

39180

34475

39702

Compound

hexachloroethane (Continued)

1 ,2-dibromoethane

1 ,2-dichloropropane

1 ,2,3-trichloropropane

1 ,2-dibromo-3-chloropropane

1,1-dichoroethene

trans- 1 ,2-dichloroethene

trichloroethene

tetrachloroethene

hexachlorobutadiene

Category of 

organism

fish

insect

invertebrate

fish

fish

insect

fish

fish

invertebrate

fish

insect

insect

amphibian

crustacean

fish

insect

invertebrate

fish

insect

invertebrate

fish

insect

invertebrate

Common 

name of 

organism

crayfish

catfish

rainbow trout

water flea

midge

snail

bluegill

bass

bluegill

fathead minnow

water flea

water flea

fathead minnow

bluegill

bass

snail

bladder snail

bluegill

fathead minnow

water flea

water flea

water flea

axolotl

toad

scud

medaka

medaka

water boatman

water flea

tubificidae family

flatworm

bluegill

rainbow trout

midge

water flea

flatworm

bluegill

fathead minnow

sowbug

sowbug

snail

LC50

(Hg/L)

>2.1xl03

1.6xl03

7.7x1 02

>1.0xl04

1.2xl03

>2.1xl03

1.8xl04

l.SxlO4

3.2x1 05

1.3xl05

9.9x1 04

5.2xl04

6.6x1 04

5.0x1 04

2.0x1 04

5.7xl04

2.4x1 04

7.4x1 04

2.9x1 04

1.2xl04

2.3xl05

2.2x1 05

4.8xl04

4.5xl04

2.4x1 04

2.7xl05

1.9xl03

l.lxlO5

2.3xl03

1.3xl05

1.7xl03

4.6x1 04

1.4xl03

5.5xl04

1.8xl03

1.4xl03

3.2xl02

9-OxlO 1

1.6xl02

1.3xl02

2.1xl02
122 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

Table 30. Minimum concentrations of 40 target analytes resulting in 50-percent mortality of various aquatic organisms (f~om Rowe and 

others, 1997); when two or more concentrations were available for an organism category, maximum and minimum values ?re presented; 

when two concentrations for an organism category were approximately the same, the minimum concentration is presented—Continued

[USGS, U.S. Geological Survey; LC50, lethal concentration at 50-percent mortality; ng/L, microgram per liter; >, greater than]

USGS 

parameter Compound 

code

34030 benzene

77128 styrene

34696 naphthalene

34010 methylbenzene

34371 ethylbenzene

77224 n-propylbenzene

77223 iso-propylbenzene

77135 1 ,2-dimethylbenzene

Category of 

organism

amphibian

crustacean

fish

insect

invertebrate

crustacean

fish

insect

invertebrate

crustacean

fish

insect

invertebrate

crustacean

fish

insect

invertebrate

fish

insect

fish

fish

insect

amphibian

fish

invertebrate

Common 

name of 

organism

axolotl

toad

copepod

scud

channel catfish

pink salmon

water flea

dragonfly

rotifer

rotifer

scud

guppy

rainbow trout

sowbug

water flea

snail

snail

copepod

scud

mosquito fish

rainbow trout

midge

water flea

snail

copepod

mosquito fish

coho and silver

salmon

water flea

midge

rotifer

catfish

rainbow trout

water flea

rainbow trout

fathead minnow

rainbow trout

water flea

toad

fathead minnow

rainbow trout

snail

LC50

(Mg/U

3.7xl05

1.9xl05

7.1xl05

4.2xl04

4.2xl05

4.6xl03

3.9xl05

l.OxlO4

>1.0xl06

>1.0xl03

6.2x1 04

7.5xl04

2.5xl03

6.5xl04

2.3xl04

5.6x1 05

l.lxlO5

6.8x1 04

3.9xl03

1.6xl05

l.lxlO2

1.3xl04

l.OxlO3

S.OxlO3

4.5xl05

1.3xl06

5.5xl03

3.1xl05

4.7xl04

l.lxlO5

2.1xl05

4.2xl03

7.5xl04

1.6xl03

6.3xl03

2.7xl03

9.5xl04

7.3xl04

1.6xl04

7.6xl03

>2.2xl04
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

123

Table 30. Minimum concentrations of 40 target analytes resulting in 50-percent mortality of various aquatic organisms (from Rowe and 

others, 1997); when two or more concentrations were available for an organism category, maximum and minimum values are presented; 

when two concentrations for an organism category were approximately the same, the minimum concentration is presented—Continued

[USGS, U.S>. Geological Survey; LC50, lethal concentration at 50-percent mortality; ng/L, microgram per liter; >, greater than]

USGS 

parameter 

code

85795

85795

77222

34301

34536

34566

34571

77613

34551

78032

34210

34215

Compound

1 ,3-dimethylbenzene

1 ,4-dimethylbenzene

1 ,2,4-trimethylbenzene

chlorobenzene

1 ,2-dichlorobenzene

1 ,3-dichlorobenzene

1 ,4-dichlorobenzene

1 ,2,3-trichlorobenzene

1 ,2,4-trichlorobenzene

methyl tertiary-butyl ether

2-propenal

2-propenenitrile

Category of 

organism

fish

fish

fish

fish

insect

fish

insect

fish

insect

fish

insect

fish

insect

crustacean

fish

insect

invertebrate

fish

amphibian

fish

insect

invertebrate

fish

insect

Common 

name of 

organism

goldfish

rainbow trout

goldfish

rainbow trout

fathead minnow

guppy

bass

water flea

water flea

fathead minnow

rainbow trout

midge

water flea

bluegill

bluegill

water flea

water flea

bluegill

rainbow trout

water flea

water flea

zebrafish

guppy

midge

crayfish

zebrafish

medaka

water flea

midge

snail

flatworm

fathead minnow

toad

brown trout

fathead minnow

midge

water flea

snail

medaka

fathead minnow

water flea

water flea

LC50

(Hg/U

1.6xl04

8.4x1 03

1.8xl04

2.6x1 03

7.7x1 03

4.6x1 04

5.0x10'

l.lxlO4

7.9x1 03

9.5x1 03

l.SxlO3

2.0x1 04

2.4x1 03

2.2x1 04

S.OxlO3

7.2xl03

1.7xl03

4.5x1 03

8.0x1 02

4.2x1 04

l.lxlO4

3.1xl03

3.5x1 02

1.7xl03

3.0xl03

6.3x1 03

l.lxlO3

2.1xl03

9.3x1 02

3.2x1 03

l.lxlO3

6.7x1 05

7.0x10°

l.SxlO3

1.4x10'

>1.5xl02

5.7x10'

>1.5xl02

5.0x1 04

2.6x1 03

1.3xl04

7.6xl03
124 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

Table 31 . Minimum concentrations of 31 target analytes that affected at the 50-percent level various functions of aquatic organisms 

(from Rowe and others, 1997); when two or more concentrations were available for an organism category, maximum and minimum 

values are presented; when two concentrations for an organism category were approximately the same, the minimum concentration 

is presented

[USGS, U.S. Geological Survey; EC50, effective concentration at 50 percent affected; ng/L, microgram per liter; >, greater than; ter, quantifiable 

occurrence of birth defects; imm, immobilization; gro, measurable change in growth characteristics; clr, chlorophyll; rep, change in the reproductive 

ability; bms, biomass productivity affected; rgn, change in ability to regenerate a body part; beh, quantifiable change in behavior; hat, change in 

hate liability; equ, equilibrium, change in ability to maintain balance; pgr, population growth; loc, locomotor behavior; avo, avoidance or attraction 

to a chemical; enz, change in enzyme activity; ptr, photoactic response, attraction to or avoidance of light; pse, photosynthesis; mor, mortal: ty; 

cyt, change in the genetic processes of the cell]

USGS 

parameter 

code

34423

32106

32102

34413

32104

32103

34506

34511

34396

34501

Compound

dichloromethane

trichloromethane

tetrachloromethane

bromomethane

tribromomethane

1 ,2-dichloroethane

1,1,1 -trichloroethane

1 , 1 ,2-trichloroethane

hexachloroethane

1,1-dichloroethene

Category of 

organism

amphibian

fish

insect

plant

amphibian

insect

plant

amphibian

insect

invertebrate

fish

insect

plant

plant

insect

fish

plant

insect

invertebrate

fish

insect

plant

plant

Common 

name of 

organism

toad

bullfrog

fathead minnow

water flea

water flea

duckweed

green algae

toad

peeper

water flea

water flea

green algae

green algae

toad

frog

water flea

flatworm

medaka

guppy

water flea

water flea

green algae

green algae

green algae

green algae

water flea

water flea

fathead minnow

green algae

water flea

water flea

snail

rainbow trout

water flea

water flea

green algae

green algae

green algae

green algae

End 

point

ter

ter

imm

imm

imm

gro

clr

ter

ter

rep

imm

gro

bms

ter

ter

imm

rgn

beh

imm

imm

beh

gro

gro

Pgr

clr

imm

imm

imm

clr

imm

rep

hat

equ

imm

imm

clr

clr

clr

gro

EC50

(Hg/L)

>3.2xl04

l.SxlO4

9.9xl04

1.7xl06

1.4xl05

2.0xl06

>5.0xl05

>4.0xl04

2.7xl02

3.4xl05

5.2xl04

9.5xl05

5.6xl05

>9.2xl04

9.0x1 02

9.7x1 04

1.5xl03

5.0xl02

6.0x1 0" 1

2.0xl03

1.7xl03

6.7xl03

2.1xl03

4.0xl04

3.9x1 04

l.SxlO5

1.6xl05

l.lxlO4

>5.0xl05

7.8xl04

3.2xl04

3.6x1 04

8.4x1 02

1.3xl04

l.SxlO3

8.7xl04

6.7x1 04

>5.6xl05

4.1xl05
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

125

Table 31 . Minimum concentrations of 31 target analytes that affected at the 50-percent level various functions of aquatic organisms 

(from Rowe and others, 1997); when two or more concentrations were available for an organism category, maximum and minimum 

values are presented; when two concentrations for an organism category were approximately the same, the minimum concentration 

is presented—Continued

[USGS, U.S. Geological Survey; EC50, effective concentration at 50 percent affected; ug/L, microgram per liter; >, greater than; ter, quantifiable 

occurrence of birth defects; imm, immobilization; gro, measurable change in growth characteristics; clr, chlorophyll; rep, change in the reproductive 

ability; bms, biomass productivity affected; rgn, change in ability to regenerate a body part; beh, quantifiable change in behavior; hat, change in 

hatchability; equ, equilibrium, change in ability to maintain balance; pgr, population growth; loc, locomotor behavior; avo, avoidance or attraction 

to a chemical; enz, change in enzyme activity; ptr, photoactic response, attraction to or avoidance of light; pse, photosynthesis; mor, mortality; 

cyt, change in the genetic processes of the cell]

USGS 

parameter 

code

39180

34475

39702

34030

34696

34010

34371

77224

77223

77342

77135

Compound

trichloroethene

tetrachloroethene

hexachlorobutadiene

benzene

naphthalene

methylbenzene

ethylbenzene

n-propylbenzene

iso-propylbenzene

n-butyl benzene

1 ,2-dimethylbenzene

Category of 

organism

fish

insect

invertebrate

plant

fish

insect

invertebrate

plant

fish

fish

insect

invertebrate

plant

insect

plant

fish

insect

plant

insect

plant

insect

plant

insect

plant

insect

fish

Common 

name of 

organism

fathead minnow

water flea

water flea

flatworm

green algae

fathead minnow

water flea

water flea

flatworm

green algae

rainbow trout

coho and silver

salmon

fathead minnow

midge

water flea

snail

green algae

green algae

water flea

water flea

green algae

diatom

coho and silver

salmon

fathead minnow

mosquito

water flea

green algae

green algae

water flea

green algae

water flea

green algae

water flea

water flea

green algae

water flea

coho and silver

salmon

coho and silver

salmon

End 

point

imm

imm

imm

rgn

gro

imm

imm

imm

rgn

clr

loc

avo

bms

imm

enz

imm

gro

gro

imm

ptr

gro

pse

avo

bms

imm

enz

gro

gro

imm

gro

imm

gro

imm

imm

gro

imm

avo

avo

EC50

(ug/L)

2.2x1 04

7.6x1 04

7.8x1 03

1.7xl03

4.5x1 05

1.4xl04

7.5x1 03

3.2xl03

9.0x1 02

>5.0xl05

1.4xl02

l.SxlO3

1.7xl02

1.4xl06

6.3xl03

9.7x1 05

>1.4xl06

2.9 *104

4.7xl03

6.9x1 02

3.3xl04

2.8xl03

2.3xl03

8.3xlO]

2.2x1 04

3.6xl03

2.4x1 05

9.4x1 03

l.SxlO3

4.6x1 03

2.0x1 03

l.SxlO3

1.4xl03

6.0x1 02

2.6x1 03

3.4x1 02

7.6x1 02

6.0x1 02
126 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

Table 31. Minimum concentrations of 31 target analytes that affected at the 50-percent level various functions of aquatic organisms 

(from Rowe and others, 1997); when two or more concentrations were available for an organism category, maximum and minimum 

values are presented; when two concentrations for an organism category were approximately the same, the minimum concentration 

is presented—Continued

[USGS, U.S. Geological Survey; EC50, effective concentration at 50 percent affected; ng/L, microgram per liter; >, greater than; ter, quantifiable 

occurrence of birth defects; imm, immobilization; gro, measurable change in growth characteristics; clr, chlorophyll; rep, change in the reproductive 

ability; bms, biomass productivity affected; rgn, change in ability to regenerate a body part; beh, quantifiable change in behavior; hat, chan?e in 

hatchability; equ, equilibrium, change in ability to maintain balance; pgr, population growth; loc, locomotor behavior; avo, avoidance or attraction 

to a chemical; enz, change in enzyme activity; ptr, photoactic response, attraction to or avoidance of light; pse, photosynthesis; mor, mortality; 

cyt, change in the genetic processes of the cell]

USGS 

parameter 

code

77135

85795

85795

77222

34301

34536

34566

34571

77613

34551

34210

Compound

1,2-dimethylbenzene (Continued)

1 ,3-dimethylbenzene

1 ,4-dimethylbenzene

1 ,2,4-trimethylbenzene

chlorobenzene

1 ,2-dichlorobenzene

1 ,3-dichlorobenzene

1 ,4-dichlorobenzene

1 ,2,3-trichlorobenzene

1 ,2,4-trichlorobenzene

2-propenal

Category of 

organism

insect

plant

insect

plant

insect

plant

insect

insect

plant

fish

insect

plant

insect

plant

fish

insect

plant

insect

plant

fish

insect

invertebrate

plant

insect

Common 

name of 

organism

water flea

water flea

green algae

green algae

water flea

water flea

green algae

green algae

water flea

water flea

green algae

green algae

water flea

water flea

water flea

diatom

green algae

rainbow trout

water flea

water flea

green algae

green algae

water flea

water flea

green algae

green algae

rainbow trout

water flea

water flea

green algae

green algae

water flea

water flea

diatom

rainbow trout

water flea

water flea

flatworm

green algae

diatom

water flea

water flea

End 

point

imm

imm

gro

gro

imm

imm

gro

gro

imm

imm

gro

gro

imm

mor

imm

cyt

Pgr

loc

imm

rep

clr

gro

imm

rep

clr

bms

equ

imm

rep

gro

gro

imm

rep

cyt

equ

imm

rep

rgn

clr

cyt

imm

imm

EC50

(ug/L)

3.2xl03

l.OxlO3

5.5xl04

4.2xl03

9.6x1 03

4.7xl03

4.9xl03

3.9xl03

8.5xl03

3.6xl03

4.4xl03

3.2xl03

3.6xl03

2.2xl04

5.8xl02

2.4x1 05

2.0xl05

1.6xl03

7.8xl02

5.5xl02

7.1xl04

1.4xl04

4.2xl03

1.4xl03

1.2xl05

1.9xl04

l.lxlO3

1.6xl03

9.3 xlO2

3.8xl04

1.6xl03

3.5xl02

2.0xl02

6.4xl03

1.3xl03

1.2xl03

4.5xl02

l.lxlO3

2.2xl04

2.8xl03

9.3xlO !

S.lxlO 1
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

127

The quotient method used for defining the effect 

of organic compounds on organisms has several 

weaknesses (Bascietto and others, 1990). It does not 

account for the effects of incremental dosages of the 

organic compound, nor can it be used to estimate 

indirect effects of toxic compounds such as food chain 

interactions. It does not account for other ecosystem 

effects such as predator-prey relationships, community 

metabolism, and structural shifts. Finally, the quotient 

method does not consider the differences between 

organisms used in the laboratory bioassay studies and 

the actual aquatic systems. Also, laboratory studies are 

usually done with single organic compounds, whereas 

organic compounds seldom, if ever, occur singularly 

in natural stream systems.

Mixtures of organic compounds are being 

emphasized in current efforts to improve methods 

for determining the effects of organic compounds 

on aquatic organisms. The toxicity of a mixture 

may be more or less than the sum of the toxicities 

of the individual compounds making up the mixture. 

Present efforts (Renner, 1996) are directed toward 

classifying chemical structures into five or six modeof-action groups. The toxicity of a mixture is often 

dominated by narcotic chemicals because the mixture 

includes many industrial chemicals such as chlorinated benzenes, methylbenzenes, and alkanes not 

designed to be biologically active (Renner, 1996). 

Laboratory tests of such chemicals indicate that the 

effects of mixtures are additive for endpoints such as 

mortality and inhibition of reproduction and growth. 

However, attempts to transfer this information to 

natural systems are hindered because the mixture 

composition begins to change, once it enters the 

environment, as a result of processes such as volatilization. Consequently, there is no factor that can be 

used to transfer laboratory results to the natural 

system.

ADVECTION AND DISPERSION

A VOC dissolved in stream water is subject to 

the same advective and dispersive forces that act upon 

the mass of water in which the VOC is dissolved. The 

VOC is advected downstream with the flow of the 

stream. This advective flow usually is characterized by 

the water velocity, which is essential in estimating the 

effects of many processes on the VOC concentration 

as the water mass moves downstream.

The water velocity is not constant over the cross 

section of a stream but varies over the water depth and 

from bank to bank as a result of shear forces at the 

interface between the water and the banks and bottom 

of the stream and, to a much lesser extent, between the 

water and the atmosphere. The water velocity used in 

previous examples accounted for these variations by 

using a velocity averaged over the cross section of 

flow. These velocity variations also result in dispersive mixing in the vertical, lateral, and longitudinal 

directions of the stream.

VERTICAL MIXING

Vertical mixing or mixing over the flow depth 

is rapid in streams. Sayre (1973) indicated that the 

distance downstream from a source of contaminart 

necessary to obtain an approximately uniform concentration over the flow depth can be estimated from

0.5 U Y* (135)

where

Ly is the downstream distance (m) required 

for vertical mixing,

U is the mean water velocity (m/s),

Y is the flow depth (m), and

ey is the vertical eddy diffusivity (nr/s)

averaged over the depth. 

Various experimental results (Fischer, 1973) 

indicated that the vertical eddy diffusivity can be 

approximated by

e,, = 0.067 Y u* (136)

where

u* is the shear velocity (m/s) given by

u* = (gYS) 0.5 (137)

where

g is the acceleration of gravity (m/s/s), and 

S is the slope of the stream channel (m/m). 

The shear velocity is related to the roughness oft\ f 

stream channel by (Fischer and others, 1979)

u* = U U

0.5

(138)
128 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

where

f is the Darcy-Weisbach friction factor. 

Using 1/15 as a typical value of (f/8)0'5 (Fischer and 

others, 1979) and combining equations 135, 136, and 

138 gives

Ly - 112 Y. (139)

Equation 139 indicates that the distance 

downstream from a contaminant source necessary to 

obtain an approximately uniform concentration over 

the flow depth is 112 times the depth. Thus, for a 

stream with a depth of 0.5 m, vertical mixing is 

achieved in about 56m, whereas for a stream with a 

depth of 10 m, about 1,120 m would be necessary. 

Equation 139 is generally consistent with the common 

estimate (Yotsukura and Sayre, 1976) that a reach 

length of 50 to 100 flow depths is necessary for an 

approximately uniform vertical concentration profile 

of a neutrally buoyant, water-soluble contaminant 

discharged at the surface of a stream.

LATERAL MIXING

Mixing over the lateral dimension of a stream 

commonly is called transverse mixing. Estimation 

of the downstream distance required for completion 

of transverse mixing depends on the location of the 

contaminant source in the stream cross section and the 

desired degree of mixing. The degree of mixing, Pm, 

in percent, can be computed from (Kilpatrick and 

Cobb, 1985)

ioo-^Y (140)

where

Cm is the mean cross-sectional concentration, 

Cj and Qj are the concentration and the water

discharge at point i in the cross

section, and

n is the number of points in the cross section. 

As the desired degree of mixing increases, the distance 

required to achieve such mixing increases rapidly. 

Yotsukura and Cobb (1972) estimated that the 

distance required to improve the degree of mixing

from 95 to 99.5 percent may be as long as the distance 

required to achieve the initial 95-percent mixing. For 

this reason, 95-percent mixing commonly is used as a 

degree of mixing that can be attained with reasonable 

mixing distances (Kilpatrick and Wilson, 1989).

The mixing distance required for 95-percent 

mixing of a point source at the center of the cross 

section can be estimated (Fischer and others, 1979) 

from the following:

0.1 UB' (141)

where

LZ 

B

is the mixing distance (m), 

is the stream width (m), and

y

is the transverse eddy diffusivity (m /s).

The transverse eddy diffusivity is given by

-,z = az Y u* (142)

where 

a. is a dimensionless parameter that varies

widely, depending on the type of channel. 

For straight, rectangular flumes, most experimental 

values are in the range of 0.1 to 0.2, with an irrigation 

canal having values of 0.24 and 0.25 (Fischer and 

others, 1979). In natural channels, bends and bank 

irregularities have a large effect on transverse mixing, 

and az values are usually in the range from 0.4 to 0.8. 

Fischer and others (1979) suggested a value of 0.6, 

with an error of ±50 percent. Such a value is 

appropriate for slowly meandering streams with 

moderate bank irregularities. For streams with sharp 

bends or rapid changes in cross-sectional geometry, a 

larger value should be used. However, there are no 

specific guidelines on how much larger the parameter 

should be. It is best to underestimate this parameter 

because such underestimation overestimates the 

distance required to achieve the commonly used 

95-percent mixing.

Using a value 0.6 for az and combining 

equations 141, 142 and 137 gives

0.053 U B' (143)
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

129

where

Lz is the distance (m) required for 95-percent 

transverse mixing of a contaminant source 

at the centerline of the stream. 

For a contaminant-source entering the stream at 

some point on one of the banks, such as a storm water 

drain, the mixing distance is four times the distance 

computed from equation 143. This equation indicates 

that the mixing distance is strongly dependent on the 

stream width and increases with the square of the 

width. This equation also indicates that the mixing 

distance decreases as the depth and slope increase. 

Conversely, the mixing length increases as the 

velocity increases because the contaminant has less 

chance for transverse mixing at the higher velocity.

An alternative form of equation 143 can be 

obtained by combining it with the Manning equation. 

For metric units, the Manning equation has the form 

(Barnes, 1967)

U =

y0.67 § 0.5

(144)

where

n is the roughness coefficient. 

Equation 144 assumes a wide stream where the 

hydraulic radius is approximately equal to the flow 

depth. Combining equations 143 and 144 gives

Lz = 0.053 [|IBY 1/6.

'JL (145)

which indicates for a constant width/depth ratio that 

the mixing distance is approximately proportional to 

the width. Equation 145 has the advantage that the 

water velocity and the stream slope do not need to be 

estimated; however, the roughness coefficient is 

needed. The roughness coefficients of streams has 

been discussed by Barnes (1967).

LONGITUDINAL MIXING

Longitudinal mixing or dispersion is the 

stretching out of the contaminant mass as the mass is 

advected downstream. Longitudinal dispersion occurs 

primarily as a result of velocity variations in the lateral 

direction of the stream cross section. That part of the

contaminant mass in the center of the cross section 

is advected downstream at a velocity higher than the 

average velocity. Conversely, that part of the contaminant mass in the low-velocity regions near the streambanks moves downstream at a velocity lower than the 

average velocity, resulting in elongation or longitudinal mixing of the contaminant mass.

The effect of longitudinal dispersion is 

dependent upon the type of contaminant source. 

For a contaminant discharged continuously into a 

stream, constant plateau concentrations are eventually 

established in the stream and the effect of longitudinal 

dispersion is eliminated. The concentration may 

decrease with distance downstream as a result of 

other processes such as volatilization. However, 

longitudinal dispersion has no effect on the plateau 

concentrations because steady-state conditions are 

established.

Longitudinal dispersion is important, however, 

for an instantaneous slug injection of contaminant 

such as might occur with a spill. It has been shown 

(Fischer, 1966) that

C(L,t) =

M

Aj4jtDL t

exp -

where

C is the contaminant concentration (^g/m3 ) at

longitudinal position L (m) at time t 0), 

M is the mass of contaminant released (^g), 

A is the cross-sectional area (m2), 

DL is the longitudinal dispersion coefficient

(nT/s), and

U is the mean water velocity (m/s). 

Equation 146 applies to a one-dimensional system and 

it assumes that the contaminant is conservative.

If the contaminant is not conservative, then 

equation 146 becomes (Camp, 1963)

C(L,t) =

M

exp

(L-Ut)

4DT t

L i= i

(147)
130 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

where

Kj is the first-order rate coefficient (s-1 )

describing loss of the contaminant from 

the stream by process i.

Longitudinal dispersion in a stream continues indefinitely. For practical purposes, longitudinal dispersion 

ceases when the analytical procedure used for 

measuring the concentration of the contaminant is no 

longer able to detect changes in concentration with 

time as the contaminant mass passes a downstream 

point. Application of equations 146 and 147 indicates 

that these equations do not predict concentration-time 

distributions well near the contaminant source, nor do 

these equations predict the long tails commonly 

observed in contaminant distributions determined in 

natural stream systems.

Application of equations 146 and 147 requires 

the longitudinal dispersion coefficient, and this coefficient has been studied extensively. Despite this 

extensive study, estimation of the longitudinal dispersion coefficient is still subject to considerable error. 

Various approaches have been used to develop 

equations for predicting this coefficient.

A shear velocity analogy (Fischer and others, 

1979) resulted in the equation

DL = aL Yu* (148)

where

aL is a dimensionless parameter. 

Unfortunately, experimental values of aL vary widely, 

ranging from 8.6 for a canal to 7,500 for the Missouri 

River (Fischer and others, 1979).

There have been numerous attempts to relate the 

longitudinal dispersion coefficient to the hydraulic and 

geometric properties of streams. Fischer and others 

(1979) proposed

0.011U2 B2 

Y

(149)

where metric units are used with seconds as the time 

unit. Predictions using equation 149 are usually within 

a factor of about four of experimental dispersion 

coefficients. Fischer and others (1979) considered this 

agreement reasonable in view of the fact that the 

change of moment method and the routing procedure 

commonly used for analyzing experimental data may

give dispersion coefficients differing by as much 

as a factor of two for the same set of data, Much 

of this difficulty results from the long tails on the 

concentration-time distributions and the effect of these 

tails on the computation of the variances of the distributions. However, it is these tails that def ne the 

dispersion being measured; consequently, it is difficult 

to decide at what point to truncate the calculations. 

McQuivey and Keefer (1974) proposed

_ 0.058 Q ~ ~SB~ (150)

where

DL is the longitudinal dispersion coefficient (m2/s),

Q is the water discharge (m3/s), 

S is the slope of the channel (m/m), and 

B is the channel width (m). 

Equation 150 applies to a steady base-flow condition 

and was developed from 37 measurements on 

18 streams. This equation has a standard error of 

estimate of about 30 percent. Measured longitudinal '•j

dispersion coefficients ranged from 5.4 m/s for the 

Monocacy River in Maryland to 1,030 nr/s for the 

Missouri River in Nebraska and Iowa.

Estimates of the longitudinal dispersion coefficient from equations 148, 149, and 150 for a specific 

set of hydraulic and geometric properties for a stream 

could differ substantially. If this occurs, properties of 

the stream can be compared with those of streams for 

which longitudinal dispersion coefficients have been 

measured (McQuivey and Keefer, 1974; Fischer and 

others, 1979) to assist in determining whrt is the best 

estimate of the longitudinal dispersion coefficient for 

the stream of interest. Other equations have been 

presented in the literature. However, a complete 

review of the literature on longitudinal dispersion is 

beyond the scope of this report. Readers n ceding more 

information on any aspects of dispersion should 

consult Rutherford (1994).

ESTIMATION OF THE IMPORTANCE 

OF LONGITUDINAL DISPERSION

The importance of longitudinal dispersion on 

the fate of the target analytes in streams was estimated 

for the deep, low-velocity stream, intermediate-depth, 

high-velocity stream, and shallow, medium-velocity 

stream examples used in previous calculations.
PROCESSES AFFECTING THE TRANSPORT, BEHAVIOR, AND FATE 

OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

131

Dichloromethane and diisopropyl ether were selected 

as examples from the 55 target analytes (table 1). 

Temperature was assumed to be 20°C, and an average 

air-film mass-transfer coefficient of 756 m/day was 

used. Volatilization coefficients were calculated using 

equations 53, 54, 55, 27, 28, 16, and 13 and the same 

procedures used in previous examples. Longitudinal 

dispersion coefficients of 2 m2/s (deep, low-velocity 

stream), 1,000 nr/s (intermediate-depth, high-velocity 

stream), and 10 m"/s (shallow, medium-velocity 

stream) were used.

Concentrations computed from equations 146 

and 147 for various distances downstream are plotted 

in figures 17, 18, and 19 as a function of elapsed time 

from the release of the contaminants. The dispersant 

was assumed to be conservative, and the VOCs were 

assumed to be subject only to longitudinal dispersion

and volatilization. The time unit in seconds used in 

equations 146 and 147 was converted to hours to 

facilitate plotting the results.

Figure 17 shows the results for the deep, lowvelocity stream characteristic of the O'ConnorDobbins (1958) equation. The peak concentration 

decreased substantially with distance downstream. 

There was, however, only a little difference between 

the curves for the conservative dispersant and 

the nonconservative VOCs. The ratio of the peak 

concentration of the conservative dispersant at 

40,000 m to the concentration at 3,000 m was 0.273. 

Corresponding ratios for dichloromethane and 

diisopropyl ether were 0.238 and 0.244, respectively, 

indicating that 87 and 89 percent of the changes in the 

concentrations of dichloromethane and diisopropyl 

ether were the result of longitudinal dispersion.

1.0

0.9

3,000 METERS

0.8

DC 

LLJ

gj°-7

CL

CO

n£ 0.6

DC 

^ 0.5

5si 0.4

LU 0.3 

O

——— DISPERSION FOR CONSERVATIVE DISPERSANT

.......... DISPERSION + VOLATILIZATION FOR NONCONSERVATIVE DISPERSANT

(DICHLOROMETHANE)

—— _ DISPERSION + VOLATILIZATION FOR NONCONSERVATIVE DISPERSANT 

(DIISOPROPYL ETHER)

DISPERSION COEFFICIENT = 2 SQUARE METERS PER SECOND

0.2 -

0.1

12,000 METERS

25,000 METERS

40,000 METERS

20 40 60 80 " 100 120 140

ELAPSED TIME FROM RELEASE, IN HOURS

160 180 200

Figure 17. Concentration of a conservative dispersant and the nonconservative dispersants dichloromethane and diisopropyl ethe*" as a 

function of time for a deep, low-velocity stream characteristic of the O'Connor-Dobbins (1958) equation; Henry's law constants ere 229 

and 167 pascals cubic meter per gram mole at 20 degrees Celsius for dichloromethane and diisopropyl ether, respectively.
132 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

1.3

1.2

1.1

O_ 

CO

0.8

0.7

O

06

O

0.5

QC

0.4

O 

§0.3

0.2

0.1

10,000 METERS

30,000 METERS

——— DISPERSION FOR CONSERVATIVE DISPERSANT

.......... DISPERSION + VOLATILIZATION FOR NONCONSERVATIVE DISPERSANT

(DICHLOROMETHANE)

——— DISPERSION + VOLATILIZATION FOR NONCONSERVATIVE DISPERSANT 

(DIISOPROPYL ETHER)

DISPERSION COEFFICIENT = 1,000 SQUARE METERS PER SECOND

80,000 METERS

8 10 12 14 16

ELAPSED TIME FROM RELEASE, IN HOURS

18 20 22 24

Figure 18. Concentration of a conservative dispersant and the nonconservative dispersants dichloromethane and diisopropyl ether as 

a function of time for an intermediate-depth, high-velocity stream characteristic of the Churchill and others (1962) equation; Henry's 

law constants are 229 and 167 pascals cubic meter per gram mole at 20 degrees Celsius for dichloromethane and diisopropyl ether, 

respectively.

Figure 18 shows the results for the intermediatedepth, high-velocity stream characteristic of the 

Churchill and others (1962) equation. The ratio of the 

peak concentration of the conservative dispersant at 

80,000 m to the concentration at 10,000 m was 0348. 

Corresponding ratios for dichloromethane and 

diisopropyl ether were 0.179 and 0.211, respectively, 

indicating that 51 and 61 percent of the changes in the 

concentrations of dichloromethane and diisopropyl 

ether were the result of longitudinal dispersion.

Figure 19 shows the results for the shallow, 

medium-velocity stream characteristic of the 

Owens and others (1964) equation. The ratio of the 

peak concentration of the conservative dispersant 

at 2,000 m to the concentration at 250 m was 

0.350. Corresponding ratios for dichloromethane 

and diisopropyl ether were 0.0133 and 0.0289,

respectively, indicating that 3.8 and 8.3 percent of 

the changes in the concentrations of dichloromethane 

and diisopropyl ether were the result of longitudinal 

dispersion.

The results in figure 19 and other examples 

described previously indicate that volatilization rates 

for shallow flows are rapid. The results in figure 17 

indicate that substantial decreases in concentration can 

occur as a result of longitudinal dispersion, even when 

changes as a result of volatilization are minimal 

because of the deep flow. The results in fgure 18 are 

for intermediate conditions, where both processes 

contribute to decreases in the concentration with 

distance downstream.

The results in figures 17, 18, and 19 are 

idealized somewhat because it is well known that 

concentration-time distributions in strearrs do
INFORMATION NEEDS 133

DISPERSION FOR CONSERVATIVE DISPERSANT

DISPERSION + VOLATILIZATION FOR NONCONSERVATIVE DISPERSANT 

(DICHLOROMETHANE)

——— DISPERSION + VOLATILIZATION FOR NONCONSERVATIVE DISPERSANT 

(DIISOPROPYL ETHER)

DISPERSION COEFFICIENT = 10 SQUARE METERS PER SECOND

0.6 0.8 1.0 

ELAPSED TIME FROM RELEASE, IN HOURS

Figure 19. Concentration of a conservative dispersant and the nonconservative dispersants dichloromethane and diisopropyl ether as a 

function of time for a shallow, medium-velocity stream characteristic of the Owens and others (1964) equation; Henry's law constants 

are 229 and 167 pascals cubic meter per gram mole at 20 degrees Celsius for dichloromethane and diisopropyl ether, respectively.

not fit exactly the normal distributions predicted 

by equations 146 and 147. Specifically, the tails 

of the concentration-time distributions commonly 

are much longer than predicted by these equations. 

Equations 146 and 147 do provide, however, a 

useful first approximation of the effects of longitudinal dispersion on contaminant distributions in 

streams.

INFORMATION NEEDS

This review of the literature and application 

of the available information to predicting the fate 

of VOCs in streams in general, and specifically the 

target analytes, has identified several areas in which

information was limited or was lacking. Specific areas 

where additional physical property data and process 

research are needed are discussed below.

VOLATILIZATION

The Henry's law constant is required to 

determine the tendency of a compound to partition 

between the water and air phases of the environment. 

For those compounds with Henry's law constants 

small enough that mixing processes in both the water 

and air phases are important, this constant also ha? a 

direct effect on the overall volatilization rate coef ~- 

cient. Generally, good coverage of the temperature 

dependencies of the Henry's law constants for the 

target analytes was found (tables 2 and 4). However,
134 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGAMC COMPOUNDS IN STREAMS

Henry's law constants were not available for 12 of the 

target analytes. These analytes were bromoethene, 1,2- 

dibromo-3-chloropropane, cis- 1,3-dichloropropene, 

trans- 1,3-dichloropropene, iso-propylbenzene, 

n-butylbenzene, styrene, ethyl tertiary-butyl ether, 

tertiary-amyl methyl ether, diisopropyl ether, 

2-propenal, and 2-propenenitrile. Constants for these 

analytes were predicted using vapor pressure and 

solubility data from the literature, although these data 

were limited, and, in many cases, of unknown quality. 

Specifically, there are currently no data for bromoethene. The solubility for this compound was calculated 

from an octanol-water partition coefficient estimated 

from the value for chloroethene. Also, there are few 

physical property data for the four ether compounds, a 

group of compounds that are important because of 

their use as automotive gasoline oxygenates. The 

Henry's law constant for methyl tertiary-butyl ether 

has been measured (Robbins and others, 1993) as a 

function of temperature. However, the temperature 

range was 25°C to 50°C, thus requiring extrapolation 

to the lower temperatures appropriate for most 

environmental conditions.

The reference-substance concept (eqs. 25 

and 26) is fundamental to predicting the volatilization 

of VOCs from streams. The water-film referencesubstance concept has been verified for a number 

of compounds. However, most of these compounds 

had Henry's law constants large enough that virtually 

all the resistance to volatilization was in the water 

film. There is little experimental data on the effect 

of an increasing air-film resistance on the volatilization process. Also, the air-film reference-substance 

concept has been verified for only one compound. 

The argument might be made that volatilization 

becomes less important and eventually negligible 

as the Henry's law constant decreases. However, 

if these solutes with low Henry's law constants 

are volatile enough to be considered as target analytes 

and if the Henry's law constant indicates an airfilm resistance, then this resistance must be considered. Present knowledge is somewhat limited in 

this respect. A target analyte of current interest 

is methyl tertiary-butyl ether, and this compound 

has a Henry's law constant such that the air-film 

resistance can be important. More information is 

needed on the volatilization characteristics of this 

compound.

ABSORPTION

The volatilization model indicates that absorption should occur when the ambient air concentration 

exceeds the air concentration that would H in equilibrium with the ambient water concentration. Absorption is most likely to be important in geographic areas 

with high atmospheric concentrations of VOCs and 

low stream-water concentrations of the same VOCs. 

However, such absorption has not been erplicitly 

demonstrated. Measurements are needed to determine 

if the atmosphere can be a significant source of VOCs 

in the streams of such areas.

WET DEPOSITION

Limited data exist on gas scavengirg ratios 

for the target analytes, and there is considerable 

variability in the available data. For most analytes, 

±1 standard deviation of the mean was a large 

percentage of the mean and, in some case^, 

approached the mean in value. This variability 

indicates the experimental difficulties involved in 

measuring small concentrations of VOCs in air and 

precipitation samples. With improved instrumentation 

and lower detection limits, more accurate determinations of these ratios should now be possible.

Henry's law constants at low temperatures are 

needed for some of the target analytes. Tl ^se data 

would permit a more thorough evaluation of the 

relationship between the gas scavenging ratio and the 

Henry's law constant, as well as an accurate determination of the temperature dependence of the gas 

scavenging ratio.

The concentrations of VOCs in streams 

predicted to result from gas scavenging and wet 

deposition were very low, indicating that this process 

probably is not a significant source. However, a 

simplistic approach was used for these predictions, 

and experimental verification is necessary. The contributions from wet deposition onto surface? and runoff 

into the stream are expected to be minimal because of 

rapid volatilization back into the atmosphere. VOCs 

have, however, been found in stormwater runoff, 

indicating that gas scavenging and wet deposition 

could be a source of VOCs in streams. Eroerimental 

determination of the possible magnitude of this source 

under a variety of conditions is needed.
INFORMATION NEEDS 135

MICROBIAL DEGRADATION

Considerable information exists in the literature 

on the microbial degradation of the target analytes. 

However, this is the one process for which it is 

probably the most difficult to predict effects on VOC 

concentrations in streams. Indeed, microbial degradation is a highly system-specific and site-specific 

process. For example, different microbial communities 

will need different periods of time to adapt to a VOC 

of interest. During this so-called lag period, the 

microbes develop the enzymes necessary to 

accomplish degradation of the VOCs. Information 

needs, therefore, include compound-dependent data 

on whether or not adaptation can occur during the 

time periods VOCs are expected to persist in streams. 

Adaptation is most likely to occur during long-term 

exposures such as would occur during continuous 

discharges of waste waters. However, because such 

discharges are gradually being reduced or eliminated, 

a more important question is whether or not adaptation 

can occur during exposure to VOCs in transient events 

such as occur during stormwater runoff, surfacewater runoff, and spills. The answer to this question 

is difficult because of the complex nature of the 

microbial degradation process and the transient nature 

of VOCs in streams.

SORPTION

The literature indicates that the extent of 

sorption of organic compounds can vary appreciably 

between soils and sediments, with the difference 

apparently dependent on the type and extent of diagenesis of the soil. A significant source of VOCs in urban 

streams may be those associated with sediment 

particles flushed from impervious surfaces into the 

streams during storm events. Information is needed on 

the organic-carbon contents and sorption characteristics of such particles, and how they compare with 

natural soils and sediments.

It has been hypothesized that VOCs sorbed to 

sediments in a stream under an aggrading condition 

could be buried and then pose a threat to humans or 

aquatic organisms at a later time long after the bulk 

of the VOC discharge has dissipated. VOCs have 

been found in bottom sediments, but there is no 

information as to when the VOCs were deposited in 

these sediments. Information is needed on the ability 

of VOCs to persist in bottom sediments.

Recent research has indicated that sorption 

isotherms may be nonlinear. Such behavior is in 

conflict with the linear isotherms commonly predicted 

by the partition model of sorption. A result of this 

apparent nonlinearity is a sorption coefficient that 

depends on concentration, and such dependence 

greatly complicates the prediction of the extent of 

sorption. Also, this nonlinearity appears to be mo?t 

important for low concentrations of the compound 

being sorbed, and this is the concentration range most 

likely to be important in environmental situations. 

Research is needed to define the extent of this 

nonlinear effect and how it affects the prediction 

of the sorption of organic compounds in the lowconcentration ranges prevailing in the environment.

HYDROLYSIS

Nine target analytes have reported hydrolysis 

half-lives of 548 days or less. These analytes are 

chloro me thane, dichloromethane, bromome thane, chloroethene, 1,1,1-trichloroethane, 1,2- 

dibromoethane, 1,1-dichloroethene, tetrachloroethene, and 1,3-dichloropropene. For several of these 

analytes, the reported half-lives varied by orders of 

magnitude. Minimum half-lives, therefore, were used 

in the current evaluation to provide a conservative 

approach. However, if the true half-life for a specific 

compound is an order of magnitude longer than 

assumed, then in reality hydrolysis may not be significant relative to volatilization. A careful reevaluation of 

the hydrolysis half-lives is needed for those target 

analytes assumed to have short half-lives.

Controversy exists on the hydrolysis properties 

of 2-propenal and 2-propenenitrile, as discussed ir 

detail in the hydrolysis section of this report. Because 

water is the medium of interest in this work, these 

questions must be answered to understand the 

behavior of these two compounds in streams.

Previous discussion in the sorption section of 

this report indicated that the target analytes with tH 

higher molecular weights and lower water solubilities 

tended to sorb to sediments, with the extent of sorp+ion 

depending on the concentration and organic-carbon 

content of the sediment. Larson and Weber (1994) 

used the hydrolysis of l,2-dibromo-3-chloropropane 

to demonstrate that hydrolysis rates and pathways may 

be different for systems with and without sediments. 

They concluded that more information is needed en
136 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

the effects of sediments on the hydrolysis of organic 

compounds. The presence of sediment also could 

affect the rates of other reactions such as microbial 

degradation, aquatic photolysis, and oxidation of the 

sorbed compounds.

AQUATIC PHOTOLYSIS

Very little information exists on the photolysis 

of the target analytes in the aquatic environment, in 

contrast to the atmosphere where photolysis is the 

major process for removal of many VOCs. Thus, there 

is a need for information on aquatic photolysis. The 

absorption spectrum of hexachlorobutadiene extends 

into the sunlight spectrum, but the belief (Callahan and 

others, 1979) is that the light absorption coefficient is 

not large enough for direct photolysis to be important. 

Experimental verification of this hypothesis is needed. 

The compound 2-propenenitrile does not absorb 

significantly in the wavelength range characteristic of 

sunlight, but this molecule may be subject to indirect 

photolysis in natural systems where the appropriate 

sensitizers are present (Callahan and others, 1979). 

The general topic of indirect photolysis has received 

little attention, and this process could be important for 

several of the target analytes. Indirect photolysis is a 

complex process, however, and information on the 

types, occurrences, and concentrations of the 

sensitizers necessary for this process is generally 

lacking.

The most important issue regarding photolysis 

is the question of the effect of water depth. It has 

been established that photolysis in a quiescent water 

column decreases as the depth increases because 

of reduced light penetration. In a flowing stream, 

however, the water is not quiescent, but there 

is a continual exchange and mixing of the water. 

Consequently, it may be reasonable to expect that 

photolysis is limited to the surface layers and can 

continue to occur in these layers as a result of replacement of surface water with water from below that 

contains the nonphotolyzed compound. If such a 

mechanism is operative, aquatic photolysis would 

be more significant than if the water column were 

quiescent and, therefore, could be significant relative 

to volatilization for deep flows. Experimental verification of this mechanism and the extent to which surface 

renewal facilitates aquatic photolysis in streams is 

needed.

OXIDATION

Much of the information available on the 

oxidation of VOCs in streams has been calculated or 

estimated based on results for molecules with similar 

structures. The available information indicates that 

oxidation can be significant for some target analytes in 

deep streams. However, the issue again is the question 

of the effect of depth, as was discussed fcr photolysis. 

It has been hypothesized that the oxidants causing 

oxidation of VOCs result from the photolysis of humic 

materials in the water, and verification of this hypothesis is necessary. Assuming that this mechanism for 

the formation of the oxidants is valid, it is important to 

know whether the production of these oxidants is 

limited by light penetration as the depth increases, 

or whether vertical mixing and replacement of the 

surface-water layers are rapid enough to preclude such 

depth limitations. If the formation of these oxidants is 

not limited as the depth increases, oxidation could be 

significant relative to volatilization for de^p flows. 

Experimental verification of this hypothecs is needed.

BIOCONCENTRATION

The potential for bioconcentration correlates 

with the octanol-water partition coefficient. The 

octanol-water partition coefficients of the target 

analytes are such that the masses of VOCs removed 

from streams by this process are likely to be small. 

However, estimation of these masses depends directly 

on the organism density in the stream, and little 

information currently is available on this density. 

Using a fish density an order of magnitude larger than 

the commonly used density results in prediction of 

significant removal of hexachlorobutadiene by bioconcentration, relative to volatilization. More information 

is needed on the organism densities likely to exist in 

streams.

SUMMARY AND CONCLUSIONS

The transport, behavior, and fate of volatile 

organic compounds (VOCs) in streams are determined 

by combinations of various chemical, physical, and 

biological processes. These processes are volatilization, absorption, wet and dry deposition, microbial 

degradation, sorption, hydrolysis, aquatic photolysis,
SUMMARY AND CONCLUSIONS 137

oxidation, chemical reaction, bioconcentration, 

advection, and dispersion. The literature on these 

processes was reviewed and used to estimate the 

effects these processes have in determining the fate 

in streams of 55 VOCs selected for emphasis by the 

U.S. Geological Survey National Water-Quality 

Assessment Program. These particular VOCs were 

selected because of their frequent occurrence in 

surface water and because they were known or 

suspected hazards to the health of humans and/or 

aquatic organisms. A few compounds were included 

because of special uses and/or environmental effects, 

such as those compounds used as oxygenates in 

reformulated automotive gasolines, those suspected 

of causing deterioration of the ozone layer, and those 

significantly bioconcentrated by aquatic organisms. 

While the focus of this report is on these 55 VOCs, 

the same principles and general conclusions based on 

these principles apply to other VOCs. Specific conclusions resulting from this study are presented in the 

following paragraphs.

1. Volatilization is the movement of a VOC from 

water across the water/air interface into the 

air. Volatilization of a VOC from a stream is 

dependent on the mixing conditions in both 

the water and air phases. The relative 

importance of mixing conditions in these 

two phases depends on the Henry's law 

constant of the VOC.

2. Average experimental Henry's law constants 

at 25°C for 43 target analytes range from 

34.3 Pa m3/g mol for 1,2,3-trichloropropane 

to 34,200 Pa m3/g mol for dichlorodifluoromethane. Corresponding values at 20°C are 

28.2 Pa m3/g mol and 31,400 Pa m3/g mol. 

Calculated Henry's law constants for 12 target 

analytes range from 7.92 Pa m3/g mol for 

2-propenal at 20°C to 1,520 Pa m3/g mol 

for n-butylbenzene at 25°C. Of the 55 target 

analytes, 47 have Henry's law constants large 

enough that 90 percent or more of the resistance 

to volatilization is predicted to be in the water 

phase for average environmental conditions. 

Consequently, volatilization of these analytes 

is controlled by the mixing conditions within 

the water phase, and the Henry's law constant 

has little direct effect on the volatilization 

coefficient.

3. The two analytes with the lowest Henry's law 

constant are 2-propenenitrile and 2-propenal. 

The resistance to volatilization of these twc 

analytes is predicted to be about equally divided 

between the water and air phases for average 

environmental conditions.

4. Estimates of the volatilization half-lives for tr'bromomethane, trichloromethane, and dichlorodifluoromethane for conditions approximating a 

deep, low-velocity stream, an intermediatedepth, high-velocity stream, and a shallow, 

medium-velocity stream indicate that the halflife decreases as the velocity increases for a 

constant depth. Conversely, the half-life 

increases as the flow depth increases for a 

constant velocity. The half-lives are longest for 

tribromomethane, which has the lowest Kerry's 

law constant of these three compounds. The 

half-lives for trichloromethane and dichlorodifluoromethane are similar, despite the fact that 

the Henry's law constant for dichlorodifluoromethane is about one-hundred times larger 

than that for trichloromethane. This lack of 

dependence on the Henry's law constant is 

attributed to the fact that these two compounds 

have Henry's law constants large enough tl at 

volatilization is controlled almost completely 

by the mixing conditions in the water. The 

three compounds were selected for this comparison because they have Henry's law constants 

near the low, middle, and high end of the distribution of Henry's law constants for the target 

analytes.

5. Volatilization half-lives for tribromomethane,

trichloromethane, and dichlorodifluoromethane 

decrease as the temperature increases for 

conditions characteristic of a deep, low-velocity 

stream, an intermediate-depth, high-velocity 

stream, and a shallow, medium-velocity 

stream. The decrease is largest for tribromomethane, which has a Henry's law constant 

low enough that the constant is important in 

contributing to the control of the volatilization 

process.

6. Absorption of VOCs from the atmosphere by a 

stream occurs when the ambient air concentration is greater than the concentration that would 

be in equilibrium with the ambient water
138 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

concentration. The importance of this process 

as a source of VOCs in streams has not been 

determined.

7. Dry deposition of VOCs with particles is not

expected to be important for the target analytes. 

Vapor pressures of the analytes are high enough 

that little sorption of the analytes to particulates 

in the atmosphere is expected.

8. Wet deposition of VOCs can occur with particles 

and precipitation, although little sorption of the 

target analytes to atmospheric particulates is 

expected. Wet deposition with precipitation 

is characterized by the gas scavenging ratio, 

which is the ratio of the concentrations of 

the organic compound in the precipitation and 

in the ambient air. Equilibrium is established 

rapidly between the air and precipitation 

concentrations, resulting in the gas scavenging 

ratio being inversely proportional to the 

Henry's law constant.

9. The gas scavenging ratio is directly dependent 

on the temperature and inversely dependent 

on the Henry's law constant. The exponential 

dependence of the Henry's law constant on 

temperature has a much greater effect on the gas 

scavenging ratio than the explicit dependence 

on temperature, resulting in the gas scavenging 

ratio increasing as the temperature decreases. 

Consequently, VOCs exhibit a greater tendency 

to partition into the precipitation as the temperature decreases. Gas scavenging ratios appear to 

depend somewhat on the type of precipitation, 

with fog having higher values than other types 

of precipitation, presumably because of the 

greater interfacial area between the air and 

water phases.

10. Microbial degradation is the use of a VOC by 

microorganisms as a source of food and 

energy. Degradation is a series of processes 

that are highly system-specific and site-specific 

and requires that the microbial community 

develop enzymes suitable for degrading the 

VOCs of interest. VOCs in some situations 

may be degraded by the process of cometabolism, where some other organic compound 

is the source of food and energy for the 

microorganisms.

11. The target analytes most likely to be subject 

to microbial degradation in streams are 

the aromatic compounds, 2-propenal, and 

2-propenenitrile.

12. Estimates of the importance of microtral degradation of naphthalene relative to volatilization 

indicate that microbial degradation could be 

important in deep, low-velocity streams, with 

lesser importance in intermediate-d^pth, highvelocity streams. Microbial degradation in 

shallow, medium-velocity streams is not 

expected to be important, relative to 

volatilization.

13. Sorption of VOCs to sediments in streams 

involves partitioning of the VOCs into the 

organic content of the sediment, with little 

physical adsorption to the mineral surfaces 

of the sediment.

14. Sorption to sediments is inversely proportional

to the water solubility and directly proportional 

to the octanol-water partition coefficient of 

the VOC. The target analytes most likely 

to sorb to sediments are hexachlorobutadiene, 

n-butylbenzene, and 1,2,3-trichlorobenzene.

15. The fraction of VOCs sorbed to a sediment

increases as the organic-carbon fraction of the 

sediment increases and as the sedirrent concentration increases. For sediment concentrations 

less than 100 milligrams per liter, estimates 

of the fractions sorbed are less than 0.1. The 

estimates of the fraction sorbed increases 

rapidly for higher sediment concentrations.

16. Hydrolysis is the reaction of a VOC with water, 

and nine target analytes have half-lives for 

hydrolysis of 548 days or less. The currently 

available experimental half-lives from the 

literature for several analytes vary by orders 

of magnitude, resulting in uncertairties in 

estimates of the importance of hydrolysis 

on concentrations of these VOCs in streams.

17. The estimates of the fractions of bromomethane, 

1,2-dibromoethane, and 1,3-dichloropropene 

removed by hydrolysis exceed estimates of the 

fraction removed by volatilization for a deep, 

low-velocity stream. The effect of hydrolysis 

also is predicted to be important for chloroethane for this type of stream. Hydrolysis of the 

target analytes is not significant relative to 

volatilization for a shallow, medium-velocity
SUMMARY AND CONCLUSIONS 139

stream. Hydrolysis is not significant relative to 

volatilization for an intermediate-depth, highvelocity stream, with the possible exceptions 

of 1,2-dibromoethane and 1,3-dichloropropene, 

for which some effect is predicted.

18. Aquatic photolysis is the reaction of a VOC with 

the sunlight incident upon the surface of a 

stream. Little information was found in the 

literature on aquatic photolysis of the target 

analytes. The analytes considered most likely 

to undergo photolysis are those with ring 

structures.

19. Estimates of the importance of aquatic photolysis 

of naphthalene relative to volatilization indicate 

that photolysis is not significant. These 

estimates are based on the assumption that 

light penetration and the resultant photolysis 

are limited by the depth of the water column. 

However, in a stream, the water at the surface 

is continuously replaced with water from below 

as a result of vertical mixing, and the depth 

may not be rate-limiting. If surface photolysis 

with continual replacement is assumed, 

estimates of the losses as a result of photolysis 

exceed volatilization losses at a depth of about 

3 m for naphthalene and at a depth of about 

5.7 m for benzene. These estimates of the 

extent of photolysis should be considered 

as upper bounds because of the assumption 

of no effect of depth on the photolysis process. 

Experimental determination of the effect of 

depth on photolysis in a stream with continual 

replacement of the surface water with water 

from the lower parts of the stream is needed.

20. Oxidation is the reaction of a VOC with an

oxidant. Oxidants most likely to be present in 

streams are the peroxy radical, photoexcited 

diradicals, hydroxy radical, singlet oxygen, 

and ozone. These oxidants are hypothesized 

to result from the photolysis of natural humic 

materials in the water.

21. Fourteen target analytes have experimental or 

calculated oxidation half-lives of less than 

1 year. These analytes are benzene, three alkyl 

benzenes, six halogenated aromatics, three 

halogenated alkenes, and 2-propenenitrile. 

Oxidation relative to volatilization is predicted 

to be significant for deep flows for methylbenzene and is predicted to be minimally important 

for benzene. The predictions for the other

12 analytes are intermediate between these 

extremes. These estimates of the extent of 

oxidation should be considered as upper bounds 

because of the assumption of no effect of d^pth 

on the oxidation process. Experimental determination of the effect of depth on oxidation 

in a stream with continual replacement of t>e 

surface water with water from the lower parts 

of the stream is needed.

22. Bioconcentration is the uptake of a VOC from the 

water of a stream by an aquatic organism. This 

process is analogous to the partitioning of ? 

VOC between water and an organic solvent 

such as octanol. Consequently, the bioconc^ntration factor is strongly dependent on the 

octanol-water partition coefficient. Logarithm 

base 10 octanol-water partition coefficients 

for the target analytes range from -0.92 for 2- 

propenenitrile to 4.78 for hexachlorobutadiene.

23. Estimates of the effect of bioconcentration on 

the concentration of hexachlorobutadiene 

in streams indicate that the fraction concentrated is generally less than 0.06 for a fish 

density typically used in the literature.

24. Analysis of aquatic toxicity values indicate that the 

concentrations of VOCs necessary to cause acute 

adverse effects on aquatic organisms are not likely 

to be routinely present in streams. Spills could 

result in short-term exposures to high concentrations. Target analytes most likely to cause adverse 

effects, should they occur, are bromomethane, 

hexachlorobutadiene, 1,2,3-trichlorobenzene, 

and 2-propenal. The effect of long-term chroric 

exposure to low concentrations of VOCs is largely 

unknown. However, persistence of VOCs in 

streams for long time periods is unlikely because 

of the volatile nature of these compounds. On the 

other hand, fairly long-term exposures could cccur 

during winter under ice cover.

25. VOCs in streams are advected downstream with a 

mean velocity averaged over the cross section 

of the channel.

26. Vertical mixing of VOCs in streams is rapid, with 

mixing over the flow depth usually complete 

within a distance downstream of about 100 flow 

depths.

27. Lateral mixing of VOCs in streams requires 

long distances downstream, with the 

distance required proportional to the square 

of the channel width. A VOC source on the
140 TRANSPORT, BEHAVIOR, AND FATE OF VOLATILE ORGANIC COMPOUNDS IN STREAMS

streambank requires four times the downstream distance for lateral mixing that a 

centerline source requires.

28. Longitudinal dispersion affects the variation 

of the concentration of VOCs with distance 

downstream for an instantaneous source of 

VOCs but has no effect on a continuous source 

because steady-state concentrations are 

established. The predicted effect of longitudinal 

dispersion relative to volatilization on the 

concentrations of dichloromethane and 

diisopropyl ether is substantial for a deep, lowvelocity stream. The predicted effect of longitudinal dispersion relative to volatilization is 

minimal for a shallow, medium-velocity stream. 

The predicted changes in the concentration as a 

result of longitudinal dispersion and volatilization for an intermediate-depth, high-velocity 

stream are approximately equal for dichloromethane and diisopropyl ether.

29. Volatilization is likely to be the dominant process 

affecting the concentrations of VOCs in 

streams. However, conclusions presented herein 

indicate that other physical as well as chemical 

and biological processes also could contribute 

to the determination of the behavior and fate of 

VOCs in streams. Consequently, each situation 

must be analyzed specifically to determine 

which processes are important and which 

processes can be neglected.

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