Environmental Protection
Office of Water
Planning and Standards IWH 5531
Washington DC 20460
                          December 1979
                          EPA-440 4-79-0291
Water-Related
Environmental Fate of
129 Priority Pollutants

Volume II
Halogenated Aliphatic Hydrocarbons
Halogenated Ethers
Monocyclic Aromatics
Phthalate Esters
Polycyclic Aromatic Hydrocarbons
Nitrosamines
Miscellaneous Compounds

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                                  DISCLAIMER
This report has been reviewed by the Office of Water Planning and Standards,
U.S. EPA, and approved for publication.   Approval does not signify that the
contents necessarily reflect the. views and policies of the U.S.  Environmental
Protection Agency, nor does mention of trade names or commercial products con-
stitute endorsement or recommendation for use.

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                                                         EPA-440/4-79-029b
                                                         December 1979
                   WATER-RELATED ENVIRONMENTAL FATE OF
                         129 PRIORITY POLLUTANTS
                               Volume II:
         Halogenated Aliphatic Hydrocarbons, Halogenated Ethers,
Monocyclic Aromatics, Phthalate Esters, Polycyclic Aromatic Hydrocarbons,
                Nitrosamines, and Miscellaneous Compounds
                                   by
                 Michael A. Callahan, Michael W. Slimak,
   Norman W. Gabel, Ira P. May, Charles F. Fowler, J. Randall Freed,
 Patricia Jennings, Robert L. Durfee, Frank C. Whitmore, Bruno Maestri,
          William R. Mabey, Buford R. Holt, and Constance Gould
                       EPA Contract No. 68-01-3852
                       EPA Contract No. 68-01-3867
                   Project Officer:  Donald J. Ehreth
              Monitoring and Data Support Division (WH-553)
                 Office of Water Planning and Standards
                         Washington, D.C. 20460
                 OFFICE OF WATER PLANNING AND STANDARDS
                  OFFICE OF WATER AND WASTE MANAGEMENT
                  U.S. ENVIRONMENTAL PROTECTION AGENCY
                         WASHINGTON, D.C. 20460
                                              p,

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                                   FOREWORD
     Effective regulatory action for toxic chemicals requires an understanding
of the human and environmental risks associated with the manufacture, use, and
disposal of the chemical.  The assessment of risk requires a scientific judg-
ment about the probability of harm to the environment resulting from known or
potential environmental concentrations.  Environmental concentrations are a
function of (1) the amount and form of the chemical released into the environ-
ment, (2) the geographic area, (3) prior accumulation, (4) time of measure-
ment, and (5) the behavior of the chemical in the environment.  The behavior,
or fate and transport characteristics, of toxic pollutants in the environment
depends on a variety of chemical, physical, and biological processes (e.g.,
photolysis, hydrolysis, volatilization, sorption, biodegradation, biotransfor-
mation).  Evaluating these processes for specific compounds and placing each
interaction into environmental perspective is the basic goal of this report.

     This two-volume report is a comprehensive review of the water-related
environmental fate and transport literature available for 129 chemical com-
pounds and elements, sometimes referred to as the 129 priority pollutants.
                                 Michael W. Slimak, Chief
                                 Exposure Assessment Section
                                 Water Quality Analysis Branch
                                 Monitoring and Data Support Division (WH-553)
                                 Office of Water Planning and Standards
                                      -i-

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                     TABLE OF CONTENTS AND LIST OF CHAPTERS
VOLUME I:  INTRODUCTION AND TECHNICAL BACKGROUND, METALS AND INORGANICS,
           PESTICIDES, AND PCBs.

Chapter    Description

           Section I;  Introduction and Technical Background
  1.               Introduction
  2.               Fate and Transport Processes
  3.               Determination of Water-Related Environmental Fate:
                     Procedures, Methods, and Report Format
  4.               Conclusions and Recommendations
           Section II;  Metals and Inorganics
  5.               Antimony
  6.               Arsenic
  7.               Asbestos
  8.               Beryllium
  9.               Cadmium
 10.               Chromium
 11.               Copper
 12.               Cyanide
 13.               Lead
 14.               Mercury
 15.               Nickel
 16.               Selenium
 17.               Silver
 18.               Thallium
 19.               Zinc
           Section III;  Pesticides
 20.               Acrolein
 21.               Aldrin
 22.               Chlordane
 23.               ODD
 24.               DDE
 25.               DDT
 26.               Dieldrin
 27.               Endosulfan and Endosulfan Sulfate
 28.               Endrin and Endrin Aldehyde
 29.               Heptachlor
 30.               Heptachlor Epoxide
 31.               Hexachlorocyclohexane (& 8,  6 isomers)
 32.               Y~Hexachlorocyclohexane (Lindane)
 33.               Isophorone
 34.               TCDD
 35.               Toxaphene
           Section IV;  PCBs and Related Compounds
 36.               Polychlorinated Biphenyls
 37.               2-Chloronaphthalene
                                      -iii-

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VOLUME II:  HALOGENATED ALIPHATIC HYDROCARBONS, HALOGENATED ETHERS, MONOCYCLIC
            AROMATICS, PHTHALATE ESTERS, POLYCYCLIC AROMATIC HYDROCARBONS,
            NITROSAMINES, AND MISCELLANEOUS COMPOUNDS.

Chapter     Description

            Section V;  Halogenated Aliphatic Hydrocarbons
 38.               Chloromethane (Methyl Chloride)
 39.               Dichloromethane (Methylene Chloride)
 40.               Trichloromethane (Chloroform)
 41.               Tetrachloromethane (Carbon Tetrachloride)
 42.               Chloroethane (Ethyl Chloride)
 43.               1,1-Dichloroethane (Ethylidine Chloride)
 44.               1,2-Dichloroethane (Ethylene Dichloride)
 45.               1,1,1-Trichloroethane (Methyl Chloroform)
 46.               1,1,2-Trichloroethane
 47.               1,1,2,2-Tetrachloroethane
 48.               Hexachloroethane
 49.               Chloroethene (Vinyl Chloride)
 50.               1,1-Dichloroethene (Vinylidine Chloride)
 51.               1,2-trans-Dichloroethene
 52.               Trichloroethene
 53.               Tetrachloroethene (Perchloroethylene)
 54.               1,2-Dichloropropane
 55.               1,3-Dichloropropene
 56.               Hexachlorobutadiene
 57.               Hexachlorocyclopentadiene
 58.               Bromomethane (Methyl Bromide)
 59.               Bromodichloromethane
 60.               Dibromochloromethane
 61.               Tribromomethane (Bromoform)
 62.               Dichlorodifluoromethane
 63.               Trichlorofluoromethane
            Section VI;  Halogenated Ethers
 64.               Bis(chloromethyl)ether
 65.               Bis(2-chloroethyl)ether
 66.               Bis(2-chloroisopropyl)ether
 67.               2-Chloroethyl vinyl ether
 68.               4-Chlorophenyl phenyl ether
 69.               4-Bromophenyl phenyl ether
 70.               Bis(2-chloroethoxy)methane
            Section VII;  Monocyclic Aromatics
 71.               Benzene
 72.               Chlorobenzene
 73.               1,2-Dichlorobenzene (j3-Dichlorobenzene)
 74.               1,3-Dichlorobenzene (m-Dichlorobenzene)
 75.               1,4-Dichlorobenzene (p_-Dichlorobenzene)
 76.               1,2,4-Trichlorobenzene
                                   -iv-

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Chapter     Description

            Section VII;  (continued)
 77.               Hexachlorobenzene
 78.               Ethylbenzene
 79.               Nitrobenzene
 80.               Toluene
 81.               2,4-Dinitrotoluene
 82.               2,6-Dinitrotoluene
 83.               Phenol
 84.               2-Chlorophenol
 85.               2,4-Dichlorophenol
 86.               2,4,6-Trichlorophenol
 87.               Pentachlorophenol
 88.               2-Nitrophenol
 89.               4-Nitrophenol
 90.               2,4-Dinitrophenol
 91                2,4-Dimethyl phenol
 92.               jD-Chloro-m-cresol
 93.               4,6-Dinitro-o-cresol
            Section VIII;  Phthalate Esters
 94.               Phthalate Esters:  Dimethyl, Diethyl,
                     Di-n-butyl, Di-n-octyl, Bis(2-ethylhexyl),
                     and Butyl benzyl
            Section IX;  Polycyclic Aromatic Hydrocarbons
 95.               Polycyclic Aromatic Hydrocarbons: Acenaphthene,
                     Acenaphthylene, Fluorene, and Naphthalene
 96.               Polycyclic Aromatic Hydrocarbons:  Anthracene,
                     Fluoranthene, and Phenanthrene
 97.               Polycyclic Aromatic Hydrocarbons:  Benzo[a]anthracene,
                     Benzo[b]fluoranthene, Benzo[k]fluoranthene, Chrysene,
                     and Pyrene
 98.               Polycyclic Aromatic Hydrocarbons:  Benzo[ghi]perylene,
                     Benzo[a]pyrene, Dibenzo[a]anthracene, and
                     Indeno[1,2,3-cd]pyrene
            Section X;  Nitrosamines and Miscellaneous Compounds
 99.               Dimethyl nitrosamine
100.               Diphenyl nitrosamine
101.               Di-n-propyl nitrosamine
102.               Benzidine
103.               3,3'-Dichlorobenzidine
104.               1,2-Diphenylhydrazine (Hydrazobenzene)
105.               Acrylonitrile
                                      -v-

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                               ACKNOWLEDGEMENTS
    A number of people have actively participated in this study, which was
begun in September of 1977.  The assistance, support, and management guidance
provided by Martin P. Halper, former Chief of the Water Quality Analysis
Branch (1976-1978), and Donald J. Ehreth, present Chief of the Water Quality
Analysis Branch, U.S. EPA, is gratefully acknowledged.

    Michael A. Callahan, U.S. EPA, conceptualized and managed the Phase I
tasks and was the principal technical editor of all fate chapters.  In addi-
tion, he performed part of the literature search for the halogenated aliphatic
hydrocarbons and participated in extensive rewrites of that section.  Michael
W. Slimak, U.S. EPA, managed the Phase II tasks, provided additional editorial
assistance, and completed the final report.  Acknowledgement is also given to
the staff members of the Water Quality Analysis Branch.  Specifically, our
thanks to:  Rod Frederick, Lynn Delpire, Charles Delos, Mark Segal, Charles
Gentry, Chris Ehret, and Richard Seraydarian.

    Most of the literature search and the writing of the report was done by
two contractors, Versar, Inc. and SRI International.  The following individ-
uals participated as shown:

     1.  VERSAR INC. (EPA Contract No. 68-01-3852)
           Program Management;  Robert L. Durfee, Donald H. Sargent, Gayaneh
           Contos, Bruno Maestri
           Technical Direction;  Michael W. Slimak (1977-78), Norman W. Gabel
           (1978-80)
           Authors;
               Metals and Inorganics:  J. Randall Freed, Ira P. May
               PCBs and Related Compounds:  J. Randall Freed, Frank C.
                  Whitmore, Charles F. Fowler
             *  Halogenated Aliphatic Hydrocarbons:  Patricia Jennings,
                  Robert L. Durfee
               Halogenated Ethers:  Norman W. Gabel
               Monocyclic Aromatics:  Patricia Jennings, Norman W. Gabel,
                  Robert L. Durfee, Charles F. Fowler
             *  Monocyclic Aromatics - Phenols and Cresols:  Norman W. Gabel
               Phthalate Esters:  Bruno Maestri, Charles F. Fowler
               Polycyclic Aromatic Hydrocarbons:  Michael W. Slimak,
                  Charles F. Fowler
               Nitrosamines and Miscellaneous Compounds:  Norman W. Gabel
           Data Gathering and Automated Literature Searches:  Theodore French,
           Michael Keller
           Editing;  Norman W. Gabel, Robert L. Durfee, Bruno Maestri, Juliet
           Crumrine
                                       -vii-

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         VERSAR INC. (continued)

         Typing;  Laura Skiba, Shirley Harrison,  Lorraine Douglas, Kathy
         Turnblora, Rebecca Brown, Nancy Downie,  Neda Helmandollar
         Graphics and Reproduction;   Kenneth Ratkiewicz,  Joseph Gillette,
         Carole Craddock, Michael Cairns.

     2.  SRI INTERNATIONAL (EPA Contract No. 68-01-3867)

         Program Management;   Stephen L. Brown,  Oscar Johnson, William R.
         Ma bey
         Technical Direction;   William R.  Mabey
         Authors;

              Pesticides:  William R. Mabey, Buford Holt, Constance
                 Gould

         Automated Literature  Searches:  Shirley Radding, Jerie Etherton
         Editing;  Lee Work
         Typing;  Kathleen Williams, Maria Buyco.
    The general assistance and guidance provided by personnel within EPA's
Athens Environmental Research Laboratory,  especially George Baughman and his
staff, is also greatly appreciated.

    Numerous individuals provided comments and suggestions on the draft
report; their efforts are appreciated.   A special note of gratitude is due
those who carefully reviewed the draft  report and provided timely and
pertinent comments.  These reviewers include:

              G.  Baughman, U.S.  EPA-Athens
              G.  Zweig, U.S. EPA-Washington
              J.  Kariya, U.S. EPA-Washington
              J.  Cohen, U.S. EPA-Cincinnati
              J.  Gillette, U.S.  EPA-Corvallis
              M.  Strier, U.S. EPA-Washington
              J.  Butler, Harvard University
              R.  Baker, U.S. Geological Survey
              K.  Irgolic, Texas A&M University
              R.  Bailey, Dow Chemical U.S.A.
              W.  Dilling, Dow Chemical  U.S .A.

    Special gratitude is due Bruno Maestri and Shirley Harrison,  Versar Inc.,
for their efforts in completing the final  document, especially their attention
to detail.
                                      -viii-

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SECTION V;  HALOGENATED ALIPHATIC HYDROCARBONS




                Chapters 38-63

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                   38.  CHLOROMETHANE (METHYL CHLORIDE)

38.1  Statement of Probable Fate

    Volatilization is the major transport process for removal of chloro-
methane from aquatic systems.  Once in the troposphere chloromethane is at-
tacked by hydroxyl radicals with the subsequent formation of formyl chlo-
ride as the principal initial product.  Any unreacted chloromethane reach-
ing the stratosphere will undergo photodissociation.

    Based on the information currently reviewed, it appears that oxidation,
hydrolysis, and biodegradation are not important fate processes of chloro-
methane in the aquatic environment.  No information was found indicating
that adsorption and bioaccumulation are important environmental transport
processes for chloromethane.

38.2  Identification

    Chloromethane is known to be ubiquitous in the environment.  It enters
the environment from natural as well as anthropogenic sources, and has been
detected in finished drinking water (Environmental Protection Agency 1975),
in seawater (Lovelock 1975; Singh et al. 1978), and in the troposphere
(Singh et al. 1978).

    The chemical structure of chloromethane is shown below.

                                        Alternate Names

                                       Methyl chloride
     Cl	 C 	H                      Mono chloromethane
           H

    Chloromethane

    CAS NO. 74-87-3
    TSL NO. PA 63000

38.3  Physical Properties

    The general physical properties of chloromethane are as follows.

    Molecular weight                          50.59
    (Weast 1977)

    Melting point                            -97.73C
    (Weast 1977)
                                  38-1

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    Boiling point at 760 torr                -24.2C
    (Weast 1977)

    Vapor Pressure at 20C                    3765 torr
    (McConnell et al. 1975)

    Solubility in water at 20C               6450 to 7250 mg/1*
    (McConnell _et al. 1975)

    Log octanol/water partition coefficient   0.91
    (Hansch et al. 1975)
*Several values for solubility of chioromethane in water at 20C were found
in the literature.  These values range from 6450 mg/1 (Dean 1973) to 7250
mg/1 (Pearson and McConnell 1975).

38.4  Summary of Fate Data

    38.4.1  Photolysis

         No information was found pertaining to the rate of photolysis of
chloromethane in the aqueous environment under ambient conditions.  Dilling
ot_ al_. (1975) studied the decomposition rates of dichloromethane and tri-
chloromethane, compounds structurally similar to chloromethane.  At a con-
centration of one ppm in water, in aerated sealed tubes, Dilling et al.
(1975) found no significant difference between decomposition rates of these
compounds in tubes exposed to sunlight and those kept in dark over periods
of 6 and 12 months.

         Due to the high vapor pressure of chloromethane, volatilization to
the atmosphere is believed to be quite rapid.  Inasmuch as the ozone layer
above the earth absorbs all wavelengths of light below about 290 nm (Roller
1965) and chloromethane does' not absorb light at wavelengths above 290 nm,
appreciable photodissociation below the ozone layer (in the troposphere)
would not be expected to occur.  Chloromethane in the troposphere is at-
tacked, via hydrogen abstraction, by hydroxyl radicals (Hanst 1978; Spence
jet a^. 1976).  The rates are such that this photooxidation reaction appears
to be the predominant fate process for chloromethane.

         The wavelength region in which most photodissociation in the stra-
tosphere occurs is from about 180 to 240 nm (Robbins 1976).  It has been
experimentally determined that chloromethane undergoes light absorption
from 174 to about 200 nm (Robbins 1976).  As a result, any urireacted chlo-
romethane reaching the stratosphere would be expected to undergo signifi-
cant photodissociation.
                                     38-2

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    38.4.2  Oxidation

         No information was found pertaining specifically to the oxidation
of chioromethane in the aqueous environment.  Billing ejt al. (1975) con-
ducted aqueous reactivity experiments on CHC12 and CHC13 Tcompounds
structurally similar to chloromethane) in which there was a large excess of
dissolved oxygen compared to the total amount of dichloromethane and tri-
chloromethane.  The air space above the solution in the closed system con-
tained approximately 90 times as much oxygen as was present in the satu-
rated solution.  Since the observed rates of decomposition were much
smaller than rates of volatilization determined by the same authors for
chloromethane, dichloromethane and tricoloromethane, and since most of the
decomposition was attributed to ionic hydrolysis rather than oxidation, it
is concluded that oxidation in the aqueous system is not an important pro-
cess for chloromethane.

         Due to the high vapor pressure of chloromethane, volatilization
to the atmosphere should be quite rapid, and once in the troposphere, chlo-
romethane is attacked by hydroxyl radicals via the mechanism of hydrogen
abstraction (Hanst 1978; Spence t _al. 1976).  The bimolecular rate con-
stant for reaction of chloromethane with *OH is reported by Cox _e_t al.
(1976) to be 8.5 x 10~14cm3sec"1, corresponding to a lifetime (time
for reduction to 1/e) of 0.37 years.  This agrees reasonably well with the
rate constant of 1.69 x lO"1^ exp (-1066/T) cm^sec"1, or 4.7 x
10~1^cm3sec~1 at 25C, reported by Yung t al. (1975) and reflect-
ing a lifetime, described as an upper limit, of 0.89 years.

         The principal initial product of this photooxidation reaction is
reported to be formyl chloride (HCOC1), with hydrogen peroxide (^02) >
carbon monoxide (CO) and hydrogen chloride (HC1) being formed in smaller
quantities (Spence ^t al. 1976).  Formyl chloride is readily hydrolyzed to
HC1 and CO (Morriso~and Boyd 1973).

         With the assumption of 30 years as the troposphere-to-stratosphere
turnover time (time for all but 1/e of tropospheric air to diffuse into the
stratosphere), the 0.37 year tropospheric lifetime for chloromethane of Cox
et al. (1976) corresponds to about one percent of tropospheric chloro-
methane eventually reaching the stratosphere.

    38.4.3  Hydrolysis

         A hydrolytic half-life of 417 days (1.14 years) has been reported
for chloromethane at pH 7 and 25C (Mabey and Mill 1978; Radding et_ al.
1977).  This corresponds to a first-order rate constant for hydrolysis of
chloromethane of 1.9 x 10~8sec~1 (Mabey and Mill 1978; Radding et al.
1977).  The solvolysis of chloromethane in water at 20C is reported to
proceed at a pseudo-first order rate of 8.9 x 10~'sec~^ corresponding
                                      38-3

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to a half-life of 2.5 years (Zafirion 1975); however, no measurement of pH
was given.  Due to the high vapor pressure and low boiling point of this
compound, it is expected that most of the chloromethane present in water
will be transported to the atmosphere more rapidly than it will react with
water.

    38.4.4  Volatilization

         Billing e al. (1975) estimated the experimental half-life for
volatilization of chloromethane originally present at 1 mg/1 to be 27
minutes when stirred at 200 rpm in water at approximately 25C in an open
container.  Removal of 90 percent of the chloromethane under the same con-
ditions required 91 minutes.  For chloroaliphatics in general, stirring
speed was found to have a marked effect on volatilization rate.  With no
stirring except 15 seconds every five minutes the time required for 50 per-
cent depletion of compounds such as trichloromethane, trichloroethene, and
1,1,1-trichloroethane, which are structurally similar to chloromethane, was
greater than 90 minutes, or on the order of three times greater than the
stirred case for these compounds.

         Evaporation appears to be the major pathway by which chloromethane
is transported from the water.  Billing et_ ai_. (1975) are careful to point
out the difficulties encountered in extrapolating their laboratory results
to real-world conditions, where the concentration of the organic solute
would probably be very much less than 1 mg/1 and where surface and bulk
agitation would be highly variable.  Although the data appear to be valid
on a relative basis (i.e., correctly illustrating the relative rates of
volatilization of chlorinated aliphatics), they cannot be used as absolute
measures of volatilization rates from natural waters.  For the purposes of
this document, the data are used as rough-order-of-magnitude indications of
the importance of volatilization relative to other transport and fate pro-
cesses, with the strong effects of agitation considered.  The validity of
this application has not been established.

         A subsequent study by Dilling (1977) used the same experimental
conditions as in the 1975 study, and an average half-life for volatiliza-
tion of 27.6 minutes for 1 mg/1 chloromethane was obtained.  The purpose of
this subsequent study was to use the experimental data obtained to test two
theoretical models formulated to predict evaporation rates of slightly
soluble organic compounds from water.  Dilling (1977) found that the the-
oretical model of Mackay and Wolkoff (1973) failed to predict evaporative
half-lives, but that the model of Mackay and Leinonen (1975) using para-
meters from Liss and Slater (1974) correlated well with experimental re-
sults.  For example, the evaporative half-life for chloromethane obtained
experimentally by Dilling (1977) was approximately 28 minutes as compared
to 14.9 minutes predicted by the Mackay and Leinonen (1975) model and 0.599
minutes predicted by the Mackay and Wolkoff (1973) model.
                                   38-4

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         Billing (1977), however, comments that the apparent numerical
agreement between his data and the values predicted by the Mackay and
Leinonen (1975) model may be fortuitous.  Estimates of volatilization rates
based on the Mackay and Leinonen (1975) model depend primarily on liquid-
gas phase exchange rate constants, whereas the experimental model of
Billing e_t _a^L. (1975) and Billing (1977) is controlled by the rate of
stirring and the wind velocity across the surface of the water.

         Pearson and McConnell (1975) suggest that the presence of chloro-
methane in ambient waters is due to absorption of chloroorganics from the
atmosphere by water.  This absorption is thought to occur most effectively
when the atmosphere is scrubbed during the precipitation process.  Aerial
transport of these chloroorganics, as well, is postulated by Pearson and
McConnell (1975) to play a major role in the wide distribution of these
compounds and accounts for their presence in upland waters.

    38.4.5  Sorption

         No data on sorption processes specifically involving chloromethane
were found.  The importance of sorption relative to other transport and
fate processes is controlled by the partitioning of the subject chemical
between the particulate, aqueous and air phases.  In the case of chloro-
methane, its high vapor pressure, only moderately low aqueous solubility
and relatively low (for chlorinated aliphatics) octanol/water partition
coefficient seem to indicate that this compound would partition primarily
into the air and water compartments rather than into sediments and sus-
pended particulates.

    38.4.6  Bioaccumulation

         According to Kopperman e_t al. (1976) not all organochlorine com-
pounds bioaccumulate to high levels.  The data suggest that polar com-
pounds are more easily biodegraded, whereas non-polar, highly lipophilic
compounds accumulate.  Neely et al. (1974) have shown that bioaccumulation
is directly related to the octanol/water partition coefficient (P) of the
compound.  The log octanol/water partition coefficient (lop P) of chloro-
methane is 0.91 (Hansch e_t al_. 1975), indicating that chloromethane is not
very lipophilic.  As a result, chloromethane would not be expected to ex-
hibit a significant tendency to bioaccumulate in organisms.

    38.4.7  Biotransformation and Biodegradation

         No information was found pertaining specifically to the biodegra-
dation of chloromethane.  If the partitioning of chloromethane in the aqu-
                                    38-5

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eous environment favors the air and aqueous compartments over particu-
lates (including biota), as postulated above in the section on sorption,
then biodegradation would not be expected to be an important fate process.
for this chemical.

38.5  Data Summary

    Table 38-1 summarizes the aquatic fate data found above.  Oxidation
rates are photooxidation rates and refer to the rate of reaction of chloro-
methane with hydroxyl radicals in the troposphere.

    The predominant fate of chloromethane is a result of the high vapor
pressure of this compound which causes it to volatilize rapidly into the
atmosphere.  Once in the troposphere, chloromethane is attacked by hydroxyl
radicals resulting in the subsequent formation of formyl chloride as the
principal product.
                                   38-6

-------



















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38.6  Literature Cited

Cox, R.A.,  R.G. Derwent,  A.E.J.  Eggleton,  and  J.E.  Lovelock.   1976.
  Photochemical oxidation of halocarbons  in the  troposphere.   Atmos.
  Envir. 10:305-308.

Dean, J.A.  (Ed).  1973.   Lange's handbook of chemistry.   llth Edition.
  McGraw-Hill Book Company,  New  York.   1576 p.

Billing, W.L.  1977.   Interphase transfer processes.   II.   Evaporation
  rates of  chloromethanes,  ethanes,  ethylenes,  propanes,  and  propylenes
  from dilute aqueous solutions.  Comparisons  with  theoretical predictions.
  Environ.  Sci. Technol.   11(4):405-409.

Dilling, W.L., N.B. Tefertiller, and G.J. Kallos.   1975.   Evaporation rates
  of methylene chloride,  chloroform, 1,1,1-trichloroethane, trichloro-
  ethylene, tetrachloroethylene  and  other chlorinated  compounds in dilute
  aqueous solutions.   Environ. Sci.  Technol. 9(9):833-838.

Environmental Protection Agency.  1975.  Preliminary  assessment of
  suspected carcinogens  in drinking  water.  Environmental  Protection Agency
  (Office of Toxic Substances),  Washington, D.C. 33 p. EPA 560/4-75-003.

Hansch, C., A. Vittoria,  C.  Silipo,  and P.Y.C. Jow.  1975.  Partition
  coefficients and the structure - activity relationship  of the anesthetic
  gases. J. Med. Chem.  18(6):546-548.

Hanst, P.L.  1978.  Part II:  Halogenated pollutants.   Noxious trace gases
  in the air.  Chemistry 51(2):6-12.

Roller, L.R.  1965.  Solar radiation in ultraviolet radiation.  2nd
  Edition.  J. Wiley and Sons, New York.

Kopperman, H.L., D.W. Kuehl, and G.E. Glass.  1976.  Chlorinated compounds
  found in waste treatment effluents and their capacity to bioaccumulate.
  Proceedings of the conference on the environmental  impact of water
  chlorination.  Oak Ridge, Tennessee, October 22-24,  1975.  345 p.

Liss,  P.S. and P.G. Slater.  1974.  Flux of gases across the air-sea
  interface.  Nature 247:181-184.

Lovelock, J.E.  1975.  Natural halocarbons in  the air and in the sea.
  Nature 256:193-194.

Mabey, W. and T. Mill.  1978.  Critical review of hydrolysis of organic
  compounds in water under environmental conditions.   Phys. Chem. Ref.  Data
  7(2):383-415.
                                   38-8

-------
Mackay, D. and P.J. Leinonen.  1975.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. Technol.
  9(13):1178-1180.

Mackay, D. and A.W. Wolkoff.  1973.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. Technol.
  7(7):611-614.

McConnell, G., D.M. Ferguson, and C.R. Pearson.  1975.  Chlorinated
  hydrocarbons and the environment.  Endeavor XXXIV: 13-18.

Morrison, R.T. and R.W. Boyd.  1973.  Organic chemistry.  3rd Edition.
  Allyn and Bacon, Inc., Boston, Mass. 1258 p.

Neely, W.B., D.R. Branson, and G.E. Blau.  1974.  Partition coefficient
  to measure bioconcentration potential of organic chemicals in fish.
  Environ. Sci. Technol. 8(13):1113-1115.

Pearson, C.R. and G. McConnell.  1975.  Chlorinated C^ and G
  hydrocarbons in the marine envirofnment.  Proc. Roy. Soc. London B
  189:305-332.

Radding, S.B., D.H. Liu, H.L. Johnson, and T. Mill.  1977.  Review of the
  environmentral fate of selected chemicals.  Environmental Protection
  Agency (Office of Toxic Substances), Washington, B.C. 147 p. EPA
  560/5-77-003.

Robbins, D.E.  1976.  Photodissociation of methyl chloride and methyl
  bromide in the atmosphere.  Geophys. Res. Lett. 3(4):213-216.

Singh, H.B., L.J. Salas, H. Shiegeishi, and A.H. Smith.  1978.  Fate of
  halogenated compounds in the atmosphere interim report - 1977.
  Environmental Protection Agency (Office of Research and Development),
  Research Triangle Park, N.C. 57 p.  EPA 600/3-78-017.

Spence, J.W., P.L, Hanst, and B.W. Gay, Jr.  1976.  Atmospheric oxidation
  of methyl chloride, methylene chloride, and chloroform.  J. Air Pollut.
  Control Assoc.  26(10):994-996.

Weast, R.C. (Ed).  1977.  Handbook of chemistry and physics.  58th Edition.
  CRC Press, Cleveland, Ohio.

Yung, Y.L., M.B. McElroy, and S.C. Wofsy.  1975.  Atmospheric halocarbons:
  A discussion with emphasis on chloroform.  Geophys. Res. Lett.
  2(9):397-399.

Zafirion, O.C.  1975.  Reaction of methyl halides with seawater and marine
  aerosols.  J. Mar. Res. 33:75-81.
                                    38-9

-------
                 39.  DICHLOROMETHANE (METHYLENE CHLORIDE)
39.1  Statement of Probable Fate

    Volatilization is the major transport process for removal of dichloro-
methane from aquatic systems.  Once in the troposphere, dichloromethane is
attacked by hydroxyl radicals with the subsequent formation of carbon di-
oxide, carbon monoxide, and phosgene as principal, initial products.  Any
unreacted dichloromethane which reaches the stratosphere will undergo pho-
todissociation.

    Based on the information reviewed, it appears that neither oxidation
nor hydrolysis is an important fate process of dichloromethane in the
aquatic environment.  No information was found to indicate that either
adsorption or bioaccumulation is an important transport process for di-
chloromethane.

39.2  Identification

    Dichloromethane (methylene chloride) has been detected in finished
drinking water (Environmental Protection Agency 1975) and in the tropo-
sphere (Singh et_ al. 1978).

    The chemical structure of dichloromethane is shown below.

            Cl
                                       Alternate Names
      /"M     p ^^^ ij
                                      Methylene chloride
                                      Methylene dichloride
            H                          Methane dichloride
                                      Methylene bichloride
    Dichloromethane

    CAS NO. 75-09-2
    TSL NO. PA 80500

39.3  Physical Properties

    The general physical properties of dichloromethane are as follows.

    Molecular weight                         84.94
    (Weast  1977)

    Melting point                            -95C
    (Weast  1977)
                                   39-1

-------
    Boiling point at 760 torr                39.75C
    (Weast 1977)

    Vapor pressure at 20C                   362.4 torr
    (Pearson and McConnell 1975)

    Solubility in water at 25C              13,200 to 20,000 mg/1*

    Log octonal/water partition coefficient  1.25
    (Hansch et al. 1975)
*Several values for solubility of dichloromethane in water at 20C were
found in the literature.  These values range from 13,200 mg/1 (Pearson and
McConnell 1975) to 20,000 mg/1 (Dean 1973)

39.4  Summary of Fate Data

    39.4.1  Photolysis

         Dilling j2_t al. (1975) studied the rate of decomposition of di-
chloromethane at a r~mg/l concentration in water in aerated sealed tubes
and found no significant difference between the half-life of dichloro-
methane in tubes exposed to sunlight and those kept in darkness over a
period of one year.  This result indicates that photochemical reactions in
aqueous media are probably not important for dichloromethane.

         Photodissociation in the aquatic environment would not be expected
to occur for dichloromethane since this compound has no chrocaophores which
absorb in the visible or near ultraviolet region of the electromagnetic
spectrum (Jaffe and Orchin 1962).  An experiment was conducted by Dilling
et al. (1976) which simulated environmental exposure to sunlight (wave-
TengThs > 290 nm and a  temperature of 27 + 1C).  Experimental results
showed that dichloromethane decomposed at a rather slow rate under tropos-
pheric conditions with  either NO or N0 present; the estimated photo-
decomposition half-life was greater than 250 hours.

    39.4.2  Oxidation

         Little information pertaining to the oxidation of dichloromethane
in the aqueous environment under ambient conditions was found.  Dilling et
al. (1975) conducted aqueous reactivity experiments in which there was a
Ta~rge excess of dissolved oxygen compared to the total amount of dichloro-
methane.  The air space above the solution in the closed system contained
approximately 90 times  as much oxygen as was present in the saturated solu-
tion.  The observed rate of disappearance, which was attributed primarily
                                      39-2

-------
to ionic hydrolysis rather than oxidation, was much lower than rates of  .
volatilization reported for dischloromethane by the same authors.  From
these results, it is concluded that oxidation in the aqueous phase probably
is not important for dichloromethane.  Due to the high vapor pressure of
diehioromethane, volatilization to the atmosphere is quite rapid.  Once in
the troposphere, the compound is attacked by hydroxyl radicals (OH') via
hydrogen abstraction (Hanst 1978- Spence 1976).  The rate of this reaction
is reported to be 1.04 x 10~13cm3sec-l corresponding to a lifetime of
0.30 years (Cox et al. 1976).  The principal product of this photooxidation
reaction is reported to be carbon dioxide (003), with carbon monoxide
(CO) and some phosgene (COC12) being formed in smaller quantities (Spence
_e_t _al^. 1976).  Phosgene is readily hydrolyzed to HC1 and C02 (Morrison
and Boyd (1973).

         For a troposphere-to-stratosphere turnover time (time for all but
1/e of tropospheric air to diffuse into the stratosphere) of 30  years, a
0.30 year tropospheric life time (Cox _e_t al. 1976) would result  in about
one percent of tropospheric dichloromethane reaching the stratosphere.
Once in the stratosphere, dichloromethane would be expected to undergo pho-
dissociation resulting from interaction with high energy ultraviolet radi-
ation.

  .  39.4.3  Hydrolysis

         A maximum hydrolytic half-life of 704 years, extrapolated from ex-
perimental data obtained at 100-150C, has been reported for dichlorome-
thane at pH 7 and 25C (Radding _e_t _al. 1977).  This corresponds  to a re-
ported first-order rate constant for hydrolysis of dichloromethane of
3.2 x 10" *lsec~l.  The validity of the extrapolation method used by
Radding &t_ al_. (1977) has not been established.  The data above  are not at
all in agreement with the results of the aqueous reactivity experiments of
Billing _et_ _a_l. (1975), described previously, which indicated a (first-
order) rate of disappearance at 25C of 0.039 months"^, corresponding to
a half-life of approximately 18 months for dichloromethane.  The authors
attributed most of the effect noted to ionic hydrolysis, although oxidation
and vapor-phase hydrolysis were also possible in the experimental system
used.  At best, the results of Billing et al. can be interpreted as a
maximum rate (minimum half-life) for hydrolysis of dichloromethane in
aqueous systems.

         A study by Fells and Moelwyn-Hughes (1958) investigated the be-
havior of dichloromethane when hydrolyzed in water under acidic  and basic
conditions in the temperature range 353-423K (80-150C).  The hydrolysis
of dichloromethane in acidic solution follows a simple, first-order rate
law.  The acid hydrolysis of dichloromethane at 373.16K (100C) is re-
ported to proceed at a rate of 5.77 + 0.05 sec~l, which corresponds to a
measured half-life period of 19,800 minutes  (13.75 days).  Experimentally,
                                     39-3

-------
it was found that hydrochloric acid and formaldehyde are produced in equi
valent concentrations from the beginning of the reaction.  The reaction
takes place in two stages, and can be represented as follows:


         (1)  CH2C12 + H20 - ~CH2C1(OH) + HC1
         (2)  CH2Cl(OH)i - ~CH20 + HC1
         The kinetics of the hydrolysis of dichloromethane in alkaline
solution were examined by Fells and Moelwyn-Hughes (1958) with potassium
hydroxide as the catalytic agent.  Reaction between dichlororaethane and
hydroxide ion in aqueous solution is not a simple second-order reaction.
Reaction with the solvent (water) occurs, and, in addition, the formal-
dehyde being formed reacts with the potassium hydroxide, giving methyl
alcohol and formate ion as products (Cannizzaro reaction).  The reaction
can be represented as follows:

                         k1
  (1)  CH2C12 + HOH - ^ - *~ ^2 + 2 H+ + 2 Cl~
  (2)  CH2C12 + 2 OH~ - k - ^ CH20 + 2 Cl~ + H20
  (3)  2 CH20 + OH~ - - HC02- + CH3OH
       (Cannizzaro Reaction)

The alkaline hydrolysis of dichloromethane at 373.16 K (100C) from re-
action (2) above is reported to be 1.61 x 10~^1 mole~lsec~^ .

    39.4.4  Volatilization

         Billing et_ a_l . (1975) estimated the experimental half-life for
volatilization of dichlororaethane originally present at 1 mg/1 to be
21 ^ 3 minutes when stirred at 200 rpm in water at approximately 25C in
an open container.  Removal of 90 percent of the dichloromethane under the
same conditions required 60 minutes.  For chloroaliphatics in general,
stirring speed was found to have a marked effect on volatilization rate.
With minimal stirring (15 seconds every five minutes), the time required
for 50 percent depletion of dichloromethane was approximately 90 minutes,
which is four times greater than observed with constant stirring.  The
presence of sodium chloride at a 3 percent concentration, as in seawater,
effected a 10 percent decrease in the evaporative rate.

         From the results of the above experiments, Billing e_t^ a]^. (1975)
conclude that evaporation is probably the major pathway by which dichloro-
methane is lost from  water.  In this context, Billing e_t aJL, (1975) are
careful to point out  the difficulties encountered in extrapolating their
laboratory results to real-world conditions, where the concentration of the
organic solute would  probably be very much less than 1 mg/1 and where sur-
                                    39-4

-------
face and bulk agitation would be highly variable.  Although the data appear
to be valid on a relative basis (i.e., correctly illustrating the relative
rates of volatilization of chlorinated aliphatics), they cannot be used as
absolute measures of volatilization rates from natural waters.  For the
purposes of this document, the data are used as rough-order-of-magnitude
indications of the importance of volatilization relative to other transport
and fate processes, with the strong effects of agitation considered.  The
validity of this application has not been established.

         A subsequent study by Billing (1977) was conducted using the same
experimental conditions as in the aforementioned investigation, and an
average half-life of 25.2 minutes with respect to evaporation of 0.99 mg/1
dichloromethane was obtained.  The purpose of this subsequent study was to
use the experimental data that was obtained to test two theoretical models
formulated to predict evaporative rates of slightly soluble organic com-
pounds from water.  Billing (1977) found that the theoretical model pro-
posed by Mackay and Wolkoff (1973) failed to predict evaporative half-
lives, but that the theoretical model of Mackay and Leinonen (1975) using
the parameters of Liss and Slater (1974) correlated well with the experi-
mental half-lives obtained.  For example, the evaporative half-life ob-
tained experimentally by Billing (1977) was 18 to 25 minutes as compared to
a 20.7 minute evaporative half-life calculated by the Mackay and Leinonen
(1975) model using the parameters of Liss and Slater (1974), whereas the
half life predicted by the Mackay and Wolkoff (1973) model was 2.23
minutes.

         Billing (1977), however, comments that the apparent numerical
agreement between his data and the values predicted by the Mackay and
Leinonen (1975) model may be fortuitous.  Estimates of volatilization rates
based on the Mackay and Leinonen (1975) model depend primarily on liquid-
gas phase exchange rate constants, whereas the experimental model of
Billing _e_t _al. (1975) and Billing (1977) is controlled by the rate of
stirring and the wind velocity across the surface of the water.

         Pearson and McConnell (1975) suggest that the presence of dichlo-
romethane and other halogenated aliphatics in ambient waters is due to ab-
sorption of chloroorganics from the atmosphere by water and return to the
earth via precipitation.  Aerial transport of these chloroorganics is
indicated by Pearson and McConnell (1975) to play a major role in the wider
distribution of these compounds and to account for their presence in upland
waters.

    39.4.5  Sorption

         Billing ejt ad. (1975) carried out two closed system experiments
where solute loss could occur only by adsorption.  Bry bentonite clay at
                                     39-5

-------
375 mg/1 was introduced into a sealed solution and in ten minutes there oc-
curred approximately a ten percent adsorption of dichloromethane and other
chloroaliphatics onto the clay.  When the amount of clay added to the
closed system was doubled (750 mg/1), 22 percent solute loss via adsorption
was noted after 30 minutes.  There was no further solute adsorption after
this time.  The authors indicated that there appeared to be little selec-
tivity (with respect to the adsorption process) among the various chlori-
nated compounds used in the experiment.  The authors observed some adsorp-
tion of dichloromethane by dry powdered dolomitic limestone, but, again, no
selectivity among the solute compounds was observed.

         In a sealed system with approximately 500 mg/1 peat moss, approxi-
mately 40 percent of the dichloromethane was adsorbed onto the peat moss
after 10 minutes.  At longer times, no further solute removal was noted.

    39.4.6  Bioaccumulation

         According to Kopperman _et^ _al (1976), not all organochlorine com-
pounds bioaccumulate to high levels.  The data suggest that polar com-
pounds are more easily biodegraded, whereas non-polar, highly lipophilic
compounds accumulate.  Neely _e_t _al. (1974) have shown that bioaccumulation
is directly related to the octanol/water partition coefficient (P) of the
compound.   The log octanol/water partition coefficient (log P) of dichlo-
romethane is 1.25 (Hansch  _a_l. 1975), indicating that dichloromethane is
not highly lipophilic and probably would not exhibit a significant tendency
to bioaccumulate in organisms.

    39.4.7  Biotransformation and Biodegradation

         No information pertaining specifically to the biodegradation of
dichloromethane was found.  Thorn and Agg (1975) have included dichloro-
methane in a list of synthetic organic chemicals which should be degradable
by biological sewage treatment, provided suitable acclimatization can be
achieved.  They note, however, that not many compounds in this list occur
in nature, and, as a result, it is unlikely that microorganisms already
possess the ability to destroy them.  Thorn and Agg (1975) consider
dichloromethane to be potentially biodegradable.

    39.4.8  Other Reactions

         Rook (1977) has suggested that dichloromethane, which has been
found in finished drinking water, may be present in such chlorinated water
due to the haloform reaction.  Stated simply, the haloform reaction is an
aqueous chlorination reaction which can occur during the treatment of water
supplies, generally in alkaline solution, with organic compounds containing
the acetyl group or with structures that may be oxidized readily to the
                                     39-6

-------
acetyl group (Morris 1975).  The three hydrogens of the methyl component of
the acetyl group are successively replaced by chlorine or other halogen,
and then the carbon bond to the carbonyl group is hydrolyzed giving rise to
a haloform and a carboxylic acid (Morris 1975).  The simplified reaction is
given below:

           CH3COR + 3HOC1	* CC13COR + 3H20     (1)
           CC13COR + H20 	** CHC13 + RCOOH      (2)

         It has been suggested that organic compounds that have a structure
equivalent to an acetyl group are present in the form of humic materials
(specifically, fulvic acids) which cause the yellow to brown stain of sur-
face waters (Rook 1974;  Rook 1977;   Morris 1975).  The carbon between two
meta-positioned OH-groups of a hydroxylated aromatic ring is proposed as
the most reactive site for haloform formation (Rook 1977).

39.5  Data Summary

    Table 39-1 summarizes the aquatic fate data discussed above.  The rate
of oxidation that is presented is a photooxidation rate and refers to the
rate of reaction of volatilized dichloromethane with hydroxyl radicals in
the troposphere.  Such photooxidation appears to be the major environmen-
tal fate of dichloromethane, and is a result of the high vapor pressure of
this compound which causes it to volatilize rapidly from water into the
atmosphere.  Any dichloromethane that remains unreacted in the troposphere
will diffuse upward and be photodissociated in the stratosphere.  Alter-
natively, atmospheric dichloromethane may be returned to the earth in
precipitation.
                                    39-7

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39.6  Literature Cited

Cox, R.A.,  R.G. Derwent,  A.E.J.  Eggleton,  and J.E.  Lovelock.   1976.
  Photochemical oxidation of halocarbons  in the  troposphere.   Atmos.
  Environ.  10:305-308.

Dean, J.A.  (ed).  1973.   Lange's handbook of chemistry,   llth Edition.
  McGraw-Hill Book Company, New  York.   1576p.

Billing, W.L.  1977.  Interphase transfer processes.   II.   Evaporation
  rates of  chloromethanes, ethanes,  ethylenes, propanes,  and  propylenes
  from dilute aqueous solutions.  Comparisons with  theoretical
  predictions.   Environ. Sci. Technol. 11(4):405-409.

Billing, W.L., C.J. Bredeweg, and N.B. Tefertiller.  1976.  Organic
  photochemistry.  Simulated atmospheric  rates of methylene chloride,
  1,1,1,-trichloroethane, trichloroethylene, tetrachloroethylene,  and other
  compounds.  Environ. Sci. Technol. 10(4):351-356.

Billing, W.L., N.B. Tefertiller, and G.J.  Kallos.  1975.   Evaporation
  rates of  methylene chloride,  chloroform, 1,1,1-trichloroethane,
  trichloroethylene, tetrachloroethylene,  and other chlorinated compounds
  in dilute aqueous solutions.   Environ.  Sci. Technol.  9(9) :833-838.

Environmental Protection Agency.  1975.  Preliminary  assessment of
  suspected carcinogens  in drinking  water.  U.S. Environmental Protection
  Agency, (Office of Toxic Substances), Washington, B.C.   33  p. EPA
  560/4-75-003.

Fells, I. and E.A. Moelwyn-Hughes.  1958.   The kinetics of  the hydrolysis
  of methylene dichloride.  J.  Chem. Soc.  (Lond.) 1326-1333.

Hansch, C., A. Vittoria,  C. Silipo,  and P.Y.C. Jow.  1975.  Partition
  coefficient and the structure-activity  relationship of  the  anesthetic
  gases.  J. Med. Chem.  18(6):546-548.

Hanst, P.L.  1978.  Noxious trace gases in the air, Part  II:   halogenated
  pollutants.  Chemistry 51(2):6-12.

Jaffe, H.H. and M. Orchin.  1962.  Theory and applications  of
  ultraviolet spectroscopy.  John Wiley and Sons, Inc., New York.   624p.

Kopperman,  H.L., B.W. Kuehl, and G.E. Glass.  1976.  Chlorinated compounds
  found in waste treatment effluents and  their capacity to  bioaccumulate.
  Proceedings of the conference  on the environmental  impact of water
  chlorination.  Oak Ridge, Tennessee, October 22-24, 1975.  327-345p.
                                       39-9

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Liss, P.S. and P.G. Slater.  1974.  Flux of gases across the air-sea
  interface.  Nature 247:181-184.

Mackay, D. and P.J. Leinonen.  1975.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. Technol.
  9(13):1178-1180.

Mackay, D. and A.W. Wolkoff.  1973.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. Technol.
  7(7):611-614.

Morris, J.C.  1975.  Formation of halogenated organics by chlorination of
  water supplies.  U.S. Environmental Protection Agency, (Office of
  Research and Development) Washington, B.C. 54p.  EPA 600/1-75-002.

Morrison, R.T. and R.N. Boyd.  1973.  Organic chemistry, 3rd Edition.
  Allyn and Bacon, Inc., Boston, Mass. 1258p.

Neely, W.B., D.R. Branson, and G.E. Blau.  1974.  Partition coefficient to
  measure bioconcentration potential of organic chemicals in fish.
  Environ. Sci. Technol. 8(13):1113-1115.

Pearson, C.R. and G. McConnell.   1975.  Chlorinated Cj_ and C^
  hydrocarbons in the marine environment.  Proc. Roy. Soc. London B
  189:305-322.

Radding, S.B., D.H. Liu, H.L. Johnson, and T. Mill.  1977.  Review of the
  environmental fate of selected chemicals.  U.S. Environmental Protection
  Agency, (Office of Toxic Substances), Washington, D.C. 147p.  EPA
  560/5-77-003.

Rook, J.J.  1974.  Formation of haloforms during chlorination of natural
  waters.  J. Water Treat. 23(Part 2):234-243.

Rook, J.J.  1977.  Chlorination reactions of fulvic acids in natural
  waters.  Environ. Sci. Technol. 11(5):478-482.

Singh, H.B., L.J. Salas, H. Shiegeishi, and A.H. Smith.  1978.  Fate of
  halogenated compounds in the atmosphere interim report - 1977.   U.S.
  Environmental Protection Agency, (Office of Research and Development),
  Research Triangle Park, N.C.  57p.  EPA 600/3-78-017.

Spence, J.W., P.L. Hanst, and B.W. Gay, Jr.  1976.  Atmospheric oxidation
  of methyl chloride, methylene chloride, and chloroform.  J. Air Pollut.
  Control Assoc. 26(10) .-994-996.
                                   39-10

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Thorn, N.S. and A.R. Agg.  1975.  The breakdown of synthetic organic
  compounds in biological processes.  Proc.  Roy.  Soc.  London B 189:347-357,

Weast, R.C. (ed.).  1977.  Handbook of chemistry  and physics.  58th
  Edition. CRC Press, Inc., Cleveland, Ohio. 2398p.
                                   39-11

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                    40.  TRICHLOROMETHANE (CHLOROFORM)


40.1  Statement of Probable Fate

    Volatilization is the major transport process for removal of trichloro-
methane (chloroform) from aquatic systems.  Once in the troposphere, tri-
chloromethane is attacked by hydroxyl radicals with the subsequent forma-
tion of phosgene (COC^) and possibly the CIO radical as principal, ini-
tial products.  The portion unreacted in the troposphere may be returned to
the earth in precipitation or on particulates, and a small amount may dif-
fuse upward to the stratosphere, where it photodissociates via interaction
with high-energy ultraviolet light.

    Based on the information reviewed, it appears that neither oxidation
nor hydrolysis is an important fate process for trichloromethane in the
aquatic environment.  No information was found to indicate that either bio-
accumulation or sorption is an important process for trichloromethane in
the aquatic environment.

40.2  Identification

    Trichloromethane is known to be ubiquitous in the environment;  its
presence has been detected in finished drinking water (Environmental Pro-
tection Agency 1975a;  Morita et _al. 1974), in marine waters and sediments
(Pearson and McConnell 1975;  McConnell _e_t _al. 1975), in rainwater (Pearson
and McConnell 1975), in wastewater effluents (Glaze and Henderson 1975),
and in the atmosphere (Singh ^t _a_l. 1978;  Environmental Protection Agency
1975b, Lillian et al. 1975).

    The chemical structure of trichloromethane is shown below.
           _.                          Alternate Name

           1                          Chloroform
      Cl	 C 	H
           C!
    Trichloromethane

    CAS NO. 67-66-3
    TSL NO. FS91000
                                   40-1

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40.3  Physical Properties

    The general physical properties of trichloromethane are given below.

    Molecular weight                         119.38
    (Weast 1977)

    Melting point                            -63.5C
    (Weast 1977)

    Boiling point at 760 torr                 61.7C
    (Weast 1977)

    Vapor pressure at 20C                   150.5 torr
    (Pearson and McConnell 1975)

    Solubility in water at 20C               8,200 mg/1
    (Pearson and McConnell 1975)

    Log octanol/water partition coefficient   1.97
    (Hansch _et al. 1975)

40.4  Summary of Fate Data

    40.4.1  Photolysis

         Relatively little information pertaining specifically to the rate
of photolysis of trichloromethane in the aqueous environment under ambient
conditions was found.  Billing _et al. (1975) studied the rate of decomposi-
tion of trichloromethane at a 1 mgTT concentration in water in aerated
sealed tubes and found little difference between the rate of disappearance
of trichloromethane in tubes exposed to  sunlight and those kept in dark-
ness.  After one year, an average (sample size = 2) of 36 percent of the
trichloromethane had disappeared from samples kept in darkness as opposed
to an average loss of 43.5 percent for samples exposed to sunlight.  An-
other study by Jensen and Rosenberg (1975) compared the rate of disappear-
ance of trichloromethane in a closed flask series in daylight and a closed
aquarium system in darkness and found no significant difference in the two
systems.  The experimental results support the findings that trichloro-
methane absorbs light only at wavelengths well below the 290 nm tropo-
spheric cutoff (Lillian e_t al.  1975) and, as a result, would not be ex-
pected to undergo direct photolysis below the ozone layer.

         Trichloromethane absorbs ultraviolet light maximally at 175 nm
(Lillian et al. 1975).  This wavelength almost coincides with the wave-
length regTon from 180 to 240 nm which is the wavelength range in which
photodissociation in the stratosphere occurs (Robbins 1976).  As a result,
                                     40-2

-------
any trichloromethane which reaches the stratosphere would be expected to
photodissociate, although the rate is unknown.

    40.4.2  Oxidation

         Billing e_t _a.L. (1975) conducted aqueous reactivity experiments in
which there was a large excess of dissolved oxygen compared to the total
amount of trichloromethane.  The air space above the solution in the closed
system contained approximately 90 times as much oxygen as was present in
the saturated solution.  The observed disappearance, which occurred at a
rate much lower than rates of volatilzation from water of trichloromethane
as reported by the same authors, was attributed primarily to ionic hydroly-
sis rather than oxidation.  On this basis it is concluded that oxidation in
the aqueous phase is not an important process for trichloromethane.

         Due to the high vapor pressure of trichloromethane, volatilization
to the atmosphere is quite rapid.  Once in the troposphere, trichloro-
methane is attacked by hydroxyl radicals via hydrogen abstraction (Hanst
1978;  Spence _e_t al. 1976;  Environmental Protection Agency 1975b; Cox _et
al. 1976).  The product of this reaction is reported to be a CCl3 radical
which reacts with oxygen to yield phosgene (COC^) and possibly chlorine
oxide (CIO) radicals (Spence et_ al. 1976; Hanst 1978;  Environmental Pro-
tection Agency 1975b) .  Neither of these initial products would be expected
to persist; phosgene is readily hydrolyzed to HC1 and C02 (Morrison and
Boyd 1973).

         Reaction with hydroxyl radicals is stated by Yung _e_t al. (1975) to
be the primary mechanism for the removal of trichloromethane from the atmo-
sphere.  The estimated atmospheric lifetime for trichloromethane is re-
ported to be about two to three months (Environmental Protection  Agency
1975b;  Yung t al. 1975).  According to Cox_et al. (1976), trichloromethane
reacts with hydroxyl radicals present in the troposphere at a rate of 16.8
x 10~-"-^cm^sec~^, which corresponds to a reported tropospheric life-
time (time for reduction to 1/e of the original concentration) of 0.19
years.  Yung et al. (1975) report a bimolecular rate constant of 1.0 x
10~12 exp(-6307T) cm3sec~1 (11.9 x 1014 cm3sec~1 at 23C),
corresponding to a tropospheric lifetime of approximately 10  sec (0.32
years).

         At a tropospheric lifetime of 0.2 to 0.3 years, and assuming a
troposphere-to-stratosphere turnover time (time for all but 1/e of
tropospheric air to diffuse into the stratosphere) of 30 years, on the
order of one percent of tropospheric trichloromethane would eventually
reach the stratosphere.
                                    40-3

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    40.4.3  Hydrolysis

         A maximum hydrolytic half-life of 3,500 years has been estimated
for trichloromethane at pH 7 and 25C (Radding j^t _al. 1977) by extrapola-
tion from data at 100-150C.  This corresponds to a first-order rate con-
stant for hydrolysis of trichloromethane of 6.9 x 10~^2sec-l (1.3 x
10"4 month -1) (Radding jrt al. 1977).  The validity of the extra-
polation method has not been established, and the estimated hydrolytic
half-life of 3,500 years is in sharp contrast to the half-life of ap-
proximately 15 months, corresponding to a first-order rate of 0.045
month~l, attributed primarily to ionic hydrolysis by Dilling et_ al.
(1975) from the aqueous reactivity experiments described previously.  Since
oxidation and vapor-phase hydrolysis could each have occurred in the system
used, the rate reported must be regarded as an upper limit for aqueous-
phase hydrolysis of trichloromethane.  It should be noted that a similar
discrepancy in values exists for dichloromethane (a compound structurally
similar to chloroform) with respect to the half-life of hydrolysis es-
timated by Radding et_ al_. (1977) by extrapolation from data at 100-150C
and the experimental value obtained by Dilling t al. (1975) at 25C.

    40.4.4  Volatilization

         Dilling et al. (1975) estimated the experimental half-life for
vplatilization of trichloromethane originally present at 1 mg/1 to be 21 +
4 minutes when stirred at 200 rpm in water at approximately 25C in an open
container.  Removal of 90 percent of the trichloromethane under the same
conditions required 62 minutes.  For chloroaliphatics in general, stirring
speed was found to have a marked effect on volatilization rate.  With no
stirring except 15 seconds every five minutes, the time required for 50
percent depletion of trichloromethane was greater than 90 minutes, which is
on  the order of four times greater than the stirred case for this compound.
The presence of sodium chloride at a three percent concentration, as in
seawater, caused about a ten percent decrease in the evaporation rate.

         From the results of the above experiments, Dilling ejt_ al. (1975)
conclude that evaporation is the major pathway by which trichloromethane,
as well as the other chlorinated compounds, is lost from water.  Dilling
_et_ al. (1975) are careful to point out the difficulties encountered in
extrapolating their laboratory results to real-world conditions, where the
concentration of the organic solute would probably be very much less than 1
rag/1 and where surface and bulk agitation would be highly variable.  Al-
though the data appear to be valid on a relative basis (i.e., correctly
illustrating the relative rates of volatilization of chlorinated alipha-
tics), they cannot be used as absolute measures of volatilization rates
from natural waters.  For the purposes of this document, the data are used
as  rough-order of magnitude indications of the importance of volatilization
                                    40-4

-------
relative to other transport and fate processes, with the strong effects of
agitation considered.  The validity of this application has not been es-
tablished.

         A subsequent study by Billing (1977) was conducted using the same
experimental conditions as in the 1975 study, and an average evaporative
half-life of 25.7 minutes for 0.96 mg/1 trichloromethane was obtained.  The
purpose of this subsequent study was to use the experimental data obtained
to test two theoretical models formulated to predict evaporative rates of
slightly soluble organic compounds from water.  Dilling (1977) found that
the theoretical model of Mackay and Wolkoff (1973) failed to predict evap-
orative half-lives, but that the calculated evaporative half-lives from the
theoretical model of Mackay and Leinonen (1975) using parameters from Liss
and Slater (1974), correlated well with the experimental evaporative half-
lives obtained by Dilling (1977).  For example, the evaporative half-life
obtained experimentally by Dilling (1977) was 19 to 26 minutes as compared
to 23.7 minutes predicted by the Mackay and Leinonen (1975) model and 1.46
minutes predicted by the Mackay and Wolkoff (1973) model.

         Dilling (1977), however, comments that the apparent numerical
agreement between his data and the values predicted by the Mackay and
Leinonen (1975) model may be fortuitous.  Estimates of volatilization rates
based on the Mackay and Leinonen (1975) model depend primarily on liquid-
gas phase exchange rate constants, whereas the experimental model of
Dilling et al. (1975) and Dilling (1977) is controlled by the rate of
stirring and the wind velocity across the surface of the water.

         Pearson and McConnell (1975) suggest that the presence of tri-
chloromethane and other halogenated aliphatics in ambient waters is due to
absorption of chloroorganics from the atmosphere by water droplets and re-
turn to the earth during precipitation.  Aerial transport of these chloro-
organics is indicated by Pearson and McConnell (1975) to play a major role
in the wider distribution of these compounds and accounts for their pre-
sence in upland waters.

         A study by Jensen and Rosenberg (1975) indicates that volatiliza-
tion of trichloromethane from water proceeds more rapidily than photolysis,
oxidation, or hydrolysis in the aqueous medium.  An initial concentration
of between 0.1 to 1 mg/1 trichloromethane in 20 liters of water was found
to decrease 50 to 60 percent after eight days in an aquarium kept in light
and partly open.  Trichloromethane initially present at the same concentra-
tion in a closed system exposed to daylight and a closed system exposed to
dark exhibited less than five percent loss in 8 days.  Assuming that all
processes involved were first-order, then the rate of volatilization
appears to be an order of magnitude larger than the summation of the rates
for degradative and other loss processes in the closed system where
volatilization could not occur.
                                     40-5

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    40.4.5  Sorption

         Dilling ej: al. (1975) carried out two closed system experiments
where the solute loss could be only by adsorption.  Dry bentonite clay at
375 mg/1 was introduced into a sealed solution, and after ten minutes there
was approximately a ten percent adsorption of trichloromethane and other
chlorinated compounds by the clay.  When the amount of clay added to the
closed system was doubled (750 mg/1) there was 22 percent solute loss via
adsorption after 30 minutes.  There was no further solute adsorption after
this time.  The authors indicated that there appeared to be little selec-
tivity of the clay among the various chlorinated compounds used.  Some
adsorption of trichloromethane and other short-chain chloroaliphatics by
dry powdered dolomitic limestone was observed, but again without selec-
tivity for specific solutes.

         Dilling et_ al_. (1975) also found that, in a sealed system with
approximately 500 mg/1 peat moss, approximately 40 percent of the tri-
chloromethane solute adsorbed in 10 minutes.  At longer times, no further
solute removal was noted.

         Pearson and McConnell (1975) found that the concentrations of
chloroorganics in Liverpool Bay (England) sediments varied from a few parts
per trillion (by mass) to a maximum of 4000 parts per trillion trichloro-
methane.  Samples of marine sediment from Liverpool Bay contained the same
compounds as the overlying waters.  There was, however, no significant cor-
relation found between the concentration of trichloromethane in marine
sediment and the concentration in the water above the sediment at the time
the samples were taken (Pearson and McConnell 1975;  McConnell et al.
1975).  In addition, there was no correlation between high concentrations
of trichloromethane and geographical features, as was noticed for several
other compounds, nor was there any correlation between the trichloromethane
concentration in the sediment and particle size.  McConnell est_ al_. (1975),
however, did note that coarse gravels had little adsorptive capacity for
trichloromethane, whereas sediments rich in organic detritus had a much
higher adsorptive capacity.  In summary, there was no clear evidence of
selective concentration of trichloromethane in sediments.

    40.4.6  Bioaccumulation

         According to Kopperman _et_ _a_l. (1976), not all organochlorine com-
pounds bioaccumulate to high levels.  The data suggest that polar compounds
are more easily biodegraded, whereas non-polar, highly lipophilic compounds
accumulate.  Neely _et^ al. (1974) have shown that bioaccumulation is di-
rectly related to the octanol/water partition coefficient (P) of the com-
                                     40-6

-------
pound.  The log octanol/water partition coefficient (log P) of trichloro-
methane is 1.97 (Hansch e_t _al. 1975;  Chiou e_t _al. 1977;  Leo e_t al. 1971),
thus indicating a possible tendency of this compound to bioaccumulate under
conditions of constant exposure.

         Evidence for weak to moderate bioaccumulation of trichloromethane
by marine organisms is reported by Pearson and McConnell (1975).  These
authors, however, have indicated that although trichloromethane is somewhat
lipophilic and tends to be found at higher concentrations in fatty tissues,
there is no evidence for biomagnif ication of trichloromethane in aquatic
food chains .

    40.4.7  Biotransf ormation and Biodegradation

         According to Pearson and McConnell (1975), only completely sealed
systems such as the standard BOD (biochemical oxygen demand) bottle tech-
nique can be used to measure biochemical degradation of volatile compounds
such as trichloromethane.  The BOD bottle experiments of Pearson and
McConnell (1975) have been unable to demonstrate any significant oxygen ab-
sorption from compounds containing only C, H, and Cl, thus indicating that
biochemical degradation of such compounds is indeed very slow.

         Thorn and Agg (1975) have included trichloromethane in a list of
synthetic organic chemicals which should be degradable by biological sew-
age treatment, provided suitable acclimatization can be achieved.

    40.4.8  Other Reactions

         Rook (1977) has suggested that trichloromethane, which has been
found in finished drinking water, may be  present in such chlorinated water
due to the haloform reaction.  Stated simply, the haloform reaction is an
aqueous chlorination reaction, connected with the treatment of water
supplies, which occurs generally in alkaline solution (pH >_ 5) with
organic compounds containing the acetyl group or with structures that may
be readily oxidized to the acetyl group (Morris 1975).  The three hydrogens
on the methyl component of the acetyl group are successively replaced by
chlorine or other halogen, and then the carbon bond to the carbonyl group
is split giving rise to a haloform and a carboxylic acid (Morris 1975).
The simplified reaction is given below:

         CH3COR + 3HOC1  - > CC13COR 4- 3H20    (1)
         CC13COR + H20 - *-CHCl3 + RCOOH        (2)
         It has been suggested that organic compounds that have a structure
equivalent to an acetyl group are present in the form of humic materials
(such as fulvic acids) which cause the yellow to brown stain of surface
                                      40-7

-------
waters (Rook 1974;  Rook 1977;  Morris 1975).  Specifically, the carbon be-
tween two meta-positioned OH-groups of a hydroxylated aromatic ring is
proposed as the most reactive site for haloform formation (Rook 1977).

40.5  Data Summary

         Table 40-1 summarizes the aquatic fate of trichloromethane.  Oxi-
dation rates given are photooxidation rates and refer to the rate of re-
action of trichloromethane with hydroxyl radicals in the troposphere.

         The primary fate of trichloromethane is believed to be photooxida-
tion in the troposphere, a direct result of the high vapor pressure of this
compound which causes it to volatilize rapidly from water into the atmo-
sphere.  Once in the troposphere, trichloromethane is attacked by hydroxyl
radicals resulting in the subsequent formation of phosgene and chlorine
oxide as principal products.  Any unreacted portion of trichloromethane in
the troposphere diffuses upward and undergoes photodissociation above the
ozone layer, or is absorbed by water and returned to the earth via
precipitation.
                                       40-8

-------

























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40.6  Literature Cited

Chiou, C.T., V.H. Freed,  D.W.  Schmedding,  and  R.L.  Kohnert.   1977.
  Partition coefficient and bioaccumulation of selected organic chemicals.
  Environ. Sci. Technol.  11(5):475-478.

Cox, R.A., R.G. Derwent,  A.E.J.  Eggleton,  and  J.E.  Lovelock.   1976.
  Photochemical oxidation of halocarbons in the troposphere.   Atmos.
  Environ. 10:305-308.

Billing, W.L.  1977.  Interphase transfer processes.   II.   Evaporation
  rates of chloromethanes, ethanes,  ethylenes, propanes, and  propylenes
  from dilute aqueous solutions.  Comparisons  with theoretical predictions.
  Environ. Sci. Technol.-11(4):405-409.

Billing, W.L., N.B. Tefertiller, and G.J.  Kallos.   1975.  Evaporation
  rates of methylene chloride, chloroform, 1,1,1-trichloroethane,
  trichloroethylene, tetrachloroethylene,  and  other chlorinated compounds
  in dilute aqueous solutions.  Environ. Sci.  Technol. 9(9):833-838.

Environmental Protection Agency.  1975a.  Preliminary assessment of
  suspected carcinogens in drinking  water.  U.S. Environmental Protection
  Agency, (Office of Toxic Substances),  Washington, B.C.  33p.  EPA
  560/4-75-003.

Environmental Protection Agency.  1975b.  Report on the problem of
  halogenated air pollutants and stratospheric ozone.  U.S. Environmental
  Protection Agency, (Office of Research and Development), Research
  Triangle Park, North Carolina.  55p.  EPA 600/9-75-008.

Glaze, W.H. and J.E. Henderson,  IV.   1975.  Formation of organochlorine
  compounds from chlorination of a municipal secondary effluent.  J.  Water
  Pollut. Control Fed. 47(10):2511-2515.

Hanst, P.L.  1978.  Noxious trace gases  in the air, Part II:   halogenated
  pollutants.  Chemistry 51(2):6-12.

Hansch, C., A. Vittoria, C. Silipo,  and P.Y.C. Jow.  1975.  Partition
  coefficients and  the structure-activity relationship of  the anesthetic
  gases.  J. Med. Chein.  18(6) :546-548.

Jensen, S. and R. Rosenberg.  1975.   Begradability of some chlorinated
  aliphatic hydrocarbons in sea water and sterilized water.  Water Res.
  9:659-661.
                                   40-10

-------
Kopperman, H.L., D.W. Kuehl, and G.E. Glass.  1976.  Chlorinated compounds
  found in waste treatment effluents and their capacity to bioaccumulate.
  Proceedings of the conference on the environmental impact of water
  chlorination.  Oak Ridge, Tennessee, October 22-24, 1975.  327-345p.

Leo, A., C. Hansch, and D. Elkins.  1971.  Partition coefficients and their
  uses.  Chem. Rev. 71(6):525-612.

Lillian, D., H.B. Singh, A. Appleby, L. Lobban, R. Arnts, R. Gumpert, R.
  Hague, J. Toomey, J. Kazazis, M. Antell, D. Hansen, and B. Scott.  1975.
  Atmospheric fates of halogenated compounds.  Environ. Sci. Technol.
  9:1042-1048.

Liss, P.S. and P.G. Slater.  1974.  Flux of gases across the air-sea
  interface.  Nature  247:181-184.

Mackay, D. and P.J. Leinonen.  1975.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. Technol.
  9(13):1178-1180.

Mackay, D. and A.W. Wolkoff.  1973.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. Technol.
  7(7):611-614.

McConnell, G., D.M. Ferguson, and C.R. Pearson.  1975.  Chlorinated
  hydrocarbons and the environment.  Endeavor  XXXIV:13-18.

Morita, M., H. Nakamura, and S. Mimura.  1974.  Analytical method and
  evaluation of organic matter in river and well water.  II.  Pollution by
  chlorinated aliphatic hydrocarbons.  Tokyo Toritsu Eisei Kenkyusho Krakyu
  Nempo  25:399-403.  (Abstract only).

Morris, J. C.  1975.  Formation of halogenated organics by chlorination
  of water supplies.  U.S. Environmental Protection Agency, (Office of
  Research and Development).  Washington, D.C.  54p.  EPA 600/1-75-002.

Morrison, R. T. and R. N. Boyd, 1973.  Organic Chemistry, 3rd Edition.
  Allyn and Bacon, Inc., Boston, Mass. 1258p.

Neely, W.B., D.R. Branson, and G.E. Blau.  1974.  Partition coefficient to
  measure bioconcentration potential of organic chemicals in fish.
  Environ. Sci. Technol. 8(13):1113-1115.

Pearson, C.R. and G. McConnell.  1975.  Chlorinated C^ and C2
  hydrocarbons in the marine environment.  Proc. Roy. Soc. London B
  189:305-322.
                                   40-11

-------
Radding, S.B., D.H. Liu, H.L. Johnson, and T.  Mill.  1977.  Review of the
  environmental fate of selected chemicals.  U.S. Environmental Protection
  Agency, (Office of Toxic Substances),   Washington, D.C. I47p.  EPA
  560/5-77-003.

Robbins, D.E.  1976.  Photodissociation of methyl chloride and methyl
  bromide in the atmosphere.  Geophys. Res. Lett.  3(4):213-216.

Rook, J.J.  1974.  Formation of haloforms during chlorination of natural
  waters.  J. Soc. Water Treat. Exam.  23(Part 2):234-243.

Rook, J.J.  1977.  Chlorination reactions of fulvic acids in natural
  waters.  Environ. Sci. Technol. 11(5):478-482.

Singh, H.B., L.J. Salas, H. Shiegeishi, and A.H. Smith.   1978.  Fate of
  halogenated compounds in the atmosphere interim report - 1977.   U.S.
  Environmental Protection Agency, (Office of Research and Development),
  Research Triangle Park, N.C.  57p.  EPA 600/3-78-017.

Spence, J.W., P.L. Hanst, and B.W. Gay, Jr.  1976.  Atmospheric oxidation
  of methyl chloride, methylene chloride, and chloroform.  J. Air Pollut.
  Control Assoc.  26(10) :994-996.

Thorn, N.S. and A.R. Agg.  1975.  The breakdown of synthetic organic
  compounds in biological processes.  Proc. Roy. Soc. London B
  189:347-357.

Weast, R.C. (ed.).  1977.  Handbook of chemistry and physics.  58th
  Edition. CRC Press, Cleveland, Ohio. 2398p.

Yung, Y.L., M.B. McElroy, and S.C. Wofsy.  1975.  Atmospheric halocarbons:
  a discussion with emphasis on chloroform.  Geophys. Res. Lett.
  2(9):397-399.
                                     40-12

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              41.  TETRACHLOROMETHANE (CARBON TETRACHLORIDE)
41.1  Statement of Probable Fate

    Volatilization is the major transport process for removal of  tetra-
chloromethane from aquatic systems.  Once in  the troposphere, tetrachloro-
methane remains stable; it exhibits an extremely slow rate of reaction with
hydroxyl radicals present in the troposphere.  Tetrachloromethane eventu-
ally diffuses into the stratosphere or is carried back  to the earth during
the precipitation process.  Once in the stratosphere tetrachioromethane  is
degraded on exposure to shorter wavelength, higher energy ultraviolet light
to eventually form phosgene as the principal  initial product.

    Based on the information reviewed, it does not appear that oxidation is
an important fate process of tetrachloromethane in the  aquatic environment.
Evidence for sorption, bioaccumulation and biotransformation, hydrolysis,
and biodegradation is inconclusive.

41.2  Identification

    Tetrachloromethane is known to be ubiquitous in the environment; its
presence has been detected in finished drinking water (Environmental Pro-
tection Agency 1975a), in fresh water and rainwater (Pearson and McConnell
1975), in marine water and sediment (Pearson  and McConnell 1975; McConnell
_eit al. 1975), and in the atmosphere (Singh _e_t &!_. 1978; Environmental Pro-
tection Agency 1975b; Lillian _et al. 1975; Lovelock _et _al. 1973).

    The chemical structure of tetrachloromethane is shown below.


              I                               Alternate  Names

        Cl	 C 	Cl                           Carbon tetrachloride
              I                               Methane tetrachloride
              '                                Perchloromethane
             Cl
                                              Benzinoform

     Te tr achlorome thane

     CAS NO. 56-23-5
     TSL NO. FG 49000
                                     41-1

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41.3  Physical Properties

    The general physical properties of tetrachloromethane are given below.

    Molecular weight                         153.82
    (Weast 1977)

    Melting point                            -22.9C
    (Weast 1977)

    Boiling point at 760 torr                76.54C
    (Weast 1977)

    Vapor pressure at 20C                   90 torr
    (Pearson and McConnell 1975)

    Solubility in water at 20C              785 mg/1
    (Pearson and McConnell 1975)

    Log octanol/water partition coefficient  2.64
    (Neely _e_t al. 1974)

41.4  Summary of Fate Data

    41.4.1  Photolysis

         No information was found pertaining to the rate of photolysis of
tetrachloromethane in the aqueous environment under ambient conditions.

         Due to the high vapor pressure of tetrachloromethane, volatiliza-
tion to the atmosphere is quite rapid.  Tetrachloromethane is, however,
tropospherically stable (Singh e_t _a_l. 1978; Environmental Protection Agency
1975b;  Hanst 1978; Lillian jst al. 1975;  Cox _et a 1. 1976) and does not react
significantly with hydroxyl radicals present in the troposphere.  In addi-
tion, tetrachloromethane does not photodissociate in the troposphere since
this compound has no chromophores which absorb in the visible or near
ultraviolet region of the electromagnetic spectrum (Jaffe and Orchin
1962).

         The tropospherically stable tetrachloromethane eventually diffuses
into the stratosphere (Hanst 1978; National Research Council 1978), al-
though some tetrachloromethane is theorized to be washed out of the
atmosphere and back to the lithosphere and hydrosphere during the precipi-
tation process (Pearson and McConnell 1975).  That portion of the tetra-
chloromethane which diffuses to the stratosphere  is degraded on ex-
posure to higher energy, shorter wavelength ultraviolet light to form
                                     41-2

-------
CC13 radicals and chlorine atoms.  The 13 radicals are oxidized to
form phosgene which also undergoes photodissociation in the stratosphere
with the subsequent release of more chlorine atoms (Hanst 1978; National
Research Council 1978).  Chlorine atoms, released by reactions such as
these, are theorized to be catalysts in the destruction of the ozone layer
of the earth (Hanst 1978; National Research Council 1978).

    41.4.2  Oxidation

         No information was found pertaining to the oxidation of tetra-
chlorome thane in the aquatic environment under ambient conditions.  In ad-
dition, tetrachloromethane is known to be tropospherically stable (Singh et
al. 1975;  Environmental Protection Agency 1975b; Hanst 1978;  Lillian _e_t al.
1975; Cox et al. 1976) and, therefore, does not react significantly with
hydroxyl radicals present in the troposphere.  For instance,  the bimole-
cular rate of reaction for tetrachloromethane with hydroxyl radicals is
reported to be less than 1 x 10~16cm3sec-l with a corresponding re-
ported tropospheric lifetime (time for reduction to 1/e of initial con-
centration) of greater than  330 years (Cox et al. 1976). As  a result, pho-
tooxidation of tetrachloromethane would not be expected to play a
significant role in the fate of this compound.

         Using the minimum tropospheric lifetime of 330 years, and assuming
a troposphere-to-stratosphere turnover time (time for all but 1/e of tropo-
spheric air to diffuse into the stratosphere) of 30 years, the fraction of
tropospheric tetrachloromethane eventually reaching the stratosphere would
be over 90 pecent.

    41.4.3  Hydrolysis

         A maximum hydrolytic half-life of 7,000 years has been calculated
tetachlorome thane at a 1 mg/1 concentration at pH 7 and 25 C; at a 1000
mg/ 1 concentration under the same conditions a hydrolytic half-life of 7
years has been calculated for tetrachloromethane (Mabey and Mill 1978;
Radding et al. 1977).  These half-life values correspond to a reported
second-order rate of hydrolysis for tetrachloromethane of 4.8 x 10~'
         l, and are based on extrapolation of data at 100-150C.
         Lovelock ^t. al. (1973) have suggested that tetrachloromethane may
undergo hydrolysis in the ocean as evidenced from the fall-off in tetra-
chloromethane concentration with depth in ocean water samples.  Thus, in
addition to the stratosphere as a major sink for tetrachloromethane, the
oceans may prove to be a significant sink as well.
                                      41-3

-------
    41.4.4  Volatilization

         Volatilization is the most important transport process for the re-
moval of tetrachloromethane from aquatic systems.  Dilling et al. (1975)
estimated the experimental half-life for volatilization of tetrachloro-
methane originally praesent at 1 mg/1 to be 29 minutes when stirred at 100
rpm in water at approximately 25C in an open container.  Removal of 90
percent of the tetrachloromethane under the same conditions required 97
minutes.  For chloroaliphatics in general, stirring speed was found to have
a marked effect on volatilization rate.  Stirring intermittently for 15
seconds every five minutes, the time required for 50 percent depletion of
trichloromethane and dichloromethane, analogues of tetrachloromethane, was
greater than 90 minutes,-or four times greater than was observed during
constant stirring.  Dilling ^t _al. (1975) are careful to point out the
diffculties encountered in extrapolating their laboratory results to
real-world conditions, where the concentration of the organic solute would
probably be very much less than 1 mg/1 and where surface and bulk agitation
would be highly variable.  Although the data appear to be valid on a
relative basis (i.e., correctly illustrating the relative rates of volatil-
ization of chlorinated aliphatics), they cannot be used as absolute
measures of volatilization rates from natural waters.  For the purposes of
this document, the data are used as rough-order-of-magnitude indications of
the importance of volatilization relative to other transport and fate
processes, with the strong effects of agitation considered.  The validity
of this application has not been established.

         A subsequent study by Dilling (1977) was conducted using the same
experimental conditions as the 1975 study, and an evaporative half-life for
0.90 mg/1 tetrachloromethane of 28.8 minutes (average of four) was ob-
tained.  The purpose of this subsequent study was to use the experimental
data previously obtained  to test two theoretical models which may be used
to predict evaporation rates of slightly soluble organic compounds from
water.  Dilling (1977) found that  the  theoretical model by Mackay and Wolk-
off (1973) failed to predict evaporation half-lives, but that the model of
Mackay and Leinonen (1975) using parameters from Liss and Slater (1974)
correlated well with the  experimental half-lives obtained by Dilling
(1977).  For example, the  evaporative half-life obtained experimentally by
Dilling (1977) was 29 minutes, compared to 25.5 minutes predicted by the
Mackay and Leinonen (1975) model and 0.20 minutes predicted by the Mackay
and Wolkoff  (1973) model.

         Dilling  (1977),  however,  comments that the apparent numerical
agreement between his data and the values predicted by  the Mackay and
Leinonen  (1975) model may be fortuitous.  Estimates of  volatilization rates
based on  the Mackay and Leinonen (1975) model depend primarily on
liquid-gas phase  exchange rate constants, whereas the experimental model of
Dilling _et _al. (1975) and  Dilling  (1977) is controlled  by the rate of
stirring and the  wind velocity across  the surface of the water.
                                     41-4

-------
         Pearson and McConnell (1975) suggest that the presence of tetra-
chloromethane, as well as other halogenated aliphatics, in ambient waters
is due to the absorption of chloroorganics from the atmosphere by water.
This process is thought to occur most effectively when the atmosphere is
scrubbed as in the precipitation process.  Aerial transport of these
chloroorganics is indicated by Pearson and McConnell (1975) to play a major
role in the wider distribution of these compounds and accounts for their
presence in upland waters.

    41.4.5  Sorption

         There is little information pertaining specifically to adsorption
of tetrachloromethane onto sediments.  Pearson and McConnell (1975) found
that the concentrations of chloroorganics in Liverpool Bay sediments in
England varied from a few ng/1 to a maximum of 5500 ng/1 1,1,1-trichloro-
ethane plus tetrachloromethane combined.  Samples of marine sediment from
Liverpool Bay contained the same compounds as the overlying waters.  Due to
the inability of the sampling technique to resolve tetrachloromethane and
1,1,1-trichloroethane, it could not be determined whether a correlation ex-
isted between the concentrations of tetrachloromethane in marine sediments
and interstitial waters above the sediments at the time the samples were
taken (Pearson and McConnell 1975; McConnell _et _al. 1975).  In addition, it
is unknown whether a correlation existed between high concentrations and
geographical features, as was noticed in the case of the water samples for
several compounds, or whether there was a relationship between the con-
centration in the sediment and particle size.  McConnell _e _a_l. (1975) did
note, however, that coarse gravels had little adsorptive capacity for
tetrachloromethane, whereas sediments rich in organic detritus had a much
higher adsorptive capacity.  In summary, there was no clear evidence of
selective concentration of tetrachlororaethane in sediments.

    41.4.6  Bioaccumulation

         According to Kopperman et al. (1976), not all organochlorine com-
pounds bioaccumulate to high levels.  The data suggest that polar compounds
are more easily biodegraded, whereas non-polar, highly lipophilic compounds
accumulate.  Neely et al. (1974) have shown that bioaccumulation is
directly related to the octanol/water partition coefficient (P) of the com-
pound.  The log octanol/water partition coefficient (log P) of tetrachloro-
methane is 2.64 (Neely _et _al. 1974; Chiou ej; _al. 1977; Leo _et al. 1971),
indicating a tendency for this compound to bioaccumulate under conditions
of constant exposure.  Pearson and McConnell (1975), however, indicate that
although tetrachloromethane, and other organochlorines examined, are some-
what lipophilic and tend to be found at higher concentrations in fatty tis-
sues, there is no evidence for the biomagnification of tetrachloro-
methane or other short-chain choroaliphatics in food chains.  The diffi-
                                     41-5

-------
culties encountered with the analytical methods in the Pearson and McCon-
nell (1975) study make estimates of bioaccumulation based on their ex-
perimental results somewhat unreliable.

    41.4.7  Biotransformation and Biodegradation

         No information was found pertaining specifically to the biodegra-
dation of tetrachloromethane.  In the sea, however, it is known that many
species of brown seaweed, molluscs, sponges, and bacteria metabolize
halogens (Anonymous 1977).

         According to Pearson and McConnell (1975) only completely sealed
systems, such as the standard BOD (biochemical oxygen demand) bottle tech-
nique, can be used to measure biochemical degradation of volatile com-
pounds such as tetrachloromethane.  The BOD bottle experiments of Pearson
and McConnell do not demonstrate any significant oxygen absorption from
compounds containing only C, H, and Cl, thus indicating that biochemical
degradation of such compounds is indeed very slow.

         Thorn and Agg (1975) have included tetrachloromethane in a list of
synthetic organic chemicals which should be degradable by biological sewage
treatment methods provided suitable acclimatization can be achieved.  They
note, however, that not many compounds on the list occur free in nature,
and, as a result, it is unlikely that microorganisms already possess the
ability to destroy them.

41.5  Data Summary

         Table 41-1 summarizes the aquatic fate information discussed
above.  The oxidation rate is a photooxidation rate and refers to the rate
at which tetrachloromethane  is attacked by hydroxyl radicals in the
troposphere.

         The ultimate fate of tetrachlororaethane is photodissociation in
the stratosphere, an end  result of the high vapor pressure of this com-
pound which causes it to  volatilize rapidly into the atmosphere.  Once in
the troposphere, tetrachloromethane remains relatively stable and eventual-
ly diffuses into the stratosphere or is carried back  to the earth during
the precipitation  process.  Once in the  stratosphere, tetrachloromethane
is degraded on exposure  to shorter wavelength, higher energy ultraviolet
light.
                                     41-6

-------
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41-7

-------
41.6  Literatures Cited

Anonymous.  1977.  Man is not the only polluter.  Environ. Sci. Technol.
  ll(5):442-443.

Chiou, C.T., V.H. Freed,  D.W. Schmedding, and R.L. Kohnert.  1977.
  Partition coefficient and bioaccumulation of selected organic chemicals.
  Environ. Sci. Technol.  11(5):475-478.

Cox, R.A., R.G. Derwent,  A.E.J.  Eggleton, and J.E. Lovelock. 1976.
  Photochemical oxidation of halocarbons in the troposphere.  Atmos.
  Environ. 10:305-308.

Billing, W.L. 1977.  Interphase  transfer processes.  II.  Evaporation of
  chloromethanes, ethanes, ethylenes, propanes, and propylenes from dilute
  aqueous solutions.  Comparisons with theoretical predictions.  Environ.
  Sci. Technol. 11(4):405-409.

Billing, W.L.,  N.B. Tefertiller, and G.J. Kallos.  1975.  Evaporation
  rates of methylene chloride, chloroform, 1,1,1-trichloroethane,
  trichloroethylene, tetrachloroethylene and other chlorinated compounds in
  dilute aqueous solutions.  Environ. Sci. Technol.  9(9):833-838.

Environmental Protection Agency.  1975a.  Preliminary assessment of
  suspected carcinogens in drinking water.  Environmental Protection
  Agency,  (Office of Toxic Substances),  Washington, B.C. 33p. EPA
  560/4-75-003.

Environmental Protection Agency.  1975b.  Report on the problem of
  halogenated air pollutants and stratospheric ozone.  Environmental
  Protection Agency, (Office of  Research and Bevelopment), Research
  Triangle Park, North Carolina.  55p.  EPA 600/9-75-008.

Hanst, P.L. 1978.  Part II:  Halogenated pollutants.  Noxious trace gases
  in the  air.  Chemistry 51(2):6-12.

Jaffa, H.H. and M. Orchin.  1962.  Theory and application of ultraviolet
  spectroscopy.  John Wiley and  Sons, Inc., New York. 624p.

Kopperman, J.L., D.W. Kuehl, and G.E. Glass.  1976.  Chlorinated compounds
  found in waste treatment effluents and their capacity to bioaccumulate.
  Proceedings of the conference on the environmental impact of water
  chlorination.  Oak Ridge, Tennessee, October 22-24, 1975.  327-345p.

Leo, A.,  C. Hansch, and B. Elkins.  1971.  Partition coefficients and
  their uses.  Chem. Rev. 71(6):525-612.
                                    41-8

-------
Lillian, D.,  H.B. Singh, A. Appleby, L. Lobban, R. Arnts, R. Gumpert, R.
  Hague, J. Toomey, J. Kazazis, M. Antell, D. Hansen, and B. Scott. 1975.
  Atmospheric fates of halogenated compounds. Environ. Sci. Technol.
  9:1042-1048.

Liss, P.S. and P.G. Slater.  1974.  Flux of gases across the air-sea
  interface.   Nature  247:181-184.

Lovelock, J.E., R.J. Maggs, and R.J. Wade.  1973.  Halogenated hydrocarbons
  in and over the Atlantic.  Nature  241:194-196.

Mabey, W, And T. Mill.  1978. Critical review of hydrolysis of organic
  compounds in water under environmental conditions.  To be published.  J.
  Phys. Chem. Ref. Data, 7:000, 103 p.

Mackay, D. and  P.J. Leinonen.  1975.  Rate of evaporation of
  low-solubility contaminants from water bodies to atmosphere.  Environ.
  Sci. Technol. 9(13):1178-1180.

Mackay, D. and A.W. Wolkoff.  1973.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. Technol.
  7(7):611-614.

McConnell, G., D.M. Ferguson, and C.R. Pearson.  1975.  Chlorinated
  hydrocarbons and- the environment.  Endeavor  XXXIV:13-18.

National Research Council.  1978.  Nonfluorinated halomethanes in the
  environment.  National Academy of Sciences, Washington, D.C. 297 p.

Neely, W.B.,  D.R. Branson, and G.E. Blau.  1974.  Partition coefficient to
  measure bioconcentration potential of organic chemicals in fish.
  Environ. Sci. Technol. 8(13) :1113-1115.

Pearson, C.R. And G. McConnell.  1975.  Chlorinated Cj and C2
  hydrocarbons in the marine environment.  Proc. Roy. Soc. London B
  189:305-322.

Radding, S.B., D.H. Liu, H.L. Johnson, and T. Mill.  1977.  Review of the
  environmental fate of selected chemicals.  Environmental Protection
  Agency, (Office of Toxic Substances), Washington, D.C. 147 p.  EPA
  560/5-77-003.

Singh, H.B.,  D. Lillian, A. Appleby, and L. Lobban.  1975.  Atmospheric
  formation of carbon tetrachloride from tetrachloroethylene.  Environ.
  Let. 10(3):253-256.

Singh, H.B.,  L.J. Salas, H. Shiegeishi, and A.H. Smith.  1978.  Fate of
  halogenated compounds in the atmosphere interim report - 1977.
  Environmental Protection Agengy, (Office of Research and Development),
  Research Triangle Park, N.C. 57 p.  EPA 600/3-78-017.
                                     41-9

-------
Thorn, N.S. and A.R. Agg.  1975.  The breakdown of synthetic organic
  compounds in biological processes.  Proc. Roy. Soc. of London B
  189:347-357.

Weast, R.C. (ed.).  1977.  Handbook of chemistry and physics.  58th
  Edition. CRC Press, Inc., Cleveland, Ohio.  2398p.
                                     41-10

-------
                    42.   CHLOROETHANE (ETHYL CHLORIDE)
42.1  Statement of Probable Fate

    Information found pertaining specifically to chloroethane is relatively
scant.  Much of the information presented herein has been extrapolated from
more thoroughly studied analogues of chloroethane.   Based on the material
presented, one can predict with low to medium confidence that the important
fate processes for chloroethane are photooxidation in the atmosphere and
hydrolysis.  The predominant fate process cannot as yet be identified with
certainty.

42.2  Identification

    Chloroethane has been detected in finished drinking water by the
National Organics Reconnaissance Survey (Environmental Protection Agency
1975a).  Chloroethane is probably present in the atmosphere but has not
been detected either because of the small amounts emitted or because of
rapid photooxidation in the troposphere (Environmental Protection Agency
1975b).

    The chemical structure of chloroethane is shown below.

          H     H
                                      Alternate Names

    H	c 	 C 	Cl                Ethyl chloride
                                      Monochloroethane
                                      Hydrochloric ether
                                      Muriatic ether

    Chloroethane

    CAS NO. 75-00-3
    TSL NO. KH 75250

42.3  Physical Properties

    The general physical properties of chloroethane are as follows.

    Molecular weight                         64.52
    (Weast 1977)

    Melting point                            -136.4C
    (Weast 1977)

    Boiling point at 760 torr                12.27C
    (Weast 1977)
                                    42-1

-------
    Vapor pressure at 20C                   1000 torr
    (Verschueren 1977)

    Solubility in water at 20C              5740 mg/1
    (Verschueren 1977)

    Log octanol/water partition coefficient  1.54
    (Leo et al.  1971)

42.4  Summary of Fate Data

    42.4.1  Photolysis

         No information pertaining specifically to the rate of photodisso-
ciation of chloroethane in the aqueous environment was found.   Dilling _et
al. (1975) reported that, within experimental error, sunlight  did not
accelerate decomposition of such compounds as dichloromethane, trichloro-
methane, and 1,1,1-trichloroethane,  all more highly chlorinated analogues
of chloroethane, intially present at 1 mg/1 concentrations in  aerated
water.

         Photodissociation in the terrestrial environment would not be ex-
pected to occur for chloroethane since this compound has no chromophores
which absorb in the visible or near ultraviolet region of the  electromag-
netic spectrum (Jaffe and Orchin 1962).  The rate of reaction  of
chloroethane with hydroxyl radicals in the troposphere is relatively rapid,
but if any unreacted material were to reach the stratosphere,  it would be
expected to undergo significant photodissociation due to the presence of
higher energy, shorter wavelength ultraviolet light (Environmental Protec-
tion Agency 1975b).  Formyl chloride, which is readily photolyzed, is re-
ported as the principal initial photodissociation product of chloroethane
(Environmental Protection Agency 1975b).

    42.4.2  Oxidation

         No information pertaining specifically to the oxidation of chloro-
ethane in the aqueous environment was found.  There is, however, some
indirect evidence of a probable low potential for oxidation in aquatic sys-
tems. For instance, Dilling et al. (1975) conducted aqueous reactivity ex-
periments in which there was a large excess of dissolved oxygen compared to
the total amount of dichloromethane, trichloromethane, and 1,1,1-trichloro-
ethane, compounds similar to dichloroethane.  The loss of material noted,
which was attributed to ionic hydrolysis rather than oxidation, was much
slower than loss by volatilization of the compounds from water reported by
the same authors.  It is concluded that oxidation in aqueous systems is not
an important process for chloroethane.
                                     42-2

-------
         Due to the relatively high vapor pressure of chloroethane, volatil-
ization to the atmosphere is expected to be quite rapid.  Once in the
troposphere, chloroethane is photooxidized relatively rapidly.  Howard and
Evenson (1976) report a bimolecular rate constant for reaction of chloro-
ethane with hydroxyl radicals of 3.9 (+ 0.7) x 10 ~^ cm^ molecule"-'-
sec~l at 23C.  Assuming the tropospheric hydroxyl radical concentration
to be 10^ cm~3, this rate corresponds to a lifetime (time to reach 1/e
of the original concentration) on the order of one month.  Other available
data (Environmental Protection Agency 1975b) indicate an atmospheric half-
life for chloroethane of three or four months, and indicate formyl chloride
to be the principal initial photooxidation product. Formyl chloride is
readily hydrolyzed (Morrison and Boyd 1973).

         The fraction of tropospheric chloroethane eventually reaching the
stratosphere via diffusion is probably less than one percent, based on a
tropospheric lifetime on the order of one month and a troposphere-to-
stratosphere turnover time (time for all but 1/e of tropospheric air to
diffuse into the stratosphere) of 30 years.  It should be noted that this
estimation neglects the effects of chloroethane removal from the tropo-
sphere by other processes such as rainfall.

    42.4.3  Hydrolysis

         A maximum hydrolytic half-life of 40 days has been estimated for
chloroethane at pH 7 and 25C by extrapolation from data at.100-150C,
corresponding to a first-order rate constant for hydrolysis of chloroethane
of 2.0 x lO^sec"1 (Radding et al. 1977).  The validity of the extra-
polation technique used has not been established.  The apparently short
hydrolytic half-life and high solubility in water of chloroethane indicate
that hydrolysis may be an important fate process for this compound.

    42.4.4  Volatilization

         Billing e_t_ al. (1975) estimated the experimental half-life for
volatilization of chloroethane originally present at 1 mg/1 to be 21
minutes when stirred at 200 rpm in water at approximately 25C in an open
container.  Removal of 90 percent of the chloroethane under the same con-
ditions required 79 minutes.  For chloroaliphatics in general, stirring
speed was found to have a marked effect on volatilization rate.  When
intermittent stirring of 15 seconds duration was provided every five
minutes, the time required for 50 percent depletion of dichloromethane,
trichloromethane, and 1,1,1-trichloroethane, analogues of chloroethane, was
greater than 90 minutes.  This is on the order of four times longer than
with constant stirring.
                                     42-3

-------
         Billing et_ _al. (1975) are careful to point out the difficulties
encountered in extrapolating their laboratory results to real-world con-
ditions, where the concentration of the organic solute would probably be
very much less than 1 mg/1 and where surface and bulk agitation would be
highly variable.  Although the data appear to be valid on a relative basis
(i.e., correctly illustrating the relative rates of volatilization of
chlorinated aliphatics), they cannot be used as absolute measures of
volatilization rates from natural waters.  For the purposes of this docu-
ment, the data are used as rough-order-of-magnitude indications of the im-
portance of volatilization relative to other transport and fate processes,
with the strong effecs of agitation considered.  The validity of this
application has not been established.

         A subsequent study by Billing (1977) was conducted using the same
experimental conditions as in the 1975 study, and an average evaporative
half-life of 23.1 minutes was obtained for 1 mg/1 chloroethane in water.
The purpose of this subsequent study was to use the experimental data
previously obtained to test two theoretical models which may be used to
predict rates of evaporation of slightly soluble organic compounds from
water.  Billing (1977) found that the theoretical model by Mackay and
Wolkoff (1973) failed to predict evaporative half-lives , ancl that the
theoretical model of Mackay and Leinonen (1975) using parameters of Liss
and Slater (1974) correlated well with the experimental evaporative half-
lives obtained.  For example, the evaporative half-life obtained experi-
mentally by Billing (1977) was approximately 23 minutes compared to a 16.7
minute evaporative half-life predicted by the Mackay and Leinonen (1975)
model and 0.50 minutes predicted by Mackay and Wolkoff (1973).

         Billing (1977), however, comments that the apparent numerical
agreement between his data and the values predicted by the Mackay and
Leinonen (1975) model may be fortuitous.  Estimates of volatilization rates
based on the Mackay and Leinonen (1975) model depend primarily on liquid-
gas phase exchange rate constants, whereas the experimental model of
Billing e_t al. (1975) and Billing (1977) is controlled by the rate of
stirring and the wind velocity across the surface of the water.

         Pearson and McConnell (1975) suggest that the presence of ana-
logues of chloroethane  (such as chloroinethane, dichloromethane, trichloro-
methane, and 1,2-dichloroethane) in ambient waters is due to the absorption
of  these compounds from the atmosphere by water.  This process is thought
to  occur most effectively when the atmosphere is scrubbed as in the pre-
cipitation process.  Aerial transport of these halogenated aliphatics is
indicated by Pearson and McConnell (1975) to play a major role in the wider
distribution of these compounds and accounts for their presence in upland
waters .
                                      42-4

-------
    42.4.5  Sorption

         No information pertaining specifically to the adsorption of
chloroethane onto sediments was found.  Billing _e_t _al. (1975) carried out
two closed system experiments on dichloromethane,  trichloromethane, and
1,1,1-trichloroethane,  analogues of chloroethane,  where solute loss could
be only by adsorption.   Dry bentonite clay at 375  mg/1 was introduced into
a sealed solution and after ten minutes approximately ten percent of the
chloroalkanes had been removed from solution, presumably through adsorption
by the clay.  When the amount of clay added to the closed system was
doubled (750 mg/1) there was 22 percent solute loss after 30 minutes and no
further solute adsorption after this time.  The authors indicate that there
appeared to be little selectivity among the various chlorinated compounds
in the adsorption process.  The authors observed some adsorption of di-
chloromethane, trichlorotnethane, and 1,1,1-trichloroethane (analogues of
chloroethane) by dry powdered dolomitic limestone, but, again, no selectiv-
ity among the solutes was noted.

         In sealed aqueous samples with approximately 500 mg/1 peat moss,
approximately 40 percent of dichloromethane, trichloromethane, and 1,1,1,-
trichloroethane (analogues of chloroethane), was removed in 10 minutes.  At
longer times than 10 minutes, no further solute removal was noted.  The
subsequent decrease in the rate of disappearance of dichloromethane, tri-
chloromethane, and 1,1,1,-trichloroethane, analogues of chloroethane, at
longer time periods than 10 minutes was suggested  to be due to gradual re-
lease or desorption of these analogues of chloroethane from the peat moss
to the solution.

         Pearson and McConnell (1975) found no clear evidence of selective
concentrations by sediments of the chloroethane analogues trichloromethane,
tetrachloromethane, and 1,1,1-trichloroethane.

    42.4.6  Bioaccumulation

         According to Kopperman e_t al. (1976) not  all organochlorine com-
pounds bioaccumulate to high levels.  The data suggest that polar compounds
are more easily biodegraded, and the non-polar (highly lipophilic) com-
pounds accumulate.  Neely et^ al_. (1974) have shown that bioaccumulation is
related to  the octanol/water partition coefficient (P) of the compound.
The log octanol/water partition coefficient (log P) of 1.54 (Leo et al.
1971) indicates that chloroethane will probably not bioaccumulate to any
significant extent (see Methods section on bioaccumulation).

    42.4.7  Biotransformation and Biodegradation

         No information pertaining specifically to the rate of biodegrada-
tion of chloroethane in aquatic systems was found.  According to Pearson
and McConnell (1975) only completely sealed systems, such as the standard
                                      42-5

-------
BOD (biochemical oxygen demand) bottle technique,  can be used to measure
biochemical degradation of volatile compounds such as chloroethane.  The
BOD bottle experiments of Pearson and McConnell have been unable to demon-
strate any significant oxygen absorption from compounds containing only C,
H, and Cl, thus indicating that biochemical degradation of such compounds
is indeed very slow.  Literature references to microbial biodegradation are
few and conflicting;  the majority indicate that low molecular weight
chloroaliphatics are not metabolized (Pearson and McConnell 1975;
McConnell et, al. 1975).

42.5  Data Summary

    Table 42-1 summarizes the aquatic fate discussed above.  The oxidation
rate presented is an atmospheric photooxidation rate and refers to the rate
of reaction of chloroethane with hydroxyl radicals in the troposphere.

    The information found pertaining specifically to chloroethane is rela-
tively scant.  Based on the information presented above, it is concluded
(with low to medium confidence) that the important fate processes for
chloroethane are photooxidation in the atmosphere and hydrolysis, or both
of these occurring concomitantly.
                                     42-6

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42.6  Literature Cited

Billing, W.L.  1977.   Interphase transfer processes.  II.  Evaporation
  rates of chioromethanes,  ethanes, ethylenes, propanes, and propylenes
  from dilute aqueous solutions.  Comparisons with theoretical predictions.
  Environ. Sci. Technol.  11:405-409.

Billing, W.L., N.B. Tefertiller, and G.J. Kallos.   1975.  Evaporation
  rates of methylene chloride, chloroform, 1,1,1-trichloroethane,
  trichloroethylene,  tetrachloroethylene, and other chlorinated compounds
  in dilute aqueous solutions.  Environ. Sci. Technol. 9(9):833-838.

Environmental Protection.Agency.  I975a.  Preliminary assessment of
  suspected carcinogens in drinking water.  U.S.  Environmental Protection
  Agency, (Office of  Toxic  Substances), Washington, B.C.  33p.  EPA
  560/4-75-003.

Environmental Protection Agency.  1975b.  Report  on the problem of
  halogenated air pollutants and stratospheric ozone.  U.S. Environmental
  Protection Agency,  (Office of Research and Bevelopment),  Research
  Triangle Park, North Carolina.  55p.  EPA 600/9-75-008.

Howard, C.J. and K.M. Evenson.  1976.   Rate constants for the reactions
  of OH with ethane and some halogen substituted  ethanes at 296K.   J.
  Chem. Phys. 64(11):4303-4306.

Jaffa, H.H. and M. Orchin.   1962.  Theory and applications  of ultraviolet
  spectroscopy.  John Wiley and Sons,  Inc. New York. 624p.

Kopperman, H.L., B.W. Kuehl, and G.E.  Glass.  1976.  Chlorinated compounds
  found in waste treatment effluents and their capacity to  bioaccumulate.
  Proceedings of the conference on the environmental impact of water
  chlorination.  Oak Ridge, Tennessee, October 22-24, 1975. 327-345.

Leo, A., C. Hansch, and B.  Elkins.  1971.  Partition coefficients and their
  uses.  Chem. Rev. 71:525-616.

Liss, P.S. and P.G. Slater.  1974.  Flux of gases across the air-sea
  interface.  Nature 247:181-184.

Mackay, B. and P.J. Leinonen.  1975.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.   Environ. Sci. Technol.
  9:1178-1180.

Mackay, B. and A.W. Wolkoff.  1973.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.   Environ. Sci. Technol.
  7:611-614.
                                    42-8

-------
McConnell, G., D.M. Ferguson, and C.R. Pearson.   1975.   Chlorinated
  hydrocarbons and the environment.  Endeavor XXXIV:13-18.

Morrison, R.T. and R.N. Boyd.  1973.  Organic chemistry.   Allyn and
  Bacon, Inc., Boston, Mass.  1258p.

Neely, W.B., D.R. Branson, and G.E. Blau.  1974.   Partition coefficient to
  measure bioconcentration potential of organic  chemicals in fish.
  Environ. Sci. Technol. 8:1113-1115.

Pearson, C.R. and G. McConnell.  1975.  Chlorinated C^  and  C2
  hydrocarbons in the marine environment.  Proc.  Roy.  Soc.  London B
  189:305-322.

Radding, S.B., D.H. Liu, H.L. Johnson, and T. Mill.  1977.   Review of  the
  environmental fate of selected chemicals.  U.S.  Environmental Protection
  Agency, (Office of Toxic Substances), Washington, D.C.  I47p.   EPA
  560/5-77-003.

Verschueren, K.  1977.  Handbook of environmental  data  on organic
  chemicals.  Van Nostrand/Reinhold Press, New York.   659p.

Weast, R.C. (ed.).  1977.  Handbook of chemistry  and  physics.  58th
  Edition.  CRC Press, Inc., Cleveland, Ohio. 2398p.
                                       42-9

-------
               43.   1,1-DICHLOROETHANE (ETHYLIDINE CHLORIDE)
43.1  Statement of Probable Fate

    Volatilization appears to be the major transport process for removal of
1,1-dichloroethane from aquatic systems.  Once in the troposphere, 1,1-di-
chloroethane is attacked by hydroxyl radicals at a relatively rapid rate,
so that the tropospheric lifetime is on the order of a month.  A very small
amount may reach the stratosphere, where it would undergo photodissocia-
tion.  Very little information was found concerning other aquatic fate pro-
cesses for this compound, but information on analogous compounds suggest
that oxidation, hydrolysis and biodegradation are probably not important
for 1,1-dichloroethane.

43.2  Identification

    1,1-Dichloroethane, or ethylidine chloride, has been detected in
finished drinking water by the National Organics Reconnaissance Survey (En-
vironmental Protection Agency 1975a).

    The chemical structure of 1,1-dichloroethane is shown below:

                                            Alternate Names
          Cl    H                           	
                                            Ethylidene Chloride
     H	 C 	 C 	H                      Ethylidene Dichloride
          Cl     H

     1,1-Dichloroethane

     CAS NO.  75-34-3
     TSL NO.  KI 01750

43.3  Physical Properties

    The general physical properties of 1,1-dichloroethane are given below.

    Molecular weight                         98.96
     (Weast 1977)

    Melting point                            -96.98C
     (Weast 1977)

    Boiling point at 760 torr                57.28C
    (Weast 1977)
                                     43-1

-------
    Vapor pressure at 20C                     180 torr
    (Verschueren 1977)

    Solubility in water at 20C              5,500 mg/1
    (Verschueren 1977)

    Log octanol/water partition coefficient  1.79
    (Hansch jet al. 1975)

43.4  Summary of Fate Data

    43.4.1  Photolysis

         No information pertaining specifically to the rate of photodis-
sociation of 1,1-dichloroethane in the aqueous environment was found.  Dil-
ling _e_t _a_l. (1975) reported that within experimental error, sunlight did
not accelerate the decomposition of dichloromethane, trichloromethane, and
1,1,1-trichloroethane (analogues of 1,1-dichloroethane) initially present
at 1 mg/1 concentrations in aerated water.

         Photodissociation in the terrestrial environment would not be ex-
pected to occur for 1,1-dichloroethane since this compound has no chromo-
phores which absorb in the visible or near ultraviolet region of the elec-
tromagnetic spectrum (Jaffa and Orchin 1962).  The rate of reaction of
1,1-dichloroethane with hydroxyl radicals in the troposphere appears to be
sufficiently rapid that very little of the unreacted compound will reach
the stratosphere where it could undergo photodissociation due to the pres-
ence of higher energy, shorter wavelength ultraviolet light.  Chloroacetyl
chloride would probably be formed as the principal initial product of pho-
todissociation (Environmental Protection Agency 1975b).

    43.4.2  Oxidation

         No information pertaining specifically to the oxidation of 1,1-
dichloroethane in the aqueous environment was found.  There is, however,
indirect evidence that oxidation of this compound in aquatic systems is not
important.  Billing _e_t _aJ.. (1975) conducted aqueous reactivity experiments
in which there was a large excess of dissolved oxygen compared to the total
amount of dichloromethane, trichlororaethane, and 1,1,1-trichloroethane ,
compounds similar to 1,1-dichloroethane.  The half-lives for the three
chloroalkanes in  the system ranged from 6 to 18 months.  The decomposition
observed may have resulted from oxidation, hydrolysis, or both, but this
range of half-lives is much longer than the 22 minute half-life reported by
Dilling _e_t _al. (1975) for volatilization of 1,1-dichloroethane from a stir-
red aqueous system.  In addition, the major portion of the disappearance
was attributed to ionic hydrolysis rather than oxidation.
                                    43-2

-------
         Due to the relatively high vapor pressure of 1,1-dichloroethane,
volatilization to the atmosphere is expected to be quite rapid.  Once in
the troposphere, 1,1-dichloroethane is attacked by hydroxyl radicals.  The
biomolecular rate of this reaction at 23C was found by Howard and Evenson
(1976) to be 2.6(+0.6)xlO~13cm3molecule~1sec~1, which, for a
tropospheric hydroxyl radical concentration of 10 cm~ , corresponds to
a lifetime (time for reduction to 1/e of original concentration) in the
troposphere of about 1.5 months.

         Based on a tropospheric lifetime of 1.5 months and a troposphere-
to-stratosphere turnover time (time for all but 1/e of tropospheric air to
diffuse into the stratosphere) of 30 years, somewhat less than one percent
of tropospheric 1,1-dichloroethane would be expected to eventually reach
the stratosphere.  It should be noted that this estimation neglects the ef-
fects of 1,1-dichloroethane removal from the troposphere by other pro-
cesses such as rainfall.

    43.4.3  Hydrolysis

         No information pertaining specifically to the hydrolytic half-life
of 1,1-dichloroethane in the aqueous environment was found.  Information
on compounds structurally similar to 1,1-dichloroethane is not suitable for
prediction of analogous behavior for this compound as is shown below:

                    Estimate of
   Compound     Hydrolytic Half-Life     Limitation       Reference

chloroethane           40 days          upper limit    Radding et al.(1977)
1,2-dichloroethane   50,000 years       upper limit    Radding et al.(1977)
1,1,1-trichloroethane  6 months         lower limit    Pilling et al.(1975)

In view of the volatility of 1,1-dichloroethane, however, it appears un-
likely that hydrolysis would be the principal fate process for 1,1-di-
chloroethane .

    43.4.4  Volatilization

         Dilling _e_t al. (1975) estimated the experimental half-life for
volatilization of 1,1-dichloroethane originally present at 1 mg/1 to be 22
minutes when stirred at 200 rpm in water at approximately 25C in an open
container.  Removal of 90 percent of the 1,1-dichloroethane under the same
conditions required 109 minutes.  For chloroaliphatics in general, stirring
                                    43-3

-------
speed was found to have a marked effect on volatilization rate.   With
intermittent stirring for 15 seconds every five minutes,  the time required
for 50 percent depletion of dichloromethane,  trichloromethane and 1,1,1-
trichloroethane,  (compounds analogous to 1,1-dichloroethane) was greater
than 90 minutes,  or on the order of four times greater than was  observed
during constant stirring.  The presence of sodium chloride at a  3 percent
concentration, as in seawater, caused about a 10 percent  decrease in the
evaporative rate.

         Dilling et al. (1975) are careful to point out the difficulties
encountered in extrapolating their laboratory results to  real-world con-
ditions, where the concentration of the organic solute would probably be
very much less than 1 mg/1 and where surface  and bulk agitation  would be
highly variable.   Although the data appear to be valid on a relative basis
(i.e., correctly illustrating the relative rates of volatilization of
chlorinated aliphatics), they cannot be used  as absolute  measures of vola-
tilization rates from natural waters.  For the purposes of this  document,
the data are used as rough-order-of-magnitude indications of the importance
of volatilization relative to other transport and fate processes, with the
strong effects of agitation considered.  The  validity of  this application
has not been established.

         A subsequent study by Dilling (1977) was conducted using the same
experimental conditions as in the 1975 study, and an average half-life of
32.2 minutes was obtained for 1,1-dichloroethane at an initial concentra-
tion of 1 mg/1. The purpose of this subsequent study was  to use  the experi-
mental data previously obtained to test two theoretical models which may be
used to predict evaporative rates of slightly soluble organic compounds
from water.  Dilling (1977) found that the theoretical model proposed by
Mackay and Wolkoff (1973) failed to predict evaporative half-lives, and
that the theoretical model of Mackay and Leinonen (1975)  using para-
meters of Liss and Slater (1974) correlated well with the experimental
half-lives obtained. For example, the evaporative half-life obtained ex-
perimentally by Dilling  (1977) was approximately 32 minutes as compared to
a 21.2 minute evaporative half-life obtained  by the Mackay and Leinonen
(1975) model and 0.98 minutes predicted by the Mackay and Wolkoff (1973)
model.

         Dilling (1977), however, comments that the apparent numerical
agreement between his data and the values predicted by the Mackay and
Leinonen (1975) model may be fortuitous.  Estimates of volatilization rates
based on the Mackay and  Leinonen (1975) model depend primarily on liquid-
gas  phase exchange rate  constants, whereas the experimental model of
Dilling _e_t _al. (1975) and Dilling (1977) is controlled by the rate of
stirring and the wind velocity across the surface of the water.
                                     43-4

-------
         Pearson and McConnell (1975) suggest that the presence of 1,2-
dichloroethane (a structural isomer of 1,1-dichloroethane) and other halo-
genated aliphatics in ambient waters is due to the absorption of chloroor-
ganics from the atmosphere by water and return to the earth during precipi-
tation.  Aerial transport of these chloroorganics is indicated by Pearson
and McConnell (1975) to play a major role in the wider distribution of
these compounds and accounts for their presence in upland waters.

    43.4.5  Sorption

         No information pertaining specifically to the adsorption of 1,1-
dichloroethane onto sediments was found.  Dilling jet _al. (1975) carried out
two closed system experiments on dichloromethane, trichloromethane, and
1,1,1-trichloroethane, analogues of 1,1-dichloroethane, where solute loss
could only occur by adsorption.  Dry bentonite clay at 375 mg/1 was intro-
duced into a solution in a sealed container, and after ten minutes approxi-
mately ten percent of the chloroalkanes studied had been removed from solu-
tion, presumably through adsorption by the clay.  When the amount of clay
added to the closed system was doubled (750 mg/1) there was 22 percent
solute loss after 30 minutes.  There was no further solute adsorption after
this time.  Little selectivity was found among the various chlorinated com-
pounds in the adsorption process.  The authors observed some adsorption of
dichloromethane, trichloromethane, and 1,1,1-tichloroethane by dry powdered
dolomitic limestone, but, again, without selectivity among the solutes
used.

         In sealed aqueous samples that contained approximately 500 mg/1
peat moss, approximately 40 percent of dichloromethane, trichloromoethane,
and 1,1,1-trichloroethane was removed in 10 minutes (Dilling _e_t _al. 1975).
At longer times, no further solute removal was noted.

         Pearson and McConnell (1975) found no clear evidence of selec-
tive concentration of 1,1,1-trichloroethane, trichloromethane, and tetra-
chlororaethane in sediments.  Similar observations would presumably be ob-
tained for 1,1-dichloroethane.

    43.4.6  Bioaccumulation

         According to Kopperman _e_t al_. (1976) not all organochlorine com-
pounds bioaccumulate to high levels.  The data suggest that polar com-
pounds are more easily biodegraded, and the non-polar (highly lipophilic)
compounds accumulate.  Neely je_t _al. (1974) have shown that bioaccumulation
is related to the octanol/water partition coefficient (P) of the compound.
The log octanol/water partition coefficient (log P) of 1.79 as calculated
                                    43-5

-------
by the method of Hansch (Tute 1971;  Hansch jat al.  1975)  indicates that
1,1-dichloroethane will probably not bioaccumulate to  any significant ex-
tent (see Methods section on bioaccumulation).

    43.4.7  Biotransformation and Biodegradation

         No information pertaining specifically to the rate of biodegrada-
tion of 1,1-dichloroethane in aquatic systems was  found.   Thorn and Agg
(1975) have included 1,2-dichloroethane,  an analogue  of  1,1-dichloro-
ethane, in a list of synthetic organic chemicals which should be degradable
by biological sewage treatment provided suitable acclimatization can be
achieved.  They note, however, that  not many compounds in this list occur
in nature and, as a result, it is unlikely that microorganisms already pos-
sess the ability to destroy them.  The authors consider  1,2-dichloro-
ethane, along with various other compounds, to be  potentially biodegrad-
able.  By analogy, 1,1-dichloroethane may also be  potentially biodgradable.

         According to Pearson and McConnell (1975) only  completely sealed
systems, such as the standard BOD (biochemical oxygen demand) bottle tech-
nique, can be used to measure biochemical degradation of volatile com-
pounds such as 1,1-dichloroethane.  The BOD bottle experiments of Pearson
and McConnell have been unable to demonstrate any  significant oxygen ab-
sorption from compounds containing only C, H, and  Cl,  thus indicating that
biochemical degradation of such compounds is indeed very slow.  Literature
references to microbial biodegradation are few and conflicting;  the major-
ity indicate that low molecular weight chloroaliphatics  are not metabolized
(Pearson and McConnell 1975; McConnell _et al. 1975).

         Pearson and McConnell (1975) found some evidence of metabolism of
1,2-dichloroethane, an analogue of 1,1-dichloroethane, in the tissues of
both fish and oysters.  The administration of carbon-14  labeled 1,2-di-
chloroethane to fish and oysters resulted in rapid accumulation of l^C to
an asymptotic level, followed by loss on transfer  to  clean sea water.  An-
alysis by gas chromatography, however, showed much lower levels of 1,2-di-
chloroethane to be present than predicted by the 14^  in the tissue, thus
indicating that some metabolism may have occurred.  It is possible that
1,1-dichloroethane may be metabolized as well.

 43.5  Data Summary

    Table 43-1 summarizes the aquatic fate information discussed above.
The oxidation rate is a photooxidation rate for the reaction of 1,1-dichlo-
roethane with hydroxyl radicals in the troposphere as well as the hydro-
                                      43-6

-------
sphere.  1,1-Dichloroethane is believed to be volatilized from aquatic sys-
tems rapidly, and subsequently to be photooxidized by hydroxyl radicals in
the troposphere.   The little information found concerning other fate
processes indicate that they play a minor role in comparison to photo-
oxidation for this compound.
                                    43-7

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43.6  Literature Cited

Billing, W.L.  1977.  Interphase transfer processes.  II.  Evaporation
  rates of chloromethanes,  ethanes, ethylenes, propanes, and propylenes
  from dilute aqueous solutions.  Comparisons with theoretical predictions.
  Environ. Sci. Technol. 11:405-409.

Billing, W.L., N.B. Tefertiller, and G.J. Kallos.  1975.  Evaporation
  rates of methylene chloride, chloroform, 1,1,1-trichloroethane,
  trichloroethylene, tetrachloroethylene, and other chlorinated compounds
  in dilute aqueous solutions.  Environ. Sci. Technol. 9(9) :833-838.

Environmental Protection. Agency.  1975a.  Preliminary assessment of
  suspected carcinogens in drinking waters.  U.S. Environmental Protection
  Agency, (Office of Toxic Substances), Washington, B.C. 33p. EPA
  560/4-75-003.

Environmental Protection Agency.  1975b.  Report on the problem of
  halogenated air pollutants and stratospheric ozone.  U.S. Environmental
  Protection Agency, (Office of Research and Bevelopment),  Research
  Triangle Park, North Carolina.  55p.  EPA 600/9-75-008.

Hansch, C., A. Vittoria, C. Silipo, and P.Y.C. Jow.  1975.   Partition
  coefficients and the structure-activity relationship of the anesthetic
  gases.  J. Med. Chem. 18(6):546-548.

Howard, C.J. and K.M. Evenson.  1976.  Rate constants for the reactions of
  OH with ethane and some halogen substituted ethanes at 296 K.  J. Chem.
  Phys. 64(11):4303-4306.

Jaffe, H.H. and M. Orchin.   1962.  Theory and applications  of
  ultraviolet spectroscopy.  John Wiley and Sons, Inc. New York. 624p.

Kopperman, H.L., B.W. Kuehl, and G.E. Glass.  1976.  Chlorinated compounds
  found in waste treatment effluents and their capacity to  bioaccumulate.
  Proceedings of the conference on the environmental impact of water
  chlorination.  Oak Ridge, Tennessee, October 22-24, 1975.  327-345p.

Liss, P.S. and P.G. Slater.  1974.   Flux of gases across the air-sea
  interface.  Nature 247:181-184.

Mackay, B. and P.J. Leinonen.  1975.  Rate of evaporation of low-solubility
  contaminants form water bodies to atmosphere.  Environ. Sci. Technol.
  9:1178-1180.
                                    43-9

-------
Mackay, D. and A.W. Wolkoff.   1973.   Rate of evaporation of low-solubility
  contaminants from water bodies to  atmosphere.  Environ. Sci.  Technol.
  7:611-614.

McConnell, G., D.M. Ferguson, and C.R. Pearson.  1975.   Chlorinated
  hydrocarbons and the environment.   Endeavor XXXIV:13-18.

Neely, W.B., D.R. Branson, and G.E.  Blau.  1974.  Partition coefficient  to
  measure bioconcentration potential of organic chemicals in fish.
  Environ. Sci. Technol. 8:1113-1115.

Pearson, C.R. and G. McConnell.  1975.  Chlorinated C^  and G
  hydrocarbons in the marine environment.  Proc. Roy.  Soc. London B
  189:305-322.

Radding, S.B., D.H. Liu, H.L. Johnson, and T. Mill.  1977.  Review of the
  environmental fate of selected chemicals.  U.S. Environmental
  Protection Agency, (Office of Toxic Substances), Washington,  D.C.
  I47p.  EPA-560/5-77-003.

Thorn, N.S. and A.R. Agg.  1975.  The breakdown of synthetic organic
  compounds in biological processes.  Proc. Roy. Soc.  London B  189:347-357.

Tute, M.S.  1971.  Principles and practice of Hansch analysis:   a guide  to
  structure-activity correlation for the medicinal chemist.  Adv. Drug Res.
  6:1-77.

Verschueren, K.  1977.  Handbook of  environmental data  on organic
  chemicals.  Van Nostrand/Reinhold Press, New York.   659 p.

Weast, R.C. (ed.).  1977.  Handbook of chemistry and physics.  58th
  Edition. CRC Press, Inc., Cleveland, Ohio.  2398p.
                                   43-10

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               44.   1,2-DICHLOROETHANE (ETHYLENE DICHLORIDE)
44.1  Statement of Probable Fate

    Volatilization appears to be the major transport process for removal of
1,2-dichloroethane from aquatic systems.  Once in the troposphere, 1,2-di-
chloroethane is attacked by hydroxyl radicals, reportedly resulting in
chloroacetyl chloride as the principal initial product.   The atmospheric
lifetime of 1,2-dichloroethane, based on its rate of reaction with hydroxyl
radicals, is reported to be on the order of several months.  Any unreacted
1,2-dichloroethane which diffuses upward to the stratosphere will undergo
photodissociation.

    Other processes do not appear to be significant for 1,2-dichloroethane.
Oxidation in the aquatic environment probably does not occur except for
limited photo oxidation.  Hydrolysis is also probably too slow to be an
important process for removal of 1,2-dichloroethane.  Although no informa-
tion was found indicating that microorganisms exist which can readily bio-
degrade 1,2-dichloroethane, information was found to indicate that this
compound may be metabolized by fish and oysters.   There is, at present, no
evidence of either bioaccumulation of 1,2-dichloroethane in aquatic organ-
isms or adsorption of this compound onto suspended solids or sediments.

44.2  Identification

    1,2-Dichloroethane, or ethylene dichloride, has been detected in
finished drinking water by the National Organics  Reconnaissance Survey (En-
vironmental Protection Agency 1975a).  In addition, 1,2-dichloroethane has
been detected in the atmosphere in unquantified concentration (Hanst 1978).

    The chemical structure of 1,2-dichloroethane  is shown below.

         Cl    Cl
                I                             Alternate Names

    Hc     C    H                        Ethylene dichloride
                j                             Glycol dichloride
         H     H

    1,2-Dichloroethane

    CAS NO. 107-06-2
    TSL NO. KI 05250
                                     44-1

-------
44.3  Physical Properties

    The general physical properties of 1,2-dichloroethane are given below.

   Molecular weight                          98.98
   (Weast 1977)

   Melting point                             -35.36C
   (Weast 1977)

   Boiling point at 760 torr                 83.47C
   (Weast 1977)

   Vapor pressure at 20C                    61 torr
   (Verschueren 1977)

   Solubility in water at 20C               8,690 mg/1
   (Verschueren 1977)

   Log octanol/water partition               1.48
   coefficient (Radding _et al. 1977)

44.4  Summary of Fate Data

    44.4.1  Photolysis

         No information pertaining specifically to the rate of photo-
dissociation of 1,2-dichloroethane in the aqueous environment was found.
Dilling et al. (1975) reported that, within experimental error, sunlight
did not accelerate  the decomposition of such compounds as dichlorotnethane,
trichloromethane, and 1,1,1-trichloroethane (analogues of 1,2-dichloro-
ethane) initially present at 1 mg/1 concentrations in aerated water.  Pho-
todissociation of 1,2-dichloroethane in the aquatic environment would not
be expected to occur since this compound has no chromophores which absorb
in the visible or near ultraviolet region of the electromagnetic spectrum
(Jaffe and Orchin 1962).

         The rate of reaction of 1,2-dichloroethane with tropospheric
hydroxyl radicals is sufficiently rapid that very little unreacted compound
would be expected to reach the stratosphere where it couuld undergo photo-
dissociation due to the presence of higher energy, shorter wavelength
ultraviolet light.   Chloroacetyl chloride would probably be an initial pho-
todissociation product (Environmental Protection Agency 1975b).

    44.4.2  Oxidation

         No information pertaining specifically to the oxidation of 1,2--
dichloroethane in the aqueous environment was found.  There is, however,
                                     44-2

-------
some indirect evidence from which the lack of potential for oxidation in
aquatic systems can be inferred.  Dilling _et _al. (1975) conducted aqueous
reactivity experiments in which there was a large excess of dissolved oxy-
gen compared to the total amount of dichloromethane ,  trichloromethane , and
1,1 ,1-trichloroethane, compounds that are similar to  1 ,2-dichloroethane.
The rates of disappearance observed for these compounds were much lower
than rates of loss by volatilization from water reported by the same
authors for 1,2-dichloroethane and for the three analogous compounds.  In
addition, Dilling _e_t a_l. (1975) attributed the major portion of the ob-
served disappearance to ionic hydrolysis rather than  to oxidation.  These
results indicate that direct oxidation in aquatic systems is probably not
important for 1,2-dichloroethane.

         Due to the relatively high vapor pressure of 1,2-dichloroethane,
volatilization to the atmosphere is expected to be quite rapid.  Once in
the troposphere, the compound reacts with hydroxyl radicals.  According to
Howard and Evenson (1976), the bimolecular rate for this reaction at 23C
is 2.2 (+0.5) x 10~13 cm^ molecule" 1 sec~l, which, assuming a
tropospheric hydroxyl radical concentration of 10 cm~^ , corresponds to
a lifetime (time for reduction to 1/e of original concentration) in the
troposphere of about 1.7 months.  Radding _e_t _al. (1977) indicate an atmo-
spheric half-life for this photooxidation reaction of  234 hours (0.3
months), corresponding to a rate constant of 0.1 x 10   M~^
         Other information on the reaction of 1,2-dichloroethane with
hydroxyl radicals indicate an atmospheric lifetime of three or four months
(Environmental Protection Agency 1975b).  Chloracetyl chloride is reported
to be the principal initial reaction products formed.  This compound is
readily hydrolyzed to HC1 and the corresponding carboxylic acid (Morrison
and Boyd 1973).

         Based on a tropospheric lifetime on the order of one month, and
assuming a troposphere-to-atmosphere turnover time (time for all but 1/e of
tropospheric air to diffuse into the stratosphere) of 30 years, somewhat
less than one percent of tropospheric 1,2-dichloroethane would be expected
to eventually reach the stratosphere.  It should be noted that the above
analysis neglects the effects of other tropospheric removal processes such
as scrubbing by rain.

    44.4.3  Hydrolysis

         A maximum hydrolytic half-life of 50,000 years has been estimated
for 1,2-dichloroethane at pH 7 and 25C from experimental data at 100-150C
(Radding et al. 1977).  This corresponds to a first-order rate constant for
hydrolysis of 1,2-dichloroethane of approximately 5 x 10~^sec~l
(Radding et_ _al. 1977).  This estimated half-life is based on the half-life
of 1 , 2-dibromoethane which, being brominated rather than chlorinated, was
                                     44-3

-------
theorized to be ten times as reactive as 1,2-dichloroethane.   The validity
of this estimation method has not been established.   It must  be noted that
50,000 years is much longer than indicated by the 6  to 18 month minimum
hydrolytic half-lives for similar compounds (dichloromethane, tri-
chloromethane, 1,1,1-trichloroethane) reported by Dilling _e_t  al. (1975).
In addition, Radding e_al. (1977) estimated the hydrolytic"half-life of
chloroethane, another compound similar to 1,1-dichloroethane, to be 40 days
or less.  While the rate of hydrolysis does not appear to be  well estab-
lished for 1,2-dichloroethane, it appears unlikely that hydrolysis is
important for this compound in aquatic systems because of the rapidity of
the volatilization loss.

    44.4.4  Volatilization

         Dilling _e_t _al. (1975) estimated the experimental half-life for
volatilization of 1,2-dichloroethane originally present at 1  mg/1 to be 29
minutes when stirred at 200 rpm in water at approximately 25C in an open
container.  Removal of 90 percent of the 1,2-dichloroethane under the same
conditions required 96 minutes.  For chloroaliphatics in general, stirring
speed was found to have a marked effect on volatilization rate.  With
intermittent stirring for 15 seconds every five minutes, the  time required
for 50 percent depletion of dichloromethane, trichiororaethane, and 1,1,1-
trichloroethane, analogues of 1,2-dichloroethane, was greater than 90
minutes, or on the order of four times greater than  was observed during
constant stirring.  The presence of sodium chloride  at a 3 percent con-
centration, as in seawater, effected a 10 percent decrease in the evapora-
tive rate of the chlorinated compound. Dilling et al. (1975)  are careful  to
point out the difficulties encountered in extrapolating their laboratory
results to real-world conditions, where the concentration of  the organic
solute would probably be very much less than 1 mg/1  and where surface and
bulk agitation would be highly variable.  Although the data appear to be
valid on a relative basis (i.e., correctly illustrating the relative rates
of volatilization of chlorinated aliphatics), they cannot be  used as
absolute measures of volatilization rates from natural waters.  For pur-
poses of this document, the data are used as rough-order-of-magnitude in-
dications of the importance of volatilization relative to other transport
and fate processes, with the strong effects of agitation considered.  The
validity of this application has not been established.

         A subsequent study by Dilling (1977) was conducted using the same
experimental conditions as the aforementioned investigation,  and an average
half-life of 28 minutes with respect to volatilization of 1 mg/1 of 1,2-
dichloroethane was obtained.  The purpose of this subsequent  study was to
use the experimental data obtained to test two theoretical models formu-
lated to predict evaporative rates of slightly soluble organic compounds
from water.  Dilling (1977) found that the theoretical model  by Mackay and
Wolkoff (1973) failed to predict evaporative half-lives, but  that the
                                     44-4

-------
theoretical model of Mackay and Leinonen (1975) using parameters of Liss
and Slater (1974) correlated well with the experimental half-lives ob- 
tained.  For example, the evaporative half-life obtained experimentally by
Billing (1977) was approximately 28 minutes as compared to a 24.5 minute
evaporative half-life obtained by the Mackay and Leinonen (1975) model,
whereas the Mackay and Wolkoff (1973) model predicted a half-life of 4.5
minutes.

         Billing (1977) however, comments that the apparent numerical
agreement between data and the values predicted by the Mackay and Leinonen
(1975) model may be fortuitous.  Estimates of volatilization rates based on
the Mackay and Leinonen (1975) model depend primarily on liquid-gas phase
exchange rate constants,, whereas the experimental model of Billing _et al.
(1975) and Billing (1977) is controlled by the rate of stirring and the
wind velocity across the surface of the water.

         Pearson and McConnell (1975) suggest that the presence of 1,2-
dichloroethane, as well as other halogenated aliphatics, in ambient waters
is due to the absorption of chloroorganics from the atmosphere by water and
their return to the earth in precipitation.  Aerial transport of these
chloroorganics is indicated by Pearson and McConnell (1975) to play a major
role in their distribution and accounts for their presence in upland
waters.

    44.4.5  Sorption

         No information pertaining specifically to the adsorption of 1,2-
dichloroethane onto sediments was found.  Billing et al. (1975) carried out
two closed system experiments on dichloromethane, trichloromethane, and
1,1,1-trichloroethane (analogues of 1,2-dichloroethane) where solute loss
could occur by adsorption.  Bry bentonite clay at 375 mg/1 was introduced
into a sealed solution of the haloaliphatic compounds and, after ten
minutes, approximately ten percent of the initial solute was removed from
solution, presumably through adsorption by the clay.  When the amount of
clay added to the closed system was doubled (750 mg/1), a solute loss of 22
percent occurred after 30 minutes with no further solute adsorption after
this time.  The authors indicated that there appeared to be little selec-
tivity among the various chlorinated compounds in the adsorption process.
Some adsorption of dichloromethane, trichloromethane, and 1,1,1-trichloro-
ethane by dry powdered dolomitic limestone was also noted, but, again,
without any selectivity among the compounds.

         In sealed aqueous samples with approximately 500 mg/1 peat moss,
approximately 40 percent of chloroaliphatic solutes (dichloromethane, tri-
chloromethane, and 1,1,1-trichloroethane) was removed in 10 minutes.  At
longer times, no further solute removal was noted.
                                     44-5

-------
         Pearson and McConnell (1975)  found no clear evidence of selec-
tive concentration of three analogues  of 1,2-dichloroethane (1,1,1-tri-
chloroethane, trichloromethane,  and tetrachloromethane)  in sediments.   This
would presumably also be the case for  1,2-dichloroethane.

    44.4.6  Bioaccumulation

         According to Kopperman  et^ l. (1976)  not all organochlorine com-
pounds bioaccumulate to high levels.  The data suggest that polar compounds
are more easily biodegraded and  that the non-polar (highly lipophilic)  com-
pounds accumulate.  Neely _e_t _al. (1974)  have shown that  bioaccumulation is
related to the octanol/water partition coefficient (P) of  the compound.
The log octanol/water partition coefficient (log P) of 1.48 (Radding et al.
1977) indicates that 1,2-dichloroethane will probably not  bioaccumulate to
any significant extent.

    44.4.7  Biotransformation and Biodegradation

         No information pertaining specifically to the rate of biodegrada-
tion of 1,2-dichloroethane in aquatic  systems  was found.  Thorn and Agg
(1975) have included 1,2-dichloroethane in a list of synthetic organic
chemicals which should be degradable by biological sewage  treatment pro-
vided that suitable acclimatization can be achieved.  They note, however,
that not many compounds in this list occur in nature and,  as a result, it
is unlikely that microorganisms already possess the ability to destroy
them.

         According to Pearson and McConnell (1975) only completely sealed
systems, such as  the standard BOD (biochemical oxygen demand) bottle tech-
nique, can be used to measure biochemical degradation of volatile compounds
such as 1,2-dichloroethane.  The BOD bottle experiments of Pearson and
McConnell have been unable to demonstrate any significant  oxygen demand for
compounds containing only C, H,  and Cl, thus indicating that biochemical
degradation of such compounds is indeed very slow.  Literature references
to microbial biodegradation are few and conflicting; the majority report
that low molecular weight chloroaliphatics are not metabolized (Pearson and
McConnell 1975; McConnell _et al. 1975).

         Pearson  and McConnell  (1975) found some evidence  of metabolism of
1,2-dichloroethane in the tissues of both fish and oysters.  The adminis-
tration of carbon-14 labelled 1,2-dichloroethane to fish and oysters re-
sulted in rapid accumulation of l^C to an asymptotic level, followed by
loss on transfer  to clean sea water.  Analysis by gas chromatography, how-
ever,  showed much lower levels  of 1,2-dichloroethane  to be present than
predicted by the  *^C accumulation, thus indicating that some metabolism
may have occurred.
                                     44-6

-------
44.5  Data Summary

    Table 44-1 summarizes the aquatic fate information discussed above.
The oxidation rate is a. photooxidation rate and refers to the rate of re-
action of 1,2-dichloroethane with hydroxyl radicals in the troposphere as
well as in the hydrosphere.

    Due to the relatively high vapor pressure of 1,2-dichloroethane, vo-
latilization from the aquatic system to the atmosphere is quite rapid.
Once in the troposphere, 1,2-dichloroethane is photooxidized relatively
rapidly, having an atmospheric lifetime of order one to four months. The
photooxidation process appears to be much more rapid than other fate
processes which could occur.
                                     44-7

-------
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-------
44.6  Literature Cited

Dilling, W.L.  L977.  Interphase transfer processes.  II.  Evaporation
  rates of chloromethanes,  ethanes, ethylenes, propanes, and propylenes
  from dilute aqueous solutions.  Comparisons with theoretical predictions.
  Environ. Sci. Technol. 11:405-409.

Dilling, W.L., N.B. Tefertiller, and G.J. Kallos.  1975.  Evaporation
  rates of methylene chloride, chloroform, 1,1,1-trichloroethane,
  trichloroethylene, tetrachloroethylene, and other chlorinated compounds
  in dilute aqueous solutions.  Environ. Sci. Technol. 9(9):833-838.

Environmental Protection Agency.  1975a.  Preliminary assessment of
  suspected carcinogens in drinking water.  U.S. Environmental Protection
  Agency, (Office of Toxic Substances), Washington, D.C. 33p. EPA
  560/4-75-003.

Environmental Protection Agency.  1975b.  Report on the problem of
  halogenated air pollutants and stratospheric ozone.  U.S. Environmental
  Protection Agency, (Office of Research and Development), Research
  Triangle Park, North Carolina.  55p. EPA 600/9-75-008.

Hanst, P.L.  1978.  Noxious trace gases in the air, Part II:   halogenated
  pollutants.  Chemistry 51(2):6~12.

Howard, C.J. and K.M. Evenson.  1976.  Rate constants for the reactions
  of OH with ethane and some halogen substituted ethanes at 296 K.
  J. Chem. Phys.  64(11):4303-4306.

Jaffa, H.H. and M. Orchin.  1962.  Theory and applications of ultraviolet
  spectroscopy.  John Wiley and Sons, Inc. New York. 624p.

Kopperman, H.L., D.W. Kuehl, and G.E. Glass.  1976.  Chlorinated compounds
  found in waste treatment effluents and their capacity to bioaccumulate.
  Proceedings of the conference on the environmental impact of water
  chlorination.  Oak Ridge, Tennessee, October 22-24, 1975.  327-345p.

Liss, P.S. and P.G. Slater.  1974.  Flux of gases across the air-sea
  interface.  Nature 247:181-184.

Mackay, D. and P.J. Leinonen.  1975.  Rate of evaporation of
  low-solubility contaminants from water bodies to atmosphere.  Environ.
  Sci. Technol.  9:1178-1180.

Mackay, D. and A.W. Wolkoff.  1973.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. Technol.
  7:611-614.
                                    44-9

-------
McConnell, G., D.M. Ferguson, and C.R.  Pearson.   1975.   Chlorinated
  hydrocarbons and the environment.  Endeavor XXXIV.'13-18.

Morrison, R.T. and R.N. Boyd.  1973.   Organic chemistry.   3rd Edition.
  Allyn and Bacon, Inc., Boston, Mass.   1258p.

Neely, W.B., D.R. Branson, and G.E. Blau.   1974.   Partition coefficient to
  measure bioconcentration potential  of organic  chemicals  in fish.
  Environ. Sci. Technol. 8:1113-1115.

Pearson, C.R. and G. McConnell.  1975.   Chlorinated C}  and C2
  hydrocarbons in the marine environment.   Proc.  Roy.  Soc.  London B
  189:305-322.

Radding, S.B., D.H. Liu, H.L. Johnson,  and T. Mill.  1977.   Review of the
  environmental fate of selected chemicals.  U.S.  Environmental Protection
  Agency, (Office of Toxic Substances), Washington, D.C.  147p.
  EPA-560/5-77-003.

Thorn, N.S. and A.R. Agg.  1975.  The  breakdown of  synthetic organic
  compounds in biological processes.   Proc. Roy.  Soc.  London B
  189:347-357.

Verschueren, K.  1977.  Handbook of environmental  data on  organic
  chemicals.  Van Nostrand/Reinhold Press, New York.   659p.

Weast, R.C. (ed.). 1977.  Handbook of chemistry  and physics.  58th Edition.
  CRC Press, Inc., Cleveland, Ohio.  2398p.
                                   44-10

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              45.  1,1,1-TRICHLOROETHANE (METHYL CHLOROFORM)
45.1  Statement of Probable Fate

    Volatilization is the major transport process for removal of 1,1,1-tri-
chloroethane from aquatic systems.  Once in the troposphere, 1,1,1-trichlo-
roethane reacts with hydroxyl radicals.  The principal tropospheric photo-
oxidation product of 1,1,1-trichloroethane is reported to be trichloroace-
taldehyde which is probably subsequently oxidized to trichloroacetic acid.
This acid, as well as some unreacted 1,1,1-trichloroethane, could be washed
out of the troposphere via precipitation.  A portion of the unreacted
1,1,1-trichloroethane probably diffuses to the stratosphere where it under-
goes photoodissociation by higher energy ultraviolet light with the subse-
quent formation of chlorine atoms and chlorine oxides.

    Based on the information found, it does not appear that either oxida-
tion or hydrolysis is an important fate process of 1,1,1-trichloroethane in
the aquatic environment.  Evidence concerning the importance of adsorption,
bioaccumulation, and biodegradation is not definitive.

45.2  Identification

    1,1,1-Trichloroethane is known to be ubiquitous in the environment.
This compound has been detected in the atmosphere (Singh e_t al. 1978;
Pearson and McConnell 1975), in finished drinking water (Environmental
Protection Agency 1975a) in freshwater, rainwater, seawater, marine
sediments (Pearson and McConnell 1975), marine organisms (Pearson and
McConnell 1975;  Dickson and Riley 1976), and in surface snow in Alaska
(Su and Goldberg 1976).

    The chemical structure of 1,1,1-trichloroethane is shown below.


           CJ     H

           I     I
      Cl	 C 	c 	H               Alternate Names

                                      Methyl chloroform
           Cl     H                    Chlorotene
                                      Genklene
    1,1,1-Trichloroethane             Baltana

    CAS NO. 71-55-6
    TSL NO. KJ 29750
                                   45-1

-------
45.3  Physical Properties

    The general physical properties of 1,1,1-trichloroethane are given be-
low.
    Molecular weight
    (Weast 1977)

    Melting point
    (Weast 1977)

    Boiling point at 760 torr
    (Weast 1977)

    Vapor pressure at 20C
    (Pearson and McConnell 1975)

    Solubility in water at 20C
133.41
-30.41C
74.1C
96.0 torr
480 to 4400 mg/1*
    Log octanol/water partition coefficient  2.17
    (Calculated from Tute 1971, See Methods
    Section on bioaccumulation)
*Several values for solubility of 1,1,1-trichloroethane in water at 20C
were found in the literature.  These values range from 480 mg/1 (Pearson
and McConnell 1975) to 4400 mg/1 (Verschueren 1977).

45.4  Summary of Fate Data

    45.4.1  Photolysis

         There was little information found pertaining specifically to the
rate of photolysis of 1,1,1-trichloroethane in the aqueous environment.
Dilling et_ a^. (1975) studied the rate of photolysis of 1,1,1-trichloro-
ethane at a concentration of 1 mg/1 in aerated water in sealed tubes and
found no significant difference between samples exposed to sunlight and
those kept in the dark.   Based on these results photolysis of 1,1,1-tri-
chloroethane in the aqueous environment under ambient conditons does not
appear to be a significant fate process.

         Photodissociation in the water or the troposphere would not be ex-
pected to occur for 1,1,1-trichloroethane since this compound has no chro-
mophores which absorb in the visible or near ultraviolet region of the
electromagnetic spectrum (Jaffe and Orchin 1962).

         Due to the high vapor pressure of 1,1,1-trichloroethane, volatil-
ization to the atmosphere is quite rapid.  1,1,1-Trichloroethane has been
                                     45-2

-------
described as relatively stable in the troposphere (McConnell and Schiff
1978;  Hanst 1978;  Lillian _et _al. 1975).  However, it appears likely that
most of the material entering the troposphere will be photooxidized by re-
action with hydroxyl radicals, even though the rates reported are relative-
ly slow in comparison with those for structurally similar chlorinated ali-
phatics.  McConnell and Schiff (1978) report that approximately fifteen
percent of the 1,1,1-trichloroethane discharged to the troposphere will be
transported to the stratosphere where it will be rapidly photodissociated
due to the presence of shorter wavelength, higher energy ultraviolet light.
Chlorine atoms and chlorine oxides are reported to be the photodissociation
products.

    45.4.2  Oxidation

         The literature reviewed indicates that oxidation probably does not
play a significant role in the aquatic fate of 1,1,1-trichloroethane.  Bil-
ling _ejt aJL. (1975) conducted laboratory experiments in which there was  ab-
out a sixfold molar excess of dissolved oxygen compared to the total amount
of 1,1,1-trichloroethane.  Thus, there was a great excess of oxygen present
which would have been available for complete oxidation of all the chlori-
nated compounds present.  The results indicated that oxidation was very
slow and, if it occurred at all, was much slower than volatilization.

         Many tropospheric half-lives, lifetimes, and rate constants for
1,1,1-trichloroethane have been reported in the literature.  Yung et al.
(1975) have reported a bimolecular rate constant for 1,1,1-trichloroethane
of 8.2 x 10~^cm^molecule~-'-sec~-'-, corresponding to a maximum trop-
ospheric lifetime (time required for reduction to 1/e of original concen-
tration) of 9.5 x 10' seconds (approximately 3.1 years).  McConnell and
Schiff (1978) have reported a tropospheric lifetime of approximately 8
years based on a tropospheric hydroxyl radical concentration of
5 x 10^cm~3.  Watson et_ _aJ. (1977) report a tropospheric lifetime of
4.39 years based on a weighted average temperature for the troposphere of
265K and a global seasonally and diurnally averaged hydroxyl radical con-
centration of 9 x 10^cm~3.  in addition, Watson _e_t _aT. (1977) report a
bimolecular rate constant for reaction of hydroxyl radicals in the tropo-
sphere with 1,1, 1-trichloroethane of (1.59 + 0.16) x ICT^cm^sec"1 at
298K.  A tropospheric lifetime for 1,1,1-trichloroethane of 1.1 years has
been reported by Cox e_t al. (1976), based on a bimolecular rate of reaction
with hydroxyl radicals of 2.8 x 10~^cm^sec~^.  This value agrees
with the rate constant of 2.8xlO~-^cm3molecule~lsec~-'- reported by
Howard and Evenson (1976). A document by the Environmental Protection Ag-
ency (1975c) reports a tropospheric half-life of 1,1,1-trichloroethane of
1.1 years; the rate of reaction on which this half-life was based was not
given, but may well have been taken from Cox _et al. (1976) or Howard and
Evenson (1976).
                                    45-3

-------
         Photooxidation products for 1,1,1-trichloroethane have been
reported in the literature.   Christiansen  _et al.  (1972)  report the
photooxidation products of 1,1,1-trichloroethane  to be hydrogen chloride,
carbon oxides, phosgene and, to  a lesser extent,  acetyl  chloride.   Both
phosgene and acetyl chloride are readily hydrolyzed (Morrison and  Boyd
(1973).  On the other hand,  a document by  the Environmental Protection
Agency (I975c) reports that  the  principal  tropospheric photooxidation
product of 1,1,1-trichloroethane is trichloroacetaldehyde.  According to
the Environmental Protection Agency (1975c), it seems likely that
trichloroacetaldehyde will subsequently be oxidized to trichloroacetic
acid.  This acid, once formed, could be washed out of the troposphere by
precipitation (Environ-mental Protection Agency 1975c).

    45.4.3  Hydrolysis

         According to the Environmental Protection Agency (1975b)  1,1,1--
trichloroethane is sensitive to  hydrolysis.  In addition, it is reported
that 1,1,1-trichloroethane is easily hydrolyzed in an excess of free water
especially at elevated temperatures and in the gas phase (Environmental
Protection Agency 1975b).  In citing the works of other  authors, Billing _et
al. (1975) reported that 1,1,1-trichloroethane is hydrolyzed at elevated
temperatures with the formation of acetic  and hydrochloric acids along with
a minor amount of vinylidene chloride.  Laboratory experiments by Billing
^t al. (1975) were conducted in closed systems in the dark to eliminate
some of the potentially competing processes such as evaporation, photoly-
sis, and oxidation.  The experimental half-life obtained was approximately
six months at 25C, corresponding to a first-order rate of 0.12 months"^
(Dilling  _al. 1975).  This may be considered a maximum rate for hydroly-
sis, although other processes might have been occurring.

    45.4.4  Volatilization

         Billing _e_t _al. (1975) estimated the experimental half-life for
volatilization of 1,1,1-trichloroethane originally present at 1 mg/1  to be
20+;3 minutes  (average of three) when stirred at 200 rpm in water at ap-
proximately 25C in an open container.  Removal of 90 percent of the 1,1,1-
trichloroethane under the same conditions required about 69 minutes (aver-
age of three).  For chloroaliphatics in general, stirring speed was found
to have a marked effect on volatilization rate.  With intermittent stirring
for 15 seconds every  5 minutes, the time required for 50 percent depletion
of 1,1,1-trichloroethane was approximately 90 minutes, or on  the order of
four times greater than was observed during constant stirring.  The pres-
ence of sodium chloride at a 3 percent concentration, as in seawater,
caused about  a ten percent decrease in the chlorinated compound evaporation
rate.
                                     45-4

-------
         Evaporation appears to be the major pathway by which 1,1,1-tri-
chloroethane is lost from water.  Dilling _e_t _al_. (1975) are careful to
point out the difficulties encountered in extrapolating their laboratory
results to real-world conditions, where the concentration of the organic
solute would probably be very much less than 1 mg/1 and where surface and
bulk agitation would be highly variable.  Although the data appear to be
valid on a relative basis (i.e., correctly illustrating the relative rates
of volatilization of chlorinated aliphatics), they cannot be used as abso-
lute measures of volatilization rates from natural waters.  For the pur-
poses of this document, the data are used as rough-order-of-magnitude indi-
cations of the importance of volatilization relative to other transport and
fate processes, with the strong effects of agitation considered.  The
validity of this application has not been established.

         A subsequent study by Dilling (1977) was conducted using the same
experimental conditions as in the 1975 study, and an average evaporation
half-life of 24.9 minutes was obtained for 0.97 mg/1 1,1,1-trichloroethane.
The purpose of this subsequent study was to use the experimental data ob-
tained to test two theoretical models formulated to predict evaporation
rates of slightly soluble organic compounds from water.  Dilling (1977)
found that the theoretical model by Mackay and Wolkoff (1973) failed to
predict evaporation half-lives, whereas the theoretical model of Mackay and
Leinonen (1975), using parameters from Liss and Slater (1974), corre-
lated well with the experimental half-lives obtained by Dilling (1977).
For example, the evaporation half-life obtained experimentally by Dilling
(1977) for 1,1,1-trichloroethane was approximately 17 to 25 minutes, in
agreement with the evaporation half-life obtained by the Mackay and
Leinonen (1975) model of 23.7 minutes, but very diffent from the value of
0.19 minutes predicted by the Mackay and Wolkoff (1973) model.

         Dilling (1977), however, comments that the apparent numerical
agreement between his data and the values predicted by the Mackay and
Leinonen (1975) model may be fortuitous.  Estimates of volatilization rates
based on the Mackay and Leinonen (1975) model depend primarily on liquid-
gas phase exchange rate constants, whereas the experimental model of  Dil-
ling J2t _al. (1975) and Dilling (1977) is controlled by the rate of stirring
and the wind velocity across the surface of the water.

         In a study by Pearson and McConnell (1975 ), rainwater collected
in Runcorn, England contained up to 9 x 10~H parts (by mass) of 1,1,1-
trichloroethane.  Municipal waters supplied to the cities of Liverpool,
Chester, and Manchester (all supplied from upland surface sources), con-
tained up to 3 x 10~^ (by mass) tetrachloromethane plus 1,1, 1-trichlo-
roethane.  While entry to the general environment appears to be predomi-
nantly evaporative, there may also be some chloroaliphatics in aqueous
industrial effluents which will pass into municipal drainage and systems
                                     45-5

-------
and rivers.  Mackay and Wolkoff (1973) have shown that the evaporation rate
of poorly soluble species from water can be quite high due to their high
activity coefficients.  Pearson and McConnell (1975) and McConnell et al.
(1975) point out that this is a reversible process since water will absorb
chloro-organics from the atmosphere, a process which will occur most effec-
tively when the atmosphere is scrubbed during periods of rainfall.  The
presence of trace organochlorines,  including 1,1,1-trichloroethane, in up-
land waters is believed to be due to aerial transport (Pearson and McCon-
nell 1975;   McConnell _et al. 1975).

    45.4.5  Sorption

         Billing _e_t al. (1975) carried out two closed system experiments
where solute loss could only be by adsorption.  Dry bentonite clay at 375
mg/1 was introduced into a sealed solution and in ten minutes there was ap-
proximately a ten percent adsorption of 1,1,1-trichloroethane, as well as
other chlorinated compounds, by the clay compared to a blank.  When the
amount of clay added to the closed system was doubled (750 mg/1) there was
22 percent solute adsorption after 30 minutes.  There was no further solute
adsorption after this time.  The authors indicated that there appeared to
be little selectivity among the various chlorinated compounds in the ad-
sorption process.  The authors observed some adsorption of 1,1,1-trichlo-
roethane by dry powdered dolomitic limestone, but, again, no selectivity
among solutes was found.

         In a sealed system with approximately 500 mg/1 peat moss, approxi-
mately 40 percent of the 1,1,1-trichloroethane was adsorbed in 10 minutes.
At longer times, no further solute removal was noted.  The subsequent de-
crease in the rate of disappearance of 1,1,1-trichloroethane at time peri-
ods longer than 10 minutes was suggested to be due to gradual release or
desorption of 1,1,1-trichloroethane from the peat moss to the solution.

         Pearson and McConnell (1975) found that the concentrations of
chloroorganics in Liverpool Bay sediments varied from a few ng/1 to a max-
imum of-5500 ng/1 of 1,1,1-trichloroethane plus tetrachioromethane com-
bined.  Samples of marine sediment from Liverpool Bay contained the same
compounds as the overlying waters.  Due to the inability of  the analytical
technique used to resolve tetrachloromethane and 1,1,1-trichloroethane,
however, it could not be determined whether a correlation existed between
the concentration of  1,1,1-trichloroethane in marine sediments and the con-
centration in the interstitial water at the time the samples were taken
(Pearson and McConnell 1975;  McConnell t a.1. 1975).  In addition, it is
unknown whether a correlation existed between high concentrations and
geographical features, as was noticed in the case of water samples for
several compounds, or whether there was a relationship between the concen-
tration in the sediment and particle size.  McConnell et_ al_. (1975), how-
ever, did note that coarse gravels had little adsorptive capacity for
                                     45-6

-------
1,1,1-trichloroethane, whereas sediments rich in organic detritus had a
much higher adsorptive capacity.  In summary, there was no clear evidence
of selective concentration of 1,1,1-trichloroethane in sediments.

    45.4.6  Bioaccumulation

         According to Kopperman et  al. (1976) not all organochlorine com-
pounds bioaccumulate to high levels.  The data suggest that polar com-
pounds are more easily biodegraded, whereas non-polar highly lipophilic
compounds accumulate.  Neely et al. (1974) have shown that bioaccumulation
is directly related to the octanol/water partition coefficient (P) of the
compound.  The log octanol/water partition coefficient (log P) as deter-
mined by calculation using the method of Tute (1971) is 2.17 for 1,1,1-tri-
chloroethane, indicating that bioaccumulation in the adipose tissues of
organisms is possible, but that bioaccumulation is probably not an impor-
tant transport mechanism for this compound.

         Pearson and McConnell (1975) attempted to determine the level of
1,1,1-trichloroethane and other chlorinated hydrocarbons in the tissues of
a wide range of organisms.  Species were chosen to represent significant
trophic levels in the marine environment.  1,1,1-Trichloroethane was well
distributed, but due to difficulties with analytical methods, no reliable
estimates of bioaccumulation could be made.

    45.4.7  Biotransformation and Biodegradation

         No information was found pertaining specifically to the rate of
biodegradation of 1,1,1-trichloroethane in aquatic systems.  Thorn and Agg
(1975) have included 1,1,1-trichloroethane in a list of synthetic organic
chemicals which should be degradable by biological sewage treatment pro-
vided suitable acclimatization can be achieved.  They note, however, that
not many compounds in this list occur free in nature and, as a result, it
is  unlikely that microorganisms already possess the ability to destroy
them.
         According to Pearson and McConnell (1975) only completely sealed
systems, such as the standard BOD (biochemical oxygen demand) bottle tech-
nique, can be used to measure biochemical degradation of volatile compounds
such as 1,1,1-trichloroethane.  The BOD bottle experiments of Pearson and
McConnell have been unable to demonstrate any significant oxygen absorp-
tion from compounds containing only C, H, and Cl, thus indicating that
biochemical degradation of such compounds is indeed very slow.

45.5  Data Summary

    Table 45-1 summarizes the aquatic fate discussed above.  The oxidation
rate is a photooxidation rate and refers  to the rate of reaction of  1,1,1-
trichloroethane with hydroxyl radicals in the troposphere.
                                      45-7

-------
    Due to the high vapor pressure of 1,1,1-trichloroethane,  volatilization
to the atmosphere is quite rapid.   Although reaction with hydroxyl radi-
cals in the troposphere is relatively slow, this photooxidation reaction
appears to be the predominant fate process  for the compound.   The unreacted
portion of tropospheric 1,1,1-trichloroethane returns to the  earth via
precipitation or diffuses to  the stratosphere where it undergoes photodis-
sociation as a result of the  shorter wavelength, higher energy ultraviolet
light present above the ozone layer.  1,1,1-Trichloroethane reaching the
lithosphere via precipitation or dry fallout would be expected to evaporate
readily.
                                      45-8

-------
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45.6  Literature Cited

Christiansen, V.O., J.A. Dahlberg, and H.F. Andersson.  1972.  On the
  nonsensitized photo-oxidation of 1,1,1-trichloroethane vapor in air.
  Acta. Chem. Scand. Series A  26:3319-3324.

Cox, R.A., R.G. Derwent, A.E.J. Eggleton, and J.E.  Lovelock.  1976.
  Photochemical oxidation of halocarbons in the troposphere.  Atmos.
  Environ.  10:305-308.

Dickson, A.G. and J.P. Riley.  1976.   The distribution of short chain
  halogenated aliphatic hydrocarbons  in some marine organisms.  Mar.
  Pollut. Bull. 7(9):167-169.

Billing, W.L.  1977.  Interphase transfer processes.  II.  Evaporation
  rates of chloromethanes, ethanes, ethylenes, propanes, and propylenes
  from dilute aqueous solutions.  Comparisons with theoretical predictions.
  Environ. Sci. Technol.  11:405-409.

Billing, W.L., N.B. Tefertiller, and  G.J. Kallos.  1975.  Evaporation rates
  of methylene chloride, chloroform,  1,1,1-trichloroethane,
  trichloroethylene, tetrachloroethylene, and other chlorinated compounds
  in dilute aqueous solutions.  Environ. Sci. Technol. 9(9):833-838.

Environmental Protection Agency.  1975a.  Preliminary assessment of
  suspected carcinogens in drinking water.  Environmental Protection Agency
  (Office of Toxic Substances), Washington, B.C.  33pi  EPA 560/4-75-003.

Environmental Protection Agency.  1975b.  Preliminary study of selected
  potential environmental contaminants - optical brighteners, methyl
  chloroform, trichloroethylene, tetrachloroethylene, ion exchange re
   sins.  Enviornmental Protection Agency (Office of Toxic Substances),
  Washington, B.C. 286p.  EPA 560/2-75-002.

Environmental Protection Agency.  1975c.  Report on the problem of
  halogenated air pollutants and stratospheric ozone.  Environmental
  Protection Agency, (Office of Research and Development), Research
  Triangle Park, North Carolina.  55p.  EPA 600/9-75-008.

Hanst, P.L.  1978.  Part II:  Halogenated pollutants.  Noxious trace gases
  in the air.  Chemistry 51(2):6-12.

Howard, C.J. and K.M. Evenson.  1976.  Rate constants for the reactions of
  OH with ethane and some halogen substituted ethanes at 296 K.  J. Chem.
  Phys. 64(11):4303-4306.

Jaffe, H.H. and M. Orchin.  1962.  Theory and applications of ultraviolet
  spectroscopy.  John Wiley and Sons, Inc. New York.  624p.
                                     45-10

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Kopperman, H.L., D.W. Kuehl, and G.E. Glass.  1976.   Chlorinated compounds
  found in waste treatment effluents and their capacity to bioaccumulate.
  Proceedings of the conference on the environmental impact of water
  chlorination.  pp.327-345 Oak Ridge, Tennessee,  October 22-24, 1975.

Lillian, D., H.B. Singh, A. Appleby, L. Lobban, R. Arnts, R. Gumpert, R.
  Hague, J. Toomey, J. Kazazis, M. Antell, D. Hansen, and B. Scott.   1975.
  Atmospheric fates of halogenated compounds.  Environ. Sci. Technol.
  9:1042-1048.

Liss, P.S. and P.G. Slater.  1974.  Flux of gases  across the air-sea
  interface.  Nature 247:181-184.

Mackay, D. And P.J. Leinonen.  1975.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. Technol.
  9:1178-1180.

Mackay, D. and A.W. Wolkoff.  1973.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. Technol.
  7:611-614.

McConnell, G., D.M. Ferguson, and C.R. Pearson.  1975.  Chlorinated
  hydrocarbons and the environment.  Endeavor XXXIV:13-18.

Morrison, R.T. and R.N. Boyd.  1973.  Organic chemistry.  3rd Edition.
  Allyn and Bacon, Inc., Boston, Mass. 1258p.

McConnell, J.C. and H.I. Schiff.  1978.  Methyl chloroform:  Impact  on
  stratospheric ozone.  Science  199:174-177.

Neely, W.B., D.R. Branson, and G.E. Blau.  1974.  Partition coefficient  to
  measure bioconcentration potential of organic chemicals in fish.
  Environ. Sci. Technol. 8:1113-1115.

Pearson, C.R. and G. McConnell.  1975.  Chlorinated  GI and G
  hydrocarbons in the marine environment.  Proc. Roy. Soc. London B
  189:305-322.

Singh, H.B., L.J. Salas, H. Shiegeishi, and A.H. Smith.  1978.  Fate of
  halogenated compounds in the atmosphere.  Interim report-1977.
  Enviornmental Protection Agency (Office of Research and Development),
  Research Triangle Park, North Carolina.  57p.  EPA   600/3-78-017.

Su, C. and E.D. Goldberg.  1976.  Environmental concentrations and  fluxes
  of some halocarbons.  in Marine pollutant transfer.  pp.353-374.   H.L.
  Windoin and R.A. Duce (eds.).  Lexington Books, B.C. Heath and Company,
  Lexington, Massachusetts.
                                     45-11

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Thorn, N.S. and A.R. Agg.   1975.   The breakdown of synthetic organic
  compounds in biological processes.  Proc.  Roy.  Soc.  London B
  189:347-357.

Tute, M.S.  1971.  Principles and practice of Hansch analysis:  A guide
  to structure-activity correlation for the  medicinal  chemist.  Adv.  Drug
  Res. 6:1-77.

Verschueren, K.  1977.  Handbook of environmental data on organic
  chemicals.  Van Nostrand/Reinhold Press, New York.  659p.

Watson, R.T., G. Machado, B.  Conaway, S. Wagner,  and D.D. Davis.  1977.  A
  temperature dependent kinetics study of the reaction of OH with CH2C1F,
  CHC12F, CHC1F2, CH3CCl3, CH3CF2C1, and CF2C1CFC12.  J.
  Phys. Chem.  81:256-262.

Weast, R.C. (ed.).  1977.  Handbook of chemisry and physics.  58th
  Edition.  CRC Press, Inc.,  Cleveland, Ohio. 2398p.

Yung, Y.L., M.B. McElroy, and S.C. Wofsy.  1975.   Atmospheric halocarbons:
  A discussion with emphasis  on chloroform.   Geophys.  Res. Lett.
  2(9):397-399.
                                   45-12

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                        46.  1.1,2-TRICHLOROETHANE


46.1  Statement of Probable Fate

    Although only a relatively small amount of information pertaining
specifically to 1,1,2-trichloroethane was found, the processes important to
the fate of this compound in the environment can be inferred from informa-
tion available on its more thoroughly studied structural isomer,  1,1,1-tri-
chloroethane, and from available information on short-chain halogenated
aliphatics in general.  By analogy to 1,1,1-trichloroethane, photooxidation
in the troposphere would be expected to be  the predominant fate process for
1,1,2-trichloroethane.  This, in turn, is contingent upon volatilization as
the primary transport mechanism for removal of the compound from  water.

46.2  Identification

    1,1,2-Trichloroethane has been detected in finished drinking  water by
the National Organics Reconnaissance Survey (Environmental Protection
Agency 1975).

    The chemical structure of 1,1,2-trichloroethane is shown below.

          Cl     Cl                    Alternate Names

                  	H               Vinyl  trichloride


          H     H

    1,1,2-Trichloroethane

    CAS NO. 79-00-5
    TSL NO.  KJ 31500

46.3  Physical Properties

    The general physical properties of 1,1,2-trichloroethane are  as  fol-
lows .

    Molecular weight                         133.41
    (Weast 1977)

    Melting point                            -36.5C
    (Weast 1977)

    Boiling point at 760 torr                133.77C
    (Weast 1977)
                                    46-1

-------
    Vapor pressure at 20C                   19 torr
    (Verschueren 1977)

    Solubility in water at 20C              4,500 mg/1
    (Verschueren 1977)

    Log octanol/water partition coefficient  2.17
    (Calculated from Tute (1971); see Methods
    Section on Bioaccumulation).

46.4  Summary of Fate Data

    46.4.1  Photolysis

         Relatively little information was found pertaining specifically to
the rate of photolysis of 1,1,2-trichloroethane in the aquatic environment.
Jensen and Rosenberg (1975), using a closed system wherein no evaporation
could occur, found no significant difference between 1,1,2-trichloroethane
in the closed system in daylight and 1,1,2-trichloroethane in the closed
system in darkness, thus indicating that little or no photodissociation
occurred.  Both seawater and deionized and boiled fresh water were utilized
in these tests.  Dilling _e_t _al. (1975) reported that, within experimental
error, sunlight did not accelerate decomposition of 1,1,1-trichloroethane,
a structural isomer of 1,1,2-trichloroethane, initially present at a concen-
tration of 1 mg/1 in aerated water.

         Photodissociation in the terrestrial environment  would not be ex-
pected to occur for 1,1,2-trichloroethane since this compound has no chro-
mophores which absorb in the visible or near ultraviolet region of the
electromagnetic spectrum (Jaffa and Orchin 1962).  In the  stratosphere,
however, 1,1,2-trichloroethane would be expected to be photodissociated due
to the presence of shorter wavelength, higher energy ultraviolet light.

    46.4.2  Oxidation

         No information pertaining specifically to the oxidation of 1,1,2-
trichloroethane in the aquatic environment was found.  There is, however,
some indirect evidence that oxidation in aquatic systems is not important.
For instance, Dilling et al. (1975) conducted aqueous reactivity experi-
ments in which there was a large excess of dissolved oxygen compared to the
total amount of 1,1,1-trichloroethane.  The results indicated; that volati-
lization was much more rapid than any oxidative process which may have
occurred.

         Once in  the  troposphere, 1,1,2-trichloroethane would be expected
to react with hydroxyl radicals.  This photooxidation reaction should
                                      46-2

-------
proceed at a rate intermediate between those for 1,2-dichloroethane and
1,1,1-trichloroethane, two structurally similar compounds, because of the
comparative number and placement of the chlorine atoms on the three
compounds.  Based on the data of Howard and Evenson (1976) the tropospheric
lifetime (time for reduction to 1/e of original concentration), of
1,2-dichloroethane, as a result of reaction with hydroxyl radicals, is on
the order of one to several months, while the troposphere half-life for
1,1,1-trichloroethane is reported to be on the order of a few years
(McConnell and Schiff 1978;  Hanst 1978;  Lillian _e_t al. 1975).

         Assuming an upper limit on the tropospheric lifetime for 1,1,2-
trichloroethane of three years, and a troposphere-to-stratosphere turnover
time (time for all but 1/e of tropospheric air to diffuse into the strato-
sphere) of 30 years, less than ten percent of the 1,1,2-trichloroethane
would be expected to reach the stratosphere.  It thus appears that photo-
oxidation in the troposphere is much more important than stratospheric
photodissociation for this compound.

    46.4.3  Hydrolysis

         No information pertaining specifically to the rate of hydrolysis
of 1,1,2-trichloroethane in the environment was found.  Billing et al.
(1975) conducted laboratory experiments in closed systems in the dark to
eliminate competing processes such as evaporation and photolysis.  The
experimental hydrolytic half-life obtained for 1,1,1-trichloroethane, a
structural isoiner of 1,1,2-trichloroethane, was approximately six months at
25C (Billing et al. 1975).  This corresponds to a first-order rate of 0.12
months"-'- (Dilling e_t al. 1975) which should be considered a maximum rate
for hydrolysis.  By analogy 1,1,2-trichloroethane may be expected to under-
go hydrolysis at a rate on the same order, that is, about 0.1 month" .

    46.4.4  Volatilization

         Dilling e_t _al_. (1975) estimated the experimental half-life for
volatilization of 1,1,2-trichloroethane originally present at 1 mg/1 to be
21 minutes when stirred at 200 rpm in water at approximately 25C in an
open container.  Removal of 90 percent of the 1,1,2-trichloroethane under
the same conditions required 102 minutes.  For chloroaliphatics in general,
stirring speed was found to have a marked effect on volatilization rate.
When intermittent stirring of 15 seconds duration was provided every five
minutes, the time required for 50 percent depletion of 1,1,1-trichloro-
ethane, a structural isomer of 1,1,2-trichloroethane, was greater than 90
minutes.  This half-life is on the order of four times longer than for
evaporation with constant stirring.  The presence of sodium chloride at
                                       46-3

-------
a 3 percent concentration, as in seawater, caused a 10 percent decrease in
the rate of evaporation.

         Billing e_t al. (1975) are careful to point out the difficulties
encountered in extrapolating their laboratory results to real-world con-
ditions, where the concentration of the organic solute would probably be
very much less than 1 mg/1 and where surface and bulk agitation would be
highly variable.  Although the data appear to be valid on a relative basis
(i.e., correctly illustrating the relative rates of volatilization of
chlorinated aliphatics), they cannot be used as absolute measures of vo-
latilization rates from natural waters.  For the purposes of this docu-
ment, the data are used as rough-order-of-magnitude indications of the
importance of volatilization relative to other transport and fate proceses,
with the strong effects of agitation considered.  The validity of this
application has not been established.

         A subsequent study by Billing (1977) was conducted using the same
experimental conditions as in the 1975 study, and an average evaportion
half-life of 35.1 minutes was obtained for 1,1,2-trichloroethane initially
present at a concentration of 0.99 mg/1.  The purpose of this subsequent
study was to use the experimental data previously obtained to test two
theoretical models which may be used to predict evaportive rates of
slightly soluble organic compounds from water.  Billing (1977) found that
the theoretical model of Mackay and Wolkoff  (1973) failed to predict
evaporative half-lives, but that the model of Mackay and Leinonen (1975)
using the parameters of Liss and Slater (1974) correlated well with the
experimental half-lives obtained. For example, the evaporative half-life
obtained experimentally by Billing (1977) was approximately 35 minutes,
compared to 30.1 minutes predicted by the Mackay and Leinonen (1975) model
and 6.1 minutes predicted by the Mackay and Wolkoff (1973) model.

         Billing (1977), however, comments that the apparent, numerical
agreement between his data and the values predicted by the Mackay and
Leinonen (1975) model may be fortuitous.  Estimates of volatilization rates
based on the Mackay and Leinonen (1975) model depend primarily on liquid-
gas phase exchange rate constants, whereas the experimental model of
Billing _e_t _al. (1975) and Billing (1977) is controlled by the rate of
stirring and the wind velocity across the surface of the water.

         Pearson and McConnell (1975) suggest that the presence of com-
pounds  similar to 1,1,2-trichloroethane (1,1,1-trichloroethane and others)
in ambient waters is due to the absorption of these compounds from the
atmosphere by water, particularly during precipitation.  Aerial transport
of halogenated aliphatics in general is indicated by Pearson and McConnell
(1975)  to play a major  role in the wider distribution of these compounds
and accounts for their  presence in upland waters.
                                     46-4

-------
    46.4.5  Sorption

         No information was found pertaining specifically to the adsorption
of 1,1,2-trichloroethane onto sediments.  Dilling et_ _al. (1975) carried out
two closed system experiments on approximately 1 mg/1 of 1,1,1-trichloro-
ethane (a structural isomer of 1,1,2-trichloroethane) where solute loss
could only be by adsorption.  Dry bentonite clay at 375 mg/1 was introduced
into a sealed solution and in ten minutes there was approximately a ten
percent adsorption of the solute by the clay.  When the amount of clay
added to the closed system was doubled (750 mg/1) there was 22 percent
solute adsorption after 30 minutes but no further solute adsorption after
this time.  The authors indicated that there appeared to be little selec-
tivity among the various chlorinated compounds in the adsorption process.
The authors observed some adsorption of 1,1,1-trichloroethane by dry
powdered dolomitic limestone, but, again, without any selectivity among the
compounds used.

         In the sealed system with approximately 500 mg/1 peat moss,
approximately 40 percent of 1,1,1-trichloroethane was adsorbed in 10
minutes. At longer times, no further solute removal was noted.  The sub-
sequent decrease in the rate of disappearance of 1,1,1-trichloroethane at
longer time periods than 10 minutes was suggested to be due to gradual
release of 1,1,1-trichloroethane from the peat moss to the solution.

         Pearson and McConnell  (1975) found no clear evidence of selective
concentrations of 1,1,1-trichloroethane in sediments; because of structural
similarities, 1,1,2-trichloroethane should behave similarly.

    46.4.6  Bioaccumulation

         According to Kopperman jst al. (1976) not all organochlorine com-
pounds bioaccumulate to high levels.  The data suggest that polar compounds
are more easily biodegraded, and the non-polar (highly lipophilic) com-
pounds accumulate.  Neely _e_t _al. (1974) have shown that bioaccumulation is
related to the octanol/water partition coefficient (P) of the compound.
The log octanol/water partition coefficient (log P) of 2.17 as calculated
by the method of Hansch (Tute 1971) indicates that 1,1,2-trichloroethane
will probably not bioaccumulate to any significant extent (See Methods sec-
tion on Bioaccumulation).

    46.4.7  Biotransformation and Biodegradation

         No information pertaining specifically to the rate of biodegrada-
tion of 1,1,2-trichloroethane in aquatic systems was found. Pearson and
McConnell (1975) stated that only completely sealed systems, such as the
                                     46-5

-------
standard BOD (biochemical oxygen demand) bottle technique, can be used to
measure biochemical degradation of volatile compounds such as 1,1,2-
trichloromethane.  The BOD bottle experiments of Pearson and McConnell have
been unable to demonstrate any significant oxygen absorption from compounds
containing only C, H, and Cl,  thus indicating that biochemical degradation
of such compounds is indeed very slow.   Literature references to microbial
biodegradation of low molecular weight  chloroaliphatics are few and con-
flicting;  the majority find that these compounds are not metabolized
(Pearson and McConnell 1975;  McConnell_et ad. 1975).

46.5  Data Summary

    Table 46-1 summarizes the  aquatic fate information discussed above.
The information found pertaining specifically to 1,1,2-trichloroethane is
relatively scant.  Based on the specific information found and information
inferred by analogy to 1,1,1-trichloroethane, one can predict with low con-
fidence that the predominant fate of 1,1,2-trichloroethane is photooxida-
tion in the troposphere.  This fate is, of course, contingent upon volatil-
ization as the dominant transport process for removal of 1,1,2-trichloro-
ethane from aquatic systems.
                                       46-6

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46.6  Literature Cited

Dilling, W.L.  1977.  Interphase transfer process.  II.   Evaporation rates
  of chloromethanes, ethanes,  ethylenes,  propanes,  and  propylenes from
  dilute aqueous solutions.  Comparisons  with theoretical predictions.
  Environ. Sci. Technol.  11:405-409.

Billing, W.L., N.B. Tefertiller, and G.J. Kallos.   1975.   Evaporation rates
  of methylene chloride, chloroform, 1,1,1-trichloroethane,
  trichloroethylene, tetrachloroethylene, and other chlorinated compounds
  in dilute aqueous solutions.  Environ.  Sci. Technol.   9(9):833-838.

Environmental Protection-Agency.  1975.   Preliminary assessment of
  suspected carcinogens in drinking water.  U.S.  Environmental Protection
  Agency, (Office of Toxic Substances),  Washington, B.C.   33p.  EPA
  560/4-75-003.
Hanst, P.L.  1978.  Noxious trace gases in the air,  Part II:
  pollutants.  Chemistry 51(2):6-12.
                                  halogenated
Howard, C.J. and K.M. Evenson.  1976.   Rate constants for the reactions of
  OH with ethane and some halogen substituted ethanes at 296K.  J. Chem.
  Phys.  64(11):4303-4306.
Jaffa, H.H. and M. Orchin.
  ultraviolet spectroscopy.
1962.   Theory and applications  of
 John Wiley and Sons,  Inc.   New York.   624p.
Jensen, S. and R. Rosenberg.  1975.  Degradability of some chlorinated
  aliphatic hydrocarbons in sea water and sterilized water.  Water Res.
  9:659-661.

Kopperman, H.L., D.W. Kuehl, and G.E. Glass.  1976.  Chlorinated compounds
  found in waste treatment effluents and their capacity to bioaccumulate.
  Proceedings of the conference on the environmental impact of water
  chlorination.  pp.327-345.  Oak Ridge, Tennessee, October 22-24, 1975.

Lillian, D., H.B. Singh, A. Appleby, L. Lobban, R. Arnts, R. Gumpert, R.
  Hague, J. Toomey, J. Kazazis, M. Antell, D. Hansen, and B. Scott.  1975.
  Atmospheric fates of halogenated compounds.  Environ. Sci. Technol.
  9:1042-1048.

Liss, P.S. and P.G. Slater.  1974.  Flux of gases across the air-sea
  interface.  Nature 247:181-184.
Mackay, D. and P.J. Leinonen.  1975.  Rate of evaporation of
  low-solubility contaminants from water bodies to atmosphere.
  Sci. Technol. 9:1178-1180.
                                    Environ.
                                    46-8

-------
Mackay, D. and A.W. Wolkoff.  1973.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. Technol.
  7:611-614.

McConnell, G., D.M. Ferguson, and C.R. Pearson.  1975.  Chlorinated
  hydrocarbons and the environment.  Endeavor XXXIV:13-18.

McConnell, J.C. and H.I. Schiff.  1978.  Methyl chloroform: impact on
  stratospheric ozone.  Science 199:174-177.

Neely, W.G., D.R. Branson, and G.E. Blau.  1974.  Partition coefficient to
  measure bioconcentration potential of organic chemicals in fish.
  Environ. Sci. Technol. 8:1113-1115.

Pearson, C.R. and G. McConnell.  1975.  Chlorinated C^ and C2
  hydrocarbons in the marine environment.  Proc. Roy. Soc. London B
  189:305-322.

Tute, M.S.  1971.  Principles and practice of Hansch analysis:  A guide to
  structure-activity correlation for the medicinal chemist.  Adv. Drug
  Res. 6:1-77.

Verschueren, K.  1977.  Handbook of environmental data on organic
  chemicals.  Van Nostrand/Reinhold Press, New York.  659p.

Weast, R.C. (ed.).  1977.  Handbook of chemistry and physics.  58th
  Edition. CRC Press, Inc., Cleveland, Ohio.  2398p.
                                    46-9

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                        47.  1,1,2,2-TETRACHLOROETHANE


47.1  Statement of Probable Fate

    Little information pertaining specifically to the environmental fate of
1,1,2,2-tetrachloroethane was found.  By analogy to a more thoroughly
studied analogue, 1,1,1-trichloroethane, both photooxidation in the tropo-
sphere and photodissociation in the stratosphere may be important fate
processes for 1,1,2,2-tetrachloroethane.  The importance of these processes
in comparison to fate processes in aqueous media depends, in turn, on
volatilization as the major process for removal from aquatic systems.  Al-
though volatilization appears to be relatively rapid, essentially no in-
formation was found concerning possibly competing processes.

47.2  Identification

    1,1,2,2-Tetrachloroethane has been detected in finished drinking water
(Shackelford and Keith 1976).

    The chemical structure of 1,1,2,2-tetrachloroethane is shown below.
             Cl     Cl
                                             Alternate Names
        H - C - C - H
                                             sym-Tetrachloroe thane
                   Cl                         Acetylene tetrachloride
     1,1,2, 2-Tetrachloroe thane

     CAS NO. 79-34-5
     TSL NO. KI 85750

47.3  Physical Properties

    The general physical properties of 1 ,1,2, 2-tetrachloroethane are as
follows .

    Molecular weight                        167.85
    (Weast 1977)

    Melting point                            -36C
    (Weast 1977)

    Boiling point at 760 torr               146. 2C
    (Weast 1977)
                                    47-1

-------
    Vapor pressure at 20C                   5 torr
    (Verschueren 1977)

    Solubility in water at 20C              2,900 mg/1
    (Verschueren 1977)

    Log octanol/water partition              2.56
    coefficient (Calculated from Tute 1971;
    see Methods section on Bioaccumulation)

47.4  Summary of Fate Data

    47.4.1  Photolysis  -

         Relatively little information pertaining specifically to the rate
of photolysis of 1,1,2,2-tetrachloroethane in the aquatic environment was
found.  Jensen and Rosenberg (1975), using a closed system in which no
evaporation could occur, found no significant difference between 1,1,2,2-
tetrachloroethane depletion in the closed system in daylight and in dark-
ness, thus indicating that no significant photolysis occurred.  Billing et
al. (1975) reported that,  within experimental error, sunlight: did not
accelerate decomposition of 1,1;l-trichloroethane, an analogue of 1,1,2,2-
tetrachloroethane , initially present at 1 mg/1 concentrations in aerated
water.  It is thus not possible to determine whether photolysis of either
1,1,1,-trichloroethane or  1,1,2,2-tetrachloroethane would occur in water at
a significant rate.  Moreover, photodissociation in the aquatic environment
would not be expected to occur since 1,1,2,2-tetrachloroethane has no chro-
mophores which absorb in the visible or near ultraviolet region of the
electromagnetic spectrum (Jaffe and Orchin 1962).  In the stratosphere,
however, 1,1,2,2-tetrachloroethane would be  expected to react analogously
to 1,1,1-trichloroethane and be photodissociated by high energy ultraviolet
light.

    47.4.2  Oxidation

         No information pertaining specifically to the oxidation of
1,1,2,2-tetrachloroethane  in the aquatic environment was found.  There is,
however, some indirect evidence for the lack of a potential for oxidation
of this compound in aquatic systems.  Billing _et _al. (1975) conducted ex-
periments in aqueous solutions where there was a large excess of dissolved
oxygen compared to the total amount of 1,1,1-trichloroethane an analogue
of 1,1,2,2-tetrachloroethane.  The results indicated that the 1,1,1-tri-
chloroethane was relatively stable under the oxidative conditions used,
exhibiting a combined half-life with respect to both oxidation and hy-
drolysis of 6 months.  Billing _e_t al. (1975) attribute the decomposition
observed to hydrolysis rather than oxidative processes.
                                     47-2

-------
         Once in the troposphere 1,1,2,2-tetrachloroethane would be ex-
pected to undergo a photooxidation reaction with hydroxyl radicals.  The
tropospheric half-life of 1,1,1-trichloroethane, a structurally similar
compound, is reported to be on the order of a few years (McConnell and
Schiff 1978;  Hanst 1978;  Lillian et al. 1975), and McConnell and Schiff
(1978) estimate that 15 percent of tropospheric 1,1,1-trichloroethane may
reach the stratosphere where it can be photodissociated.  If the rate of
tropospheric photooxidation for 1,1,2,2-tetrachloroethane is similar to
that of 1,1,1-trichloroehane, then it appears that both tropospheric photo-
oxidation and photodissociation in the stratosphere are probably impor-
tant fate processes for 1,1,2,2-tetrachloroethane.  Because of the lack of
data specific to 1,1,2,2-tetrachloroethane, however, the relative impor-
tance of these two processes cannot be determined.

    47.4.3  Hydrolysis

         No information pertaining specifically to the rate of hydrolysis
of 1,1,2,2-tetrachloroethane in the environment was found.  In the re-
activity experiments of Billing et al. (1975) in which light was excluded
from closed systems, the experimental half-life obtained for 1,1,1-tri-
chloroethane , an analogue of 1,1,2,2-tetrachloroethane, was approximately
six months at 25C.  The decomposition was primarily attributed to hydroly-
sis.  This half-life corresponds to a first-order rate of 0.12 months"*
(Billing _e_t _al. 1975), which is probably a maximum rate with respect to
hydrolysis under environmental conditions.  Analogously, 1,1,2,2-tetra-
chloroethane may be expected to undergo hydrolysis at a rate corresponding
to a half-life of several months to a few years.

    47.4.4  Volatilization

         Billing _ejt al. (1975) estimated the experimental half-life for
volatilization of 1,1,2,2-tetrachloroethane originally present at 1 mg/1 to
be 56 minutes when stirred at 200 rpm in water at approximately 25C in an
open container.  Removal of 90 percent of the compound under the same con-
ditions required over 120 minutes.  For chloroaliphatics in general,
stirring speed was found to have a marked effect on volatilization rate.
When intermittent stirring of 15 seconds duration was provided every five
minutes, the time required for 50 percent depletion of 1,1,1-trichloro-
ethane, an analogue of 1,1,2,2-tetrachloroethane, was greater than 90
minutes, which was four times greater than evaporation with constant
stirring for this compound. The presence of sodium chloride at a 3 percent
concentration, as in seawater, caused a decrease in the rate of evaporation
of short-chain chloroaliphatics.  Billing _e_t _al. (1975) are careful to
point out the difficulties encountered in extrapolating their laboratory
results to real-world conditions, where  the concentration of the organic
                                     47-3

-------
solute would probably be very much less than 1 mg/1 and where surface and
bulk agitation would be highly variable.   Although the data appear to be  .
valid on a relative basis (i.e., correctly illustrating the relative rates
of volatilization of chlorinated aliphatics),  they cannot be used as
absolute measures of volatilization rates from natural waters.  For the
purposes of this document, the data are used as rough-order-of-magnitude in-
dications of the importance of volatilization relative to other transport
and fate proceses, with the strong effects of agitation considered.  The
validity of this application has not been established.

         A subsequent study by Billing (1977)  was conducted using the same
general experimental conditions as in the previous investigation, and an
average half-life of 55.2 minutes was obtained for an initial concentration
of 0.92 mg/1 1,1,2,2-tetrachloroethane in water.  The purpose of this sub-
sequent study was to use the experimental data previously obtained to test
two theoretical models which may be used to predict evaporative rates of
slightly soluble organic compounds from water.  Billing (1977) found that
the theoretical model of Mackay and Wolkoff (1973) failed to predict evap-
orative half-lives, whereas the theoretical model of Mackay and Leinonen
(1975) using the parameters of Liss and Slater (1974) correlated well with
the experimental half-lives obtained by Billing (1977).  For example, the
evaporative half-life obtained experimentally by Billing (1977) was
approximately 55 minutes, similar to a 40.5 minute evaporative half-life
predicted by the Mackay and Leinonen (1975) model, but very different from
the value of 12 minutes predicted by the Mackay and Wolkoff (1973) model.

         Billing (1977), however, comments that the apparent numerical
agreement between his data and the values predicted by the Mackay and
Leinonen (1975) model may be fortuitous.  Estimates of volatilization rates
based on the Mackay and Leinonen (1975) model depend primarily on liquid-
gas phase exchange rate constants, whereas the experimental model of
Billing jet al. (1975) and Billing (1977) is controlled by the rate of
stirring and the wind velocity across the surface of the water.

         Pearson and McConnell (1975) suggest that the presence of 1,1,1-
trichloroethane (an analogue of 1,1,2,2-tetrachloroethane) and other halo-
genated aliphatics in ambient waters is due to the absorption of these com-
pounds from the atmosphere by water and return to the earth in precipita-
tion.  Aerial transport of halogenated aliphatics is indicated by Pearson
and McConnell (1975) to play a major role in the wider distribution of
these compounds and accounts for their presence in upland waters.

    47.4.5  Sorption

         No information pertaining specifically to the adsorption of
1,1,2,2-tetrachloroethane onto sediments was found.  Billing et jal. (1975)
                                    47-4

-------
carried out two closed system experiments on 1,1,1-trichloroethane,  an
analogue of 1,1,2,2-tetrachloroethane, wherein solute loss could occur only
by adsorption.   Dry bentonite clay at 375 mg/1 was introduced into a sealed
solution, and in ten minutes approximately ten percent of solute was ad-
sorbed by the clay.  When the amount of clay added to the closed system was
doubled (750 mg/1), there occurred 22 percent solute adsorption after 30
minutes, and no further solute adsorption after this time.  The authors
indicated that there appeared to be little selectivity among the various
chlorinated compounds in the adsorption process.1 The authors also observed
some adsorption of 1,1,1-trichloroethane by dry powdered dolomitic lime-
stone, but, again, without any selectivity among the chloroaliphatics being
studied.  In the sealed system with approximately 500 mg/1 peat moss, ap-
proximately 40 percent of 1,1,1-trichloroethane as a solute was adsorbed in
10 minutes.  No further solute removal was noted upon further sampling.

         Pearson and McConnell (1975) found no clear evidence of selective
concentration of 1,1,1-trichloroethane in sediments.  It is anticipated
that observations under the same conditions for 1,1,2,2-tetrachloroethane
would be similar to those reported above for 1,1,1-trichloroethane.   How-
ever, these data are inconclusive as to the importance of sorption in the
aquatic transport of 1,1,2,2-tetrachloroethane.

    47.4.6  Bioaccumulation

         According to Kopperman _e_t _al. (1976) not all organochlorine com-
pounds bioaccumulate to high levels.  The data suggest that polar compounds
are more easily biodegraded, and the non-polar (highly lipophilic) com-
pounds accumulate.  Neely et al. (1974) have shown that bioaccumulation is
related to the octanol/water partition coefficient (P) of the compound.
The log octanol/water partition coefficient (log P) of 2.56 calculated by
the method of Hansch (Tute 1971) indicates that bioaccumulation of
1,1,2,2-tetrachloroethane is possible (see Methods section on bioaccumula-
tion), but its importance cannot be assessed at this time.

    47.4.7  Biotransformation and Biodegradation

         No information pertaining specifically to the rate of biodegrada-
tion of 1,1,2,2-tetrachloroethane in aquatic systems was found. According
to Pearson and McConnell (1975), only completely sealed systems, such as
the standard BOD (Biochemical Oxygen Demand) bottle technique, can be used
to measure biochemical degradation of volatile compounds such as 1,1,2,2-
tetrachloroethane.  The BOD bottle experiments of Pearson and McConnell
have been unable to demonstrate any significant oxygen absorption when com-
pounds that contained only C, H, and Cl were being evaluated, thus indi-
cating that biochemical degradation of such compounds is indeed very slow.
Literature references to microbial biodegradation are few and conflicting;
the majority find that low molecular weight chloroalilphatics are not
metabolized (Pearson and McConnell 1975;  McConnell et al. 1975).
                                      47-5

-------
47.5  Data Summary

    Table 47-1 summarizes the aquatic fate information discussed above.
The information found pertaining specifically to 1,1,2,2-tetrachloroethane
is relatively scant.  Based on the specific information found and informa-
tion obtained by analogy to 1,1,1-trichloroethane and other halogenated
aliphatic compounds, one can predict (with low confidence) that volatili-
zation is the major transport process for removal from aquatic systems and
that photooxidation in the troposphere and photodissociation in the
stratosphere are each important fate processes.  The predominante fate
process cannot be predicted from the information presented.
                                     47-6

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                                       47-7

-------
47.6  Literature Cited

Dilling, W.L.  1977.  Interphase transfer processes.  II.  Evaporation
  rates of chlororaethanes, ethanes, ethylenes, propanes, and propylenes
  from dilute aqueous solutions.  Comparisons with theoretical
  predictions.  Environ. Sci. Technol. 11:405-409.

Dilling, W.L., N.B. Tefertiller, and G.J. Kallos.  1975.  Evaporation rates
  of methylene chloride, chloroform, 1,1,1-trichloroethane,
  trichloroethylene, tetrachloroethylene, and other chlorinated compounds
  in dilute aqueous solutions.  Environ. Sci. Technol. 9(9):833-838.

Hanst, P.L.  1978.  Noxious trace gases in the air. Part II:  Halogenated
  pollutants.  Chemistry 51(2):6-12.

Jaffe, H.H. and M. Orchin.  1962.  Theory and applications of
  ultraviolet spectroscopy.  John Wiley and Sons, Inc. New York.  624p.

Jensen, S. and R. Rosenberg.  1975.  Degradability of some chlorinated
  aliphatic hydrocarbons in seawater and sterilized water.  Water Res.
  9:659-661.

Kopperman, H.L., D.W. Kuehl, and G.E. Glass.  1976.  Chlorinated compounds
  found in waste  treatment effluents and their capacity to bioaccumulate.
  Proceedings of the conference on the environmental impact of water
  chlorination.  pp.327-345.  Oak Ridge, Tennessee, October 22-24, 1975.

Lillian, D., H.B. Singh, A. Appleby, L. Lobban, R. Arnts, R. Gurapert, R.
  Hague, J. Toomey, J. Kazazis, M. Antell, D. Hansen, and B. Scott.   1975.
  Atmospheric fates of halogenated compounds.  Environ. Sci. Technol.
  9:1042-1048.

Liss, P.S. and P.G. Slater.  1974.  Flux of gases across the air-sea
  interface.  Nature 247:181-184.

Mackay, D. and P.J. Leinonen.  1975.  Rate of evaporation of
  low-solubility contaminants from water bodies to atmosphere.  Environ.
  Sci. Technol.   9:1178-1180.

Mackay, D. and A.W. Wolkoff.  1973.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. Technol.
  7:611-614.

McConnell, G., D.M. Ferguson, and C.R. Pearson.  1975.  Chlorinated
  hydrocarbons and  the environment.  Endeavor XXXIV:13-18.
                                    47-8

-------
McConnell,  J.C. and H.I. Schiff.  1978.   Methyl chloroform:  impact  on
  stratospheric ozone.  Science  199:174-177.

Neely, W.B., D.R. Branson, and G.E. Blau.  1974.  Partition  coefficient to
  measure bioconcentration potential of  organic chemicals in fish.
  Environ.  Sci. Techno1. 8:1113-1115.

Pearson, C.R. and G. McConnell.  1975.   Chlorinated C]_ and $2
  hydrocarbons in the marine environment. Proc. Roy.  Soc. London B
  189:305-322.

Shackelford, W.M. and L.H. Keith.  1976.  Frequency of organic compounds
  identified in water.  U.S. Environmental Protection Agency, Athens,
  Ga.  618p. EPA 600/4-76-062.

Tute, M.S.   1971.  Principles and practice of  Hansch  analysis:  a guide to
  structure-activity correlation for the medicinal chemist.   Adv. Drug
  Res. 6:1-77.

Verschueren, K.  1977.  Handbook of environmental data on organic
  chemicals.    Van Nostrand/Reinhold Press,  New York.  659p.

Weast, R.C. (ed.).  1977.  Handbook of  chemistry and  physics.  58th
  Edition.  CRC Press, Inc., Cleveland,  Ohio.   2398p.
                                       47-9

-------
                           48.  HEXACHLOROETHANE


48.1  Statement of Probable Fate

    Very little information pertaining specifically to the aquatic fate of
hexachloroethane was found.  It is, therefore, not possible to determine
the aquatic fate of hexachloroethane at this time.

48.2  Identification

    Hexachloroethane has been detected in finished drinking water (Shackel-
ford and Keith 1976; Environmental Protection Agency 1975).

    The chemical structure of hexachloroethane is shown below.

          C!     Cl

          I      '                             Alternate Names
    a c	c ci                       	
                                             Perchloroethane
          Cl     Cl                             Carbon hexachloride

     Hexachloroethane

     CAS NO. 67-72-1
     TSL NO. KI 40250

48.3  Physical Properties

    The general physical properties of hexachloroethane are given below.

     Molecular weight                        236.74
     (Weast 1977)

     Melting point                           Sublimes
     (Weast 1977)

     Boiling point at 777  torr               186C
     (Weast 1977)

     Vapor pressure at 20C                  0.4  torr
      (Verschueren 1977)

     Solubility in water at 22C             50 mg/1
      (Verschueren 1977)

     Log  octanol/water partition coefficient 3.34
      (Calculated from Tute 1971; see Methods section
     on Bioaccumulation)
                                     48-1

-------
48.4  Summary of Fate Data

    48.4.1  Photolysis

         No specific information pertaining to the rate of photolysis of
hexachloroethane in the aqueous environment under ambient conditions was
found.

         Photolysis in the atmosphere is important if hexachloroethane vo-
latilizes appreciably; unfortunately, the relative importance of volatili-
zation in comparison to other processes is unknown for this compound.  In
the event that a portion of the hexachloroethane should enter the atmo-
sphere, it would be expected to behave as its more thoroughly studied
analogue, tetrachlororaethane.  Tetrachloromethane is relatively stable in
the troposphere with respect to reaction with hydroxyl radicals and does
not directly photodissociate since it has no chromophores which absorb in
the visible or near ultraviolet region of the electromagnetic spectrum
(Jaffe and Orchin 1962).  Like tetrachloromethane, some portion of
hexachloroethane may diffuse unreacted to the stratosphere where signifi-
cant photodissociation could occur.

    48.4.2  Oxidation

         No information was found pertaining specifically to the oxidation
of hexachloroethane in the aquatic environment under ambient conditions.
Hexachloroethane probably behaves similarly to other halogenated aliphatics
in that it is not easily oxidized in aquatic systems since there are no
functional groups present on the compound which react strongly with
hydroxyl radicals (OH-).

         No information was found pertaining specifically to the oxidation
of hexachloroethane in the atmospheric environment under ambient condi-
tions.  Inferences can be made by analogy of hexachloroethane to tetra-
chloromethane;  the reaction of tetrachloromethane with hydroxyl radicals
in the troposphere, if it occurs, is very slow.

    48.4.3  Hydrolysis

         No information on hydrolysis of hexachloroethane was found.

    48.4.4  Volatilization

         Billing et al. (1975) reported the experimental half-life with re-
spect  to volatilization of 1 mg/1 hexachloroethane in water to be 45
minutes when stirred at 200 rpm at approximately 25C in an open container.
Removal of 90 percent hexachloroethane required greater than 120 minutes.
                                    48-2

-------
The stirring speed had a marked effect on the evaporative rate.  With
intermittent stirring for 15 seconds every 5 minutes, the time required for
50 percent depletion of 1,1,1-trichloroethane and trichloromethane, two an-
alogues of hexachloroethane, was approximately 90 minutes or on the order
of four times greater than was observed during constant stirring for these
two compounds.  The presence of sodium chloride at a 3 percent concentra-
tion, as in seawater, effected a ten percent decrease in evaporative rate
for short-chain chloroaliphatics.

         It is uncertain whether evaporation is the major pathway by which
hexachloroethane is lost from water.  Dilling ejt al. (1975) are careful to
point out the difficulties encountered in extrapolating their laboratory
results to real-world conditions, where the concentration of the organic
solute would probably be .very much less than 1 mg/1 and where surface and
bulk agitation would be highly variable.  Although the data appear to be
valid on a relative basis (i.e., correctly illustrating the relative rates
of volatilization of chlorinated aliphatics), they cannot be used as abso-
lute measures of volatilization rates from natural waters.  For the pur-
poses of this document, the data are used as rough-order-of-magnitude
indications of the importance of volatilization relative to other transport
and fate proceses, with the strong effects of agitation considered.  The
validity of this application has not been established.

         A subsequent study by Dilling (1977) was conducted using similar
experimental conditions as in the previous investigation, and an average
evaporative half-life of 40.7 minutes was obtained for 0.72 mg/1 hexa-
chloroethane in stirred water.

         The purpose of this subsequent study was to use the experimental
data that was obtained to test two theoretical models formulated to predict
evaporative rates of slightly soluble organic compounds from water.
Dilling (1977) found that the theoretical model by Mackay and Wolkoff
(1973) failed to predict evaporative half-lives, whereas the theoretical
model of Mackay and Leinonen (1975) using the parameters of Liss and Slater
(1974) correlated well with the experimental half-lives obtained by Dilling
(1977).  For example, the evaporative half-life of 40.7 minutes obtained
experimentally by Dilling (1977) is in relatively close agreement with the
evaporative half-life predicted by the Mackay and Leinonen model of 38
minutes but very different from the value of 4 minutes predicted by the
Mackay and Wolkoff (1973) model.

         Dilling (1977), however, comments that the apparent numerical
agreement between his data and the values predicted by the Mackay and
Leinonen (1975) model may be fortuitous.  Estimates of volatilization rates
based on the Mackay and Leinonen (1975) model depend primarily on liquid-
gas phase exchange rate constants, whereas the experimental model of
Dilling ej: al. (1975) and Dilling (1977) is controlled by the rate of
stirring and the wind velocity across the surface of the water.
                                       48-3

-------
         Pearson and McConnell (1975) theorize that the presence of organo-
chlorine compounds in upland waters is a result of atmospheric transport
and precipitation from the atmosphere by mechanisms such as rainfall and
dew-like condensation.

    48.4.5  Sorptlon

         No information was found pertaining specifically to the adsorption
of hexachloroethane.  Pearson and McConnell (1975), however, have found
other low molecular weight organochlorine compounds in Liverpool Bay
(England) sediments at concentrations of a few nanograms per kilogram.
Those compounds contained in the samples of marine sediment were the same
as those compounds present in the overlying waters.  There has been very
little information published which indicates the presence of hexachloro-
ethane in bodies of water, and no information was found concerning its ex-
istence in aquatic sediments.

    48.4.6  Bioaccumulation

         According to Koppennan et al. (1976), not all organochlorine com-
pounds bioaccumulate to high levels.  The data suggest that polar compounds
are more easily biodegraded, whereas the non-polar (highly lipophilic) com-
pounds accumulate.  In addition, Neely _et _al. (1974) have theorized that
bioaccumulation is directly related to the partition coefficient (P) of the
compound.  The log octanol/water partition coefficient (log P) of 3.34 cal-
culated from Tute (1971) indicates that hexachloroethane may exhibit a ten-
dency to bioaccumulate in organisms (see Methods section on bioaccumula-
tion).

         Significant bioaccumulation of hexachloroethane would require an
absence of biodegradation and a large environmental loading.  Not enough
information was found to determine whether hexachloroethane might be
present in the aquatic environment under these conditions.

    48.4.7  Biotransformation and Biodegradation

         No information was found pertaining specifically to the biodegra-
dation of hexachloroethane.  Literature references to microbial biodegra-
dation are few and conflicting; the majority find  that low molecular weight
chloroaliphatics are not metabolized  (Pearson and McConnell 1975; McConnell
_e_t al. 1975).  Although Pearson and McConnell (1975) were unable to
demonstrate microbial degradation of  the chlorinated ethanes, they did
observe the occurrence of chemical degradation of  similar compounds.

48.5  Data Summary

    Table 48-1 summarizes the aquatic fate information discussed above.
Not enough information was  found to predict the most likely aquatic fate of
this compound.
                                     48-4

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-------
48.6  Literature Cited

Billing, W.L.  1977.  Interphase transfer processes.  II.  Evaporation
  rates of chlororaethanes,  ethanes, ethylenes, propanes, and propylenes
  from dilute aqueous solutions.  Comparisons with theoretical
  predictions. Environ.  Sci. Technol. 11:405-409.

Billing, W.L., N.B. Tefertiller, and G.J. Kallos.  1975.  Evaporation rates
  of methylene chloride, chloroform, 1,1,1-trichloroethane, trichloro
  ethylene, tetrachloroethylene, and other chlorinated compounds in
  dilute aqueous solutions.  Environ. Sci. Technol. 9(9):833-838.

Environmental Protection Agency.  1975.   Preliminary assessment of
  suspected carcinogens in drinking water.  U.S. Environmental Protection
  Agency, (Office of Toxic  Substances),  Washington, D.C. 33p. EPA
  560/4-75-003.

Jaffe, H.H. and M. Orchin.   1962.  Theory and applications of ultraviolet
  spectroscopy.  John Wiley and Sons, Inc. New York. 624p.

Kopperman, H.L. D.W. Kuehl, and G.E. Glass.  1976.  Chlorinated compounds
  found in waste treatment effluents and their capacity to bicaccumulate.
  Proceedings of the conference on the environmental impact of water
  chlorination.  pp.327-345.  Oak Ridge, Tennessee, October 22-24, 1975.

Liss, P.S. and P.G. Slater.  1974.  Flux of gases across the air-sea
  interface.  Nature 247:181-184.

Mackay, D. and P.J. Leinonen.  1975.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. Technol.
  9:1178-1180.

Mackay, D. and A.W. Wolkoff.  1973.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. Technol.
  7:611-614.

McConnell, G., D.M. Ferguson, and C.R. Pearson.  1975.  Chlorinated
  hydrocarbons and  the environment.  Endeavor XXXIV:13-18.

Neely, W.B., D.R. Branson, and G.E. Blau.  1974.  Partition coefficient to
  measure bioconcentration potential of organic chemicals in fish.
  Environ. Sci. Technol. 8:1113-1115.

Pearson, C.R. and G. McConnell.  1975.  Chlorinated GI and C2
  hydrocarbons in the marine environment.  Proc. Roy. Soc. London B
  189:305-322.
                                     48-6

-------
Shackelford,  W.M. and L.H. Keith.  1976.   Frequency of organic  compounds
  identified in water.  U.S. Environmental Protection Agency, Athens,
  Ga.  618p.  EPA 600/4-76-062.

Tute, M.S. 1971.  Principles and practice of Hansch analysis:   a guide  to
  structure-activity correlation for the  medicinal chemist.   Adv.  Drug
  Res. 6:1-77.

Verschueren,  K.  1977.  Handbook of environmental data on organic
  chemicals.   Van Nostrand/Reinhold Press, New York.  659p.

Weast, R.C. (ed.).  1977.  Handbook of chemistry and physics.   58th
  Edition. CRC Press, Inc., Cleveland, Ohio.  2398 p.
                                       48-7

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                    49.  CHLOROETHENE (VINYL CHLORIDE)
49.1  Statement of Probable Fate

    Chloroethene introduced into aquatic systems will most probably be
quickly transferred to the atmosphere through volatilization.  In fact, re-
sults from model simulations indicate that chloroethene should not remain
in an aquatic ecosystem under most natural conditions.  Once in the tropo-
sphere, chloroethene reacts at an extremely rapid rate with hydroxyl radi-
cals, exhibiting a half-life on the order of a few hours with the subse-
quent formation of hydrogen chloride or formyl chloride as possible pro-
ducts.  Formyl chloride,-if formed, is reported to decompose thermally at
ambient temperatures with a half-life of about 20 minutes, yielding carbon
monoxide and hydrogen chloride.  As a result, chloroethene in the tropo-
sphere should be decomposed within a day or two of release.

    Based on the information found, it does not appear that oxidation,
hydrolysis, and biodegradation are important fate processes for chloro-
ethene in the aquatic environment, nor do sorption and bioaccumulation
appear to be important transport processes.

49.2  Identification

    Chloroethene has been detected in finished drinking water (Environ-
mental Protection Agency 1975a), in the atmosphere (Environmental Pro-
tection Agency 1974;  Lillian _e_t _al. 1975;  Gay _e_t _al. 1976), and in in-
dustrial waterborne discharges (Environmental Protection Agency 1974).

    The chemical structure of chloroethene is shown below.
   C'\         /H
     \  _   /                      Alternate Names

     /         ^V                    Vinyl chloride
                                      Monochloroethylene
                                      Mo novinyl chloride
                                      MVC
    Chloroethene                      Chloroethylene

    CAS NO. 75-01-4
    TSL NO. KU 96250
                                     49-1

-------
49.3  Physical Properties

    The general physical properties of chloroethene are given below.

    Molecular weight                         62.50
    (Weast 1977)

    Melting point                            -153.8C
    (Weast 1977)

    Boiling point at 760 torr                -13.37C
    (Weast 1977)

    Vapor pressure at 25C                   2,660 torr
    (Verschueren 1977)

    Solubility in water at 10C
    (Pearson and McConnell 1975)             60 mg/1
    (Verschueren 1977)                       1.1 mg/1 at 25C

    Log octanol/water partition coefficient  0.60
    (Radding _et _al. 1977)

49.4  Summary of Fate Data

    49.4.1  Photolysis

         Atmospheric ozone prevents essentially all sunlight of wavelengths
shorter than 290 run from reaching the earth's surface. Chloroethene, in the
vapor phase, does not absorb light of wavelengths greater than 220 nra, and
in water it does not absorb above 218 nm (Hill ^t _al. 1976).  As a result,
direct photolysis of chloroethene would be expected, at best, to be a very
slow process due to lack of overlap between chloroethene absorption and
sunlight radiation spectra.  To confirm this experimentally, Hill et al.
(1976) exposed solutions of 10 mg/1 chloroethene in water to filtered light
of greater than 300 nm from a mercury lamp.  No photolysis occurred over a
90-hour period.

         It is, however, possible that light-induced transformations of
chloroethene could occur through indirect photolysis.  Photolysis experi-
ments were conducted by Hill et_ _al. (1976) in natural water and in dis-
tilled water containing photosensitizers that absorb light of wavelengths
greater than 300 nm.  It was found that chloroethene in solution decomposed
rapidly when irradiated with ultraviolet light in the presence of acetone,
a high energy triplet sensitizer, or hydrogen peroxide, a free radical
source.  Photosensitization via energy transfer for the disappearance of
chloroethene monomers occurs efficiently only when the triplet state energy
                                    49-2

-------
of the sensitizer equals or exceeds that of the acceptor (which is, in this
case, chloroethene monomer).  Since the triplet state energies of mono-
olefins are very high (78-82 kcal mole"-"-), only high energy sensitizers
such as acetone (triplet energy 80 kcal mole"1) are effective.  Little is
known, however, about the concentrations and distribution of high-energy
sensitizers in the aquatic environment.  Polyvinyl chloride plants are
likely to use small amounts of peroxides in the polymerization process,
and, as a result, sunlight-induced decomposition of chloroethene monomers
is considered as a possible pathway for its disappearance from polyvinyl
chloride plant effluent water (Hill et _al. 1976).

         Atmospheric photodissociation of chloroethene appears to be much
less important than photochemical oxidation.  Rapid photochemical oxidation
is reported to remove the compound from the troposphere with a half-life of
a few hours (Environmental Protection Agency 1975b).  As a result, neither
the chlorine in chloroethene nor chloroethene itself is likely to diffuse
to the stratosphere.

    49.4.2  Oxidation

         Experiments at elevated temperatures (85C) have shown that
chloroethene is degraded in the presence of hydrogen peroxide, the chloro-
ethene reacting with hydroxyl radicals (HO") from peroxide decomposi-
tion (Hill __! 1976).  The decomposition of hydrogen peroxide is prob-
ably the rate determining step.  The significance of this reaction pathway
is difficult to assess because no data are available concerning the concen-
tration and species of free radicals in natural waters.  These experiments
do indicate, however, that if reactive radicals are present in natural
waters at significant concentrations, they may degrade chloroethene. Ex-
perimental results show that chloroethene will not be significantly de-
graded by molecular oxygen at temperatures and oxygen concentrations pre-
sent in natural waters (Hill _e_t _al. 1976).

         Chloroethene is tropospherically reactive and, as a result, will
not accumulate in the troposphere (Environmental Protection Agency I975b;
Lillian je_t _al. 1975).  In the rapid photochemical oxidation reaction,
chlorine is removed from the chloroethene with the subsequent formation of
hydrogen chloride (HC1) or formyl chloride (HCOC1) as possible products
(Environmental Protection Agency 1975b).  Formyl chloride, if formed, is
reported to decompose at room temperature with a half-life of about 20
minutes, yielding carbon monoxide (CO) and hydrogen chloride (HC1).  As a
result, almost all atmospheric chloroethene should be destroyed within a
day or two of release.  The hydrogen chloride formed is reported to be
removed from the troposphere during precipitation.
                                    49-3

-------
    49.4.3  Hydrolysis

         Hydrolysis over a pH range of 4.3 to 9.4 does not appear to be an
important pathway for loss of chloroethene from water (Environmental Pro-
tection Agency 1974).  This statement is based on results obtained from
studies of chloroethene levels in samples of clarifier effluent from a
chloroethene plant stored at 50C for 57 hours at pH 4.3, 8.0,  and 9.4 in
sealed septum vials.  Concentrations of chloroethene remaining in these
vials indicated that chloroethene decreased at the same rate at each pH
condition (Environmental Protection Agency 1974).  This lack of pH depen-
dence indicates that chloroethene loss probably occurred due to volatili-
zation rather than hydrolysis (Environmental Protection Agency 1974).

         The hydrolytic half-life of chloroethene has been estimated to be
less than 10 years at 25C (Hill _et jal. 1976).  Since the volatilization
rate of chloroethene is much more rapid than the predicted rate of hydroly-
sis, hydrolysis should not be a significant aquatic fate.

    49.4.4  Volatilization

         Hill _et zA. (1976) indicate that loss of chloroethene from the
aquatic environment by volatilization appears to be the most significant
process affecting its distribution.  This high rate of volatilization is
due to the high vapor pressure (2,660 torr at 25C) and the low solubility
in water of 1.1 mg/1 at 25C (Verschueren 1977).  On the basis of various
model simulations formulated by Hill _e _al. (1976), it appears that chloro-
ethene should not remain in an aquatic ecosystem under most natural condi-
tions.  The loss of chloroethene at constant temperature and pressure is a
function of water turbulence and mixing efficiency (Hill et al. 1976).

         Dilling et al. (1975) estimated the experimental half-life of
volatilization of chloroethene originally present at 1 mg/1 to be 26
minutes when stirred at 200 rpm in water at approximately 25 C in an open
container.  Removal of 90 percent of the chloroethene under the same con-
ditions required 96 minutes.  For chloroaliphatics in general, stirring
speed was found to have a marked effect on volatilization rate.

         Dilling et_ al_. (1975) are careful to point out the difficulties
encountered in extrapolating their laboratory results to real-world con-
ditions, where the concentration of the organic solute would probably be
very much less than 1 mg/1 and where surface and bulk agitation would be
highly variable.  Although the data appear to be valid on a relative basis
(i.e., correctly illustrating the relative rates of volatilization of
chlorinated aliphatics), they cannot be used as absolute measures of vola-
tilization rates from natural waters.  For the purposes of this document,
the data are used as rough-order-of-magnitude indications of the importance
of volatilization relative to other transport and fate proceses, with the
strong effects of agitation considered.  The validity of this application
has not been established.
                                    49-4

-------
         A subsequent study by Billing (1977) was conducted using the same
experimental conditions as in the i975 study, and a half-life of 27.6
minutes was obtained for 0.89 mg/1 chloroethene in water. The purpose of
this subsequent study was to use the experimental data to test two theore-
tical models used to predict evaporative rates of slightly soluble organic
compounds from water.  Dilling (1977) found that the model of Mackay and
Wolkoff (1973) failed to predict evaporative half-lives, but that the model
of Mackay and Leinonen (1975), using parameters from Liss and Slater
(1974), correlated well with experimental results.  For example, the eva-
porative half-life obtained experimentally by Dilling (1977) was approxi-
mately 28 minutes for chloroethene, as compared to 16.1 minutes predicted
by the Mackay and Leinonen (1975) model and 0.0054 minutes predicted by the
Mackay and Wolkoff (1973) model.

         Dilling (1977), however, comments that the apparent numerical
agreement between his data and the values predicted by the Mackay and
Leinonen (1975) model may be fortuitous.  Estimates of volatilization rates
based on the Mackay and Leinonen (1975) model depend primarily on liquid-
gas phase exchange rate constants, whereas the experimental model of
Dilling et al. (1975) and Dilling (1977) is controlled by the rate of
stirring and the wind velocity across the surface of the water.

         The Environmental Protection Agency (1974) reported the experimen-
tal decrease of 16 mg/1 chloroethene to be 96 percent in two hours when
stirred rapidly at 22C in an open beaker of distilled water.  In contrast,
quiescent water under the same conditions yielded a chloroethene concentra-
tion loss over two hours of only 25 percent (Environmental Protection
Agency 1974).

         Assuming that all processes involved are strictly first order, the
volatilization loss data above (Environmental Protection Agency 1974)
yields half-lives of 25.8 minutes for the stirred case and 290 minutes for
the quiescent case.   The half-life of 25.8 minutes for the stirred case
agrees very closely with the 27.6 minute evaporative half-life with
stirring reported by Dilling (1977).  Volatilization  rate data for
chloroethene based on the above references can be summarized as follows:
       Agitation                   Evaporative Half-life (Minutes)

Rapid Continual Stirring                      25.8
Continual Stirring at 200 rpm                 27.6
Discontinuous Stirring (5% of the time)    -  80-90
Quiescent; No Stirring                        290
                                      49-5

-------
         Hill et al. (1976) found that the rate of bulk exchange of gaseous
chloroethene between atmosphere and water is about twice that of oxygen.
As a result, Hill et _al. (1976) conclude that the loss of chloroethene by
volatilization from water is probably the most significant process in its
distribution.

    49.4.5  Sorption

         There is little information pertaining specifically to the rate of
adsorption of chloroethene onto particulate matter.  In a study on the be-
havior of chloroethene in water, the Environmental Protection Agency (1974)
found no significant difference in the rate of chloroethene loss from dis-
tilled water, river water, or effluent from a chloroethene plant stirred at
the same rate, thus indicating negligible adsorption of chloroethene onto
particulate matter.  Hill et_ al. (1976) indicate that aquatic sediments
could exhibit long-term storage of low levels of chloroethene if extreme
environmental conditions, such as continual high levels of chloroethene
input, were present.

    49.4.6  Bioaccumulation

         According to Kopperman ^t _al. (1976), not all organochlorine com-
pounds bioaccumulate to high levels.  The data suggest that polar compounds
are more easily biodegraded, whereas non-polar highly lipophilic compounds
accumulate.  Neely _e_t _al. (1974) have shown that bioaccumulation is di-
rectly related to the octanol/water partition coefficient (P) of the com-
pound.  The log octanol/water partition coefficient (log P) as reported in
Radding t _al. (1977) is 0.60, indicating little tendency for bioaccumula-
tion in the adipose tissues of organisms.

         Some authors (Lu _e_t al. 1977;  Environmental Protection Agency
1974) contend that chloroethene is too readily volatilized to undergo bio-
accumulation, except perhaps in the most extreme exposure conditions.
Studies by Hill ej: _al. (1976) on five bacterial, three fungal, and two
single organism cultures from natural aquatic systems did not show bio-
accumulation to be an appreciable process.

    49.4.7  Biotransformation and Biodegradation

         Existing evidence indicates that chloroethene is resistant to
microbial degradation.  Studies by Hill ejt al. (1976) using an isolated
bacterial culture containing two species of bacteria and three mixed fungal
populations were unable to show that chloroethene was biodegraded over a
five week period at concentrations of 20-120 mg/1.
                                    49-6

-------
    49.4.8  Other Reactions

         Hill et al. (1976) indicate that conditions could possibly exist
where chloroethene could be converted to more highly chlorinated compounds.
In environments such as municipal water chlorination facilities, high con-
centrations of chloride would exist.  Under certain conditions, chloro-
ethene may be converted to more highly chlorinated compounds based on the
reactivity of carbon-carbon double bonds with chlorine and hypohalous acid.

         Dissolved chloroethene in water will readily escape into the gas
phase, but chemical reactions can occur with water impurities which may
inhibit its release.  Many salts have the ability to form complexes with
chloroethene and can increase its solubility.  Therefore, the amounts of
chloroethene in water could be influenced significantly by the presence of
salts (Environmental Protection Agency 1975c).

49.5  Data Summary

    Table 49-1 summarizes the aquatic fate discussed above.  Oxidation
rates are photooxidation rates and refer to the rate of reaction of
chloroethene with hydroxyl radicals in the troposphere.

    Due to the high vapor pressure of chloroethene, volatilization to the
atmosphere is quite rapid.  In fact, from the results of model simulations
it appears that chloroethene should not remain in an aquatic ecosystem
under most natural conditions (Hill _e_t _al. 1976).  Once in the troposphere,
chloroethene reacts at an extremely rapid rate with hydroxyl radicals.
                                    49-7

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49.6  Literature Cited

Billing, W.L.   1977.   Interphase transfer processes.  II.  Evaporation
  rates of chioromethanes,  ethanes, ethylenes,  propanes and propylenes from
  dilute aqueous solutions.  Comparisons with theoretical predictions.
  Environ. Sci. Technol.  11:405-409.

Billing, W.L., N.B. Tefertiller, and G.J. Kallos.  1975.  Evaporation
  rates of methylene chloride, chloroform, 1,1,1-trichloroethane,
  trichloroethylene, tetrachloroethylene and other chlorinated compounds in
  dilute aqueous solutions.  Environ. Sci. Technol. 9(9):833-838.

Environmental Protection-Agency.  1974.  Preliminary assessment of the
  environmental problems associated with vinyl  chloride and polyvinyl
  chloride (appendices).  Environmental Protection Agency, Washington, D.C.
  Report on the Activities and Findings of the  Vinyl Chloride Task Force.
  67p.

Environmental Protection Agency.  1975a.  Preliminary assessment of
  suspected carcinogens in drinking water.  Environmental Protection
  Agency, (Office of Toxic Substances), Washington, D.C.  33p.  EPA
  560/4-75-003.

Environmental Protection Agency.  1975b.  Preliminary study of selected
  potential environmental contaminants - optical brighteners, methyl
  chloroform, trichloroethylene, tetrachloroethylene, ion exchange resins.
  Environmental Protection Agency, (Office of Toxic Substances),
  Washington, D.C.  286p.  EPA 560/2-75-002.

Environmental Protection Agency.  1975c.  Scientific and technical
  assessment report on vinyl chloride and polyvinyl chloride.
  Environmental Protection Agency, (Office of Research and Development),
  Washington, D.C. 115p. EPA 600/6-75-004.

Gay, B.W., Jr., P.L. Hanst, J.J. Bufalini, and  R.C. Noonan.   1976.
  Atmospheric oxidation of chlorinated ethylenes.  Environ. Sci. Technol.
  10:58-67.

Hill, J., IV, H.P. Kollig,  D.F. Paris, N.L. Wolfe, and R.G. Zepp.  1976.
  Dynamic behavior of vinyl chloride in aquatic ecosystems.  Environmental
  Protection Agency, (Office of Research and Development), Athens, Georgia.
  64p. EPA 600/3-76-001.

Kopperman, H.L., D.W. Kuehl, and G.E. Glass.  1976.  Chlorinated compounds
  found in waste treatment effluents and their  capacity to bioaccumulate.
  Proceedings of the conference on the environmental impact of water
  chlorination.  pp.327-345.  Oak Ridge, Tennessee, October 22-24, 1975.
                                    49-9

-------
Lillian, D., H.B. Singh, A.  Appleby,  L.  Lobban,  R.  Arnts,  R.  Gumpert,  R.
  Hague, J. Toomey, J. Kazazis,  M.  Antell,  D.  Hansen,  and  B.  Scott.   1975.
  Atmospheric fates of halogenated  compounds.   Environ.  Sci.  Technol.
  9:1042-1048.

Liss, P.S. and P.G. Slater.   1974.   Flux of gases across the  air-sea
  interface.  Nature 247:181-184.

Lu, P., R.L. Metcalf, N. Plummer, and D. Mandel. 1977.  The  environmental
  fate of three carcinogens:   benzo(a)pyrene,  benzidine, and  vinyl  chloride
  evaluated in laboratory model  ecosystems. Arch.  Environ. Contam.
  Toxicol. 6:129-142.

Mackay, D. and P.J. Leinonen.   1975.   Rate  of  evaporation  of  low-solubility
  contaminants from water bodies to atmosphere.  Environ.  Sci.  Technol.
  9:1178-1180.

Mackay, D. and A.W. Wolkoff.   1973.  Rate of evaporation of
  low-solubility contaminants  from  water bodies  to  atmosphere.   Environ.
  Sci. Technol. 7:611-614.

Neely, W.B., D.R. Branson, and G.E. Blau.  1974. Partition coefficient  to
  measure bioconcentration potential  of  organic  chemicals  in  fish.
  Environ. Sci. Technol. 8:1113-1115.

Pearson, C.R. and G. McConnell.   1975.   Chlorinated C]_ and G
  hydrocarbons in the marine environment.  Proc. Roy.  Soc. London B
  189:305-322.

Radding, S.B., D.H. Liu, H.L.  Johnson, and  T.  Mill.  1977. Review  of  the
  environmental fate of selected chemicals. Environmental Protection
  Agency, (Office of Toxic Substances),  Washington, D.C.  147p.  EPA
  560/5-77-003.

Verschueren, K.  1977.  Handbook of environmental data on  organic
  chemicals.  Van Nostrand/Reinhold Press,  New York.   659p.

Weast, R.C. (ed.).  1977.  Handbook of chemistry and physics.  58th
  Edition. CRC  Press, Inc., Cleveland,  Ohio.   2398p.
                                   49-10

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               50.  1,1-DICHLORQETHENE (VINYLIDINE CHLORIDE)


50.1  Statement of Probable Fate

    Volatilization appears to be the major transport process for removal of
1,1-dichloroethene from aquatic systems.  Once in the troposphere, the com-
pound is attacked by hydroxyl radicals at the double bond, reportedly re-
sulting in such products as chloroacetyl chloride, phosgene, formic acid,
hydrochloric acid, carbon monoxide, and formaldehyde.  The tropospheric
lifetime of 1,1-dichloroethene, based on its rate of reaction with hydroxyl
radicals, appears to be substantially less than one day.  This is an esti-
mated lifetime extrapolated from the reported lifetime of trichloroethene,
a more highly chlorinated analogue of 1,1-dichloroethene.  Due to the pro-
jected reactivity of 1,1-dichloroethene with hydroxyl radicals in the
troposphere, it is unlikely that unreacted 1,1-dichloroethene will reach
the stratosphere.

    Neither hydrolysis nor oxidation in the aquatic environment appears to
be a significant fate process for 1,1-dichloroethene.  No information was
found indicating that microorganisms exist which can readily biodegrade
1,1-dichloroethene.  In addition, no evidence of bioaccumulation of 1,1-
dichloroethene in aquatic organisms or selective adsorption of this com-
pound onto suspended sediments was found.

50.2  Identification

    1,1-Dichloroethene, or vinylidene chloride,  has been detected in
finished drinking water (Environmental Protection Agency I975a).

    The chemical structure of 1,1-dichloroethene is shown below:


      c'v          P                        Alternate Names
          C - C
         s        \                          Vinylidene chloride
      Ci            H                       Vinylidine chloride
                                            1,1-Dichloroethylene
                                            1,1-DCE
     1,1-Dichloroethene

     CAS NO. 75-35-4
     TSL NO. KV 92750
                                    50-1

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50.3  Physical Properties

    The general physical properties of 1,1-dichloroethene are given be-
low:

    Molecular weight                         96.94
    (Weast 1977)

    Melting  point                           -122.1C
    (Weast 1977)

    Boiling point at 760 torr                37C
    (Weast 1977)

    Vapor pressure at 25C                   591 torr
    (Verschueren 1977)

    Solubility in water at 20C              400 mg/1
    (Pearson and McConnell 1975)

    Log octanol/water partition              1.48
    coefficient (Tute 1971; See
    Methods section on Bioaccum-
    ulation)

50.4  Summary of Fate Data

    50.4.1  Photolysis

         No information pertaining specifically to the rate of photodisso-
ciation of 1,1-dichloroethene in the aquatic environment was found.  Jensen
and Rosenberg (1975), using a closed system where no evaporation could
occur, found no significant difference between samples stored in daylight
and in darkness for tetrachloroethene and trichloroethene,  two analogues of
1,1-dichloroethene.  From these results, it was assumed that no significant
photolysis occured for these two compounds and that similar results would
be obtained for 1,1-dichloroethene.

         1,1-Dichloroethene is quite volatile and, as a result, is rapidly
transported to the troposphere.  Photodissociation in water or in the
troposphere would not be expected to occur for 1,1-dichloroethene since
this compound has no chromophores which absorb in the visible or near
ultraviolet region of the electromagnetic spectrum (Jaffe and Orchin
1962).  In addition, 1,1-dichloroethene would not be expected to be avail-
able for photodissociation above the ozone layer because of rapid photo-
oxidation in the troposphere.
                                    50-2

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    50.4.2  Oxidation
                                                        
         No information pertaining specifically to the rate of oxidation of
1,1-dichloroethene in the aquatic environment was found.  In a study uti-
lizing closely related analogues, trichloroethene and tetrachloroethene,
Billing et_ _al. (1975) conducted laboratory experiments in which there was
about a sixfold molar excess of dissolved oxygen compared to the total
amounts of trichloroethene and tetrachloroethene.  The air space above the
solutions in closed systems contained approximately 90 times as much oxygen
as  was present in the saturated solutions.  Thus, there was a great excess
of  oxygen present which would have been available for complete oxidation of
trichloroethene and tetrachloroethene.  The experimental results indicate
that volatilization is much more rapid than any oxidative process which may
have occurred.

         Billing _e_t _al. (1975) did, however, find evidence of significant
photooxidation of tetrachloroethene and trichloroethene in aqueous media in
the presence of sunlight.  The amounts of tetrachloroethene and trichloro-
ethene remaining in aerated water in the dark after 12 months is approxi-
mately 13 and 20 percent greater, respectively, than the amount of tetra-
chloroethene and trichloroethene remaining in similar samples stored in the
presence of sunlight.  The finding that sunlight had a significant effect
on the reactivity of tetrachloroethene and trichloroethene is consistent
with predictions based on vapor phase photolysis studies showing that
tetrachloroethene and trichloroethene disappeared when irradiated with long
wavelength light in the presence of nitric oxide or nitrogen dioxide
(Altshuller and Bufalini 1971).  Billing _et _al. (1975) attribute the dif-
ference in disappearance of tetrachloroethene and trichloroethene in the
presence and absence of sunlight to free radical oxidation.  It seems
likely that 1,1-dichloroethene would also exhibit significant indirect pho-
tooxidation in aqueous media in the presence of sunlight since this com-
pound is less chlorinated than tetrachloroethene and trichloroethene and,
thus, would probably be even more amenable to attack at the double bond.

         1,1-Dichloroethene is quite volatile and, thus, would be expected
to be transported rapidly to the troposphere.  Once in the troposphere,
1,1-dichloroethene should undergo fairly rapid photooxidation via hydroxyl
radical attack on the double bond, in similar fashion to the photooxidation
reported for tetrachloroethene and trichloroethene (Environmental Protec-
tion Agency 1975b).  Gay _e_t _a_l. (1976) have reported that photooxidation
of 1,1-dichloroethene is even more rapid than photooxidation of trichloro-
ethene and tetrachloroethene under conditions of fairly constant halocarbon
to N0 ratios.  Gay _et al. (1976) maintain that, since this ratio is
approximately constant, the reactivities of the compounds studied are at
least semi-quantitative.  The greater reactivity of the less chlorinated
alkanes is consistent with the results of Yung et al. (1975).
                                     50-3

-------
         According to Dilling _et_ _al. (1976), 1 ,l-dichloroet:hene, cis-
dichloroethene antf trans-dichloroethene, trichloroethene arid chloroethene
decomposed at moderate rates when compared with a series of standard hy-
drocarbons such as ethene, trans-2-butene, and others under simulated
atmospheric conditions.  These compounds were reported by Dilling et al.
(1976) to have estimated half-lives of 5 to 12 hours under bright sunlight
in the presence of nitric oxide.

         Pearson and McConnell (1975) exposed quartz flasks that contained
air and small amounts of halogenated aliphatic compounds (10"' to 10"^
by mass) to the diurnal and climatic variations of incident radiation and
temperature.  The air used for the experiment was not subjected to any
pretreatment and was, thus, not identical in terms of its humidity or in
its more detailed composition for all the experiments.  Samples of air were
withdrawn periodically for analysis of residual halogenated aliphatics.
Half-lives for halogenated aliphatic compounds obtained by interpolation
have a reproducibility of JL50% and are dependent on the climatic condi-
tions prevailing at the time of the experiment.  An interpolated tropo-
spheric half-life of 8 weeks was obtained for 1,1-dichloroethene.  This
half-life, however, is probably not very accurate for two reasons. First,
experiments were not set up and monitored simultaneously and, therefore,
samples were not subject to the same changes of solar flux and temperature.
Second, the quartz flasks were not filled with the same composition of air.

         Gay _e_t _al. (1976) reported the photooxidation products of 1,1-di-
chloroethene to be chloroacetyl chloride, phosgene, formic acid, hydro-
chloric acid, carbon monoxide, and formaldehyde.  Pearson and McConnell
(1975) reported the photooxidation products of 1,1-dichloroethene irradi-
ated in air by xenon arc (filtered to remove radiation of wavelengths below
290 run) to be carbon monoxide, carbon dioxide, and hydrochloric acid.  The
lack of agreement between the products found by Gay et al. (1976) and
Pearson and McConnell (1975) may be due to differences in experimental
conditions.

    50.4.3  Hydrolysis

         No information pertaining specifically to the rate of hydrolysis
of 1,1-dichloroethene in the aquatic environment was found.  A study of two
closely related analogues, trichloroethene and tetrachloroethene was in-
cluded in work by Dilling et _al. (1975) who conducted reactivity experi-
ments in closed aqueous systems in  the absence of light to eliminate com-
peting processes such as volatilization and  photolysis.  The experimen-
tal half-lives obtained for  tetrachloroethene and trichloroethene were 8.8
and 10.7 months, respectively.  In  terms of hydrolysis, these values can,
at best, be regarded only as maximum half-lives, since the relative contri-
butions of hydrolysis and oxidation in the experiments cannot be ascer-
tained .
                                    50-4

-------
    50.4.4  Volatilization

         Billing et al. (1975) estimated the experimental half-life for
volatilization of 1,1-dichloroethene originally present at 1 mg/1 to be 22
minutes when stirred at 200 rpm in water at approximately 25C in an open
container.  Removal of 90 percent of the 1,1-dichloroethene under the same
conditions required 89 minutes.  For chloroaliphatics in general, stirring
speed was found to have a marked effect on volatilization rate.  With
intermittent stirring for 15 seconds every five minutes, the time required
for 50 percent depletion of tetrachloroethene and trichloroethene, an-
alogues of 1,1-dichloroethene, was greater than 90 minutes, or four times
greater than was observed for these compounds during constant stirring.
The presence of sodium chloride at a 3 percent concentration, as in sea-
water, caused about a 10 percent decrease in the chlorinated compound
evaporation rate.   Dilling _e_t al. (1975) are careful to point out the
difficulties encountered in extrapolating their laboratory results to real-
world conditions, where the concentration of the organic solute would
probably be very much less than 1 mg/1 and where surface and bulk agitation
would be highly variable.  Although the data appear to be valid on a rela-
tive basis (i.e., correctly illustrating the relative rates of volatiliza-
tion of chlorinated aliphatics), they cannot be used as absolute measures
of volatilization rates from natural waters.  For the purposes of this
document, the data are used as rough-order-of-magnitude indications of the
importance of volatilization relative to other transport and fate pro-
cesses, with the strong effects of agitation considered.  The validity of
this application has not been established.

         A subsequent study by Dilling (1977) was conducted using the same
experimental conditions as the 1975 study, and an average evaporative
half-life of 27.2 minutes was obtained for 1 mg/1 of 1,1-dichloroethene.
The purpose of this subsequent study was to use the experimental data
previously obtained to test two theoretical models which may be used to
predict evaporative rates of slightly soluble organic compounds from water.
Dilling (1977) found that the theoretical model of Mackay and Wolkoff
(1973) failed to predict evaporative half-lives, but that the theoretical
model of Mackay and Leinonen (1975) using the parameters of Liss and Slater
(1974) correlated well with the experimental half-lives obtained.  For
example, the 27.7 minute evaporative half-life experimentally obtained by
Dilling (1977) for 1,1-dichloroethene compared favorably to a 20.1 minute
evaporative half-life predicted by the Mackay and Leinonen (1975) model but
did not agree with the 0.029 minute half-life predicted by the Mackay and
Wolkoff (1973) model.

         Dilling (1977), however, comments that the apparent numerical
agreement between his data and the values predicted by the Mackay and
                                    50-5

-------
Leinonen (1975) model may be fortuitous.  Estimates of volatilization rates
based on the Mackay and Leinonen (1975) model depend primarily on liquid-
gas phase exchange rate constants,  whereas the experimental model of
Billing e_t jil. (1975) and Billing (1977) is controlled by the rate of
stirring and the wind velocity across the surface of the water.

         Pearson and McConnell (1975) suggest that the presence of 1,1-di-
chloroethene, as well as other halogenated aliphatics, in ambient waters is
due to the absorption of chloroorganics from the atmosphere by water.
Aerial transport of these chloroorganics in indicated by Pearson and
McConnell (1975) to play a major role in the wider distribution of these
compounds and accounts for their presence in upland waters.

    50.4.5  Sorption

         No information pertaining specifically to the adsorption of 1,1-
dichloroethene onto sediments was found.  Billing et al. (1975) carried out
two closed system experiments on trichloroethene and tetrachloroethene
(analogues of 1,1-dichloroethene),  wherein solute loss could occur only
through adsorption.  Bry bentonite clay at 375 mg/1 was introduced into a
sealed solution and in ten minutes there was approximately a ten percent
adsorption of the solute by the clay.  When the amount of clay added to the
closed system was doubled (750 mg/1), 22 percent of the solute was removed
by adsorption in 30 minutes.  The authors indicated that there appeared to
be little selectivity among the various chlorinated compounds in the
adsorption process.  Some adsorption of trichloroethene and tetrachloro-
ethene onto dry powdered dolomitic limestone was observed, but, again,
without any selectivity among the compounds.

         In the sealed system with approximately 500 mg/1 peat moss,
approximately 40 percent of the solutes tetrachloroethene and trichloro-
ethene was adsorbed in 10 minutes.   At longer times, no further solute
removal was noted.  Pearson and McConnell (1975) found no clear evidence of
selective concentration of tetrachloroethene and trichloroethene in sedi-
ments; presumably, this would also be the case for 1,1-dichloroethene.

    50.4.6  Bioaccumulation

         According to Kopperman et_ aJ.. (1976), not all organochlorine com-
pounds bioaccumulate to high levels.  The data suggest that polar compounds
are more easily biodegraded, and the non-polar (highly lipophilic) com-
pounds accumulate.  Neely _e_t _al. (1974) have shown that bioaccumulation is
related to the octanol/water partition coefficient (P) of the compound.
The log octanol/water partition coefficient (log P) of 1.48, as calculated
by the method of Hansch (Tute 1971), indicates that 1,1-dichloroethene will
probably not bioaccuraulate to any significant extent (see Methods section
on bioaccumulation).
                                     50-6

-------
    50.4.7  Biotransformation and Biodegradation

         No information pertaining specifically to the rate of biodegrada-
tion of 1,1-dichloroethene in aquatic systems was found.  According to
Pearson and McConnell (1975) only completely sealed systems, such as the
standard BOD (biochemical oxygen demand) bottle technique, can be used to
measure biochemical degradation of volatile compounds such as 1,1-dichloro-
ethene.  The BOD bottle experiments of Pearson and McConnell have been
unable to demonstrate any significant oxygen absorption from compounds
containing only C, H, and Cl, thus indicating that biochemical degradation
of such compounds is indeed very slow.  Literature references to microbial
biodegradation are few and conflicting; the majority indicate that low
molecular weight chloroaliphatics are not metabolized (Pearson and
McConnell 1975; McConnell et_ _al. 1975).

50.5  Data Summary

    Table 50-1 summarizes the aquatic fate discussed above.  The oxidation
lifetime given is a photooxidation lifetime and refers to the rate of re-
action of 1,1-dichloroethene with hydroxyl radicals in the troposphere.
This rate is extrapolated from the reactivity of tetrachloroethene and
trichloroethene with hydroxyl radicals in the troposphere.

    Due to the relatively high vapor pressure of 1,1-dichloroethene,
volatilization from the aquatic system to the atmosphere is quite rapid.
Once in the troposphere the compound is oxidized by hydroxyl radicals. The
tropospheric lifetime of 1,1-dichloroethene is expected to be somewhat less
than that of its analogues, trichloroethene and tetrachloroethene (less
than 1 day and 10 days, respectively).
                                  50-7

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50.6  Literature Cited

Altshuller, A.P. and J.J. Bufalini.  1971.  Photochemical aspects of
  pollution: a review.  Environ. Sci. Technol. 5:39-64.

Billing, W.L., 1977.  Interphase transfer processes.  II.  Evaporation
  rates of chloromethanes, ethanes, ethylenes, propanes, and propylenes
  from dilute aqueous solutions.  Comparisons with theoretical predictions.
  Environ. Sci. Technol.  11:405-409.

Dilling, W.L., C.J. Bredeweg, and N.B. Tefertiller.  1976.  Simulated
  atmospheric photodecomposition rates of methylene chloride,
  1,1,1-trichloroethane,.trichloroethylene, tetrachloroethylene, and other
  compounds.  Environ. Sci. Technol.  10:351-356.

Dilling, W.L., N.B. Tefertiller, and G.J. Kallos.  1975.  Evaporation rates
  of methylene chloride, chloroform, 1,1,1-trichloroethane,
  trichloroethylene, tetrachloroethylene and other chlorinated compounds in
  dilute aqueous solutions.  Environ. Sci. Technol. 9(9):833-838.

Environmental Protection Agency. 1975a.  Preliminary assessment of
  suspected carcinogens in drinking water.  U.S. Environmental Protection
  Agency, (Office of Toxic Substances), Washington, D.C. 33p.
  EPA-560/4-75-003.

Environmental Protection Agency.  1975b.  Report on the problem of
  halogenated air pollutants and stratospheric ozone.  U.S. Environmental
  Protection Agency, (Office of Research and Development), Research
  Triangle Park, North Carolina.  55p. EPA 600/9-75-008.

Gay, B.W., Jr., P.L. Hanst, J.J. Bufalini, and R.C. Noonan.  1976.
  Atmospheric oxidation of chlorinated ethylenes.  Environ. Sci. Technol.
  10:58-67.

Jaffe,  H.H. and M. Orchin.  1962.  Theory and applications of
  ultraviolet spectroscopy.  John Wiley and Sons, Inc., New York.  624p.

Jensen, S. and R. Rosenberg.  1975.  Degradability of some chlorinated
  aliphatic hydrocarbons in sea water and sterilized water.  Water Res.
  9:659-661.

Kopperman, H.L., D.W. Kuehl, and G.E. Glass.  1976.  Chlorinated compounds
  found in waste treatment effluents and their capacity to bioaccumulate.
  Proceedings of the conference on the environmental impact of water
  chlorination.  pp.327-345.  Oak Ridge, Tennessee, October 22-24, 1975.
                                    50-9

-------
Liss, P.S. and P.G. Slater.  1974.   Flux of gases across the air-sea
  interface.  Nature 247:181-184.

Mackay, D. and P.J. Leinonen.  1975.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. Technol. 9:
  1178-1180.

Mackay, D. and A.W. Wolkoff.  1973.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. Technol. 7:
  611-614.

McConnell, G., D.M. Ferguson, and C.R. Pearson.  1975.  Chlorinated
  hydrocarbons and the environment.  Endeavor XXXIV:13-18.

Neely, W.B., D.R. Branson, and G.E. Blau.  1974.  Partition coefficient to
  measure bioconcentration potential of organic chemicals in fish.
  Environ. Sci. Technol.  8:1113-1115.

Pearson, C.R. and G. McConnell.  1975.  Chlorinated Cj and G
  hydrocarbons in the marine environment.  Pro. Roy. Soc. London B
  189:305-322.

Tute, M.S. 1971.  Principles and practice of Hansch analysis:  a guide to
  structure-activity correlation for the medicinal chemist.  Adv. Drug
  Res. 6:1-77.

Verschueren, K. 1977.  Handbook of environmental data on organic chemicals.
  Van Nostrand/Reinhold Press, New York, 659p.

Weast, R.C., (ed.). 1977.  Handbook of chemistry and physics.  58th
  Edition. CRC Press Inc., Cleveland, Ohio.  2398p.

Yung, Y.L., M.B. McElroy, and S.C. Wofsy.  1975.  Atmospheric halocarbons:
  a discussion with emphasis on chloroform.  Geophys. Res. Lett.
  2(9):397-399.
                                   50-10

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                       51.  1,2-trans-DICHLOROETHENE


51.1  Statement of Probable Fate

    Volatilization appears to be the major transport process for removal of
1,2-trans-dichloroethene from aquatic systems.  Once in the troposphere,
the compound is attacked at the double bond by hydroxyl radicals, resulting
in products such as formic acid, hydrochloric acid, and carbon monoxide.
The tropospheric lifetime of 1,2-trans-dichloroethene based on its rate of
reaction with hydroxyl radicals alone is probably much less than one day.
This is an estimated lifetime extrapolated from the reported lifetime of
the more highly chlorinated analogue of 1,2-trans-dichloroethene, tri-
chloroethene.  Due to the relatively high estimated reactivity of 1,2-
trans-dichloroethene with hydroxyl radicals in the troposphere, it is
unlikely that unreacted 1,2-trans-dichloroethene will diffuse upward to the
stratosphere to undergo photodissociation.

    Based on the information found, it does not appear that photodissocia-
tion is a significant fate of 1,2-trans-dichloroethene in the aquatic or
the atmospheric environment.  This is a result of the relatively rapid
volatilization from aquatic systems and the rapid rate of hydroxyl radical
attack in the troposphere.  Oxidation and hydrolysis in the aquatic en-
vironment do not appear to be significant for this compound, and no in-
formation was found indicating that microorganisms can readily biodegrade
1,2-trans-dichloroethene.

    It has already been stated that volatilization is probably the most
significant transport process of 1,2-trans-dichloroethene.  There is, at
present, no evidence of bioaccumulation of 1,2-trans-dichloroethene in
aquatic organisms.  In addition, no evidence was found pertaining to ad-
sorption of this compound onto suspended solids or sediments.

51.2  Identification

    1,2-trans-Dichloroethene has been detected in finished drinking water
(Environmental Protection Agency 1975a).

    The chemical structure of 1,2-trans-dichloroethene is shown below:

                                             Alternate Names
          H            Cl
            N  	  /                       trans-l,2-Dichloroethene
              C ^^z C
             >        v                       trans-1,2-Dichloroethylene
          _.'          >                      trans-Acetylene dichloride
                                             Dioform

     1,2-trans-Dichloroethene

     CAS NO. 540-59-0
     TSL NO. KV94000
                                    51-1

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51.3  Physical Properties

    The general physical properties of 1,2-tran.s-dichloroetherie are given
below:

    Molecular Weight                         96.94
    (Weast 1977)

    Melting point                            -50C
    (Weast 1977)

    Boiling point                            47.5C
    (Weast 1977)

    Vapor pressure at 14C                   200 torr
    (Verschueren 1977)

    Solubility in water at 20C              600 tng/1
    (Verschueren 1977)

    Log octanol/water partition              1.48
    coefficient (Calculated by the
    method of Tute (1971); see Methods
    section on bioaccumulation)

51.4  Summary of Fate Data

    51.4.1  Photolysis

         No information was found pertaining specifically to the rate of
photodissociation of 1,2-trans-dichloroethene in the aquatic environment.
Jensen and Rosenberg (1975), however, using a closed aqueous system from
which no evaporation could occur, found no significant difference in the
rate of disappearance between daylight exposure and darkness for dissolved
tetrachloroethene and trichloroethene, more thoroughly studied analogues of
1,2-trans-dichloroethene.  From these results, it is inferred that photoly-
sis of this compound is insignificant.

         1,2-trans-Dichloroethene is quite volatile and, as a result, is
rapidly transported to the troposphere.  1,2-trans-Dichloroethene is not
known to undergo photodissociation in the terrestrial environment since
this compound has no chromophores which absorb in the visible or near ultra-
violet region of the electromagnetic spectrum (Jaffe and Orchin 1962).  In
addition, 1,2-trans-dichloroethene would not be expected to be available
for photodissociation above the ozone layer because oxidation of 1,2-trans-
dichloroethene in the troposphere is expected to occur so rapidly that no
unreacted portion of the compound will survive to diffuse upward to the
stratosphere.
                                    51-2

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    51.4.2  Oxidation

         No information was found pertaining specifically to the rate of
oxidation of 1,2-trans-dichloroethene in the aquatic environment.  In a
study of closely related analogues, trichloroethene and tetrachloroethene,
Dilling _e_t_ _al_. (1975) conducted laboratory experiments in which there was
about a sixfold molar excess of dissolved oxygen compared to the total
amounts of trichloroethene and tetrachloroethene.  The air space above the
solutions in closed systems contained approximately 90 times as much oxygen
as was present in the saturated solutions.  Thus, there was a large excess
of oxygen present which would have been available for complete oxidation of
trichloroethene and tetrachloroethene.  The results obtained from samples
held in the dark for 6 and 12 months indicated half-lives of 10.7 months
for trichloroethene and 8.8 months for tetrachloroethane; Dilling et al.
(1975) attributed the disappearance of these organochlorine compounds to
oxidation.

         Dilling et_ al. (1975) also found evidence of significant photo-
oxidation of tetrachloroethene and trichloroethene, analogues of 1,2-trans-
dichloroethene, in aqueous media in the presence of sunlight.  The amounts
of tetrachloroethene and trichloroethene remaining in aerated water in the
dark after 12 months were approximately 13 and 20 percent greater, respec-
tively, than the amount of tetrachloroethene and trichloroethene remaining
in aerated water in the presence of sunlight.  The finding that sunlight
had a significant effect on the reactivity of tetrachloroethene and tri-
chloroethene is consistent with predictions based on vapor phase photolysis
studies which showed that tetrachloroethene and trichloroethene disappeared
when irradiated with long wavelength light in the presence of nitric oxide
or nitrogen dioxide (Altshuller and Bufalini 1971).  Dilling _et al. (1975)
attribute the difference in disappearance of tetrachloroethene and tri-
chloroethene in sunlight and dark to oxidation by free radicals generated
by solar irradiation.  It seems likely that 1,2-trans-dichloroethene would
also exhibit significant photooxidation in aqueous media in the presence of
sunlight since this compound is less chlorinated than tetrachloroethene and
trichloroethene and, thus, may be even more amenable to attack at the
double bond.

         1,2-trans-Dichloroethene is quite volatile and would be expected
to be transported rapidly to the troposphere.  With regard to tropospheric
attack of hydroxyl radicals, inferences have been drawn from the more
thoroughly studied chlorinated olefins, trichloroethene and tetrachloro-
ethene.  Tetrachloroethene and trichloroethene undergo fairly rapid photo-
oxidation in the troposphere as a result of hydroxyl radical attack at the
double bond (Environmental Protection Agency 1975b).  In fact, the photo-
oxidation of tetrachloroethene and trichloroethene is so rapid that no
unreacted tetrachloroethene or trichloroethene remains to diffuse to the
stratosphere to undergo photodissociation by shorter wavelength, higher
                                     51-3

-------
energy ultraviolet light (Environmental Protection Agency 1975b).
According to the Environmental Protection Agency (1975b) tetrachloroethene.
is attacked by hydroxyl radicals more slowly than most other olefin
pollutants due to the presence of four chlorine atoms.  Trichloroethene,
having one less chlorine atom, has been shown to be attacked more rapidly
by hydroxyl radicals in the troposphere than tetrachloroetheae (Gay et al.
1976; Environmental Protection Agency 1975b; Yung _e_t _al. 1975).  1,2-trans-
Dichloroethene, having one less chlorine atom than trichloroethene, would
be expected to be attacked even more rapidly by hydroxyl radicals.

         Gay _e_t al_. (1976) reported photooxidation of 1,2-trans-dichloro-
ethene to be even more rapid than photooxidation of trichloroethene and
tetrachloroethene under conditions of fairly constant halocarbon to N02
ratios.  Gay et al. (1976-) maintain that, since this ratio is approximately
constant, the reactivities of the compounds studied are at least semi-
quantitative.  According to Dilling et al. (1976), 1,2-trans-dichloro-
ethene, cis-dichloroethene, trichloroethene, and chloroethene decomposed at
moderate rates when compared with a series of standard hydrocarbons such as
ethene, trans-2-butene, and others under simulated atmospheric condi-
tions.  These chlorinated compounds were reported by Dilling et al. (1976)
to have estimated half-lives of 5 to 12 hours under bright sunlight in the
presence of nitric oxide.

         Pearson and McConnell (1975) exposed quartz flasks containing air
to the diurnal and climatic variations of incident radiation and tempera-
ture.   The air used for the experiment was not subjected to any pretreat-
ment and was, thus, not identical in terms of its humidity or in its more
detailed composition for all the experiments.  Samples of air were with-
drawn periodically for analysis of residual halogenated aliphatics, and
approximate ( 50%) half-lives for organochlorines were obtained by inter-
polation.  An interpolated tropospheric half-life of 8 weeks was obtained
for  1,1-dichloroethene, an analogue of 1,2-trans-dichloroethene.  This
half-life, however, is probably not very accurate for two reasons.  First,
experiments were not set up and monitored simultaneously and, therefore,
were not subject to identical conditions of solar flux and temperature.
Second, the quartz flasks were not filled with the same composition of air.

         Gay et al. (1976) reported the photooxidation products of
1,2-trans-dichloroethene to be formic acid, hydrochloric acid, and carbon
monoxide.  Pearson and McConnell (1975) reported the photooxidation
products of an analogue of 1,2-trans-dichloroethene, 1,1-dichloroethene,
irridated in air by a xenon arc and filtered to remove radiation of
wavelengths below 290 nm, to be carbon monoxide, carbon dioxide, and
hydrochloric acid.
                                      51-4

-------
    51.4.3  Hydrolysis

         No information was found pertaining to the rate of hydrolysis of
1,2-trans-dichloroethene.  Billing et_ _al. (1975), in summarizing work by
other researchers state that trichloroethene and tetrachloroethene, which
are structurally similar to 1,2-trans-dichloroethene, are resistant to
hydrolysis at temperatures in the range of 100C.  The presence of oxygen
was reported to accelerate the decomposition of both compounds.

    51.4.4  Volatilization

         Billing t _al. (1975) reported the experimental half-life of
volatilization of 1 mg/1' 1,2-trans-dichloroethene from water to be 22
minutes when stirred at 200 rpm at approximately 25C in an open container.
Removal of 90 percent of the 1,2-trans-dichloroethene required 89 minutes.
With stirring intermittently for 15 seconds every five minutes, the time
required for 50 percent depletion of tetrachloroethene and trichloroethene,
analogues of 1,2-trans-dichloroethene, was greater than 90 minutes, or on
the order of four times greater than was observed during constant stir-
ring. The presence of sodium chloride at a 3 percent concentration, as in
seawater, caused about a 10 percent decrease in the chlorinated compound
evaporation rate.   Billing _et^ _al. (1975) are careful to point out the
difficulties encountered in extrapolating their laboratory results to
real-world conditions, where the concentration of the organic solute would
probably be very much less than 1 mg/1 and where surface and bulk agitation
would be highly variable.  Although the data appear to be valid on a
relative basis (i.e., correctly illustrating the relative rates of
volatilization of chlorinated aliphatics), they cannot be used as absolute
measures of volatilization rates from natural waters.  For the purposes of
this document, the data are used as rough-order-of-magnitude indications of
the importance of volatilization relative to other transport and fate
proceses, with the strong effects of agitation considered.  The validity of
this application has not been established.

         A subsequent study by Billing (1977) was conducted using the same
experimental conditions as in the 1975 study; from the later study an aver-
age half-life of 24.0 minutes was obtained for 1 mg/1 trans-1,2-dichloro-
ethene.  The purpose of this subsequent study was to use the experimen-
tal data previously obtained to test two theoretical models which may be
used to predict evaporative rates of slightly soluble organic compounds
from water.  Billing (1977) found that the theoretical model by Mackay and
Wolkoff (1973) failed to predict evaporative half-lives  while the theore-
tical model of Mackay and Leinonen (1975), using  parameters from Liss and
Slater (1974), correlated well with experimental half-lives obtained.  For
example, the evaporative half-life obtained experimentally for 1,2-trans-
dichloroethene by Billing (1977) was approximately 24 minutes as compared
                                     51-5

-------
to 20.8 minutes predicted by the Mackay and Leinonen (1975) model and 0.85
minutes predicted by the Mackay and Wolkoff (1973) model.   Billing (1977),
however, comments that the apparent numerical agreement between his data
and the values predicted by the Mackay and Leinonen (1975) model may be
fortuitous.  Estimates of volatilization rates based on the Mackay and
Leinonen (1975) model depend primarily on liquid-gas phase exchange rate
constants, whereas the experimental model of Dilling et_ _al. (1975) and
Billing (1977) is controlled by the rate of stirring and the wind velocity
across the surface of the water.

          Pearson and McConnell (1975) suggest that the presence of an
analogue of 1,2-trans-dichloroethene, 1,1-dichloroethene,  as well as other
halogenated aliphatics, in ambient waters is due to the absorption of
chloroorganics from the atmosphere by water.  This process is thought to
occur most effectively when the atmosphere is scrubbed as  in the precipita-
tion process.  Aerial transport of these chloroorganics is indicated by
Pearson and McConnell (1975) to play a major role in their wider distribu-
tion and accounts for their presence in upland waters.

    51.4.5  Sorption

         No information was found pertaining specifically  to the adsorption
of 1,2-trans-dichloroethene onto sediments.  Dilling _et _al. (1975) carried
out two closed system experiments on trichloroethene and tetrachloroethene
(analogues of 1,2-trans-dichloroethene) and other chloroaliphatics, where
solute loss could only be by adsorption.  Dry bentonite clay at 375 mg/1
was introduced into a sealed solution and after ten minutes there was
approximately a ten percent adsorption by the bentonite clay compared to a
blank.  When the amount of clay added to the closed system was doubled (750
mg/1) there was 22 percent solute adsorption after 30 minutes.  There was
no further solute adsorption after this time.  The authors indicated that
there appeared to be little selectivity among the various  chlorinated com-
pounds in the adsorption process.  The authors also observed some adsorp-
tion of trichloroethene and tetrachloroethene (and other chloroaliphatics)
by dry powdered dolomitic limestone, but, again, without selectivity among
the solutes.

         In the sealed system with approximately 500 mg/1  peat moss,
approximately 40 percent of tetrachloroethene and trichloroethene was re-
moved via adsorption in 10 minutes.  At longer times, no further solute re-
moval was noted.  Pearson and McConnell (1975) found no clear evidence of
selective concentration of tetrachloroethene and trichloroethene, analogues
of 1,2-trans-dichloroethene, onto suspended solids or sediments.
                                    51-6

-------
    51.4.6  Bioaccumulation

         According to Kopperraan et al (1976), not all organochlorine com-
pounds bioaccumulate to high levels.  The data suggest that polar com-
pounds are more easily biodegraded, and the non-polar (highly lipophilic)
compounds accumulate.  Neely et al. (1974) have shown that bioaccumulation
is related to the octanol/water partition coefficient (P) of the compound.
The log octanol/water partition coefficient (log P) of 1.48 as calculated
by the method of Hansch (Tute 1971) indicates that 1,2-trans-dichloro-
ethene will probably not bioaccumulate to any significant extent (see
Methods section on bioaccumulation).

    51.4.7  Biotransformation and Biodegradation

         No information was found pertaining specifically to the rate of
biodegradation of 1,2-trans-dichloroethene in aquatic systems.  According
to Pearson and McConnell (1975) only completely sealed systems, such as the
standard BOD (biochemical oxygen demand) bottle technique, can be used to
measure biochemical degradation of volatile compounds such as trans-di-
chloroethene.  The BOD bottle experiments of Pearson and McConnell have
been unable to demonstrate any significant oxygen absorption from compounds
containing only C, H, and Cl, thus indicating that biochemical degradation
of such compounds is indeed very slow.  Literature references to microbial
biodegradation are few and conflicting; the majority find that low mole-
cular weight chloroaliphatics are not metabolized (Pearson and McConnell
1975; McConnell et, al. 1975).

51.5  Data Summary

    Table 51-1 summarizes the aquatic fate information discussed above.
The oxidation lifetime presented is a photooxidation lifetime and refers to
the rate of reaction of 1,2-trans-dichloroethene with hydroxyl radicals an
estimated from data on reactivity of tetrachloroethene and trichloroethene
with hydroxyl radicals in the troposphere.

    Due to the relatively high vapor pressure of 1,2-trans-dichloroethene,
volatilization from the aquatic system to the atmosphere is quite rapid.
Once in the troposphere, the compound is attacked by hydroxyl radicals at
the double bond, resulting in the subsequent formation of formic acid,
hydrochloric acid, carbon monoxide, and formaldehyde as reported products.
The tropospheric lifetime of 1,2-trans-dichloroethene is expected to be
somewhat less than that of its analogues, trichloroethene and tetrachloro-
ethene (less than 1 day and 10 days, respectively), since 1,2-trans-
dichloroethene has fewer chlorines and would be expected to be attacked
more readily by hydroxyl radicals than these two analogues.  Due to the
relatively high extrapolated reactivity of 1,2-trans-dichloroethene with
hydroxyl radicals in the troposphere, it is unlikely that unreacted
1,2-trans-dichloroethene will reach the stratosphere.
                                    51-7

-------
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-------
51.6  Literature Cited

Altshuller, A.P. and J.J. Bufalini 1971.  Photochemical aspects of
  pollution: A review.  Environ. Sci. Technol. 5:39-64.

Billing, W.L.  1977.  Interphase transfer processes.  II.  Evaporation
  rates of chloromethanes, ethanes, ethylenes, propanes, and propylenes
  from dilute aqueous solutions.  Comparisons with theoretical predictions.
  Environ. Sci. Technol.  11:405-409.

Dilling. W.L., C.J. Bredeweg, and N.B. Tefertiller. 1976,  Simulated
  atmospheric photodecomposition rates of raethylene chloride,
  1,1,1-trichloroethane, trichloroethylene, tetrachloroethylene, and other
  compounds.  Environ. Sci. Technol. 10:351-356.

Dilling, W.L., N.B. Tefertiller, and G.J. Kallos.  1975.  Evaporation rates
  of methylene chloride, chloroform, 1,1,1-trichloroethane,
  trichloroethylene, tetrachloroethylene and other chlorinated compounds in
  dilute aqueous solutions.  Environ. Sci. Technol. 9(9) :833-838.

Environmental Protection Agency. 1975a.  Preliminary assessment of
  suspected carcinogens in drinking water.  U.S. Environmental Protection
  Agency,(Office of Toxic Substances), Washington, D.C. 33p.
  EPA-560/4-75-003.

Environmental Protection Agency.  1975b.  Report on the problem of
  halogenated air pollutants and stratospheric ozone.  U.S. Environmental
  Protection Agency, (Office of Research and Development), Research
  Triangle Park, North Carolina.  55p. EPA 600/9-75-008.

Gay, B.W., Jr., P.L. Hanst, J.J. Bufalini, and R.C. Noonan. 1976.
  Atmospheric oxidation of chlorinated ethylenes.  Environ. Sci. Technol.
  10:58-67.

Jaffe, H.H. and M. Orchin.  1962.  Theory and applications of ultraviolet
  spectroscopy.  John Wiley and Sons, Inc. New York.  624p.

Jensen, S. and R. Rosenberg.  1975.  Degradability of some chlorinated
  aliphatic hydrocarbons in sea water and sterilized water.  Water Res.
  9:659-661.

Kopperman, H.L., D.W. Kuehl, G.E. Glass. 1976.  Chlorinated compounds
  found in waste treatment effluents and their capacity to bioaccumulate.
  Proceedings of the conference on the environmental impact of water
  chlorination.  pp.327-345.  Oak Ridge, Tennessee, October 22-24, 1975.
                                     51-9

-------
Liss, P.S. and P.G. Slater.  1974.  Flux of gases across the air-sea
  interface.  Nature 247:181-184.

Mackay, D. and P.J. Leinonen.  1975.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. Technol.
  9:1178-1180.

Mackay, D. and A.W. Wolkoff.  1973.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. Technol.
  7:611-614.

McConnell, G., D.M. Ferguson, and C.R. Pearson.  1975.  Chlorinated
  hydrocarbons and the environment.  Endeavor XXXIV:13-18.

Neely, W.B., D.R. Branson, G.E. Blau.  1974.  Partition coefficient to
  measure bioconcentration potential of organic chemicals in fish.
  Environ. Sci. Technol.  8:1113-1115.

Pearson, C.R. and G. McConnell, 1975.  Chlorinated QI> and C2
  hydrocarbons in the marine environment.  Proc. Roy. Soc. London B
  189:305-322.

Tute, M.S. 1971.  Principles and practice of Hansch analysis.  A guide to
  structure-activity correlation for the medicinal chemist.  Adv. Drug
  Res. 6:10-77.

Verschueren, K. 1977.  Handbook of environmental data on organic chemicals.
  Van Nostrand/Reinhold Press, New York.  659p.

Weast, R.C., (ed.).  1977.  Handbook of chemistry and physics.  58th
  Edition.  CRC Press Inc., Cleveland, Ohio.  2398p.

Yung, Y.L., M.B. McElroy, and S.C. Wofsy.  1975.  Atmospheric halocarbons:
  A discussion with emphasis on chloroform.  Geophys. Res. Lett.
  2(9):397-399.
                                      51-10

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                           52.   TRICHLOROETHENE
52.1  Statement of Probable Fate

    Volatilization appears to be the primary transport process for removal
of trichloroethene from aquatic systems.  Once the compound enters the
troposphere, hydroxyl radicals attack the double bond resulting in the sub-
sequent formation of dichloroacetyl chloride and phosgene as the principal
initial products.  The tropospheric lifetime (time for reduction to 1/e of
initial concentration) of trichloroethene based on its rate of reaction
with hydroxyl radicals is reported to be about 4 days.  Due to the rela-
tively high reactivity of trichloroethene with hydroxyl radicals in the
troposphere, it is unlikely that unreacted trichloroethene will diffuse
upward to the stratosphere.  The photooxidation products of trichloroethene
are readily hydrolyzed.

    Based on the information found, it does not appear that direct photo-
dissociation of carbon-carbon or carbon-chlorine bonds contributes to the
fate of trichloroethene in the aquatic or the atmospheric environment.
This is a result of the relatively rapid volatilization from aquatic sys-
tems and the rapid rate of hydroxyl radical attack in the troposphere.

    Oxidation in the aquatic environment does not appear to be a signifi-
cant fate process, although there is evidence of some oxidation of tri-
chloroethene in aqueous, closed systems in the presence of sunlight.  In
addition, hydrolysis, another potential fate process for compounds in the
aquatic environment, probably does not occur at a sufficient rate to be
important for trichloroethene.  Finally, no information was found indi-
cating that microorganisms exist which can readily degrade trichloroethene.
Some information was found, however, reporting that this compound is meta-
bolized by higher organisms leading to chlorinated acetic acid products.
Chlorinated acetic acids, in turn, have been shown to be susceptible to
degradation by microorganisms in seawater.

    Although volatilization appears to be the most significant transport
process of trichloroethene, there is some evidence of bioaccumulation of
trichloroethene in marine organisms.  There is, however, no evidence for
biomagnification in aquatic food chains.  In addition, no evidence was
found of selective concentrations of trichloroethene in marine sediments,
thus indicating that adsorption may not be an important transport process.

52.2  Identification

    Trichloroethene is known to be ubiquitous in the environment; it has
been detected in finished drinking water (Environmental Protection Agency
                                     52-1

-------
1975a) in drinking water supplies, in marine water, in rain water, food,
human tissues (Pearson and McConnell 1975;  McConnell _e_t _al. 1975), in the
atmosphere (Pearson and McConnell 1975;  McConnell et al. 1975; Singh _et
al. 1978), and in marine organisms (Pearson and McConnell 1975; Dickson and
Riley 1976).

    The chemical structure of trichloroethene is shown below.

                                      Alternate Name
    Cl            C,

        c - c                       Ethylene trichloride
       /        S.                     Trichloroethylene
    Ci            H                   Ethinyl trichloride
                                        (plus numerous commercial names)
                                      Tri-Clene
    Trichloroethene

    CAS NO. 79-01-6
    TSL NO. KX 45500

52.3  Physical Properties

    The general physical properties of trichloroethene are given below.

    Molecular weight                         131.39
    (Weast 1977)

    Melting point                            -73C
    (Weast 1977)

    Boiling point at 760 torr                87C
    (Weast 1977)

    Vapor pressure at 20C                   57.9 torr
    (Pearson and McConnell 1975)

    Solubility in water at 20C              1100 mg/1
    (Pearson and McConnell 1975)

    Log octanol/water partition coefficient  2.29
    (Leo et al. 1971)
                                    52-2

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52.4  Summary of Fate Data

    52.4.1  Photolysis

         Little information was found pertaining specifically to the rate
of direct photolysis of trichloroethene in the aqueous environment under
ambient conditions.  Jensen and Rosenberg (1975), using a closed system
where no evaporation could occur, found no significant decrease in tri-
chloroethene in the closed system in either daylight or darkness, thus
indicating that no photolysis occurred.

         Trichloroethene is quite volatile and, as a result, is rapidly
transported to the troposphere.  Photodissociation in the lower troposphere
would not be expected to occur for trichloroethene since this compound has
no chromophores which absorb in the visible or near ultraviolet region of
the electromagnetic spectrum (Jaffa and Orchin 1962).  Trichloroethene,
however, undergoes rapid photooxidation in the troposphere;  photooxidation
apparently is so rapid that trichloroethene never reaches the stratosphere.

    52.4.2  Oxidation

         Billing et_ _al. (1975) found possible evidence of photooxidation of
trichloroethene in aqueous media in the presence of sunlight.  Billing et
al. (1975) conducted reactivity experiments in water where there was a
sixfold molar excess of dissolved oxygen compared to the total amount of
trichloroethene and other solutes (all chloroaliphatics).  In addition, the
air space above the solution in the closed system contained approximately
90 times as much oxygen as was present in the solution.  Thus, there was a
great excess of oxygen present which would have been available for complete
oxidation of all the chlorinated compounds present.

         The amount of trichloroethene remaining in aerated water stored in
the dark for 12 months was approximately 20 percent greater than the amount
of trichloroethene remaining in aerated water stored in the presence of
sunlight. The fact that sunlight had a significant effect on the re-
activity of trichloroethene is consistent with predictions based on vapor
phase photolysis studies which showed that trichloroethene disappeared when
irradiated with long wavelength light in the presence of nitric oxide or
nitrogen dioxide (Altshuller and Bufalini 1971).  Billing &t_ _al. (1975)
attribute the more rapid disappearance of trichloroethene in the presence
of sunlight to free radical oxidation.

         Trichloroethene is quite volatile and, as a result, is rapidly
transported to the troposphere.  Once the compound enters the troposphere,
hydroxyl radicals attack the double bond, resulting in the subsequent
formation of dichloroacetyl chloride and phosgene as the principal initial
products (Environmental Protection Agency 1975c;  Gay _e_t al. 1976;  Hanst
                                    52-3

-------
1978).  These compounds are rapidly hydrolyzed at ambient conditions, re-
sulting in HC1, CO, CC>2 and the carboxylic acid corresponding to the acid
chloride (Morrison and Boyd 1973).

         Yung _et _al. (1975) report the tropospheric lifetime (time for re-
duction to 1/e of initial concentration) of trichloroethene as 3.7 x 10^
seconds, or about 4 days, based on its rate of reaction with hydroxyl
radicals.   The bimolecular rate constant for this reaction is reported to
be 3 x ICT^cm^sec"1 (Yung _et al. 1975).  This rate agrees closely
with the experimental data reported by Howard (1976) indicating a biraole-
cular rate for the reaction of approximately 2 x 10    cm^ molecule"^-
sec"-'-.  The Environmental Protection Agency (1975c) has reported a half-
life for trichloroethene-of less than 1 day.  Lillian _et _al. (1975) contend
that the relatively low frequency of detection of trichloroethene with
respect to other halogenated aliphatics is probably due to a relatively
small source strength and a relatively high tropospheric reactivity.  Due
to the relatively high reactivity of trichloroethene with hydroxyl radicals
in the troposphere, it appears unlikely that unreacted trichloroethene will
reach the  stratosphere.

    52.4.3  Hydrolysis

         In citing the works of other authors, Dilling _et al. (1975) re-
ported that trichloroethene resisted hydrolysis at 100C.  Hydrolysis of
trichloroethene was reported to be accelerated by the presence of oxygen,
but the products of hydrolysis from dilute solution have not been reported.
Laboratory experiments by Dilling _ejt _al. (1975) were conducted in closed
systems in the dark to eliminate competing processes such as volatilization
and photolysis.  The observed experimental half-life, which may be regarded
as resulting from oxidation as well as hydrolysis, was about 10.7 months at
25C, corresponding to a first-order rate of 0.065 months"!  (Dilling et_
al. 1975).  The relative contributions of oxidation and hydrolysis are not
known; thus, the rate, at best, is a maximum rate for hydrolysis. According
to the Environmental Protection Agency (1975b), trichloroethene is not hy-
drolyzed by water under normal conditions.

    52.4.4  Volatilization

         Dilling _e_t al. (1975) reported the experimental half-life with re-
spect to volatilization of 1 mg/1 trichloroethene from water to be about 21
minutes (average of three) when stirred at 200 rpm at approximately 25C in
an open container.  Removal of 90 percent of the trichloroethene required
about 69 minutes (average of three samples).  When intermittent stirring of
15 seconds duration was provided every five minutes, the time required for
50 percent depletion of trichloroethene was greater than 90 minutes.  This
rate is on the order of four times slower than evaporation with constant
stirring.   The presence of sodium chloride at a 3 percent concentration, as
in seawater, caused about a 10 percent decrease in the rate of evaporation.
                                    52-4

-------
         Evaporation appears to be the major pathway by which trichloro-
ethene is lost from water.  Billing ^e_t _al. (1975) are careful to point out
the difficulties encountered in extrapolating their laboratory results to
real-world conditions, where the concentration of the organic solute would
probably be very much less than 1 mg/1 and where surface and bulk agitation
would be highly variable.  Although the data appear to be valid on a re-
lative basis (i.e., correctly illustrating the relative rates of volatili-
zation of chlorinated aliphatics), they cannot be used as absolute measures
of volatilization rates from natural waters.  For the purposes of this do-
cument, the data are used as rough-order-of-magnitude indications of the
importance of volatilization relative to other transport and fate proceses,
with the strong effects of agitation considered.  The validity of this
application has not been.established.

         A subsequent study by Billing (1977) was conducted using the same
experimental conditions as in the 1975 study, and an evaporative half-life
range for trichloroethene of 17.7 to 23.5 minutes was obtained for 0.94 to
0.96 mg/1.  The purpose of this subsequent study was to use the experimen-
tal data obtained to test two theoretical models formulated to predict
rates of evaporation of slightly soluble organic compounds from water.
Billing (1977) found that the theoretical model by Mackay and Wolkoff
(1973) failed to predict evaporative half-lives, but that evaporative
half-lives predicted by the model of Mackay and Leinonen (1975), using
parameters of Liss and Slater (1974), correlated well with experimental
data.  For example, the evaporative half-life of 23.8 minutes for tri-
chloroethene predicted by the Mackay and Leinonen model is very close to
the range of 17.7 to 23.5 minutes determined by Billing (1977), in contrast
to the much lower half-life of 0.47 minutes predicted by the Mackay and
Wolkoff (1973) model.

         Billing (1977), however, comments that the apparent numerical
agreement between his data and the values predicted by the Mackay and
Leinonen (1975) model may be fortuitous.  Estimates of volatilization rates
based on the Mackay and Leinonen (1975) model depend primarily on liquid-
gas phase exchange rate constants, whereas the experimental model of
Billing _e_t _al. (1975) and Billing (1977) is controlled by the rate of
stirring and the wind velocity across the surface of the water.

         In a study by Pearson and McConnell (1975), rainwater collected in
Runcorn, England contained up to 1.5 x 10~10 parts (by mass) of trichloro-
ethene.  Municipal waters supplied to the cities of Liverpool, Chester, and
Manchester (all supplied from upland surface sources) contained up to 6 x
10~9 parts (by mass) trichloroethene.  While entry to the environment
appears to be predominantly evaporative, there may also be some chloro-
organics in aqueous industrial effluents which will pass into municipal
drainage systems and rivers.  Mackay and Wolkoff (1973) have shown that the
evaporation rate of poorly soluble species from water can be quite high due
                                    52-5

-------
to Cheir high activity coefficients.  Pearson and McConnell (1975) and
McConnell _et al^. (1975) point out that this is a reversible process since
water will absorb chloroorganics from the atmosphere, a process which will
occur most effectively when the atmosphere is scrubbed during periods of
rainfall.  The presence of trace organochlorines, including trichloro-
ethene, in upland waters is believed to be due to aerial transport (Pearson
and McConnell 1975;  McConnell _et al. 1975).

         A study by Jensen and Rosenberg (1975) indicates that the rate of
volatilization of trichloroethene proceeds more rapidly than photolysis,
oxidation, or hydrolysis.  An initial concentration between 0.1 to 1 mg/1
trichloroethene in a 20 liter volume of water was found to decrease 80 per-
cent after 8 days in an aquarium kept in light and partly open.  Tetra-
chloroethene initially present at the same concentrations in a closed sys-
tem exposed to daylight and a closed system exposed to dark showed less
than about a 5 percent disappearance in 8 days.

         Assuming that all processes involved are first order, the data of
Jensen and Rosenberg (1975) indicate the evaporative half-life for tri-
chloroethene in the partly open (and supposedly quiescent) aquarium to be
3.44 days, which is apparently comparable to the data of Billing et al.
(1975), indicating a 21 minute half-life under continuous stirring at 200
rpm and over 90 minutes under discontinuous stirring (15 seconds per five
minutes).  These results illustrate the marked effect of system agitation
on volatilization rate from water for compounds such as trichloroethene.

    52.4.5  Sorption

         Billing _e_t _al. (1975) carried out two closed system experiments
where solute loss could only be by adsorption.  Bry bentonite clay at 375
rag/1 was introduced into a sealed solution, and, in ten minutes, there was
approximately a ten percent adsorption of trichloroethene, as well as four
other chlorinated compounds, by the clay compared to a blank.  When the
amount of clay added to the closed system was doubled (750 mg/1) there was
22 percent solute adsorption after 30 minutes.  There was no further solute
adsorption after this  time.  The authors indicated that there appeared to
be little selectivity  among the various chlorinated compounds in the
adsorption process.  The authors observed some adsorption of trichloro-
ethene by dry powdered dolomitic limestone, but, again, without selectivity
among the solutes used.

         In the sealed system with approximately 500 mg/1 peat moss,
approximately 40 percent of the trichloroethene was adsorbed in 10 minutes.
At longer times, no further solute removal was noted.  The subsequent de-
crease in the rate of  disappearance of trichloroethene at longer time peri-
ods than 10 minutes was suggested to be due to gradual release or desorp-
tion of  trichloroethene from  the peat moss to  the solution.
                                    52-6

-------
         Pearson and McConnell (1975) found that the concentration of tri-
chloroethene in Liverpool Bay (England) sediments varied from a few parts
      I >-\             *            *-'         	Q
per 10   (by mass) to a maximum of 9.9 x 10  .  Samples of marine sedi-
ment from Liverpool Bay contained the same compounds as the overlying
waters,  but there was no correlation between the chloroorganic content of
the sediment and that of the raid-depth water from the same point on the
sampling grid (Pearson and McConnell 1975;   McConnell et_ al. 1975).  In
addition, Pearson and McConnell (1975) reported that there was no apparent
relation between high concentrations and geographical features, as was
noticed in the case of the water samples, nor was there any correlation
between concentration in the sediment and particle size.  In general, there
was no clear evidence of selective concentration of trichloroethene in
sediments.  McConnell et- al. (1975), however, did note that coarse gravels
have little adsorptive capacity for trichloroethene, whereas sediments rich
in organic detritus have a much higher adsorptive capacity.  They also
noted that, averaged over all samples, the concentrations of chlorinated
(-'l/(-'2 compounds (including trichloroethene) in the sediments were
similar to those in the water.

    52.4.6  Bioaccumulation

         According to Kopperman j_t _a_l. (1976) not all organochlorine com-
pounds bioaccumulate to high levels.  The data suggest that polar com-
pounds are more easily biodegraded, and the non-polar (highly lipophilic)
compounds accumulate.  In addition, Kopperman et al. (1976) point out that
the incorporation of chlorine compounds during the disinfection of waste
effluents with chlorine is an undesirable end result of effluent treatment
in that chlorinated compounds become more persistent and bioaccumulate to a
greater extent.  In this study it was found that the concentration of tri-
chloroethene in water from sewage treatment plants was 40.2 yg/1 in influ-
ent before treatment.  The effluent before chlorination contained 8.6 jjg/1,
whereas the effluent after chlorination contained 9.8 yg/1 trichloroethene.
Neely _e_t _al. (1974) have shown that bioaccumulation is directly related to
the octanol/water partition coefficient (P) of the compound.  The log
octanol/water partition coefficient (log P) of 2.29 for trichloroethene as
determined by Leo _e_t _al. (1971) indicates that bioaccumulation is possible
for this compound, but the process is probably not important in comparison
to volatilization as a removal mechanism.

         Pearson and McConnell (1975) determined the level of trichloro-
ethene and other chlorinated hydrocarbons in the tissues of a wide range of
organisms.  Species were chosen to represent significant trophic levels in
the marine environment.  Due to the occurrence of high levels of PCB and
DDT in fish-eating birds, special attention has been paid to marine and
brackish water birds.  The maximum overall increase in concentration be-
                                    52-7

-------
tween seawater and the tissues of animals at the top of food chains (such
as fish liver, sea bird eggs, and sea seal blubber) was less than 100-fold
for trichloroethene (from 0.5 x 10~* in water to 50 x 10~9 in tissues).

         Although evidence for weak to moderate bioaccumulation of tri-
chloroethene by marine organisms exists (Pearson and McConnell 1975), there
is no evidence for biomagnification of trichloroethene in aquatic food
chains.

         McConnell t _a_l. (1975) report that there is evidence of the pre-
sence of trichloroethene in human tissue at extremely low concentrations.
It is indicated, however, that there is no evidence for significant accumu-
lation in human tissue.  According to McConnell _e_t _al. (1975), these data
indicate a general background, at the yg/1 level, which pervades the
atmosphere, hydrosphere, and biosphere.
    52.4.7  Biotransformation and Biodegradation

         No information was found pertaining specifically to the rate of
biodegradation of trichloroethene in aquatic systems.  Thorn and Agg (1975)
have included trichloroethene in a list of synthetic organic chemicals
which should be degradable by biological sewage treatment provided suitable
acclimatization can be achieved.  They note, however, that not many com-
pounds in this list .occur free in nature and, as a result, it is unlikely
that microorganisms already possess the ability to destroy them.  The
authors consider trichloroethene, along with various other compounds, to be
potentially biodegradable.

         According to Pearson and McConnell (1975) only completely sealed
systems, such as the standard BOD (biochemical oxygen demand) bottle tech-
nique, can be used to measure biochemical degradation of volatile compounds
such as trichloroethene.  The BOD bottle experiments of Pearson and McCon-
nell have been unable to demonstrate any significant oxygen absorption from
compounds containing only C, H, and Cl, thus indicating that biochemical
degradation of such compounds is indeed very slow.

         There is some evidence from mammal studies  that trichloroethene
can be metabolized by higher organisms (Pearson and McConnell 1975;
McConnell et_ _a_l. 1975).  Metabolism of trichloroethene results in the
formation of dichloroacetic acid (Pearson and McConnell 1975).

         Literature references to microbial biodegradation of compounds
such as trichloroethene are few and conflicting;  the majority report that
low molecular weight chloroaliphatics are not metabolized (Pearson and
McConnell 1975;  McConnell et al. 1975).  In mammals, the metabolic path-
                                      52-8

-------
ways of trichloroethene lead to chlorinated acetic acids, either directly
or via chloroethanols.  Chlorinated acetic acids have all been shown to be
susceptible to further degradation by microorganisms in seawater (Pearson
and McConnell 1975).  McConnell et_ _al. (1975) conclude that the simple
chloroaliphatic compounds are not particularly persistent, and that their
degradation products are simple species commonly found in the environment.

52.5  Data Summary

    Table 52-1 summarizes the aquatic fate information discussed above.
The oxidation rate presented is a photooxidation rate and refers to the
rate of reaction of trichloroethene with hydroxyl radicals in the
troposphere.

    Due to the relatively high vapor pressure of trichloroethene, volatili-
zation from the aquatic system to the atmosphere is quite rapid.  Once in
the troposphere, hydroxyl radicals attack the double bond, resulting in the
subsequent formation of dichloroacetyl chloride and phosgene as the princi-
pal products.  The tropospheric lifetime of trichloroethene based on its
rate of reaction with hydroxyl radicals is reported to be 3.7 x 10^
seconds, corresponding to a lifetime of about 4 days.  Due to the rela-
tively high reactivity of trichloroethene with hydroxyl radicals in the
troposphere, it is unlikely that unreacted trichloroethene will reach the
stratosphere.
                                    52-9

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52.6  Literature Cited

Altshuller, A.P. and J.J. Bufalini.  1971.  Photochemical aspects of
  pollution:  a review.  Environ. Sci. Technol. 5:39-64.

Dickson, A.G. and J.P. Riley.  1976.  The distribution of short-chain
  halogenated aliphatic hydrocarbons in some marine organisms.  Mar.
  Pollut. Bull. 7(9):167-169.

Billing W.L.  1977.  Interphase transfer processes.  II.  Evaporation rates
  of chloromethanes, ethanes, ethylenes, propanes, and propylenes from
  dilute aqueous solutions.  Comparisons with theoretical predictions.
  Environ. Sci. Technol. 11:405-409.

Billing, W.L., N.B. Tefertiller, and G.J. Kallos.  1975.  Evaporation rates
  of methylene chloride, chloroform, 1,1,1-trichloroethane,
  trichloroethylene, tetrachloroethylene and other chlorinated compounds in
  dilute aqueous solutions.  Environ. Sci. Technol. 9(9):833-838.

Environmental Protection Agency.  1975a.  Preliminary assessment of
  suspected carcinogens in drinking water.  Environmental Protection
  Agency, (Office of Toxic Substances), Washington, D.C.  33p.  EPA
  560/4-75-003.

Environmental Protection Agency.  1975b.  Preliminary study of selected
  potential environmental contaminants - optical brighteners, methyl
  chloroform, trichloroethylene, tetrachloroethylene, ion exchange resins.
  Environmental Protection Agency, (Office of Toxic Substances),
  Washington,   D.C.  286p.  EPA 560/2-75-002.

Environmental Protection Agency.  1975c.  Report on the problem of
  halogenated air pollutants and stratospheric ozone.  Environmental
  Protection Agency, (Office of Research and Development), Research
  Triangle Park, North Carolina.  55p.  EPA 600/9-75-008.

Gay, B.W., Jr., P.L. Hanst, J.J. Bufalini, and R.C. Noonan.  1976.
  Atmospheric oxidation of chlorinated ethylenes.  Environ. Sci. Technol.
  10:58-67.

Hanst, P.L.  1978.  Part II:  Halogenated pollutants.  Noxious trace gases
  in the air.  Chemistry 51(2):6-12.

Howard, C.J.  1976.  Rate constants for the gas-phase reactions of OH
  radical with ethylene and halogenated ethylene compounds.  J. Chem.
  Phys.  65(ll):4771-4777.
                                    52-11

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Jaffe, H.H. and M. Orchin.  1962.  Theory and applications of ultraviolet
  spectroscopy.  John Wiley and Sons, Inc.  New York.  624p.

Jensen, S. and R. Rosenberg.  1975.  Degradability of some chlorinated
  aliphatic hydrocarbons in seawater and sterilized water.  Water Res.
  9:659-661.

Koppennan, H.L., D.W. Kuehl, and G.E. Glass.  1976.  Chlorinated compounds
  found in waste treatment effluents and their capacity to bioaccumulate.
  Proceedings of the conference on the environmental impact of water
  chlorination.  pp.327-345.  Oak Ridge, Tennessee, October 22-24, 1975.

Leo, A., C. Hansch, and D. Elkins.  1971.  Partition coefficients and their
  uses.  Chem. Rev. 71:525-616.

Lillian, D., H.B. Singh, A. Appleby, L. Lobban, R. Arnts, R. Gumpert, R.
  Hague, J. Toomey, J. Kazazis, M. Antell, D. Hansen and B. Scott.  1975.
  Atmospheric fates of halogenated compounds.  Environ. Sci. Technol.
  9:1042-1048.

Liss, P.S. and P.G. Slater.  1974.  Flux of gases across the air-sea
  interface.  Nature 247:181-184.

Mackay, D. and P.J. Leinonen.  1975.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. Technol.
  9:1178-1180.

Mackay, D. and A.W. Wolkoff.  1973.  Rate of evaportion of low-solubility
  contaminants from water bodies- to atmosphere.  Environ. Sci. Technol.
  7:611-614.

McConnell, G., D.M. Ferguson, and C.R. Pearson.  1975.  Chlorinated
  hydrocarbons and the environment.  Endeavor XXXIV:13-18.

Morrison, R.T and R.N. Boyd.  1973.  Organic chemistry.  3rd Edition.
  Allyn and Bacon, Inc., Boston, Mass.  1258p.

Neely, W.B., D.R. Branson, and G.E. Glau.  1974.  Partition coefficient to
  measure bioconcentration potential of organic chemicals in fish.
  Environ. Sci. Technol. 8:1113-1115.

Pearson, C.R. and G. McConnell.  1975.  Chlorinated C^ and 2
  hydrocarbons in the marine environment.  Proc. Roy. Soc. London B
  189:305-322.
                                   52-12

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Singh, H.B., L.J. Salas, H. Shiegeishi, and A.H. Smith.   1978.   Fate of
  halogenated compounds in the atmosphere.  Interim report-1977.
  Environmental Protection Agency, (Office of Research and Development),
  Research Triangle Park, North Carolina.  57p.  EPA 600/3-78-017.

Thorn, N.S. and A.R. Agg.  1975.  The breakdown of synthetic organic
  compounds in biological processes.  Proc. Roy. Soc. London B  189:347-357,

Weast, R.C. (ed.).  1977.  Handbook of chemistry and physics.  58th
  Edition.  CRC Press, Inc., Cleveland, Ohio.  2389p.

Yung, Y.L., M.B. McElroy, and S.C. Wofsy.  1975.  Atmospheric halocarbons:
  a discussion with emphasis on chloroform.  Geophys. Res. Lett.
  2(9):397-399.
                                   52-13

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                53.   TETRACHLOROETHENE (PERCHLOROETHYLENE)
53.1  Statement of Probable Fate

    Volatilization appears to be the major transport process for removal of
tetrachloroethene from aquatic systems.  Once in the troposphere, the com-
pound reportedly reacts with hydroxyl radicals which attack the double
bond, resulting in the formation of trichloroacetylchloride as the princi-
pal product, with phosgene being produced to a lesser extent.  The tropo-
spheric lifetime of tetrachloroethene based on its rate of reaction with
hydroxyl radicals is reported to be about 10 days.  On this basis, it is
unlikely that unreacted tetrachloroethene will diffuse upward to the
stratosphere.

    Based on the information found it does not appear that photodissocia-
tion is a significant fate of tetrachloroethene in the aquatic or the
atmospheric environment. Oxidation in the aquatic environment does not
appear to be significant, although there is evidence of some oxidation of
tetrachloroethene in aqueous, closed systems in the presence of sunlight.
In addition, hydrolysis, another potential fate process for compounds in
the aquatic environment, probably does not occur at a significant rate to
be important for tetrachloroethene.  Finally, no information was found
indicating that microorganisms exist which can readily biodegrade tetra-
chloroethene.  Some information, however, was found reporting this compound
to be metabolized by higher organisms leading to chlorinated acetic acid
products.  Chlorinated acetic acids, in turn, have been shown to be sus-
ceptible to degradation by microorganisms in seawater.

    Although volatilization appears to be the most significant transport
process, there is some evidence of bioaccumulation of tetrachloroethene in
marine organisms.  There is, however, no evidence for biomagnification in
aquatic food chains.  In addition, no evidence was found of selective con-
centration of tetrachloroethene in marine sediments, thus indicating that
adsorption is probably not an important transport process for this com-
pound.

53.2  Identification
      	                f

    This compound is known to be ubiquitous in the environment.  Tetra-
chloroethene has been detected in finished drinking water (Environ-
mental Protection Agency 1975a), in marine and rain water, in food, in hu-
man tissues, in the atmosphere (Pearson and McConnell 1975;  McConnell et
al. 1975;  Singh _e_t _a_l. 1978), and in marine organisms (Dickson and Riley
1976).
                                     53-1

-------
     The chemical structure of tetrachloroethene is shown below:
      Cl            Cl

                                      Alternate Names
        /
       '
      ~            C)                  Perchloroethylene
                                      Ethylene tetrachloride
                                      Tetrachloroethylene
     Tetrachloroethene                Perchloroethene

     CAS NO.  127-18-4
     TSL NO.  KX 38500

53.3  Physical Properties

    The general physical properties of tetrachloroethene are given below.

    Molecular weight                         165.83
    (Weast 1977)

    Melting point                            -22.7C
    (Verschueren 1977)

    Boiling point at 760 torr                121C
    (Weast 1977)

    Vapor pressure at 20C                   14 torr
    (Pearson and McConnell 1975)

    Solubility in water at 20C              150 to 200 mg/1*

    Log octanol/water partition coefficient  2.88
    (Neely et al. 1974)
*Several values for solubility of tetrachloroethene in water at 20C were
found in the literature.  These values range from 150 mg/1 (Pearson and
McConnell 1975) to 200 mg/1 (Chiou _e_t a|. 1977).

53.4  Summary of Fate Data

    53.4.1  Photolysis

         Relatively little information was found pertaining specifically to
the rate of photodissociation of tetrachloroethene in the aqueous environ-
                                     53-2

-------
ment under ambient conditions.  Jensen and Rosenberg (1975), using a closed
system where no evaporation could occur, found no significant difference in
loss of tetrachloroethene between daylight exposure and darkness, thus in-
dicating that no photolysis occurred.  These tests utilized both seawater
and deionized and boiled fresh water.

         Tetrachloroethene is quite volatile and, as a result, is rapidly
transported to the troposphere.  Tetrachloroethene is attacked by hydroxyl
radicals more slowly than most other olefin pollutants due to the presence
of four chlorine atoms (Environmental Protection Agency 1975c).  Accord-
ing to the Environmental Protection Agency (1975c), there is only a remote
possibility that reaction of tetrachloroethene with hydroxyl radicals is
sufficiently slow to allow a portion of the tetrachloroethene in the
troposphere to diffuse upward to the stratosphere.  Photodissociation in
the terrestrial environment would not be expected to occur for tetra-
chloroethene since this compound has no chromophores which absorb in the
visible or near ultraviolet region of the electromagnetic spectrum (Jaffe
and Orchin 1962).

    53.4.2  Oxidation

         The literature reviewed indicates that oxidation probably does not
play a significant role in the aquatic fate of tetrachloroethene.  Dilling
_e_t_ _al. (1975) conducted laboratory experiments in which there was about a
sixfold molar excess of dissolved oxygen compared to the total amount of
tetrachloroethene.  In addition, the air space above the solution in the
closed system contained approximately 90 times as much oxygen as was pre-
sent in the saturated solution.  Thus, there was a great excess of oxygen
present which would have been available for complete oxidation of all the
chlorinated compounds present.  Their results indicated that volatilization
was much more rapid than any oxidative process which may have occurred.

         Dilling t_ _a_l. (1975) did find evidence of significant photooxi-
dation of tetrachloroethene in aqueous media in the presence of sunlight.
The amount of tetrachloroethene remaining in aerated water in the dark
after 12 months was approximately 13 percent greater than the amount of
tetrachloroethene remaining in aerated water in the presence of sunlight.
The fact that sunlight had a significant effect on the reactivity of tetra-
chloroethene is consistent with predictions based on vapor phase photolysis
studies which showed that tetrachloroethene disappeared when irridiated
with long wavelength light in the presence of nitric oxide or nitrogen
dioxide (Altshuller and Eufalini 1971).  Dilling _e_t _a_l. (1975) attribute
the difference in disappearance of tetrachloroethene between samples
exposed to  sunlight and those held in the dark to oxidation by free
radicals.
                                    53-3

-------
         Tetrachloroethene is quite volatile and, as a result, is rapidly
transported to the troposphere.  Once in the troposphere, hydroxyl radicals
attack the double bond, resulting in the subsequent formation of trichloro-
acetylchloride as the principal initial product, with phosgene being pro-
duced to a lesser extent (Andersson et al. 1975; Hanst 1978; Environ-
mental Protection Agency 1975c;  Gay t _al_. 1976).  Both of these compounds
are readily hydrolyzed at ambient conditions (Morrison and Boyd 1973).
Tetrachloroethene, however, is attacked by hydroxyl radicals more slowly
than most other olefin pollutants due to the presence of four chlorine
atoms (Environmental Protection Agency 1975c).  According to Yung et al.
(1975) the tropospheric lifetime of tetrachloroethene based on its rate of
reaction with hydroxyl radicals is reported to be 8.5 x 10^ seconds,
corresponding to a lifetime of about ten days.  The bimolecular rate con-
stant for this reaction is reported to be 1.3 x 10~12Cm^sec~^ (Yung
_ejt l. 1975).  The experimental data of Howard (1976) indicate a bimole-
cular rate of about 1.7 x 10~^3 cm^ molecule"^ sec"^-.  Lillian et_
al. (1975) contend that the low concentrations of tetrachloroethene found
in the atmosphere serve as evidence for high tropospheric reactivity.

    53.4.4  Hydrolysis

         In citing the works of other authors, Billing _e_t _al. (1975) re-
ported that tetrachloroethene was unreactive hydrolytically at 150C in the
absence of oxygen.  The decomposition of tetrachloroethene, however, was
reported to be accelerated by the presence of oxygen.  Trichloroacetic and
hydrochloric acids have been reported as products.  Laboratory experiments
by Billing et al. (1975) were conducted in closed systems in the dark to
eliminate competing processes such as volatilization and photolysis.  The
experimental hydrolytic half-life obtained was about 8.8 months at 25C,
corresponding to a first-order rate of 0.079 months"! which should be
considered a maximum rate for hydrolysis.

         According to the Environmental Protection Agency (1975b) pure
tetrachloroethene in contact with water for long periods of time decomposes
to trichloroacetic acid and hydrogen chloride.

    53.4.4  Volatilization

         Billing jet _al. (1975) reported the experimental volatilization
half-life of 1 mg/1 tetrachloroethene in water to be 26 minutes (average of
three samples) when stirred at 200 rpm at approximately 25C in an open
container.  Removal of 90 percent of the tetrachloroethene required 83
minutes (average of three samples).  With intermittent stirring for 15
seconds every five minutes the time required for 50 percent depletion of
tetrachloroethene was greater than 90 minutes, which is on the order of
three times greater than that observed during constant stirring.   The
                                   53-4

-------
presence of sodium chloride at a 3 percent concentration, as in seawater,
caused about a 10 percent decrease in the chlorinated compound evaporation
rate.

         Evaporation appears to be the major pathway by which tetrachloro-
ethene is lost from water.   Dilling et_ al. (1975) are careful to point out
the difficulties encountered in extrapolating their laboratory results to
real-world conditions, where the concentration of the organic solute would
probably be very much less than 1 mg/1 and where surface and bulk agitation
would be highly variable.  Although the data appear to be valid on a rela-
tive basis (i.e., correctly illustrating the relative rates of volatiliza-
tion of chlorinated aliphatics) , they cannot be used as absolute measures
of volatilization rates from natural waters.  For the purposes of this
document, the data are used as rough-order-of-magnitude indications of the
importance of volatilization relative to other transport and fate proceses,
with the strong effects of agitation considered.  The validity of this
application has not been established.

         A subsequent study by Dilling (1977) was conducted using the same
experimental conditions as in the 1975 study, and an average half-life
range of 20.2 to 27.1 minutes was obtained for 0.92 mg/1 tetrachloroethene.
The purpose of this subsequent study was to use the experimental data ob-
tained to test two theoretical models formulated to predict evaporative
rates of slightly soluble organic compounds from water.  Dilling (1977)
found that the theoretical model of Mackay and Wolkoff (1973) failed to
predict evaporative half-lives but that the model of Mackay and Leinonen
(1975) using the parameters of Liss and Slater (1974) correlated well with
the experimental half-lives obtained by Dilling (1977).  For example, the
evaporative half-life obtained experimentally by Dilling (1977) was
approximately 20 to 27 minutes, compared to 26.5 minutes predicted by the
Mackay and Leinonen (1975) model and 0.20 minutes predicted by the Mackay
and Wolkoff (1973) model.  Dilling (1977), however, comments that the
apparent numerical agreement between his data and the values predicted by
the Mackay and Leinonen (1975) model may be fortuitous.  Estimates of
volatilization rates based on the Mackay and Leinonen (1975) model depend
primarily on liquid-gas phase exchange rate constants, whereas the experi-
mental model of Dilling _e_t ^al. (1975) and Dilling (1977) is controlled by
the rate of stirring and the wind velocity across the surface of the water.

         In a study by Pearson and McConnell (1975), rainwater collected in
Runcorn, England contained up to 1.5 x 10~^ (by mass) of tetrachloro-
ethene.  Municipal waters supplied to the cities of Liverpool, Chester, and
Manchester (all supplied from upland surface sources) contained up to 3.8 x
1Q-10 (by mass) tetrachloroethene.  While entry to the general environ-
ment appears to be predominantly evaporative, there may also be some
                                    53-5

-------
chloroorganics in aqueous industrial effluents which will pass into muni-
cipal drainage systems and rivers.   Mackay and Wolkoff (1973) have shown
that the evaporation rate of poorly soluble species from water can be quite
high due to their high activity coefficients.   Pearson and McConnell (1975)
and McConnell _ejt jtl^ (1975) point out that this is a reversible process
since water will absorb chloroorganics from the atmosphere, a process which
will occur most effectively when the atmosphere is scrubbed during periods
of rainfall.  The presence of trace organochlorines, including tetrachloro-
ethene, in upland waters is believed to be due to aerial transport (Pearson
and McConnell 1975;   McConnell _e_t _al. 1975).

         A study by Jensen and Rosenberg (1975) indicates that the rate of
volatilization of tetrachloroethene proceeds more rapidly than photolysis,
oxidation, or hydrolysis.  An initial concentration between 0.1 to 1 mg/1
tetrachloroethene in 20 liters of water was found to decrease 50 percent
after 8 days in a partially open, well lit aquarium.  Tetrachloroethene
initially present at the same concentration in closed systems either
exposed to daylight or held in the  dark was reduced only 5 percent after 8
days demonstrating that volatilization was more rapid than photolysis.

    53.4.5  Sorption

         Billing j^t _al. (1975) carried out two closed system experiments
where solute loss could be only by adsorption.  Dry bentonite clay at 375
mg/1 was introduced into a sample,  and in ten minutes there was
approximately a ten percent adsorption of tetrachloroethene, as well as
other chlorinated compounds, by the clay (compared to a blank).  When the
amount of clay added to the closed  system was doubled (750 mg/1) there was
22 percent solute loss via adsorption after 30 minutes.  There was no
further solute adsorption after this time.  The authors indicated that
there appeared to be little selectivity among the various chlorinated
compounds in the adsorption process.

         The authors observed some  adsorption of tetrachloroethene by dry
powdered dolomitic limestone, but,  again, without any significant selec-
tivity among haloaliphatic species.

         In the sealed system to which had been added approximately 500
mg/1 peat moss, approximately 40 percent of the tetrachloroethene was lost
via adsorption within 10 minutes.  At longer times, no further solute re-
moval was noted.

         Pearson and McConnell (1975) found that the concentrations of
tetrachloroethene in Liverpool Bay sediments varied from a few parts per
10^2 (by mass) to a maximum of 4.8 x 10"^.  Samples of marine sediment
from Liverpool Bay contained the same compounds as the overlying waters,
                                   53-6

-------
but there was no correlation between the chloroorganic content of the sedi-
ment and that of the mid-depth water from the same point on the sampling
grid (Pearson and McConnell 1975;  McConnell et al. 1975).  In addition,
Pearson and McConnell (1975) reported that there was no apparent relation
between high concentrations and geographical features, as was noticed in
the case of the water samples, nor was there any correlation between con-
centration in the sediment and particle size.  In general, there was no
clear evidence of selective concentration of tetrachloroethene in sedi-
ments.  McConnell et_ al_. (1975), however, did note that coarse gravels have
little adsorptive capacity for tetrachloroethene, whereas sediments rich in
organic detritus have a much higher adsorptive capacity.  They also note
that, averaged over all samples, the concentrations of chlorinated
CI/GO compounds (including tetrachloroethene) in the sediments were
similar to those in the water.

    53.4.6  Bioaccumulation

         According to Kopperman _e_t _al. (1976) not all organochlorine com-
pounds bioaccumulate to high levels.  The data suggest that polar compounds
are more easily biodegraded, and the non-polar (highly lipophilic) com-
pounds accumulate.  In addition, Kopperman _et_ _al. (1976) point out that the
formation of chlorine compounds during the disinfection of waste effluents
with chlorine is an undesirable end result of effluent treatment in that
chlorinated compounds become more persistent and bioaccumulate to a greater
extent.  In this study it was found that the concentration of tetrachloro-
ethene in water from sewage treatment plants was 6.2 Ug/1 in influent be-
fore treatment.  The effluent before chlorination contained 3.9 ug/1
tetrachloroethene, whereas the effluent after chlorination contained 4.2
Ug/1 tetrachloroethene.

         Neely _ejt _a!L. (1974) have shown that bioaccumulation is directly
related to the octanol/water partition coefficient (P) of the compound.
The log octanol/water partition coefficient (log P) of 2.88 for tetra-
chloroethene as determined by Neely _e_t _al_. (1974) indicates that this com-
pound may have the potential to bioaccumulate.

         Pearson and McConnell (1975) determined the level of tetrachloro
ethene and other chlorinated hydrocarbons in the tissues of a wide range of
organisms.  Species were chosen to represent significant trophic levels in
the marine environment.  Due to the occurrence of high levels of PCS and
DDT in fish-eating birds, special attention was paid to marine and brackish
water birds.  The maximum overall increase in concentration between sea-
water and the tissues of animals at the top of food chains (such as fish
liver, sea bird eggs, and seal blubber) was on the order of 100-fold for
tetrachloroethene (from 0.5 x 10"^ in water to 50 x 10"^ in tissue).
                                    53-7

-------
         Pearson and McConnell (1975) conducted laboratory experiments to
assess bioaccumulation of tetrachloroethene and other chloroorganics via
direct diffusion across wetted membranes.  The first experiments by Pearson
and McConnell (1975) have indicated that bioaccumulation occurs, but that
this bioaccumulation is not accompanied by any detected ill effects.  From
experiments in which dabs were exposed to tetrachloroethene it was found
that the accumulation factor is less than 10 in the flesh, arid more than
100 in the liver.  In general, it was found that concentrations in fatty
tissues such as liver are much higher than those in muscle and that the
ratios of concentration in different organs are proportional to fat con-
tent.

         Although evidence for weak to moderate bioaccumulation of tetra-
chloroethene by marine organisms exists as discussed above, there is no
evidence for biomagnification of tetrachloroethene in aquatic food chains.

         McConnell et al. (1975) report that there is evidence of the pre-
sence of tetrachloroethene in human tissue at extremely low concentrations.
It is indicated, however, that there is no evidence for significant accumu-
lation in human tissue.  According to McConnell e l. (1975), these data
indicate a general background, at the ]Jg/l level, which pervades the
atmosphere, hydrosphere, and biosphere.

    53.4.7  BiotransformatIon and Biodegradation

         No information was found pertaining specifically to the rate of
biodegradation of tetrachloroethene in aquatic systems.  Thorn and Agg
(1975) have included tetrachloroethene in a list of synthetic organic
chemicals which should be degradable by biological sewage treatment pro-
vided suitable acclimatization can be achieved.  They note, however, that
not many compounds in this list occur free in nature and, as a result, it
is unlikely that microorganisms already possess the ability to destroy
them.  Thorn and Agg (1975) consider tetrachloroethene to be potentially
biodegradable.

         According to Pearson and McConnell (1975) only completely sealed
systems, such as the standard BOD (biochemical oxygen demand) bottle tech-
nique, can be used to measure biochemical degradation of volatile com-
pounds such as tetrachloroethene.  The BOD bottle experiments of Pearson
and McConnell have been unable to demonstrate any significant oxygen
absorption from compounds containing only C, H, and Cl, thus indicating
that biochemical degradation of such compounds is indeed very slow.

         There is some evidence from mammal studies that tetrachloroethene
can be metabolized by higher organisms (Pearson and McConnell 1975;
McConnell _e_t _al_. 1975).  Metabolism of tetrachloroethene results in the
formation of trichloroacetic acid.
                                       53-8

-------
         Literature references to microbial biodegradation of compounds
similar to tetrachloroethene are few and conflicting;  the majority find
that low molecular weight chloroaliphatics are not metabolized (Pearson and
McConnell 1975;  McConnell e_t _a_l. 1975).

         In mammals, the metabolic pathways of tetrachloroethene lead to
chlorinated acetic acids, either directly or via chloroethanols.  Chlor-
inated acetic acids have all been shown to be susceptible to further degra-
dation by microorganisms in seawater (Pearson and McConnell 1975).
McConnell et al. (1975) concluded that the simple chloroaliphatic compounds
are not particularly persistent, but that their degradation products are
simple chemical species commonly found in the environment.

53.5  Data Summary

    Table 53-1 summarizes the aquatic fate discussed above.  The oxidation
rates presented are a photooxidation rates and refer to the rate of
reaction of tetrachlorethene with hydroxyl radicals in the troposphere.

    Due to the relatively high vapor pressure of tetrachloroethene, vola-
tilization from the aquatic system to the atmosphere is quite rapid.  Once
in the troposphere, the compound reacts with hydroxyl radicals at the
double bond, resulting in the subsequent formation of trichloroacetyl
chloride as the principal product with phosgene being produced to a lesser
extent.  The tropospheric lifetime of tetrachloroethene based on its rate
of reaction with hydroxyl radicals is reported to be 8.5 x 10^ seconds,
corresponding to a lifetime of about 10 days.  Due to the relatively high
reactivity of tetrachloroethene with hydroxyl radicals in the troposphere,
it appears unlikely that unreacted tetrachloroethene will reach the strato-
sphere.
                                    53-9

-------

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-------
53.6  Literature Cited

Altshuller, A.P. and J.J. Bufalini.  1971.  Photochemical aspects of
  pollution: a review.  Environ. Sci. Technol. 5:39-64.

Andersson, H.F., J.A. Dahlberg, and R. Wettstrom.  1975.  On the formation
  of phosgene and trichloroacetylchloride in the non-sensitized
  photooxidation of perchloroethylene in air.  Acta Chem. Scand. A
  29:473-474.

Chiou, C.T., V.H. Freed, D.W. Schmedding, and R.L. Kohnert.  1977.  ,
  Partition coefficient and bioaccumulation of selected organic chemicals.
  Environ. Sci. Technol. 11:475-478.

Dickson, A.G. and J.P. Riley.  1976.  The distribution of short-chain
  halogenated aliphatic hydrocarbons in some marine organisms.   Mar.
  Pollut. Bull. 7(9):167-169.

Billing, W.L.  1977.  Interphase transfer processes.  II.  Evaporation
  rates of chlororaethanes, ethanes, ethylenes, propanes, and propylenes
  from dilute aqueous solutions.  Comparisons with theoretical  predictions.
  Environ. Sci. Technol. 11:405-409.

Billing, W.L., N.B. Tefertiller, and G.J. Kallos.  1975.  Evaporation
  rates of methylene chloride, chloroform, 1,1,1-trichloroethane,
  trichloroethylene, tetrahloroethylene, and other chlorinated  compounds in
  dilute aqueous solutions.  Environ. Sci. Technol. 9(9):833-838.

Environmental Protection Agency.  1975a.  Preliminary assessment of
  suspected carcinogens in drinking water.  U.S. Environmental  Protection
  Agency, (Office of Toxic Substances), Washington, B.C.  33p.   EPA
  560/4-75-003.

Environmental Protection Agency.  1975b.  Preliminary study of  selected
  potential environmental contaminants - optical brighteners, methyl
  chloroform, trichloroetheylene, tetrachloroethylene, ion exchange resins.
  U.S. Environmental Protection Agency, (Office of Toxic Substances),
  Washington, D.C.  286p.  EPA 560/2-75-002.

Environmental Protection Agency.  1975c.  Report on the  problem of
  halogenated air pollutants and stratospheric ozone.  U.S. Environmental
  Protection Agency, (Office of Research and Development), Research
  Triangle Park, North Carolina.  55p.  EPA 600/9-75-008.
                                   53-11

-------
Gay, B.W.,  Jr., P.L. Hanst, J.J. Bufalini,  and R.C.  Noonan.  1976.
  Atmospheric oxidation of chlorinated ethylenes.   Environ. Sci. Technol.
  10:58-67.

Hanst, P.L.  1978.  Noxious trace gases in  the air,  Part II:   Halogenated
  pollutants. Chemistry 51(2):6-12.

Howard, C.J.  1976.  Rate constants  for the gas-phase reactions of  OH
  radicals  with ethylene and halogenated ethylene  compounds.   J. Chem.
  Phys.  65(ll):4771-4777.

Jaffe, H.H. and M. Orchin.  1962.  Theory and applications of ultraviolet
  spectroscopy.  John Wiley and Sons,  Inc.   New York. 624p.

Jensen, S.  and R. Rosenberg.  1975.   Degradability of some chlorinated
  aliphatic hydrocarbons in seawater and sterilized water.  Water Res.
  9:659-661.

Kopperman,  H.L., D.W. Kuehl, and G.E.  Glass.  1976.   Chlorinated compounds
  found in waste treatment effluents and their capacity to bioaccumulate.
  Proceedings of the conference on the environmental impact of water
  chlorination.  pp.327-345.  Oak Ridge, Tennessee,  October 22-24,  1975.

Lillian, D., H.B. Singh, A. Appleby, L. Lobban, R. Arnts, R.  Gumpert, R.
  Hague, J. Toomey, J. Kazazis, M. Antell,  D. Hansen, and B.  Scott.  1975.
  Atmospheric fates of halogenated compounds.  Environ. Sci.  Technol.
  9:1042-1048.

Liss, P.S.  and P.G. Slater.  1974.  Flux of gases  across the air-sea
  interface.  Nature 247:181-184.

Mackay, D.  and P.J. Leinonen.  1975.  Rate  of evaporation of low-solubility
  contaminants from water bodies to  atmosphere.  Environ. Sci. Technol.
  9:1178-1180.

Mackay, D.  and A.W. Wolkoff.  1973.   Rate of evaporation of low-solubility
  contaminants from water bodies to  atmosphere.  Environ. Sci. Technol.
  7:611-614.

McConnell,  G., D.M. Ferguson, and C.R. Pearson.  1975.  Chlorinated
  hydrocarbons and the environment.   Endeavor XXXIV:13-18.

Morrison, R.T. and R.N. Boyd.  1973.  Organic chemistry.  3rd Edition.
  Allyn and Bacon, Inc., Boston, Mass.  1258p.

Neely, W.B., D.R. Branson, and G.E.  Blau. 1974.  Partition coefficient  to
  measure bioconcentration potential of organic chemicals in fish.
  Environ.   Sci. Technol. 8:1113-1115.
                                   53-12

-------
Pearson, C.R.  and G. McConnell.  1975.  Chlorinated Cj and 2
  hydrocarbons in the marine environment.   Proc. Roy.  Soc. London B
  189:305-322.

Singh, H.B., L.J. Salas, H. Shiegeishi, and A.H. Smith.  1978.  Fate of
  halogenated  compounds in the atmosphere.  Interim report-1977.    U.S.
  Environmental Protection Agency, (Office of Research and Development),
  Research Triangle Park, North Carolina.   57p.  EPA 600/3-78-017.

Thorn, N.S. and A.R. Agg.  1975.  The breakdown of synthetic organic
  compounds in biological proceses.  Proc. Roy. Soc. London B
  189:347-357.

Verschueren, K.  1977.  Handbook of environmental data on organic
  chemicals.  Van Nostrand/Reinhold Press, New York.  659p.

Weast, R.C. (ed.).  1977.  Handbook of chemistry and physics.  58th
  Edition.  CRC Press, Inc., Cleveland, Ohio.  2398p.

Yung, Y.L., M.B. McElroy, and S.C. Wofsy.   1975.  Atmospheric halocarbons:
  a discussion with emphasis on chloroform.  Geophy. Res. Lett.
  2(9):397-399.
                                   53-13

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                         54.   1,2-DICHLORQPRQPANE


54.1  Statement of Probable Fate

    1,2-Dichloropropane exhibits the general chemical stability typical of
the short-chain chloroaliphatics.  The most rapid process by which this
chemical is transported from the aquatic environment appears to be volatil-
ization.  Although the molecules may be destroyed in the troposphere
through reaction with hydroxyl radicals, it is expected that a portion will
be reintroduced to the surface waters through precipitation.  The aquatic
fate of this compound is probably hydrolysis.

54.2  Identification

    1,2-Dichloropropane has been detected in finished drinking water, sur-
face water, and industrial effluents (Shackelford and Keith 1976).  The
chemical structure of 1,2-dichloropropane is shown below.

        H    Cl     H                        Alternate Names
                                             Propylene chloride
        C 	 C  	  C 	H                    Propylene dichloride
              H     H


    1,2-Dichloropropane

    CAS NO. 78-87-5
    TSL NO. TX 96250

54.3  Physical Properties

    The general physical properties of 1,2-dichloropropane are as follows.

    Molecular weight                         112.99
    (Verschueren 1977)

    Melting point                            -100C
    (Verschueren 1977)

    Boiling point at 760 torr                96.8C
    (Verschueren 1977)

    Vapor pressure at 20C                   42 torr
    (Verschueren 1977)
                                    54-1

-------
    Solubility in water at 20C              2,700 mg/1
    (Verschueren 1977)

    Log octanol/water partition coefficient  2.28
    (Calculated as per Tute 1971)

54.4  Summary of Fate Data

    54.4.1  Photolysis

         1,2-Dichloropropane does not have any chromophores that will
absorb electromagnetic radiation in the ultraviolet or visible spectral re-
gions (Jaffe and Orchin 1962).  Direct photolysis is thus not expected to
be an important aquatic fate process, although no specific experimental
data were found to support this view.

    54.4.2  Oxidation

         No data were found concerning the oxidation of 1,2-dichloropro-
pane.  Based upon the results of Dilling je _al. (1975) with short-chain
(one and two carbon) chloroaliphatics, oxidation does not appear to be
important as an aquatic fate for this type of compound.  This does not
preclude the possibility of oxidation in the atmosphere after volatiliza-
tion; tropospheric photooxidation by hydroxyl radicals is probably impor-
tant

    54.4.3  Hydrolysis

         No environmentally relevant kinetic data were found for the
hydrolysis of this compound.  Dilling _e_t al. (1975) showed that the
half-lives with respect to hydrolysis for one and two carbon chloro-
aliphatics are six months to several years.  1,2-Dichloropropane should be
expected to have a similar uncatalyzed rate of hydrolysis.  If this
compound is adsorbed by clay, base catalyzed hydrolysis could ultimately
lead to 1,2-propanediol.  This reaction sequence shown below is based upon
standard preparative procedures.  Clay surfaces are known to facilitate
both acid and base catalyzed reactions in an aqueous environment (Gabel and
Ponnamperuma 1967).

                      Cl

             Cl "~CHrt"""~ CH  "" ""CH*s  "  *  Cl"'CHrt"*^""" CH      3  ^
                                                 i
                                                OH
          H-C	CH	 CH3      -    HOCH2	CH

            Xr/                                '
              0                                  OH
                                    54-2

-------
    54.4.4  Volatilization

         According to Billing et al. (1975), the half-lives with respect to
volatilization from water for 1,2,3-trichloropropane and 1,2,2,3-tetra-
chloropropane are 51 minutes and 47 minutes, respectively.  The initial
concentration of each was 1 mg/1 in an open, stirred container.  Analo-
gously, 1,2-dichloropropane, which is more volatile than either of the more
highly halogenated chloroaliphatics mentioned above, should be expected to
have a half-life equal to or less than 50 minutes under the conditions used
by Billing ^_t _al_. (1975).  Once volatilized, this compound will be returned
to the soil and surface water via precipitation unless oxidized in the
atmosphere.  Billing et_ _al. (1975) have pointed out the difficulty in
extrapolating their laboratory data to environmental conditions, but vol-
atilization half-lives can be expected to be on the order of one to several
hours, depending on conditions.

    54.4.5  Sorption

         No data were found concerning adsorption of 1,2-dichloropropane
onto particulates and sediments.  A log P (log octanol/water partition
coefficient) of 2.28 can be calculated for this compound based on the
method of Tute (1971).  This implies a potential for removal by adsorption
from water for 1,2-dichloropropane.  Additionally, a log soil/water parti-
tion coefficient of 1.81 can be calculated based on the method of Briggs
(1973);  this further indicates a potential for adsorptive transport,
although data are lacking in this area.  To whatever extent this compound
is adsorbed by clay, the importance of hydrolysis will increase.

    54.4.6  Bioaccumulation

         No specific data on bioaccumulation of this chemical were found.
Here, again, the sizeable partition coefficients may indicate some poten-
tial for bioaccumulation.

    54.4.7  Biotransformation and Biodegradation

         1,2-Bichloropropane can be used as a carbon source by several soil
bacteria at concentrations up to 1,000 mg/1 (Altman and Lawlor 1966).
Biodegradation is also thought to occur in sewage treatment facilities
(Thorn and Agg 1975).  According to Roberts and Stoydin (1976), biodegrada-
tion in soil is very slow compared to volatilization.

54.5  Bata Summary

         The aquatic fate data for 1,2-dichloropropane are summarized in
Table 54-1.  The most important aquatic fate process for this compound is
probably hydrolysis, and tropospheric photooxidation by hydroxyl radicals
is probably also important.
                                    54-3

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54.6  Literature Cited

Altman, J. and S. Lawlor.  1966.  The effects of some chlorinated
  hydrocarbons on certain soil bacteria.  J.  Appl. Bact.  29(2):260-265.

Briggs, G.C.  1973.  A simple relationship between soil adsorption of
  organic chemicals and their octanol/water partition coefficient.
  Proceedings 7th British Insecticide and Fungicide Conferences.

Dilling, W.L., N.B. Tefertiller and G.J. Kallos.  1975.  Evaporation rates
  and reactivities of raethylene chloride, chloroform,
  1,1,1-trichloroethane, trichloroethylene, tetrachloroethylene and other
  chlorinated compounds in dilute aqueous solution.  Environ. Sci. and
  Technol. 9:833-838.   '

Gabel, N.W. and C. Ponnamperuma.  1967.  Model for origin of
  monosaccharides.  Nature  216:453-455.

Jaffa", H.H. and M. Orchin.  1962.  Theory and application of ultraviolet
  spectroscopy.  John Wiley and Sons, New York. 253p.

Roberts, T.R. and G. Stoydin.  1976.  The degradation of (Z)- and
  (E)-l,3-dichloropropenes and 1,2-dichloropropane in soil.  Pestic. Sci.
  7:325-335.

Shackelford, W.M. and L.H. Keith.  1976.  Frequency of organic compounds
  identified in water.  U.S. Environmental Protection Agency, (ERL),
  Athens, Ga.  617p.  (EPA-600/4-76-062).

Thorn, N.S. and A.R. Agg.  1975.  The breakdown of synthetic organic
  compounds in biological processes.  Proc. R. Soc. Lond. B  189:347-357.

Tute, M.S.  1971.  Principles and practice of Hansch analysis:  a guide to
  structure - activity correlation for the medicinal chemist.  Adv. Drug
  Res. 5:1-77.  Academic Press, New York.

Verschueren, K.  1977.  Handbook of environmental data on organic
  chemicals.  Van Nostrand Reinhold, New York.  659p.
                                       54-5

-------
                         55.  1,3-DICHLOROPROPENE
55.1  Statement of Probable Fate

    1,3-Dichloropropene is sufficiently volatile to be transported into the
atmosphere as a vapor.  It is expected that the compound will be photo-
oxidized by hydroxyl radicals in the troposphere.  1,3-Dichloropropene is
hydrolyzed in the environment to 3-chloroallyl alcohol.

55.2  Identification

    1,3-Dichloropropene has been detected in finished drinking water
(Shackelford and Keith 1976).  The chemical structure of 1,3-dichloro-
propene is shown below.
     Cl.
                                             Alternate Names

      H
           \
                       H
                   i



                 \
                           (TRANS ISOMER)
1,3-Dichloropropylene
                       Cl
     Cl
       \
      H'
           \
                           (CIS ISOMER)
                   Cl

            = C

      /        \
 1,3-Dichloropropene

 CAS NO. 542-75-6
 TSL NO. UC 83100
55.3  Physical Properties

    The general physical properties of 1,3-dichloropropene are as follows.

                                             110.98
Molecular weight
(Windholz 1976)

Melting point

Boiling point at 760 torr
(Weast 1977)
                                             No data found

                                             104.3C (cis isomer)
                                             112C (trans isomer)
                                    55-1

-------
    Vapor pressure at 20C                   25 torr
    (Dreisbach 1952)

    Solubility in water at 25C              2700 tng/1 (cis isomer)
    (Billing 1977)                           2800 mg/1 (trans isomer)

    Log octanol/water partition coefficient  1.98
    (Calc. by method  of Tute 1971)

55.4  Summary of Fate Data

    55.4.1  Photolysis

         No data were found on the photolysis of this compound in water.
Based on the data of  Billing _et al. (1975), direct photolysis is relatively
slow for chlorinated  ethenes; presumably,  this should also be the case for
chlorinated propenes.

    55.4.2  Oxidation

         No specific  data were found on oxidation of 1,3-dichloropropene.
The work of Billing et^ al. (1975) with chloroethenes indicated that oxida-
tion of these compounds is typically not as important as volatilization or
hydrolysis.  This also may be true of 1,3-dichloropropene. Atmospheric
oxidation, however, may be important after volatilization, although no
specific information pertaining to this compound was found.  Chlorinated
ethenes are photooxidized relatively rapidly in the troposphere;  Yung t
al. (1975) report a tropospheric lifetime  for trichloroethene with respect
to reaction with hydroxyl radicals of about four days.  The lifetime of
1,3-dichloropropene may be similarly short.

    55.4.3  Hydrolysis

         Billing _e_t _al. (1975) determined  half-lives with respect to hy-
drolysis of 10.7 months for trichloroethylene and 8.8 months for tetra-
chloroethylene, both chlorinated alkenes.   Analogously, the half-life of
hydrolysis for 1,3-dichloropropene should be at least several months.  This
conclusion agrees well with the first-order rate for hydrolysis of 1,3-di-
chloropropene in soil of 0.01 sec"* to 0.035 sec~l reported by Leistra
(1970), which corresponds to a half-life of 20-70 days.  The product of
this hydrolysis reaction is the corresponding 3-chloroallyl alcohol, a po-
tent biocide (Belser and Castro 1971).  The 3-chloroallyl alcohol is prob-
ably degraded further to 3-chloroacrolein, which can then undergo hydroly-
sis to malondialdehyde as shown below.  Oxidation of the dialdehyde pro-
duces malonic acid, a metabolic intermediate of many different food-chains.

             CI-CH= CHCH2OH     *-  C1-CH = CH CHO-
              a
               \
                CH	CH2	CHO      	** OCH	CH2	CHO

             H0/                     55-2

-------
    55.4.4  Volatilization

         According to Billing jet al. (1975), the evaporative half-life of
1,3-dichloropropene,  originally present at 1 mg/1 in water in a stirred
container, was determined to be 31 minutes (both cis and trans isomers
present).  Loss of 90 percent required 98 minutes.  This behavior is con-
sistent with the reported vapor pressure of this material.  Billing et al.
(1975) have pointed out the difficulty in extrapolating their laboratory
data to environmental conditions, but the evaporative half-life in the
environment is expected to be on the order of one-half hour to several
hours, depending on conditions.

    55.4.5  Sorption

         When 1,3-dichloropropene is applied as a soil fumigant, it is
readily adsorbed and persists for long periods of time (Leistra 1970;
Thomason and McKenry 1973).  In many cases, the adsorption of 1,3-di-
chloropropene is proportional to the content of organic matter in the soil
(Leistra 1970).  Degradation rates can be variable, but dissappearance
generally is more rapid in clay soils than in sandy ones (Van Dijk 1973).
The products of abiotic hydrolysis in moist soil and in water are the cor-
responding cis- and trans-3-chloroallyl alcohols (Castro and Belser 1966).
It is very likely that base-catalyzed hydrolysis on the clay surface
facilitates the formation of these products.  The relative importance of
adsorption as an aquatic fate process cannot be ascertained from the re-
viewed literature.

    55.4.6  Bioaccumulation

         Based on a calculated value for log octanol/water partition co-
efficient of 1.98, using the method of Tute (1971), bioaccumulation of this
compound may be possible under conditions of chronic exposure. However, no
experimental data concerning bioaccumulation were found.

    55.4.7  Biotransformation and Biodegradation

         Biodegradation of this compound does occur, according to Belser
and Castro (1971) and Altman and Lawlor (1966).  Belser and Castro (1971)
studied the microbiological metabolism of the toxic hydrolysis product,
3-chloroallyl alcohol.  Altman and Lawlor (1966) found that some soil
                                     55-3

-------
bacteria can utilize 1,3-dichloropropene as a carbon source.  Rates applic-
able to water were not found, but are expected to be slower than those for
other removal processes.

55.5  Data Summary

    The aquatic fate data, reported above for 1,3-dichloropropene, are
summarized on Table 55-1.  The primary transport  process for removal from
the aquatic environment appears to be volatilization.  Once in the tropo-
sphere, 1,3-dichloropropene is probably readily attacked by hydroxyl radi-
cals, and this photooxidation process appears to  be the predominant fate of
the compound.  Hydrolysis may also be of importance as a fate process in
aquatic systems.
                                    55-4

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-------
55.6  Literature Cited

Altraan, j. and S. Lawlor.  1966.   The effects of some chlorinated
  hydrocarbons on certain soil bacteria.   J.  Appl.  Bact.   29(2):260-265.

Belser, N.O. and C.E. Castro.  1971.  Biodegradation - the metabolism of
  the nematocides cis and trans-3-chloroallyl alcohol by a bacterium
  isolated from soil.  J. Agr. Food. Chem.  19(l):23-26.

Castro, C.E. and N.O. Belser.  1966.  Hydrolysis of cis- and
  trans-1,3-dichloropropene in wet soil.   J.  Agr. Food Chem.  14(1):69-70.

Billing, W.L.  1977.  Interphase  transfer processes.  II.  Evaporation
  rates of chloromethanes,  ethanes, ethylenes, propanes,  and propylenes
  from dilute aqueous solutions.   Comparisons with theoretical predictions.
  Environ. Sci. Technol.  11:405-409.

Dilling, W.L., N.B. Terfertiller  and G.J. Kallos.  1975.   Evaporation rates
  and reactivities of methylene chloride,chloroform, 1,1,1-trichloroethene,
  trichloromethylene, tetrachloroethylene, and other chlorinated compounds
  in dilute aqueous solutions.  Environ.  Sci. Technol.  9:833-838.

Dreisbach, R.R.  1952.  Pressure-volume-temperature relationships of
  organic compounds.  Handbook Publishers, Inc. Sandusky, Ohio. 349p.

Leistra, M.  1970.  Distribution  of 1,3-dichloropropene over the phases in
  soil.  J. Agr. Food. Chem.  18(6):1124-1126.

Shackelford, W.M. and L.H.  Keith.  1976.   Frequency of organic compounds
,  identified in water.  U.S. Environmental Protection Agency, (ERL),
  Athens, Ga.  617p.  (EPA-600/4-76-062).

Thomason, I.J. and M.V. McKenry.   1973.  Part I.  Movement and fate as
  affected by various conditions  in several soils.   Hilgardia
  42(11):393-421.

Tute, M.S.  1971.  Principles and practice of Hansch analysis:  a guide to
  structure-activity correlation  for the  medicinal chemist.  Adv. Drug
  Res. 6:1-77.  Academic Press, New York.

Van Dijk, H.  1973.  Degradation of 1,3-dichloropropenes in the soil.
  Agro-Ecosystems  1:193-204.

Weast, R.C. (ed.).  1977.  Handbook chemistry and physics.  CRC Press,
  Inc., Cleveland, Ohio.  2398p.
                                    55-6

-------
Windholz, M. (ed).  1976.   The Merck index,  9th edition.   Merck and  Co.,
  Rahway, N.J.   1313p.

Yung, Y.L., M.B. McElroy,  and S.C. Wofsy.   1975.   Atmospheric  halocarbons:
  a discussion with emphasis on chloroform.   Geophys.  Res. Lett.
  2(9):397-399.
                                   55-7

-------
                     56.  HEXACHLOROBUTADIENE
56.1  Statement of Probable Fate

    Although the literature reviewed indicates that adsorption onto sedi-
ments is an important aquatic transport process for this compound, the po-
tential contribution of bioaccumulation, volatilization, and quite possibly
other processes, should not be discounted.  Hexachlorobutadiene is a very
persistent environmental pollutant.  Its ultimate fate cannot, as yet, be
determined.

56.2  Identification

    Hexachlorobutadiene is present in domestic drinking water supplies (EPA
1975a), in aquatic organisms (EPA 1975b; Pearson and McConnell 1975;
McConnell et_ al. 1975), in aquatic sediments (EPA 1975b; Pearson and
McConnell 1975;  McConnell e_al. 1975), and is thought to be present in
ambient waters as a significant by-product of production wastes from the
manufacture of tetrachloroethene, trichloroethene,  and tetrachloromethane
(EPA 1975b).  The chemical structure of hexachlorobutadiene is given below.

   \        /
      C= C             C)
    /        \        /                   Alternate Names
   Cl             C  C
              /        \                   HCBD
             Cl           Cl                  Hexachloro-1,3-butadiene
     Hexachlorobutadiene

     CAS NO. 87-68-3
     TSL NO. EJ 07000

56.3  Physical Properties

    The general physical properties of hexachlorobutadiene (HCBD) are as
follows.
    Molecular weight                         260.76
    (Weast 1977)

    Melting point                            -21C
    (Weast 1977)

    Boiling point at 760 torr                215C
    (Weast 1977)
                                    56-1

-------
    Vapor pressure at 20C                   0.15 torr
    (Pearson and McConnell 1975)

    Solubility in water at 20C              2 mg/1
    (Pearson and McConnell 1975)

    Log octanol/water partition coefficient  3.74
    (Calculated from method of Tute 1971)

56.4  Summary of Fate Data

    56.4.1  Photolysi^

         Although the absorption spectrum of hexachlorobutadiene extends
into the electromagnetic spectrum of sunlight, the absorption coefficient
in this spectral region may be insufficient to make direct photolysis an
important aquatic fate.  A study by the Environmental Protection Agency
(1976) showed that photolysis of hexachlorobutadiene in benzene yielded
numerous products with a molecular weight higher than HCBD in quantities
greater than the initial amount of HCBD present.  This occurred after 15
minutes of irradiation at a wavelength of 273 nm.  It was not apparent,
however, whether HCBD was being photoactivated directly.

    56.4.2  Oxidation

         No information was found in the reviewed literature which pertains
specifically to the oxidation of hexachlorobutadiene.  Perhalogenated
hydrocarbons are typically very stable with respect to oxidation.

    56.4.3  Hydrolysis

         No information was found in the reviewed literature pertaining
specifically to the hydrolysis of hexachlorobutadiene.  If hydrolysis of
HCBD takes place on the clay surfaces of recently deposited sediments, the
first expected product would be a chlorinated butenoic acid, based on re-
sults of studies by Roedig and Bernemann (1956) on solvolysis of HCBD in
ethanol.

    56.4.4  Volatilization

         Unlike most short chain halogenated aliphatics, hexachlorobuta-
diene has a low vapor pressure and, thus, may not volatilize rapidly from
the aqueous environment to the atmosphere.  Hexachlorobutadiene has been
reported to be present in domestic drinking water supplies in low concen-
trations (EPA 1975a) and has been detected at concentrations of 1.9 and 4.7
yg/1 in water at two areas near Geismar, Louisiana.  These concentra-
tions indicate that HCBD may be quite persistent in natural waters.  How-
ever, hexachloroethane, which is structurally somewhat similar to HCBD and
exhibits a vapor pressure of 0.4 torr at 20C (Verschueren 1977) in cora-
                                    56-2

-------
parison to 0.15 torr for HCBD (Pearson and McConnell 1975), appears to be
volatilized rather rapidly from water.  Dilling (1977) determined a vo-
latilization half-life in water of 40.7 minutes for hexachloroethane ini-
tially present at 0.72 mg/1 in an open system stirred constantly at 200
rpm.  Although no specific rate data were found for HCBD, volatilization
may be an important transport process for this compound in aqueous systems.

    56.4.5  Sorption

         The currently reviewed literature contains an appreciable amount
of information pertaining specifically to the adsorption of HCBD onto
sediments.  In a study of the Mississippi Delta region it was found that
the level of hexachlorobutadiene in water was less than 2 yg/1 while the
concentration of hexachlorobutadiene in mud or soil samples exceeded 200
yg/1 (EPA 1976).  In this same study, water samples from the waste of an
industrial company in Geismar, Louisiana, was found to contain from <0.1
yg/1 to 4.5 yg/1 HCBD.  Levels of HCBD in the mud, however, reached a max-
imum of 2,370 yg/1, indicating selective concentration of several orders of
magnitude.  Leeuwangh _e_t l. (1975) found that the concentration of HCBD in
uncontaminated sediment after equilibration with water that contained HCBD
was a 100 x that found in the water.

         Samples taken from Liverpool Bay (England) showed the presence of
HCBD, but rarely at levels greater than one yg/1 (Pearson and McConnell
1975).  McConnell _et al. (1975) noted that coarse gravels have little ad-
sorptive capacity for chlorinated aliphatics, whereas sediments rich in
organic detritus have a much higher adsorptive capacity.  The calculated
log P (octanol/water partition coefficient) of 3.74 implies that hexa-
chlorobutadiene should be strongly adsorbed by humus material.
    56.4.6  Bioaccumulation

         According to Kopperman .et _!; (1976) polar compounds are more
easily biodegraded, whereas non-polar compounds tend to accumulate.  The
log octanol/water partition coefficient (log P) of 3.74, which was calcu-
lated using the method of Tute (1971), indicates that HCBD may bioaccumu-
late significantly.

         Results of a study conducted for the Environmental Protection
Agency (1976) showed that HCBD did not accumulate to high levels in test
animals and that rates of uptake, as well as distribution in organs, were
irregular.  A flow-through system was used to determine accumulation of the
compounds by algae and sediment.   Concentration factors remained below 300
for both during relatively short-term experiments.  A food-chain study com-
pared the relative accumulation effects on a predator feeding upon HCBD-
contaminated food-fish and a similar predator taking up the compound both
through its food and through contaminated water.  Uptake of HCBD was
                                    56-3

-------
erratic.  These laborarory experiments were compared with patterns of
accumulation under field conditions.   HCBD-free crayfish were caged, placed
at a contaminated field site,  and removed periodically for GC analysis.
HCBD was not concentrated to a great  extent and was accumulated irregu-
larly.  Concentration factors  for crayfish left in the field site for 17
days ranged from 7.8 to 300 (33.7 to  1,290.3 yg/l HCBD).

    56.4.7  Biotransformation  and Biodegradation

         No information was found in  the currently reviewed literature
pertaining specifically to the biodegradation of HCBD.

    56.5  Data Summary

         A summary of the aquatic fate information discussed above for HCBD
is presented on Table 56-1.  It appears that adsorption onto sediments is
an important transport process and that bioaccumulation and volatilization
may occur as well.  The ultimate fate of HCBD cannot, as yet, be
determined.
                                    56-4

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-------
56.6  Literature Cited

Billing, W.L.  1977.  Interphase transfer processes.   II.   Evaporation
  rates of chloromethanes, ethanes, ethylenes, propanes, and propylenes
  from dilute aqueous solutions.  Comparisons with theoretical predictions.
  Environ. Sci. Technol.  11:405-409.

Environmental Protection Agency.  1975a.  Preliminary assessment of
  suspected carcinogens in drinking water.  EPA 560/4-75-003.  U.S.
  Environmental Protection Agency (Office of Toxic Substances),
  Washington,   B.C.  33p.

Environmental Protection Agency.  1975b.  Survey of industrial processing
  data.  Task I - hexachlorobenzene and hexachlorobutadiene pollution from
  chlorocarbon processing.  EPA 560/3-75/003.  U.S. Environmental
  Protection Agency (Office of Toxic Substances), Washington, B.C.  176p.

Environmental Protection Agency.  1976.  An ecological study of
  hexachlorobutadiene.  EPA 560/6-76-010.  U.S. Environmental Protection
  Agency (Office of Toxic Substances), Washington, B.C.  61p.

Kopperman, H.L., B.W. Kuehl, and G.E. Glass.  1976.  Chlorinated compounds
  found in waste-treatment effluents and their capacity to bioaccumulate.
  Proceedings of the conference on the environmental impact of water
  chlorination.  pp.327-347.  Oak Ridge, Tennessee, October 22-24, 1975.

Leeuwangh, P., H. Suit and L. Schneiders.  1975.  Toxicity of
  hexachlorobutadiene in aquatic organisms.  Sublethal effects of toxic
  chemicals on aquatic animals.  167-176p.  Proc. of Swedish-Netherlands
  Symp., Sept. 2-5.  Elsevier Scientific Publishing Company, Inc., New
  York.  (Abstract only).

McConnell, G., B.M. Ferguson, and C.R. Pearson.  1975.  Chlorinated
  hydrocarbons and the environment.  Endeavor XXXIV:13-18.

Pearson, C.R. and G. McConnell.  1975.  Chlorinated Cj and fy
  hydrocarbons in the marine environment.  Proc. Royal Soc. Lond. B
  189:305-332.

Roedig, A. and P. Bernemann.  1956.  Highly chlorinated vinyl ethers.
  1-ethoxypentachloro-l,3-butadiene from perchlorobutadiene.  Justus
  Liebig's Ann. Chem. 600:1-11.

Tute, M.S.  1971.  Principles and practice of Hansch analysis:  a guide to
  structure-activity correlation for the medicinal chemist.  Adv. Brug
  Res.  6:1-77.
                                    56-6

-------
Verschueren, K.  1977.  Handbook of environmental data on organic
  chemicals.  Van Nostrand Reinhold, New York.  659p.

Weast, R.C. (ed.).  1977.  Handbook of chemistry and physics.   CRC   Press,
  Inc., Cleveland, Ohio.  2398p.
                                    56-7

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                      57.  HEXACHLOROCYCLOPENTADIENE
57.1  Statement of Probable Fate

    Several transport and fate processes appear to operate at significant
rates to remove hexachlorocyclopentadiene from aquatic systems.  The re-
lative importance of these processes is thought to depend strongly on the
characteristics of the individual water body, so that there is no clear
indication that one process is predominant on an overall basis.  The most
important fate processes appear to be hydrolysis and near-surface photoly-
sis, while transport occurs via the water column (as dissolved species), by
volatilization to the atmosphere, and through adsorption onto particulates
(perhaps to a lesser extent).  The fate of HCCPD in the troposphere is not
known.

57.2  Identification

    Hexachlorocyclopentadiene has been detected in industrial effluents
(Shackelford and Keith 1976).  The chemical structure of hexachlorocyclo-
pentadiene is shown below.
     Cl
Alternate Names
         .
         c =  c          ci

         I           c
         c =  c          ci

                    "Cl
HCCPD
Perchlorocyclopentadiene
Hex
    Hexachlorocyclopentadiene

    CAS NO. 77-47-4
    TSL NO. GY 12250

57.3  Physical Properties

    The general physical properties of hexachlorocyclopentadiene are as
follows:
    Molecular weight
    (Weast 1977)

    Melting point
    (Weast 1977)
272.77


-9.9C
                                    57-1

-------
    Boiling point                                239C
    (Weast 1977)

    Vapor pressure at 25C                       0.081 torr
    (Ungnade and McBee 1957)

    Solubility in water                          0.805 mg/1
    (Lu et al. 1975)
    (Zepp et al. 1979)                           1.8 mg/1 (25C)

    Log octanol/water partition coefficient      3.99
    (Zepp et_ al. 1979)

57.4  Summary of Fate Data

    57.4.1  Photolysis

         Zepp ejt _al. (1979) conducted photolysis experiments with hexa-
chlorocyclopentadiene (HCCPD) in distilled water which utilized a variety
of solar conditions.  The compound was found to be highly photoreactive,
exhibiting near-surface half-lives of less than ten minutes in all experi-
ments.  Photolysis rate constants were also determined for HCCPD in Aucilla
(Fla.) River water, the results indicating that photosensitized reactions,
as well as direct photolysis, occurred in the river water.

         From the above data, plus the absorption spectrum and the pho-
tolytic reaction quantum efficiency from controlled laboratory experiments,
Zepp _e_t _al. (1979) computed a mean near-surface rate constant for direct
photolysis of HCCPD at 40N latitude of about 3.9 hour'1.  This value,
which was obtained by averaging over both dark and light periods for a
year, corresponds to a half-life of about 11 minutes.  Through the use  of a
mathematical model to predict behavior of organic chemicals in various
types of water bodies, Zepp ^_t jut. (1979) conclude that photolysis is the
predominant fate process for HCCPD in ponds and eutrophic lakes but not in
turbid rivers.

         Zepp ^_t j.1. (1979) indicate that tetrachlorocyclopentadienone
(TCPD), a highly reactive compound, is the primary photoproduct of HCCPD in
water  and it is likely to exist predominately in the hydratecl form in  the
aquatic environment.

    57.4.2  Oxidation

         No information specifically concerning oxidation of HCCPD was
found in the reviewed literature.  Zepp _e_t _al. (1979) reported evidence of
                                    57-2

-------
photosensitized reaction of HCCPD in natural waters which may have been
photooxidation.

    57.4.3  Hydrolysis

         Hydrolysis rate constants for hexachlorocyclopentadiene at several
temperatures and pH values were determined by Zepp js_t al. (1979).  The re-
action, in both distilled and natural water samples, was found to be inde-
pendent of pH over the range 5 to 9 and dependent on temperature according
to:

         kobs = 5.85 x 1011 exp(-2.460 x 104/RT).

From the equation, the hydrolysis rate constant at 25C is 5.6 x
10""7Sec~l, corresponding to a half-life of 14 days.  Hydrolysis of
HCCPD adsorbed onto natural pond sediments was found to be at least as
rapid as in the water.  Acid catalyzed hydrolysis of HCCPD leads to
tetrachlorocyclopentadienone hydrate (Ungnade and McBee 1957):
                                                        CI
    57.4.4  Volatilization

         Zepp et al. (1979) determined the ratio of the volatilization rate
to the reaeration rate constant for HCCPD in natural waters to be 0.58.
The experimental method used was that described by Hill e_t _al_. (1976).
This ratio was then utilized in a mathematical model to indicate relative
importance of volatilization and other processes in the transport and fate
of HCCPD in several types of aquatic systems.  The results indicated that
on the order of 15 percent of the HCCPD load in a turbid river would be re-
moved by volatilization, as compared to less than five percent for a pond
or a eutrophic lake.
                                     57-3

-------
    57.4.5  Sorption

         The log octanol/water partition coefficient of 3.99 (Zepp et al.
1979) indicates that adsorption of hexachlorocyclopentadiene onto oTganTc-
laden sediments may be appreciable.  HCCPD held in sediments may be hydro-
lyzed or biodegraded; according to Zepp et_ _al. (1979), hydrolysis is Che
more important process.  The results obtained by Zepp _et al. (1979), from
analysis of HCCPD transport and fate by means of a mathematical model, indi-
cate that sorption is not as important for this compound as competing
processes.

    57.4.6  Bioaccumulation

         In a model ecosystem studied by Lu _et_ _a_l. (1975), hexachlorocyclo-
pentadiene demonstrated bioaccumulation in all test organisms, ecological
magnification (or EM) factors were as follows:  algae 340; snail 929; mos-
quito 1634; fish 448.

    57.4.7  Biotransformation and Biodegradation

         The extent to which their test organisms were capable of biode-
grading hexachlorocyclopentadiene in a model aquatic ecosystem was esti-
mated by Lu _e_t al. (1975) as:  alga, 4 percent; snail, 10 percent; mosqui-
to, 2 percent; fish, 27 percent.  (The figures refer to the percentage of
radioactivity present as metabolized HCCPD in comparison to unchanged HCCPD
within the organism)  Insofar as perchlorinated carbon-carbon double bonds
are inert to biotic or abiotic oxidation, the primary product of biodegra-
dation is probably tetrachlorocyclopentadienone hydrate.  Biological de-
toxification of the hydrate by compound formation with glucuronic acid or
oligosaccharides could account for the unidentified biodegradation products
reported by Lu _e al. (1975).

         The results of Zepp _e_t al. (1979), from the study of HCCPD by a
mathematical model of fate and transport processes, indicate that biode-
gradation is probably not an important fate of this compound, in that one
percent or less of HCCPD in aquatic systems was predicted to be microbially
transformed.

57.5  Data Summary

    The data on the environmental fate of HCCPD are summarized in Table
57-1 .  Near-surface photolysis and hydrolysis (both in the water column
and in sediments) appear to be important fate processes for HCCPD.  Trans-
port can occur by volatilization, as dissolved species in the water  column,
and through adsorption onto particulates (perhaps to a lesser extent).
                                     57-4

-------
According to Zepp et al. (1979) the relative importance of these processes
depends strongly on the characteristics of the individual aquatic system.
Their analysis indicates that photolysis is predominant in ponds and lakes,
but that water-borne export and volatilization are each more important than
photolysis in a turbid river.

    Because a number of aquatic transport and fate processes appear to com-
pete, in the case of HCCPD, and because the accuracy of the method used by
Zepp _e_t_ al. (1979) to mathematically analyze them has not been verified, a
single predominant fate process cannot be selected at this time.
                                      57-5

-------
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-------
57.5  Literature Cited

Hill, J.W., H.P. Kollig, D.F. Paris, N.L. Wolfe and R.G. Zepp.  1976.
  Dynamic behavior of vinyl chloride in aquatic ecosystems.  U.S.
  Environmental Protection Agency (ERL), Athens, Ga.  64p. (EPA
  600/3-76-001).

Lu, P.Y., and R.L. Metcalf, A.S. Hirwe and J.W. Williams.  1975.
  Evaluation of environmental distribution and fate of hexachlorocyclo-
  pentadiene, chlordane, heptachlor, and heptachlor epoxide in a
  laboratory model ecosystem.  J. Agr. Food Chem.  23(5):967-973.

Shackelford, W.M. and L.H. Keith.  1976.  Frequency of organic compounds
  identified in water.  U.S. Environmental Protection Agency (ERL),  Athens,
  Ga. 617p.  (EPA 600/4-76-062).

Ungnade, H.E. and E.T. McBee.  1957.  The chemistry of
  perchlorocyclopentenes and cyclopentadienes.  Chem. Rev.  58:249-320.

Weast, R.E. (ed.).  1977.  Handbook of chemistry of physics.  58th Ed.
  Chemical Rubber Publishing Co., Cleveland, Ohio.  2398p.

Zepp, R.G., N.L. Wolfe, G.L. Baughman, P.F. Schlotzhauer and J.N.
  MacAllister.  1979.  Dynamics of processes influencing the behavior  of
  hexachlorocyclopentadiene in the aquatic environment.  U.S. Environmental
  Protection Agency (ERL), Athens, Ga. (Paper presented before the
  Division of Environmental Chemistry, American Chemical Society,
  Washington, D.C. September 9-14, 1979).
                                    57-7

-------
                    58.  BROMOMETHANE (METHYL BROMIDE)
58.1  Statement of Probable Fate

    The two most probable fates of bromomethane in the environment are pho-
tooxidation in the troposphere and hydrolysis in the aquatic environment.

    Volatilization is the major transport process of bromomethane from
aquatic systems.  Once in the troposphere, bromomethane reacts with hy-
droxyl radicals to form bromine atoms and inorganic bromides, which may re-
turn to earth during the precipitation process.  Any unreacted portion of
bromomethane which diffuses upward to the stratosphere would be expected to
undergo significant photodissociation.

    A maximum hydrolytic half-life of 20 days has been reported for bromo-
methane at pH 7 and 25C.  As a result, hydrolysis may serve as an impor-
tant fate of this compound.

    Based on the information found, it appears that oxidation is not an
important fate process of bromomethane in the aquatic environment.  In ad-
dition, the literature reviewed contained no specific information indicat-
ing that biodegradation, adsorption, and bioaccumulation are important pro-
cesses for this chemical in the aquatic environment.

58.2  Identification

    Bromomethane may be ubiquitous in the environment due to natural as
well as anthropogenic sources.  Its presence has been detected in finished
drinking water (Environmental Protection Agency 1975), in seawater
(Lovelock 1975;  Singh t_ jQ. 1978), and in the troposphere (Singh et al.
1978).

    The chemical structure of bromomethane is shown below.


                                      Alternate Names

    H     C     Br                      Methyl bromide
                                      Terabol
          H                            Monobromomethane
                                      Embafume
    Bromomethane

    CAS NO.  74-83-9
    TSL NO.  PA 49000
                                   58-1

-------
58.3  Physical Properties

    The general physical properties of bromomethane are given below.

    Molecular weight                         94.94
    (Weast 1977)

    Melting point                            -93.6C
    (Weast 1977)

    Boiling point at 760 torr                4.6C
    (Verschueren 1977)

    Vapor pressure at 20C                   1420 torr
    (Environmental Protection Agency 1976)

    Solubility in water at 20C              900 mg/1
    (Verschueren 1977)

    Log octanol/water partition coefficient  1.1
    (Calculated from Tute 1971, See Methods
    Section on Bioaccumulation)

58.4  Summary of Fate Data

    58.4.1  Photolysis

         No information was found pertaining to the photolysis of bromo-
methane in the aqueous environment.  Due to the high vapor pressure of
bromomethane, volatilization to the atmosphere is quite rapid.  The domin-
ant loss mechanism for bromomethane from the lower troposphere is said to
be upward diffusion (Robbins 1976).  Bromomethane is not known to undergo
significant light absorption at wavelengths above 290 nm (Robbins 1976).
As a result, photodissociation cannot play an. important role in the loss of
bromomethane below the ozone layer since the ozone layer effectively ab-
sorbs wavelengths of light less than 290 nm (Hanst 1978).  The wavelength
region in which significant photodissociation in the stratosphere occurs is
from about 180 to 240 nm (Robbins 1976).  That portion of broraomethane
which diffuses upward to the stratosphere would be expected to undergo
significant photodissociation.

    58.4.2  Oxidation

         No information was found pertaining specifically to the oxidation
of bromomethane in the aqueous environment under ambient conditions.  Ac-
cording to Radding et_ _al. (1977), oxidation of alkyl halides in water by RO
radical is not significant, having a half-life of greater than 100 years.
                                      58-2

-------
         In the upper troposphere and lower stratosphere loss of bromo-
raethane occurs by reaction with hydroxyl radicals (Robbins 1976).  Wofsy et
_al. (1975) contend that bromomethane is removed as follows:

               OH + CH3Br 	-*~H20 + -CH2Br

         The bromine atom is subsequently released and inorganic bromide is
reported to be carried out of the lower atmosphere by rain (Wofsy e al.
1975).  The rate  constant for the above reaction is reported to be
2 x lO'^cm^molecule'i-sec"1   exp (-1200/T).  By substituting
300K  (27C) for T, a bimolecular rate of 3.6 x 10~14 cm3sec~1 is
obtained for reaction of bromomethane with hydroxyl radicals at 27C.

         The photooxidation rate data of Wofsy et_ al. indicates a tropos-
pheric lifetime (1/k) on the order of one year for bromoraethane.  Assum-
ing a  troposphere-to-stratosphere turnover time (time required for all but
1/e of tropospheric air to diffuse into the stratosphere) of 30 years,  ab-
out three percent of tropospheric bromomethane would be expected to eventu-
ally reach the stratosphere.

    58.4.3  Hydrolysis

         A maximum hydrolytic half-life of 20 days has been reported for
bromomethane at pH 7 and 25C (Mabey and Mill 1978;  Radding .et al. 1977).
This corresponds to a first order rate constant for hydrolysis of bromo-
raethane of 4.0 x 10~7sec~1 (Mabey and Mill 1978;  Radding .et al. 1977).
The short hydrolytic half-life of bromomethane makes it likely that hydro-
lysis  is an important fate process for this compound.

    58.4.4  Volatilization

         Volatilization is the dominant process for the removal of bromo-
methane from the aqueous environment.  Although no rates for volatilization
of bromomethane were found, one can extrapolate information from its more
thoroughly studied analogue, chloromethane, and other halogenated alipha-
tics.   Billing e jd. (1975) estimated the experimental half-life for
volatilization of chloromethane originally present at 1 mg/1 to be 27
minutes when stirred at 200 rpm in water at approximately 25 C in an open
container.  Removal of 90 percent of the chloromethane under the same con-
ditions required 91 minutes.  For chloroaliphatics in general, stirring
speed  was found to have a marked effect on volatilization rate.  With no
stirring except 15 seconds every five minutes, the time required for 50
percent depletion of trichloromethane, trichloroethene, and 1,1,1-tri-
chloroethane, compounds similar to chloromethane, was greater than 90
minutes or on the order of from times greater than the stirred case for
these  compounds.
                                    58-3

-------
         Billing jit _al. (1975) are careful to point out the difficulties
encountered in extrapolating their laboratory results to real-world condi-
tions, where the concentration of the organic solute would probably be very
much less than 1 mg/1 and where surface and bulk agitation would be highly
variable.  Although the data appear to be valid on a relative basis (i.e.,
correctly illustrating the relative rates of volatilization of chlorinated
aliphatics), they cannot be used as absolute measures of volatilization
rates from natural waters.  For the purposes of this document, the data are
used as rough-order-of-magnitude indications of the importance of volatili-
zation relative to other transport and fate processes, with the strong ef-
fects of agitation considered.  The validity of this application has not
been established.

    58.4.5  Sorption

         No data pertaining to adsorption of bromomethane onto sediments
were found.  Since bromomethane volatilizes rapidly from the aqueous en-
vironment, it is unlikely that adsorption onto sediment is a significant
transport process for this chemical.

    58.4.6  Bioaccumulation

         The log octanol/water partition coefficient (log P) for bromo-
methane is 1.10 (calculated from Tute 1971), indicating that bromomethane
is not very lipophilic and, thus, probably does not exhibit a significant
tendency to bioaccumulate in organisms.  Laboratory animals given doses of
bromomethane have demonstrated no indications of significant accumulation
of the compound (Environmental Protection Agency 1976).  In addition, there
has been no indication of significant accumulation in soils from annual ap-
plication (Environmental Protection Agency 1976).

    58.4.7  Biotransformation and Biodegradation

         No data were found pertaining to biological degradation of bromo-
methane by aquatic organisms.

58.5  Data Summary

    Table 58-1 summarizes the aquatic fate of bromomethane.  The oxidation
rate is a photooxidation  rate and refers to the rate of reaction of
bromomethane with hydroxyl radicals in the troposphere.  The two most
likely fates of bromoraethane in the environment are photooxidation in the
troposphere and aquatic hydrolysis.
                                    58-4

-------

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-------
58.6  Literature Cited

Oilling, W.L., N.B. Tefertiller,  and G.J.  Kallos.   1975.   Evaporation
  rates of methylene chloride,  chloroform, 1,1, 1-trichlororaethane,
  trichloroethylene, tetrachloroethylene and other chlorinated compounds
  in dilute aqueous solutions.   Environ. Sci.  Technol. 9(9):833-838.

Environmental Protection Agency.   1975.  Preliminary assessment of sus-
  pected carcinogens in drinking  water.  Environmental Protection Agency
  (Office of Toxic Substances), Washington, D.C.  33p.  EPA 560/4-75-003.

Environmental Protection Agency.   1976.  A literature survey oriented
  towards adverse environmental effects resultant  from the use of azo com-
  pounds, brominated hydrocarbons, EDTA, formaldehyde resins, and o-nitro-
  chlorobenzene.  Environmental Protection Agency  (Office of Toxic Sub-
  stances), Washington, D.C.  480p.  EPA 560/2-76-005.

Hanst, P.L.  1978.  Part II.  Halogenated pollutants.  Noxious trace gases
  in the air.  Chemistry 51(2):6-12.

Liss, P.S. and P.G. Slater.  1974.  Flux of gases  across the air-sea inter-
  face.  Nature  247:181-184.

Lovelock., J.E.  1975.  Natural halocarbons in the  air and in the sea.
  Nature 256:193-194.

Mabey, W. and T. Mill.  1978.  Critical review of  hydrolysis of organic
  compounds in water under environmental conditions.  J. Phys. Chem. Ref.
  Data 7(2):383-415.

Radding, S.B., D.H. Liu, H.L. Johnson, and T.  Mill.  1977.   Review of the
  environmental fate of selected chemicals.  Environmental Protection
  Agency (Office of Toxic Substances), Washington, D.C.  147p.  EPA
  560/5-77-003.

Robbins, D.E. 1976.  Photodissociation of methyl chloride and methyl
  bromide in the atmosphere.  Geophys. Res. Lett.  3(4) : 213-2 16 ,.

Singh, H.B., L.J. Salas, H. Shiegeishi, and A.H. Smith.  1978.  Fate of
  halogenated compounds in the atmosphere interim report - 1977.  Environ-
  mental Protection Agency (Office of Research and Development), Research
  Triangle Park, N.C.  57p. EPA 600/3-78-017.

Tute, M.S.  1971.  Principles and practice of Hansch analysis:  A guide to
  structure-activity correlation for the medicinal chemist.  Adv. Drug
  Res. 6:1-77.
                                    58-6

-------
Verschueren, K.  1977.  Handbook of environmental data on organic chemi-
  cals.  Van Nostrand/Reinhold Press, New York.  659p.

Weast, R.C. (ed.).  1977.  Handbook of chemistry and physics.  58th
  Edition.  CRC Press, Cleveland, Ohio.  2398p.

Wofsy, S.C., M.B. McElroy, and Y.L. Yung.  1975.  The chemistry of atmos-
  pheric bromine.  Geophys. Res. Lett. 2(6):215-218.
                                    58-7

-------
                         59.  BROMODICHLOROMETHANE
59.1  Statement of Probable Fate

    The information found is not sufficient to determine the aquatic fate
of bromodichloromethane.

59.2  Identification

    Bromodichloromethane has been detected in finished drinking water
(Kleopfer and Fairless 1972;  Environmental Protection Agency 1975), and in
sources of drinking water (Environmental Protection Agency 1975).  Bromodi-
chloromethane is hypothesized to be present in water supplies as a result
of the haloform reaction which occurs during the chlorination of such water
(Rook  1974;  Environmental Protection Agency 1975;  Glaze and Henderson
1975).

    The chemical structure of bromodichloromethane is shown below.

             Br                        Alternate Names
                                     Dichlorobromomethane
        Cl	 C 	Cl
             H

    Bromodichloromethane

    CAS NO. 75-27-4
    TSL NO. Not Assigned

59.3  Physical Properties

    The general physical properties of bromodichloromethane are given be-
low.
    Molecular weight                         163.83
    (Weast 1977)

    Melting point                            -57.1C
    (Weast 1977)

    Boiling point at 760 torr                90C
    (Weast 1977)

    Vapor pressure at 20C                   50 torr
    (Dreisbach 1952)
                                    59-1

-------
    Solubility in water                      Not available

    Log octanol/water partition coefficient  1.88
    (Calculated from Tute 1971, See Methods
    Section on Bioaccumulation)

59.4  Summary of Fate Data

    59.4.1  Photolysis

         No information was found pertaining specifically to the rate of
photolysis of bromodichloromethane in either the aquatic or atmospheric en-
vironments.

    59.4.2  Oxidation

         No information was found pertaining specifically to the rate of
oxidation of bromodichloromethane in the aquatic environment,.  Bromodi-
chloromethane is probably like other halogenated aliphatics in that it is
not easily oxidized in aquatic systems since there are no functional groups
which react strongly with hydroxyl radicals.

    59.4.3  Hydrolysis

         A maximum hydrolytic half-life of 137 years has been reported for
bromodichloromethane at pH 7 and 25C (Mabey and Mill 1978).  This corre-
sponds to a first-order rate constant for hydrolysis of bromodichloro-
methane of 1.6 x Kr^sec'1 (Mabey and Mill 1978).

    59.4.4  Volatilization

         No information was found pertaining specifically to the rate of .
volatilization of bromodichloromethane from water.  Volatilization may very
well be an important transport process for bromodichloromethane from water
to the atmosphere.  It must be noted, however, that the vapor pressure of
bromodichloromethane is much lower than that for chloroform and other
chloroalkanes.  As a result, bromodichloromethane would not be expected to
volatilize as rapidly as chlorinated methanes such as chloroform.  The con-
centration of bromodichloromethane present in water supplies has been re-
ported to decrease as a result of volatilization while flowing through open
channels (Rook 1974).

    59.4.5  Sorption

         No information was found pertaining specifically to the adsorption
of bromodichloromethane onto suspended sediments.  The concentration of
bromodichloromethane produced  by chlorination can be reduced by treatment
of the water with powdered activated carbon (Rook 1974).
                                    59-2

-------
    59.4.6  Bioaccumulation

         Neely _et _al. (1974) have shown that bioaccumulation is directly
related to the octanol/water partition coefficient (P) of the compound.
The log octanol/water partition coefficient (log P) as calculated by the
method of Hansch is 1.88 (Tute 1971) indicating that bromodichloromethane
is somewhat lipophilic (see Methods section on bioaccumulation).  As a re-
sult, bioaccumulation in organisms may be possible for this compound.  No
information was found, however, reporting the bioaccumulation of bromodi-
chloromethane in organisms.

    59.4.7  Biotransf ormation and Biodegradation

         No information was found pertaining specifically to the
biodegradation of bromodichl or ome thane.  In the sea, many species of brown
seaweed, molluscs, sponges, and bacteria metabolize halogens (Anonymous
1977).  The red algae, or Rhodophyta, are known to use mostly bromine and
they also synthesize several different organic substances, some containing
iodine and chlorine.

    59.4.8  Other Reactions

         Rook (1977) suggests that bromodichl orome thane , which has been
found in finished drinking water, may be present in such chlorinated water
due to the haloform reaction.  The haloform reaction, connected with the
chlorination of water supplies, occurs generally in alkaline solution with
organic compounds containing the acetyl group or with structures that may
be readily oxidized to the acetyl group (Morris 1975).  The three hydrogens
of the methyl component of the acetyl group are successively replaced by
chlorine or other halogens , and subsequently the carbon bond to the
carbonyl group is hydrolyzed giving rise to a. haloform and a carboxylic
acid (Morris 1975).  The simplified reaction is given below using
production of chloroform as an example:

         CH3COR + 3HOC1 - * CC13COR + 3H20    (1)
         CC13COR + H20  - -CHCl3 + RCOOH    ' (2)
         It is suggested that organic compounds containing the acetyl
group, needed for the haloform reaction to occur, are present in the form
of humic materials which cause the yellow to brown coloration of surface
waters (Rook 1974;  Rook 1977;  Morris 1975).   For the most part humic
materials have a molecular weight in the 10^-10^ range and fall into
the group of compounds designated as fulvic acids (Rook 1977; Morris 1975)
                                    59-3

-------
         There are several conditions which must be met in order for the
haloform reaction to occur.   These conditions are discussed below.

         First, the haloform reaction generally occurs at pH of 5 or
greater.  Chlorine dispersed in water at pH greater than 5, as employed in
water treatment in concentrations up to 100 mg/1 or about 10-3 molar, is
hydrolyzed instantly and entirely to HOC1 and OC1~ (Morris 1975).  The
overall reactions are as follows :
               HC03~ - *-HOCl + Cl~ + CC>2
         HOC1                            "   H+ + OC1~

Thus, it is necessary to consider the reactions of hypochlorous acid rather
than those of Cl in describing the potential reactions of dilute aqueous
chlorine with organic compounds.

         Secondly, the haloform reaction must occur at sites on the fulvic
acids which are the most conducive to the formation of haloforms.  Although
the carbon content of fulvic acid is about 50 percent, there are very few
carbons which have the chemical constituents necessary for the haloform re-
action to occur.  It is for this reason that the yield of haloforms pro-
duced upon aqueous chlorination is fairly low (Rook 1977).  The carbon be-
tween two meta positioned OH-groups of a hydroxylated aromatic ring is
proposed as the most reactive site for haloform formation (Rook 1977).

         The bromide ion, even at very low concentrations (as low as 0.1
mg/1), greatly influences chlorination reactions (Rook 1974;  Morris 1975).
Bromide is very rapidly oxidized to hypobromous acid by aqueous chlorine.
The resulting HOBr reacts more rapidly and completely with organic matter
than does HOC1 (Morris 1975;  Rook 1974).  The bromide regenerated in the
reduction of HOBr is readily reoxidized by excess HOC1 and thus can serve
as a catalyst in  oxidation reactions (Morris 1975;  Rook 1974).

59.5  Data Summary

    Table 59-1 summarizes the aquatic fate data discussed above.  The
information found is insufficient  to predict the aquatic  fate of bromo-
dichloromethane.
                                    59-4

-------
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-------
59.6  Literature Cited

Anonymous.  1977.  Man is not the only polluter.  Environ.  Sci. Technol.
  ll(5):442-443.

Dreisbach, R.R.  1952.  Pressure-volume-temperature relationships of
  organic compounds.  Handbook Publishers, Inc., Sandusky,  Ohio.  135p.

Environmental Protection Agency.   1975.  Preliminary assessment of sus-
  pected carcinogens in drinking  water.  Environmental Protection Agency,
  (Office of Toxic Substances), Washington, D.C.  33p.  EPA 560/4-75-003.

Glaze, W.H. and J.E. Henderson, IV.  1975.  Formation of organochlorine
  compounds from the chlorination of a municipal secondary effluent.
  J. Water Pollut. Control Fed. 47:2511-2515.

Kleopfer, R.D. and B.J. Fairless.  1972.  Characterization of organic
  components in a municipal water supply.  Environ. Sci. Technol. 6(12):
  1036-1037.

Mabey, W. and T. Mill.  1978.  Critical review of hydrolysis of organic
  compounds in water under environmental conditions.  Phys. Chem. Ref. Data
  7(2):383-415.

Morris, J.C.  1975.  Formation of halogenated  organics by chlorination
  of water supplies.  Environmental Protection Agency (Office of Research
  and Development) 54p.  EPA-600/1-75-002.

Neely, W.B., D.R. Branson, and G.E. Blau.  1974.  Partition coefficient
  to measure bioconcentration potential of organic chemicals in fish.
  Environ. Sci. Technol. 8(13):1113-1115.

Rook, J.J.  1974.  Formation of haloforms during chlorination of natural
  waters.  J. Soc. Water Treat. Exam.  23(Part 2):234-243.

Rook, J.J.  1977.  Chlorination reactions of fulvic acids in natural
  waters.  Environ. Sci. Technol. 11(5):478-482.

Tute, M.S.  1971.  Principles and practices of Hansch analysis:  a guide
  to structure-activity correlation for the medicinal chemist.  Adv. Drug
  Res. 6:1-77.

Weast, R.C. (ed.).  1977.  Handbook of chemistry and physics.  58th
  Edition.  CRC Press, Cleveland, Ohio. 2398p.
                                    59-6

-------
                         60.  DIBROMOCHLOROMETHANE
60.1  Statement of Probable Fate

    Very little data on physical properties or reactions of dibromochloro-
methane were found in the literature.  It is, therefore, not possible to
determine the aquatic fate of dibromochloromethane at this time.

60.2  Identification

    Dibromochloromethane has been detected in finished drinking water
(Kleopfer and Fairless 1972;  Environmental Protection Agency 1975), in
drinking water supplies (Environmental Protection Agency 1975), and in
wastewater effluents (Glaze and Henderson 1975).  Dibromochloromethane is
hypothesized to be present in water supplies as a result of the haloform
reaction which takes place during the chlorination of such water (Rook
1974;  Environmental Protection Agency 1975;  Glaze and Henderson 1975).

    The chemical structure of dibromochloromethane is shown below.
             Br

              I                        Alternate Names
        CJ	 C 	Br
                                      Chi orodibromomethane
             H

    Dibromochloromethane

    CAS NO. 124-48-1
    TSL. NO. None Assigned

60.3  Physical Properties

    The general physical properties of dibromochloromethane are as follows,

    Molecular weight                         208.29
    (Weast 1977)

    Melting point                           <-20C
    (Verschueren 1977)

    Boiling point at 748 torr               119-120C
    (Weast 1977)

    Vapor pressure at 10.5C                15 torr
    (Dreisbach 1952)
                                    60-1

-------
    Solubility in water                     Not available

    Log octanol/water partition coefficient 2.09
    (Calculated from Tute 1971, see Methods
    Section on Bioaccumulation)

60.4  Summary of Fate Data

    60.4.1  Photolysis

         No information was found pertaining to the rate of photolysis of
dibromochloromethane in either the aquatic or atmospheric environments
under ambient conditions.

    60.4.2  Oxidation

         No information was found pertaining to the rate of oxidation of
dibromochloromethane in either the aquatic or atmospheric environments.
Dibromochloromethane is probably like other halogenated aliphatics in that
it is not easily oxidized in aquatic systems since there are no functional
groups which react strongly with OH.

    60.4.3  Hydrolysis

         A maximum hydrolytic half-life of 274 years has been reported for
dibromochloromethane at pH 7 and 25C (Mabey and Mill 1978).  This corres-
ponds to a first-order rate constant for hydrolysis of dibromochloromethane
of 8.0 x 10~11sec~1 (Mabey and Mill 1978).

    60.4.4  Volatilization

         No information was found pertaining specifically to the rate of
volatilization of dibromochloromethane from water.  Volatilization may be
an important transport process for dibromochloromethane from water to the
atmosphere.  It must be noted, however, that the vapor pressure of dibromo-
chloromethane is much lower than that for chloroform and other chloro-
alkanes.  As a result, dibromochloromethane would not be expected to
volatilize as rapidly as chlorinated methanes such as chloroform.  The con-
centration of dibromochloromethane present in water supplies has, however,
been reported to decrease as a result of volatilization while flowing
through open channels (Rook 1974).

    60.4.5  Sorption

         No information was found pertaining specifically to the adsorption
of dibromochloromethane onto suspended sediments.  The concentration of
dibromochloromethane produced by chlorination can be reduced by treatment
of water containing this compound with powdered activated carbon (Rook
1974).
                                    60-2

-------
    60.4.6  Bioaccumulation

         Neely _e_t al_. (1974) have shown that bioaccumulation is directly
related to the octanol/water partition coefficient (P) of the compound.
The log octanol/water partition coefficient (log P) as calculated by the
method of Hansch is 2.09 (Tute 1971) indicating that dibromochloromethane
is somewhat lipophilic (see Methods section on bioaccumulation).  As a re-
sult, dibromochloromethane may exhibit a tendency to bioaccumulate in
organisms.  No information was found, however, reporting the bioaccumula-
tion of dibromochloromethane in organisms.

    60.4.7  Biotransf ormation and Biodegradation

         No information was found pertaining specifically to the biodegra-
dation of dibromochloromethane.  In the sea, many species of brown seaweed,
molluscs, sponges, and bacteria metabolize halogens (Anonymous 1977).  The
red algae, or Rhodophyta, are known to use mostly bromine and they also
synthesize several different organic substances, some containing iodine and
chlorine.

    60.4.8  Other Reactions

         Rook (1977) suggested that dibromochloromethane, which has been
found in finished drinking water, may be present in such chlorinated water
due to the haloform reaction.  The haloform reaction, connected with the
treatment of water supplies, occurs generally in alkaline solution with
organic compounds containing the acetyl group or with structures that may
be readily oxidized to the acetyl group (Morris 1975).  The three hydrogens
of the methyl component of the acetyl group are successively replaced by
chlorine or other halogen, and then the carbon bond to the carbonyl group
is hydrolyzed giving rise to a haloform and a carboxylic acid (Morris
1975).  The simplified reaction is given below using production of chloro-
form as an example:
         CH3COR + 3 HOC1 - *>CCl3COR + 3 t^O    (1)
         CC13COR + H20 -  - i^CHCl3 + RCOOH        (2)
         It is suggested that organic compounds that have an acetyl group,
needed for the haloform reaction to occur, are present in the form of humic
materials which cause the yellow to brown coloration of surface waters
(Rook 1974;  Rook 1977;   Morris 1975).  For the most part humic materials
have a molecular weight in the 10^ - 10^ range and fall into the group
of compounds designated as fulvic acids (Rook 1977;   Morris 1975).

         There are several conditions which must be met in order for the
haloform reaction to occur.  These conditions are as follows.
                                    60-3

-------
          First,  the  haloform  reaction  generally  occurs  at  pH of  5  or
 greater.   Chlorine dispersed  in  water  at  pH  greater  than 5,  as employed  in
 water  treatment  in concentrations  up to 100  mg/1 or  about  103 molar,  is
 hydrolyzed  instantly and entirely  to HOC1 and OC1~ (Morris  1975).   The
 overall  reactions are  as follows:

          C12 + HC03	- HOC1 + Cl + C02

          HOC1                       H+ + OC1-

 Thus,  it  is necesasary  to consider the reactions  of  hypochlorous acid
 rather than those of C12 in describing the potential reactions of  dilute
 aqueous chlorine with  organic compounds.

          Secondly, the  haloform  reaction must occur  at  sites  on the  fulvic
 acids which are the most conducive to the formation  of  haloforms.  Although
 the carbon content of  fulvic acid  is about 50 percent,  there  are very few
 carbons which have the  chemical constituents necessary  for the haloform re-
 action to occur.  It is for this reason that the  yield  of haloforms
 produced upon aqueous chlorination is fairly low  (Rook  1977).  The carbon
 between two meta positioned OH groups of a hydroxylated aromatic ring is
 proposed as the most reactive site for haloform formation (Rook 1977).

         The bromide ion, even at very low concentrations (as low as 0.1
mg/1), greatly influences chlorination reactions  (Rook  1974;  Morris 1975).
Bromide is very rapidly oxidized to hypobromous acid  by aqueous chlorine.
The resulting HOBr reacts more rapidly and completely with organic matter
than does HOC1 (Morris  1975;  Rook 1974).   The bromide regenerated in the
reduction of HOBr is readily reoxidized by excess HOC1 and thus can serve
as a catalyst in oxidation reactions (Morris 1975;  Rook 1974).

60.5  Data Summary

         Table 60-1  summarizes the aquatic fate data  discussed above.  The
information found is insufficient to predict the  aquatic fate of  dibromo-
chloromethane.
                                     60-4

-------
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                             60-5

-------
60.6  Literature Cited

Anonymous.  1977.  Man is not the only polluter.   Environ. Sci. Technol.
  ll(5):442-443.

Dreisbach, R.R.  1952.  Pressure-volume-temperature relationships of
  organic compounds.  Handbook Publishers,  Inc.  Sandusky, Ohio.  135p.

Environmental Protection Agency.  1975.  Preliminary assessment of sus-
  pected carcinogens in drinking water.  Environmental Protection Agency,
  (Office of Toxic Substances), Washington, D.C.   33p.  EPA 560/4-75-003.

Glaze, W.H. and J.E. Henderson, IV.  1975.   Formation of organochlorine
  compounds from the chlorination of a municipal  secondary effluent.
  J. Water Pollut. Control. Fed.  47:2511-2515.

Kleopfer, R.D. and B.J. Fairless.  1972.  Characterization of organic
  components in a municipal water supply.  Environ. Sci. Technol.  6(12):
  1036-1037.

Mabey, W. and T. Mill.  1978.  Critical review of hydrolysis of organic
  compounds in water under environmental conditions.  J. Phys. Chetn. Ref.
  Data 7(2):383-415.

Morris, J.C.  1975.  Formation of halogenated organics by chlorination of
  water supplies.  Environmental Protection Agency (Office of Research and
  Development).  54p.  EPA 600/1-75-002.

Neely, W.B., D.R. Branson, and G.E. Blau.  1974.   Partition coefficient to
  measure bioconcentration potential of organic chemicals in fish.  En-
  viron. Sci. Technol. 8(13):1113-1115.

Rook, J.J.  1974.  Formation of haloforms during  chlorination of natural
  waters.  J. Soc. Water Treat. Exam. 23(Part 2):234-243.

Rook, J.J.  1977.  Chlorination reactions of fulvic acids in natural
  waters.  Environ. Sci. Technol. 11(5):478-482.

Tute, M.S.  1971.  Principles and practices of Hansch analysis:  a guide
  to structure-activity correlation for the medicinal chemist.  Adv. Drug
  Res.  6:1-77.

Verschueren, K.  1977.  Handbook of environmental data on organic chemi-
  cals.  Van Nostrand/Reinhold Press, New York.  659p.

Weast, R.C. (ed.).  1977.  Handbook of  chemistry and physics.  58th
  Edition.  CRC Press, Cleveland, Ohio. 2398p.
                                    60-6

-------
                     61.  TRIBROMOMETHANE (BROMOFORM)
61.1  Statement of Probable Fate

    The currently reviewed literature contains insufficient information to
indicate the aquatic fate of tribromomethane.

61.2  Identification

    Tribromoraethane has been detected in finished drinking water (Kleopfer
and Fairless 1972;  Environmental Protection Agency 1975) and in drinking
water supplies (Environmental Protection Agency 1975).  Tribromomethane is
hypothesized to be present in water supplies as a result of the haloform
reaction which takes place during the chlorination of such water (Rook
1974;  Environmental Protection Agency 1975;  Glaze and Henderson 1975).

    The chemical structure of tribromomethane is shown below.

          Br
          I                            Alternate Names

     r           r                    Bromoform
                                     Methenyl tribromide
          H

   Tribromomethane

   CAS NO. 75-25-2
   TSL NO. PB 56000

61.3  Physical Properties

    The general physical properties of tribromomethane are as follows.

    Molecular weight                         252.75
    (Weast 1977)

    Melting point                            8.3C
    (Weast 1977)

    Boiling point at 760 torr                149.5C
    (Weast 1977)

    Vapor pressure at 34C                   10 torr
    (Weast 1977)
                                    61-1

-------
    Solubility in water                      3010 mg/1 at 15C
    (Seidell 1941)                           3190 mg/1 at 30C

    Log octanol/water partition coefficient  2.30
    (Calculated from Tute 1971; see
    methods section on bioacccumulation)

61.4  Summary of Fate Data

    61.4.1  Photolysis

         No information pertaining to the rate of photolysis of tribromo-
methane in either the aquatic or atmospheric environments was found.

    61.4.2  Oxidation

         No information pertaining specifically to the rate of oxidation of
tribromomethane in the aquatic environment was found.  According to Radding
_e_t al. (1977), oxidation of tribromomethane by hydroxyl (OH") radicals in
the troposphere will be relatively rapid with a one or two month half-life.
The photooxidation product of tribromomethane is likely to be COBr2
which, in turn, may be removed by rain and be hydrolyzed to C02 and HBr
(Radding _et _al. 1977).  Assuming a troposphere-to-stratosphere turnover
time  (time for all but 1/e of tropospheric air to diffuse into the
stratosphere) of 30 years, a tropospheric half-life of one or two months
would result in less than one percent of tropospheric tribromoraethane
reaching the stratosphere.

    61.4.3  Hydrolysis

         A maximum hydrolytic half-life of 686 years has been estimated for
tribromomethane at pH 7 and 25 C by extrapolation from experimental data at
100-150C (Radding _et al. 1977).  This corresponds to a first-order rate
constant for hydrolysis of tribromomethane of 3.2 x 10~^sec~^ (Radding
e_t al. 1977;  Mabey and Mill 1978).

    61.4.4  Volatilization

         No information pertaining specifically to the rate of volatiliza-
tion  of tribroraomethane from water was found.  Volatilization may very well
be an important transport process for tribromomethane from water to the
atmosphere.  It must be noted, however, that the vapor pressure of tri-
bromomethane is much lower than that for chloroform (trichloromethane) and
many  other chloroalkanes.  As a result, tribromomethane would not be ex-
pected to volatilize as rapidly as chlorinated methanes.  The concentration
of tribroraomethane present in water supplies has been reported to decrease,
while flowing through open channels, as a result of volatilization (Rook
1974).
                                    61-2

-------
    61.4.5  Sorption

         No information pertaining to the adsorption of tribromomethane
onto sediments was found.

    61.4.6  Bioaccumulation

         Neely t al. (1974) have shown that bioaccumulation is related to
the octanol/water partition coefficient (P) of the compound.  The log oc-
tanol/water partition coefficient (log P) ,  as calculated by the method of
Hansch (Tute 1971), is 2.30, indicating that tribromomethane is some-
what lipophilic and that its bioaccumulation in organisms is possible.  No
laboratory or field information was found,  however, to verify the potential
bioaccumulation of tribromomethane.

    61.4.7  Biotransformation and Biodegradation

         No information pertaining specifically to the biodegradation of
tribromomethane was found.  In the sea many species of brown seaweed,
molluscs, sponges, and bacteria metabolize  halogens (Anonymous 1977).  The
red algae, or rhodophyta^ which are known to use mostly bromine, synthesize
several different halogenated organic substances.   A red seaweed of the
genus Asparagopsis produces tribromomethane at a concentration of 1% of the
total plant composition on a dry weight basis.  Since tribromomethane
occurs naturally in seaweed, it is possible that some species of micro-
organisms could metabolize tribromomethane  specifically.

    61.4.8  Other Reactions

         Thus far it has not been demonstrated whether mixed bromochloro-
forms and bromoforms arise as a result of the haloform reaction.  Nonethe-
less, bromide, even at very low concentrations (as low as 0.1 mg/1),
greatly influences chlorination reactions (Rook 1974;  Morris 1975).
Bromide is very rapidly oxidized to hypobromous acid by aqueous chlorine.
The resulting HOBr reacts more rapidly and  completely with organic matter
than does HOC1 (Morris 1975;  Rook 1974).  The bromide regenerated in the
reduction of HOBr is readily reoxidized by excess HOC1 and can thus serve
as a catalyst in oxidation reactions (Morris 1975;  Rook 1974).

         Rook (1977) has suggested that tribromoraethane (bromoform), which
has been found in finished drinking water,  may be present in such chlorin-
ated water due to the haloform reaction.  Stated simply, the haloform re-
action is an aqueous chlorination reaction which occurs generally in alka-
line solution with organic compounds containing the acetyl group or with
structures that may be readily oxidized to  the acetyl group.  In the
reaction, the hydrogens on the methyl component of the acetyl group are
                                    61-3

-------
successively replaced by chlorine or other halogen and then the carbon bond
to the carbonyl group is split via hydrolysis to produce a haloforra and a
carboxylic acid (Morris 1975). The simplified reaction is given below:

              CH3COR + 3HOC1  	*  CC13COR + 3H20    (1)
              CC13COR + H20 	+- CHC13 + RCOOH       (2)

         It is suggested that organic compounds having a structure equiva-
lent to an acetyl group are present in the form of humic materials (speci-
fically, fulvic acids) which cause the yellow to brown stain of surface
waters (Rook 1974;  Rook 1977;  Morris 1975).  Specifically, the carbon
between two meta-positioned OH-groups of a hydroxylated aromatic ring is
proposed as the most reactive site  for haloform formation (Rook 1977).

61.5  Data Summary

    Table 61-1 summarizes the aquatic fate data discussed above.  The
half-life for oxidation is a photooxidation half-life and is based on the
rate of reaction of tribromomethane with hydroxyl radicals in the tropo-
sphere.  The available data are insufficient to indicate the aquatic fate
of tribromomethane.
                                     61-4

-------
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No information on aqueous o
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                                61-5

-------
61.6  Literature Cited

Anonymous.  1977.   Man is not the only polluter.  Environ. Sci. Technol.
  ll(5):442-443.

Environmental Protection Agency.  1975.  Preliminary assessment of
  suspected carcinogens in drinking water.  U.S. Environmental Protection
  Agency, (Office of Toxic Substances), Washington, D.C.  33p.  EPA
  560/4-75-003.

Glaze, W.H. and J.E. Henderson, IV.  1975.  Formation of organochlorine
  compounds from the chlorination of a municipal secondary effluent.   J.
  Water Pollut. Control Fed.  47:2511-2515.

Kleopfer, R.D. and B.J. Fairless.  1972.  Characterization of organic
  components in a municipal water supply.  Environ. Sci. Technol.
  6(12):1036-1037.

Mabey, W. and T. Mill.  1978.  Critical review of hydrolysis of organic
  compounds in water under environmental conditions.  J. Phys. Chem. Ref.
  Data 7(2):383-415.

Morris, J.C.   1975.  Formation of halogenated organics by chlorination of
  water supplies.  U.S. Environmental Protection Agency, (Office of
  Research and Development), Washington, D.C. 54p.  EPA 600/1-75-002.

Neely, W.B., D.R. Branson, and G.E. Blau.  1974.  Partition coefficient  to
  measure bioconcentration potential of organic chemicals in fish.
  Environ. Sci. Technol. 8(13):1113-1115.

Radding, S.B., D.H. Liu, H.L. Johnson, and T. Mill.   1977.  Review of the
  environmental fate of selected chemicals.  U.S. Environmental Protection
  Agency, (Office of Toxic Substances), Washington, D.C. 147p.  EPA
  560/5-77-003.

Rook, J.J.  1974.  Formation of haloforms during chlorination of natural
  waters.  J.  Soc. Water Treat. Exam.  23(Part  2):234-243.

Rook, J.J.  1977.  Chlorination reactions of fulvic acids in natural
  waters.  Environ. Sci. Technol.  11(5):478-482.

Seidell, A. 1941.  Solubility of organic compounds.  3rd Edition, volume 2,
  Van Nostrand, New York.  1914p.

Tute, M.S.  1971.  Principles and  practice of Hansch analysis:  a guide  to
  structure-activity correlation for the medicinal chemist.  Adv. Drug
  Res.  6:1-77.

Weast, R.C.   (ed.).  1977.  Handbook of chemistry and physics.  58th
  Edition. CRC Press,  Inc., Cleveland, Ohio. 2398p.
                                    61-6

-------
                        62.  DICHLORODIFLUOROMETHANE
62.1  Statement of Probable Fate

    Dichlorodifluororaethane introduced into aqueous systems will most
likely volatilize to the atmosphere.  Once in the troposphere, dichlorodi-
fluoromethane remains stable.  It eventually diffuses into the stratosphere
or is carried back to the earth during the precipitation process.  Once in
the stratosphere dichlorodifluoromethane is photolyzed by shorter wave-
length, higher energy ultraviolet light with the subsequent formation of
chlorine atoms.

    Based on the information found it appears that oxidation is not an
important fate process for dichlorodifluoromethane in the aquatic environ-
ment.  No evidence was found for significant adsorption, bioaccumulation,
hydrolysis, or biodegradation.

62.2  Identification

    Dichlorodifluoromethane has been detected in surface snow in Alaska (Su
and Goldberg 1976) and in the atmosphere (Singh t _al. 1978;  Lillian et
a.1. 1975;  Cox ^t al. 1976;  Hanst 1978;  Environmental Protection Agency
1975;  National Research Council 1976;  Howard ej: _al. 1975;  Howard and
Durkin 1973;  Su and Goldberg 1976).

    The chemical structure of dichlorodifluoromethane is shown below.

                                      Alternate Names

                	Cl                 Fluorocarbon-12
                                      Freon-12

              F
    Dichlorodifluoromethane

    CAS NO. 75-71-8
    TSL NO. PA 82000

62.3  Physical Properties

    The general physical properties of dichlorodifluoromethane are as
follows.

    Molecular weight                         129.91
    (Weast 1977)
                                    62-1

-------
    Melting point                            -158C
    (Weast 1977)

    Boiling point at 760 torr                -29.8C
    (Weast 1977)

    Vapor pressure at 20C                   4306 torr
    (Pearson and McConnell 1975)

    Solubility in water at 25C              280 mg/1
    (Pearson and McConnell 1975)

    Log octanol/water partition coefficient  2.16
    (Hansch t _al. 1975)

62.4  Summary of Fate Data

    62.4.1  Photolysis

         No information was found pertaining specifically to the rate of
photolysis of dichlorodifluoromethane in the aquatic environment under
ambient conditions.

         Due to the high vapor pressure of dichlorodifluoromethane, volati-
lization to the atmosphere is quite rapid.  The compound is tropospheri-
cally stable (Environmental Protection Agency 1975;  Hanst 1975; Howard and
Durkin 1973); it does not react readily with hydroxyl radicals, nor does it
photodissociate in the tropsphere since it exhibits no absorption of light
greater than 200 nm (Hanst 1978; Howard et al. 1975).  Lovelock et al.
(1973) have suggested a tropospheric residence time of 30 years for di-
chlorodifluoromethane before diffusion to the stratosphere.

         In the stratosphere, dichlorodifluoromethane is broken down by the
absorption of higher energy, shorter wavelength ultraviolet light (Hanst
1978;  Rebbert and Ausloos 1975; Jayanty _et al. 1975).  The initial step in
photodissociation is the abstraction of a chlorine atom (Environmental
Protection Agency 1975):  CC12F2	^-	*- CC1F2 "*" C1"   Eventu-
ally, the photodissociation proceeds as follows:

CF2C12  hv  CF2C1 + Cl .frv  > :CF2 + C12 .hv> :CF2 + 2C1-

         Thus, according to Rebbert and Ausloos (1975) the photodissocia-
tion of dichlorodifluoromethane results in the release of two chlorine
atoms since less energy is required for the cleavage of the C-C1 bond than
for the cleavage of the OF bond.  According to Jayanty et al. (1975), the
                                    62-2

-------
photolysis of dichlorodifluoromethane in the presence of Oo at 213.9 nm
and 25C leads to the production of CF20 and d2 and, potentially,
chlorine atoms.  Chlorine atoms, released by reactions such as these, are
theorized to be catalysts in the destruction of the stratospheric ozone
layer (Hanst 1978;  Environmental Protection Agency 1975).

    62.4.2  Oxidation

         No information was found pertaining to the oxidation of dichloro-
difluoromethane in the aquatic environment under ambient conditons.  In
addition, dichlorodifluoromethane is known to be relatively stable with
respect to attack by hydroxyl radicals present in the troposphere (Environ-
mental Protection Agency 1975;  Hanst 1978;  Lillian ^t al. 1975; Cox et
al. 1976;  Howard t al. 1975;  Howard and Durkin 1973).  For instance, the
bimolecular rate of reaction for dichlorodifluoromethane with hydroxyl
radicals is less than 1 x lO'^cm^sec"!  with a corresponding life-
time (time for reduction to 1/e of the original concentration) of greater
than 330 years (Cox^t_al. 1976).  According to Howard ejt _al. (1975),
fluorocarbon compounds are highly resistant to attack by conventional ox-
idizing agents at temperatures below 200C.

         Assuming a troposphere-to-stratosphere turnover time (time for all
but 1/e of tropospheric air to diffuse into the stratosphere) of 30 years,
a tropospheric lifetime of 30 years would result in over 90 percent of
tropospheric dichloridifluoromethane eventually reaching the stratosphere.
Jesson et_ _al. (1977) and Sze and Wu (1976) have postulated a shorter tropo-
spheric lifetime, on the order of 20 years, based on analysis of the data
of other workers in the field, but no known destructive mechanisms in the
troposphere (sinks) are identified as being capable of removing dichloro-
difluoromethane.  A tropospheric lifetime of 20 years would indicate that
about 40 percent of tropospheric dichlorodifluoromethane would reach the
stratosphere.

    62.4.3  Hydrolysis

         The fluorocarbons as a group exhibit a low rate of hydrolysis in
comparison to other halogenated compounds, and the rates of hydrolysis are
greatly affected by temperature, pressure, and the presence of catalytic
materials such as metals (Howard _e_t _al. 1975).  The rate of hydrolysis of
dichlorodifluoromethane at 1 atmosphere of pressure and 30C was reported
to be not discernable by the analytical technique used in the work cited
(Howard _e_t _a_l. 1975), although the compound is known to hydrolyze under
more severe conditions such as high pressure, high temperature, presence of
high concentrations of  metals, or any combination of these (Howard ^ al.
1975; Hagen and Elphingstone 1974).  On the basis that conditions in the
aquatic environment typically would more nearly correspond to pure water
                                    62-3

-------
than to the more severe hydrolytic conditions mentioned above, it is
postulated that hydrolysis of dichlorodifluoromethane, if it occurs, would
proceed at a negligible rate in comparison to the rate of volatilization
and subsequent photodissociation.

    62.4.4  Volatilization

         The high vapor pressure and low aqueous solubility of dichlorodi-
fluoromethane are conducive to rapid volatilization from an aquatic en-
vironment.  Although no specific data were found, volatilization is thought
to be the major transport process for dichlorodifluoromethane from aqueous
systems.

    62.4.5  Sorption

         No information was found pertaining to the adsorption of dichloro-
dif luoromethane onto sediments.  Although the high vapor pressure indicates
that volatilization is the major transport process for dichlorodifluoro-
methane in the aquatic environment, the log octanol/water partition coeffi-
cient (log P) of 2.16 for this compound (Hansch et al. 1975) indicates that
adsorption onto organic particulates may be possible.  In brief, data are
inconclusive as to whether or not dichlorodifluoromethane significantly
adsorbs onto sediments.

    62.4.6  Bioaccumulation

         Relatively little information was found pertaining specifically to
the bioaccuraulation of dichlorodif luoromethane.  Neely e_t _al_. (1974) have
shown that bioaccumulation is directly related to the octanol/water parti-
tion coefficient (P) of the compound.  The experimentally determined log
octanol/water partition coefficient (log P) of dichlorodifluoromethane is
reported to be 2.16 (Hansch ejt al.  1975) indicating that dichlorodif luoro-
methane is lipophilic and that bioaccumulation in organisms may be possible
under conditions of constant exposure.  Evidence that dichlorodifluoro-
methane does not significantly bioaccuraulate in organisms was given by
Howard et_ _a_l. (1975).  The fact that dichlorodifluororaethane has a high
vapor pressure (4306 mm Hg at 20C), and consequently is very volatile,
seems to exclude persistence in the aquatic environment.

    62.4.7  Biotransformation and Biodegradation

         No information was found pertaining to the biodegradability of
dichlorodif luoromethane.  According to Howard &t_ a]^. (1975) the volatility
of dichlorodifluoromethane, as well as other fluorocarbons, would limit, if
not preclude, biodegradation.  Su and Goldberg (1976), however, report that
dichlorodifluoromethane, as well as other synthetic organic compounds, is
persistent in natural waters.
                                    62-4

-------
62.5  Data Summary

    Table 62-1 summarizes the aquatic fate discussed above.  Oxidation
rates are photooxidation rates and refer to the rate at which dichloro-
difluoromethane is attacked by hydroxyl radicals in the troposphere. Di-
chlorodifluoromethane is rapidly volatilized, and its predominant fate is
photolysis in the atmosphere. Once in the troposphere, dichlorodifluoro-
methane remains stable as a result of its extremely slow rate of reaction
with hydroxyl radicals, so that much of the compound in the troposphere
eventually diffuses upward to the stratosphere or is carried back to the
earth during the precipitation process.  Once in the stratosphere, di-
chlorodif luoromethane is photolyzed by shorter wavelength, higher energy
ultraviolet light.
                                     62-5

-------
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                                     62-6

-------
62.6  Literature Cited

Cox, R.A.,  R.G. Derwent, A.E.J. Eggleton, and J.E.Lovelock.   1976.   Photo-
  chemical  oxidation of halocarbons in the troposphere.   Atmos.  Environ.
  10:305-308.

Environmental Protection Agency.  1975.  Report on the problem of halo-
  genated air pollutants and stratospheric ozone.  Environmental Protec-
  tion Agency (Office of Research and Development), Research Triangle
  Park, North Carolina.  55p.  EPA 600/9-75-008.

Hagen, A.P. and E.A. Elphingstone.  1974.  The high pressure hydrolysis of
  CFC13 and CF2Cl2.  J. Inorg. Nucl. Chem. 36:509-511.

Hansch, C., A. Vittoria, C. Silipo, and P.Y.C. Jow.  1975.  Partition co-
  efficients and the structure-activity relationship of  the  anesthetic
  gases.  J. Med. Chem.  18(6):546-548.
Hanst, P.L.  1978.  Part II:  Halogenated pollutants.
  in the air.  Chemistry  51(2):6-12.
Noxious trace gases
Howard, P.H. and P.R. Durkin.  1973.  Preliminary environmental hazard
  assessment of chlorinated naphthalenes, silicones, fluorocarbons,
  benzenepolycarboxylates,  and chlorophenols.   Environmental Protection
  Agency (Office of Toxic Substances), Washington, D.C.  263p.
  EPA 560/2-74-001.

Howard, P.H., P.R. Durkin,  and A. Hanchett.  1975.  Environmental hazard
  assessment of one and two carbon fluorocarbons.   Environmental Protec-
  tion Agency (Office of Toxic Substances), Washington, D.C.  246p.
  EPA 560/2-75-003.

Jayanty, R.K.M., R. Simonaitis, and J. Heicklen.  1975.  The photolysis of
  chlorofluoromethanes in the presence of 02 or 03 at 213.9 nm and
  their reactions with O^D).  J. Photochem. 4:381-398.

Jesson, J.P., P. Meakin, and L.C. Glasgow.  1977.   The fluorocarbon-ozone
  theory-II.  Tropospheric  lifetime-an estimate of the tropospheric
  lifetime of CC13F.  Atmos. Environ.  11(6):499-508.

Lillian, D., H.B. Singh, A. Appleby, L. Lobban, R. Arnts, R. Gumpert,  R.
  Hague, J. Toomey, J. Kazazis, M. Antell, D.  Hansen, and B. Scott.    1975.
  Atmospheric fates of halogenated compounds.   Environ. Sci. Technol.
  9:1042-1048.
                                    62-7

-------
Lovelock, J.E., R.J. Maggs, and R.J. Wade.   1973.   Halogenated hydro-
  carbons in and over the Atlantic.  Nature  241:194-196.

National Research Council.  1976.  Environmental effects of chloro-
  fluoromethane release.  National Academy  of Sciences, Washington,B.C.
  125p.

Neely, W.B., D.R. Branson, and G.E. Blau.   1974.  Partition coefficient
  to measure bioconcentration potential of  organic chemicals in fish.
  Environ. Sci. Technol. 8(13):1113-1115.

Pearson, C.R. and G. McConnell.  1975.  Chlorinated Cj and C2
  hydrocarbons in the marine environment.   Proc. Roy.  Soc. London B
  189:305-322.

Rebbert, R.E. and P.J. Ausloos.  1975.  Photodecomposition of CFC13
  and CF2Cl2.  J. Photochem.  4:419-434.

Singh, H.B., L.J. Salas, H. Shiegeishi, and A.H. Smith.  1978.  Fate of
  halogenated compounds in the atmosphere  interim report-1977.  Environ-
  mental Protection Agency (Office of Research and Development), Re-
  search Triangle Park, N.C.  57 p. EPA 600/3-78-017.

Su, C. and E.D. Goldberg.  1976.  Environmental concentrations and fluxes
  of some halocarbons.  In: "Marine Pollutant Transfer".  H.L. Windom and
  R.A. Duce (Eds).  Lexington Books, D.C.  Health and Company, Lexington,
  Massachusetts,  p. 353-374.

Sze, N.D. and M.F. Wu.  1976.  Measurements of fluorocarbons 11 and 12 and
  model validation: an assessment.  Atraos.  Environ. 10(12):1117-1125.

Weast, R.C. (ed.).  1977.  Handbook of chemistry and physics.  58th
  Edition.  CRC Press, Cleveland, Ohio. 2398p.
                                    62-8

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                        63.  TRICHLOROFLUOROMETHANE
63.1  Statement of Probable Fate

    Volatilization is the major transport process for removal of trichloro-
fluoromethane from aquatic systems.  In the troposphere, trichlorofluoro-
methane remains stable, but it eventually diffuses into the stratosphere or
is carried back to the earth during the precipitation process.  Once in the
stratosphere, trichlorofluoromethane probably reacts with shorter wave-
length, higher energy ultraviolet light to produce chlorine atoms. These
chlorine atoms are theorized to be catalysts in the destruction of the
stratospheric ozone layer.

    Based on the information reviewed it does not appear that oxidation is
an important fate process of trichlorofluororaethane in the aquatic environ-
ment.  Evidence for adsorption, bioaccumulation, hydrolysis, and biodegra-
dation is inconclusive.

63.2  Identification

    Trichlorofluoromethane has been detected in surface waters (Su and
Goldberg 1976), in surface snow in Alaska (Su and Goldberg 1976), in sea-
water (Lovelock et al. 1973), and in the atmosphere (Singh jst _al. 1978;
Lillian &t_ al. 1975;   Cox t al. 1976;  Hanst 1978;  Environmental Pro-
tection Agency 1975;  National Research Council 1976;  Lovelock 1971; Howard
et _al. 1975;  Howard and Durkin 1973;  Lovelock 1975;  Su and Goldberg
1976).

    The chemical structure of trichlorofluoromethane is shown below.


              Cl                       Alternate Names

                 Cl                  Fluorocarbon-11
                                      Freon-11

              Cl

    Trichlorofluoromethane

    CAS NO. 75-69-4
    TSL NO. PB 61250
                                    63-1

-------
63.3  Physical Properties

    The general physical properties of trichlorofluoromethane are as
follows.

    Molecular weight                         137.4
    (Verschueren 1977)

    Melting point                            -111C
    (Verschueren 1977)

    Boiling point at 760 torr                23.8C
    (Verschueren 1977)

    Vapor pressure at 20C                   667.4 torr
    (Pearson and McConnell 1975)

    Solubility in water                      1,100 mg/1*

    Log octanol/water partition coefficient  2.53
    (Hansch et al. 1975)
*More than one value for the solubility of trichlorofluoromethane in water
were found in the literature.  The values are  reported to be 1,100 mg/1 at
20C (Pearson and McConnell 1975) and 1,100 mg/1 at 25C (Verschueren
1977).

63.4  Summary of Fate Data

    63.4.1  Photolysis

         No information was found pertaining to the rate of photolysis of
trichlorofluoromethane in the aquatic environment.

         Due to the high vapor pressure of trichlorofluororaethane, volatil-
ization to the atmosphere is quite rapid.  Since trichlorofluromethane has
been found in rainwater (Su and Goldberg 1976), it seems likely that some
trichlorofluoromethane is washed out of the troposphere onto the  earth
during the precipitation process.  Trichlorofluoromethane is reported to be
tropospherically stable (Environmental Protection Agency 1975;  Hanst 1978;
Lillian et_ _al. 1975; Cox et_ al. 1976;  Howard _et al. 1975;  Howard and
Durkin 1973) and does not react readily with hydroxyl radicals present in
the troposphere.  Lovelock et al. (1973) have suggested a residence time of
10 years, assuming no significant tropospheric degradation and complete
destruction in the stratosphere.  Trichlorofluororaethane does not photo-
dissociate in the troposphere since  it does not significantly absorb
wavelengths of light greater than 200 nm (Hanst 1978;  Howard el: _al. 1975).
                                    63-2

-------
         The tropospherically stable trichlorofluoromethane eventually dif-
fuses into the stratosphere (Hanst 1978;  National Research Council 1976)
where it is broken down by the attack of higher energy, shorter wavelength
ultraviolet light (Hanst 1978;  Environmental Protection Agency 1975;
Rebbert and Ausloos 1975;  Jayanty et al . 1975).  Lillian et_ al. (1975)
estimate a minimum stratospheric half-life for trichlorofluoromethane of
69 hours.  The initial step in the photodissociation of trichlorofluoro-
methane is the abstraction of a chlorine atom (Environmental Protection
Agency 1975):  CC^F  ..   > -CC^F + Cl-   Eventually, the
photodissociation proceeds as follows:
              CFC13    -CFCl2 + Cl- -   :CFC1

                             C12 -^ 2C1-

Thus, according to Rebbert and Aasloos (1975) the photodissociation of tri-
chlorofluoromethane results in the release of two chlorine atoms since less
energy is required for the cleavage of the C-C1 bond than for the cleavage
of the C-F bond.  According to Jayanty e^ al. (1975) the photolysis of tri-
chlorofluoromethane in the presence of 02 at 213.9 nm and 25C leads to
the production of CFC10 and, potentially, chlorine molecules (Cl2)
Chlorine atoms, released by reactions such as these, are theorized to be
catalysts in the destruction of the stratospheric ozone layer (Hanst 1978;
Environmental Protection Agency 1975).

    63.4.2   Oxidation

         No information was found pertaining to the oxidation of trichloro-
fluoromethane in the aquatic environment under ambient conditions. Tri-
chlorofluoromethane does not react significantly with hydroxyl radicals
present in the troposphere (Environmental Protection Agency 1975;  Hanst
1978;  Lillian e^, al.  1975;  Cox et_ ail. 1976;  Howard e al. 1975;  Howard
and Durkin 1973) and, as a result, is tropospherically stable.  For in-
stance, the bimolecular rate of reaction for trichlorofluoromethane with
hydroxyl radicals is reported to be less than 1.0 x 10~* cm^s~* which
corresponds to a reported tropospheric lifetime (time for reduction to 1/e
of the original concentration) of much greater than 1000 years (Cox et al .
1976).  According to Howard et_ al_. (1975), fluorocarbon compounds are
highly resistant to attack by conventional oxidizing agents at temperatures
below 200C.

         In contrast,  Jesson e_t a_l. (1977) interpret the available data as
indicating a tropospheric lifetime for trichlorofluoromethane of 15 to 20
years, although no mechanism of tropospheric destruction (sink) is identi-
fied.  Sze and Wu (1976) state that this shorter tropospheric lifetime
                                     63-3

-------
cannot be ruled out.  Assuming a tropospheric-to-stratospheric turnover
time (time for all but 1/e of the tropospheric air to diffuse into the
stratosphere) of 30 years, the shorter lifetime corresponds to about 40
percent of tropospheric  trichlorofluoromethane reaching the stratosphere,
in comparison to over 95 percent for a tropospheric lifetime of 1000 years.

    63.4.3  Hydrolysis

         No specific information was found indicating that trichlorofluoro-
methane hydrolyzes sufficiently rapidly under ambient conditions for this
process to be an important environmental fate of the compound.  The fluoro-
carbons as a group exhibit a low rate of hydrolysis in comparison to other
halogenated compounds (Howard _e_t _al. 1975).  The rate of hydrolysis is
greatly affected by temperature, pressure, and the presence of other
materials, such as high concentrations of metals which act as catalysts.
In pure water, the rate of hydrolysis of trichlorofluoromethane at atmo-
spheric pressure and 30C is reported to be not discernable by the analy-
tical method used in the work cited (Howard _et _al. 1975).  Trichlorofluoro-
methane  hydrolyzes more rapidly under conditions of high pressure, high
temperature, presence of catalyses, or any combination of these (Howard et
al. 1975;  Hagen and Elphingstone 1974).  In summary, trichlorofluoro-
methane would be expected to undergo hydrolysis under ambient conditions at
a rate of volatilization and subsequent photodissociation.

    63.4.4  Volatilization

         The high vapor pressure (667.4 torr at 20C), low solubility
(1,100 mg/1 reported at 20 and 25C) and low boiling point (23.8C at 760
torr) of trichlorofluoromethane are conducive to rapid volatilization from
an aquatic environment.  Though no specific data were found, volatilization
is thought to be the major transport process for removal of trichloro-
fluoromethane from aqueous systems.

    63.4.5  Sorption

         No information was found pertaining to the adsorption of tri-
chlorofluoromethane onto sediments.  Although volatilization is believed to
be the major transport process for this compound, the octanol/water parti-
tion coefficient (log P=2.53) of trichlorofluoromethane  indicates that
adsorption onto sediments may be possible.  In brief, data are inconclusive
as to whether or not trichlorofluoromethane significantly adsorbs onto
sediments.

    63.4.6  Bioaccumulation

         Relatively little information was found pertaining specifically to
the bioaccumulation of trichlorofluoromethane.  Dickson  and Riley (1976)
                                     63-4

-------
have found trichlorofluoromethane at levels of 0.1 to 5.0 ppb, (dry weight
basis) in various organs of fish and molluscs.  These levels, however, do
not necessarily indicate a potential for bioaccumulation.

         Neely et al. (1974) have theorized that bioaccumulation is
directly related to the octanol/water partition coefficient  (P) of the com-
pound.  The experimentally determined log octanol/water partition coeffi-
cient (log P) of trichlorofluoromethane is reported to be 2.53 (Hansch et
jjj.. 1975), indicating that trichlorofluoromethane is lipophilic and may
possibly bioaccumulate in organisms.

         Howard _et al. (1975) suggest that the high volatility of tri-
chlorofluoromethane (vapor pressure 667.4 torr at 20C) seems to make
persistence in the aquatic environment unlikely.

    63.4.7  Biotransformation and Biodegradation

         No information was found pertaining to the biodegradability of
trichlorofluoromethane.  According to Howard _et _al. (1975) the rapid rate
of volatilization expected for trichlororaethane, as well as other
fluorocarbons, would limit, if not preclude, biodegradation.

63.5  Data Summary

    Table 63-1 summarizes the aquatic fate discussed above.  The oxidation
rate presented is a photooxidation rate and refers to the rate at which
trichlorofluoromethane is attacked by hydroxyl radicals in the troposphere.

    The primary fate of trichlorofluoromethane is photodissociation in the
stratosphere, a direct result of the rapid volatilization of this compound
into the atmosphere.  Once in the troposphere, trichlorofluoromethane re-
mains stable and eventually diffuses upward to the stratosphere or is
carried back to earth during the precipitation process.  Once in the
stratosphere, trichlorofluoromethane reacts with shorter wavelength, higher
energy ultraviolet light to eventually form chlorine atoms which are
theorized to serve as a catalyst in destruction of the stratospheric ozone
layer.
                                    63-5

-------
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63.6  Literature Cited

Cox, R.A., R.G. Derwent, A.E.J. Eggleton and J.E.  Lovelock.   1976.
  Photochemical oxidation of halocarbons in the troposphere.   Atmos.
  Environ. 10:305-308.

Dickson, A.G. and J.P. Riley.  1976.  The distribution of short-chain
  halogenated aliphatic hydrocarbons in some marine organisms.  Mar.
  Pollut. Bull. 7(9):167-169.

Environmental Protection Agency.  1975.  Report on the problem of
  halogenated air pollutants and stratospheric ozone.  Environmental
  Protection Agency, (Office of Research and Development), Research
  Triangle Park, North Carolina.  55p.  EPA 600/9-75-008.

Hagen, A.P. and E.A. Elphingstone.  1974.  The high presure hydrolysis of
  CFC13 and CF2C12.  J. Inorg. Nucl. Chem.  36:509-511.

Hansch, C., A. Vittoria, C. Silipo, and P.Y.C. Jow.  1975.  Partition
  coefficients and the structure-activity relationship of the anesthetic
  gases.  J. Med. Chera. 18(6):546-548.

Hanst, P.L. 1978.  Part II:  Halogenated pollutants.  Noxious trace gases
  in the air.  Chemistry 51(2):6-12.

Howard, P.H. and P.R. Durkin.  1973.  Preliminary environmental hazard
  assessment of chlorinated naphthalenes, silicones, fluorocarbons,
  benzenepolycarboxylates, and chlorophenols.  Environmental  Protection
  Agency (Office of Toxic Substances), Washington, D.C. 263p.  EPA
  560/2-74-001.

Howard, P.H., P.R. Durkin, and A. Hanchett.  1975.  Environmental hazard
  assessment of one and two carbon fluorocarbons.   Environmental Protection
  Agency, (Office of Toxic Substances), Washington, D.C.  246 p.  EPA
  560/2-75-003.

Jayanty, R.K.M., R. Simonaitis, and J. Heicklen.  1975.  The  photolysis of
  chlorofluoromethanes in the presence of 02 or 03 at 213.9 nm and
  their reactions with 0(^-0).  J. Photochem. 4:381-398.

Jesson, J.P., P. Meakin, and L.C. Glasgow.  1977.   The fluorocarbon-ozone
  theory-II.  Tropospheric lifetime-an estimate of the tropospheric
  lifetime of CC^F.  Atmos. Environ.   11(6) :499-508.
                                    63-7

-------
Lillian, D. ,  H.B. Singh, A. Appleby, L. Lobban, R. Arnts, R. Gumpert, R.
  Hague, J. Toomey, J. Kazazis, M. Antell,  D. Hansen, and B. Scott.  1975.
  Atmospheric fates of halogenated compounds.  Environ.  Sci. Technol
  9:1042-1048.

Lovelock, J.E.  1971.  Atmospheric fluorine compounds as indicators of air
  movements.   Nature 230:379.

Lovelock, J.E.  1975.  Natural halocarbons  in the air and in the sea.
  Nature 256:193-194.

Lovelock, J.E. ,  R.J. Maggs, and R.J. Wade.   1973.  Halogenated hydrocarbons
  in and over the Atlantic.  Nature 241:194-196.

National Research Council.  1976.  Environmental effects of
  chlorofluoromethane release.  National Academy of Sciences, Washington,
  D.C.  125p.

Neely, W.B. ,  D.R. Branson, and G.E. Blau.  1974.  Partition coefficient to
  measure bioconcentration potential of organic chemicals in fish.
  Environ. Sci.  Technol. 8( 13) : 1113-1 115.

Pearson, C.R. and G. McConnell.  1975.   Chlorinated C^ and ^2
  hydrocarbons in the marine environment.  Proc. Roy. Soc. London B
  189:305-322.

Rebbert, R.E. and P.J. Ausloos.  1975.   Photodecomposition of CFC13 and
           J. Photochem. 4:419-434.
Singh, H.B., L.J. Salas, H. Shiegeishi, and A.H. Smith.  1978.  Fate of
  halogenated compounds in the atmosphere interim report - 1977.
  Environmental Protection Agency, (Office of Research and Development),
  Research Triangle Park, N.C.  57 p.  EPA 600/3-78-017.

Su, C. and E.D. Goldberg.  1976.  Environmental concentrations and fluxes
  of some halocarbons.  ir Marine Pollutant Transfer,  pp. 353-374.  H.L.
  Windom and R.A.  Duce (eds.).  Lexington Books, D.C. Heath and Company,
  Lexington, Massachusetts.

Sze, N.D. and M.F. Wu.  1976.  Measurements of fluorocarbons 11 and 12 and
  model validation: an assessment.  Atmos. Environ. 10(12) :1117-1125.

Verschueren, K.  1977.  Handbook of environmental data on organic
  chemicals.  Van Nostrand/Reinhold Co. , New York.  659p.
                                    63-8

-------
SECTION VI:  HALOGENATED ETHERS
        Chapters 64-70

-------
                        64.   BIS(CHLOROMETHYL)ETHER


64.1  Statement of Probable Fate

    Bis(chloromethyl)ether (BCME) hydrolyzes very rapidly (half-life 10-40
seconds) on contact with water.  Although BCME also may be volatilized or
oxidized, these processes proceed at a much slower rate, and are thus not
competitive with hydrolysis.  Hydrolytic half-lives of BCME in organisms
are probably comparably short thus precluding the existence of food chain
transfers.

64.2  Identification

    The chemical structure of bis(chloromethyl)ether is shown below.
         H           H

         I            I                        Alternate Names
  a  c    o  c  ci
         I            I                        BCME
         '            I                        Bis-CME
                                             sym-Dichloromethyl ether
    Bis(chloromethyl)ether                   Oxybis(chloromethane)

    CAS NO. 542-88-1
    TSL NO. KN 15750

64.3  Physical Properties

    The general physical properties of bis(chloromethyl)ether are given be-
low.
    Molecular weight                         114.96
    (Weast 1977)

    Melting point                            -41.5C
    (Weast 1977)

    Boiling point at 760 torr                104C
    (Weast 1977)

    Vapor pressure at 22C                   30 torr
    (Dreisbach 1952)

    Solubility in water at 25C              22,000 mg/1
    (Calc. by method of Moriguchi 1975
    using the data of Quayle 1953)

    Log octanol/water partition coefficient  -0.38
    (Calc. by Radding ej: _al. 1977)
                                   64-1

-------
64.4  Summary of Fate Data

    64.4.1  Photolysis

         Direct photolysis would not be expected to occur in surface waters
or the troposphere since bis(chloromethyl)ether does not possess any chro-
mophores that absorb radiation in the visible or near ultraviolet regions
of the electromagnetic spectrum (Jaffe and Orchin 1962).  No information
was found that would suggest photolysis as an environmental fate process,

    64.4.2  Oxidation

         Radding _e_t al. (1977) have calculated a rate constant of <10~^
1.  mole"* sec~l, corresponding to a half-life of >230 hours, for the
reaction of hydroxyl radical with bis(chloromethyl)ether.  Although oxida-
tion is a mechanism by which BCME can be degraded, it obviously is not
kinetically competitive with hydrolysis in aqueous systems.  No information
pertaining to the reaction of BCME with molecular oxygen in water was
found.  In the atmosphere, however, oxidation may compete with hydrolysis
as a mechanism for BCME degradation at low humidity levels inasmuch as Tou
and Kallos (1976) have reported a hydrolytic half-life of greater than 20
hours in moist air.  The estimated half-life for oxidation by atmospheric
hydroxyl radical is less than one day;  and for oxidation by either ozone
or' alkylperoxyl radical, it is more than one year (Radding _e_t _a_l. 1977).

    64.4.3  Hydrolysis

         Tou et al. (1974), by careful measurements with 1 ppm BCME dis-
solved in water, showed that BCME hydrolyzes in circumneutral water inde-
pendently of pH, and that the rate constant at 20C is 0.018 sec"1,
corresponding to a half-life of approximately 38 seconds.  Several other
investigations corroborate this value (Van Duuren et_ _al. 1972;  Nichols and
Merritt 1973;  Tou and Kallos 1974;  Alvarez and Rosen 1976;  Tou and
Kallos 1976).  BCME can be formed by a reverse reaction between aqueous
formaldehyde and HC1 (Frankel t al. 1974).  The yield after 12-24 hours is
given by the empirical expression:

            (BCME) = K(CH20)(HC1)

where K is between 1.3 x 10"^ and 7 x 10"^ for concentrations of for-
maldehyde from 100 to 4000 ppm and HC1 from 100 to 40,000 ppm at 40% re-
lative humidity and 26C,  It should be noted that these concentrations of
formaldehyde and HCl are unlikely to be encountered in environmental sur-
face waters.
                                     64-2

-------
    64.4.4  Volatilization

         The rate of hydrolysis of bis(chloromethyl)ether is extremely
rapid when the material is dissolved in water or a mixture of an organic
solvent and water.  Any bis(chloromethyl)ether in an industrially generated
waste stream apparently would never reach the natural environment if the
stream contained water.  Presumably, BCME could volatilize rapidly from an
aquatic system only if it were discharged in a water-immiscible solvent
with a high vapor pressure.

    64.4.5  Sorption

         There is little information on the adsorption of BCME onto parti-
culates.  Radding e_t _al_.  (1977) report a calculated log octanol/water
partition coefficient (log P) value of -0.38 for bis(chloromethyl)ether.
On the basis of this value it is suggested that very little adsorption is
indicated for BCME on humus.  No conclusion can be drawn about adsorption
on clay or other minerals,

    64.4.6  Bioaccumulation, Biotransformation and Biodegradation

         No information on bioaccumulation or biodegradation has been found
in the reviewed literature.  BCME is much too toxic and hydrolytically un-
stable to permit such observations.

64.5  Data Summary

    Table 64-1 summarizes the aquatic fate data discussed above for bis-
(chloromethyl)ether.  The aquatic fate is very rapid hydrolysis.  Oxida-
tion by photolytically produced hydroxyl radical becomes a competing
process only in the atmosphere.
                                      64-3

-------
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-------
64,6  Literature Cited

Alvarez, M. and R.T. Rosen.  1976.  Formation and decomposition of
  bis(chloromethyl)ether in aqueous media.  Int.  J.  Environ.  Anal.  Chem.
  4(3):241-246.

Dreisbach, R.R.  1952.  Pressure-volume-temperature  relationships of
  organic compounds.  Handbook Publishers Inc.  Sandusky,  Ohio 349p.

Frankel, L.S., K.S. McCallum and L. Collier.  1974.   Formation of
  bis(chloromethyl)ether from formaldehyde and hydrogen chloride.  Environ.
  Sci. Technol.  8(4):356-359.

Jaffe, H.H. and M. Orchin.  1962.  Theory and application of  ultraviolet
  spectroscopy.  John Wiley and Sons, New York.  253p.

Moriguchi, I.  1975.  Quantitative structure activity studies.  Parameters
  relating to hydrophobicity.  Chem. Pharm. Bull.  23(2):247-257.

Nichols, R.W. and R.F. Merritt.  1973.  Relative  solvolytic reactivities  of
  chloromethyl ether and bis(chloromethyl)ether.   J. Nat.  Cancer Inst.
  50:1373-1374.

Quayle, O.R.  1953.  The parachors of organic compounds.   Chem. Rev.
  53:439-589.

Radding, S.B., D.H. Liu, H.L. Johnson, and T. Mill.   1977.  Review of  the
  environmental fate of selected chemicals.  U.S. Environmental Protection
  Agency, Office of Toxic Substances, Washington, B.C.   147p. (EPA
  560/5-77-003).

Tou, J.C. and G.J. Kallos.  1974.  Aqueous HC1 and formaldehyde mixtures
  for formation of bis(chloromethyl)ether.  Amer. Ind.  Hyg. Assoc.  J.
  35(7):419-422.

Tou, J.C. and G.J. Kallos.  1976.  Possible formation of
  bis(chloromethyl)ether from reactions of formaldehyde and chloride  ion.
  Anal. Chem.  48(7) :958-963.

Tou, J.C., L.B. Westover, and L.F. Sonnabend.  1974.  Kinetic studies  of
  bis(chloromethyl)ether hydrolysis by mass spectrometry.   J. Phys. Chem.
  78(11):1096-1098.

Van Duuren, B.L., C. Katz, B.M. Goldschmidt, K. Frenkel,  and  A. Sivak.
  1972.  Carcinogencity of halo ethers.  II.  Structure-activity
  relationships of analogs of bis(chloromethyl)ether.  J.  Nat. Cancer  Inst.
  48:1431-1439.

Weast, R.C.  1977.  CRC handbook of chemistry and physics.  58th edition.
  CRC Press Inc., Cleveland, Ohio.  2398 p.
                                   64-5

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                       65.  BIS(2-CHLOROETHYL)ETHER
65.1  Statement of Probable Fate

    Based on available information, it is not possible to determine the
most probable aquatic fate of bis(2-chloroethyl)ether.  The relative impor-
tance of volatilization in comparison to other processes is not known for
this compound.  In the event that a portion of bis(2-chloroethyl)ether
should enter the atmosphere, it will probably undergo photodestruction in
the troposphere.  Slow hydrolysis of the carbon-chlorine bonds may provide
the greatest contribution to the aquatic fate of this pollutant.  No infor-
mation was found from which any conclusion regarding biodegradation in sur-
face waters can be drawn.

65.2  Identification

    Bis(2-chloroethyl)ether has been detected in raw and finished drinking
water, and industrial effluents (Shackelford and Keith 1976).  The chemical
structure of bis(2-chloroethyl)ether is shown below.
            H           H     H

                               '              Alternate Names
o  c  c  o  c  c  a
       II           II              l,l'-0xybis(2-chloroethane)
      /,     H           H     H              Bis(B-chloroethyl)ether
                                             Chlorex
                                             l-Chloro-2-(&-chloroethoxy)-
                                               ethane
Bis(2-chloroethyl)ether

CAS No. 111-44-4
TSL No. KN 08750

65.3  Physical Properties

    The general physical properties of bis(2-chloroethyl)ether are as
follows.

    Molecular weight                         143.02
    (Weast 1977)

    Melting point                            -46.8C
    (Weast 1977)

    Boiling point at 760 torr                178C
    (Weast 1977)
                                    65-1

-------
    Vapor pressure at 20C                   0.71 torr
    (Verschueren 1977)

    Solubility in water*                     10,200 mg/1
    (Verschueren 1977)

    Log octanol/water partition coefficient  1.58
    (Gale, by Leo et al. 1971)
Experimental data generated at room temperature; no specific temperature
reported.

65.4  Summary of Fate Data

    65.4.1  Photolysis

         Direct photolysis would not be expected to occur in surface waters
or the troposphere since bis(2-chloroethyl)ether does not possess any chro-
mophores that absorb radiation in the visible or near ultraviolet regions
of the electromagnetic spectrum (Jaffa and Orchin 1962).  No information
was found that would suggest photolysis as an environmental fate process,

    65.4.2  Oxidation

         Although water may have an inhibitory effect on the formation of
ether peroxides, it apparently does not prevent their formation (Patai
1967).  Since no information discussing the formation of peroxides from
ethers and molecular oxygen in dilute aqueous solutions was found, it is
uncertain whether ether peroxides form in the aquatic environment.  In-
direct photolysis, involving abstraction of alkyl hydrogens by the hydroxyl
radicals normally present in surface waters, is considered to be too slow
to be environmentally relevant (Dorfman and Adams 1973).

         The relative importance of volatilization in comparison to other
processes is unknown for this compound.  In the event that a portion of
bis(2-chloroethyl)ether should enter the atmosphere, it will probably
undergo photodestruction in the troposphere.  From the smog chamber studies
of Altshuller et_ al. (1962) and Laity et. al. (1973), it can be inferred
that  the half-life with respect to photodestruction in a smog chamber for
ethyl ether should be four hours.  Since oxidation reactions of alkyl
ethers involve carbon-hydrogen scission at the carbon atom adjacent to the
ether linkage, it can be expected that bis(2-chloroethyl)ether will also
have  a half-life of about four hours under similar conditions.  It must be
emphasized, however, that half-lives based on smog chamber data, do not
take  into account all of the meteorological variables encountered in a
natural environmental airshed.
                                65-2

-------
    65.4.3  Hydrolysis

         A reaction medium of concentrated mineral acid is usually re-
quired for the solvolysis of dialkyl ethers to proceed at a measurable rate
(Fieser and Fieser 1956).  Hydrolytic cleavage of dialkyl ethers is thus
environmentally irrelevant.  The only other covalent bonds that are capable
of hydrolytic cleavage are the carbon-chlorine bonds.  Bohme and Sell
(1948) report a first order rate constant of chloride hydrolysis for bis-
(2-chloroethyl)ether in aqueous dioxane at 100C as 1.5 x 10   min"-*-.
No environmentally relevant kinetic data were found for the hydrolysis of
this compound.  Dilling ^t al. (1975) reported that the half-lives with
respect to hydrolysis for one and two carbon chloroaliphatic compounds are
six months to several years.  Bis(2-chloroethyl)ether may have a similar
hydrolytic half-life.

    65.4.4  Volatilization

         The vapor pressure of bis(2-chloroethyl)ether (0.71 torr at 20C)
suggests that it might be sufficiently volatile to be transported into the
atmosphere.  Using the approach of Mackay and Wolkoff (1973), Durkin et al.
(1975) calculated the half-life with respect to volatilization for bis(2-
chloroethyl)ether from a body of water to be 5.78 days.  (It should be
noted that this method of calculating evaporative half-lives is not uni-
versally accepted). In view of this haloether's solubility in water (10,200
mg/1), it appears likely that it could be precipitated from the atmosphere
with rain and, in this manner, continuously recycle between surface water
and atmosphere until it is destroyed.  Because of its water solubility,
some migration through the soil may occur.

         Some information on the volatility of bis(2-chloroethyl)ether in a
natural aquatic environment has been provided in an indirect way by
Kleopfer and Fairless (1972).  These investigators monitored the concen-
tration of bis(2-chloroisopropyl)ether in the Ohio River 150 miles from an
industrial outfall, and found that the pollutant was present at approxi-
mately the expected level calculated from the dilution factors that would
obtain during river transport.  (The calculated concentration of pollutant
was 1.8 yg/1 and the measured concentrations were within the range of 0.5
to 5.0 yg/1).  This observation suggests that neither sedimentary sorption,
volatilization, nor biodegradation were overtly operative during transport
of this haloether over that particular 150 miles of the Ohio River.  The
vapor pressures at 20C of bis(2-chloroethyl)ether and bis(2-chloroiso-
propyDether are 0.71 torr and 0.85 torr, respectively, and their solubili-
ties are 10,200 mg/1 and 1,700 mg/1, respectively.  By inference, bis(2-
chloroethyl)ether would be expected to be similarly  transported without
appreciable evaporation.
                                      65-3

-------
    65.4.5  Sorption

         With the exception of the field study of Kleopfer and Fairless
(1972) on the transport of bis(2-chloroisopropyl)ether in the Ohio River,
no information specifically pertaining to the relevance of sorption pro-
cesses for beta-haloalkyl ethers within the aquatic environment was found.
The results of this field study suggests that low molecular weight beta-
haloalkyl ethers do not readily become immobilized as part of the bed sedi-
ment in a river system.  The solubility of bis(2-chloroethyl)ether (10,200
mg/1) and the value of its log octanol/water partition coefficient (1.58,
Leo e_t al. 1971) indicate little potential for adsorption on suspended
organic matter.

    65.4.6  Bioaccumulation

         No information indicating that bis(2-chloroethyl)ether will bio-
accumulate was found.  Moreover, Metcalf and Sanborn (1975) maintain that
compounds with solubilities of 50 mg/1 or more generally have little poten-
tial for aquatic bioaccumulation.  This statement is based on the direct
dependence of aquatic partitioning processes upon relative solubilities;
compounds soluble to the extent of 50 mg/1 have been empirically observed
to have little partitioning preference for biological systems or organic
particulates.

    65.4.7  Biotransformation and Biodegradation

         No information was found from which any conclusion regarding
biodegradation in surface waters can be reached with any degree of confi-
dence.  Ludzack and Ettinger (1963) found that significant degradation of
bis(2-chloroethyl)ether, which had been added to Ohio River water supple-
mented with settled sewage, occurred only after a 25-30 day period of accli-
mation.  Kleopfer and Fairless (1972) reported no degradation of bis(2-
chloroisopropyl)ether five days after it had been added to Ohio River
water.

65.5  Data Summary

    Table 65-1 summarizes the preceding discussion.  Atmospheric photo-
oxidation contributes an indeterminate amount to the pollutant's destruc-
tion, and hydrolysis of the carbon-chlorine bonds is a slow but perhaps
significant process.  The uncertain relationship of volatilization to the
destructive fate processes precludes a determination of the probable
aquatic fate  of bis(2-chloroethyl)ether.
                                       65-4

-------
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65.6  Literature Cited

Altshuller, A.P., I.R. Cohen, S.F. Sleva, and S.L. Kopczynski.  1962.  Air
  pollution:  photooxidation of aromatic hydrocarbons.  Science
  !38(3538):442-443.

Bb'hme, H. and K. Sell.  1948.  Die hydrolyse halogenierter ather und
  thioather in dioxan-wasser-gemischen.  Chem. Ber.  81(2):123-130.

Dilling, W.L., N.B. Tefertiller, and G.J. Kallos.  1975.  Evaporation rates
  of methylene chloride, chloroform, 1,1,1-trichloroethane, trichloro-
  ethylene, tetrachloroethylene, and other chlorinated compounds in dilute
  aqueous solutions.  Environ. Sci. Technol.  9(9):833-838.

Dorfman, L.M. and G.E. Adams.  1973.  Reactivity  of the hydroxyl radical  in
  aqueous solution.  NSRDS-NBS-46.  NTIS:COM-73-50623.  Springfield, Va.

Durkin, P.R., P.H. Howard, and J. Saxena.  1975.  Investigation of
  selected potential environmental contaminants:  haloethers.  U.S.
  Environmental Protection Agency, Office of Toxic Substances, Washington,
  B.C.  168p.  (EPA 560/2-75-006).

Fieser, L.F. and M. Fieser.   1956.  Organic chemistry.  3rd Edition.  D.C.
  Heath and Co., Boston, Mass.  1112p.

Jaffa, H.H. and M. Orchin.   1962.  Theory and application of  ultraviolet
  spectroscopy.  John Wiley and Sons, New York.   253p.

Kleopfer, R.D. and B.J. Fairless.   1972.  Characterization of organic
  compounds in a municipal water supply.  Environ. Sci. Technol.
  6(12)1036-1037.

Laity, J.L., I.G. Burstain, and B.R. Appel.  1973.  Photochemical smog and
  the atmospheric reactions of solvents.  Chap.  7, pp. 95-122.  Solvents
  Theory and Practice.  R.W. Tess (ed.).  Advances in Chemistry Series  124.
  Am. Chem. Soc., Washington, D.C.

Leo, A., C. Hansch and D. Elkins.   1971.  Partition coefficients and their
  uses.  Chem. Rev.  71:525-612.

Ludzack, F.J. and M.B. Ettinger.  1963.  Biodegradability of  organic
  chemicals isolated from rivers.  Purdue Univ.,  Eng. Bull..,  Ext. Ser. No.
  115:278-282.  (Abstract only).  CA.   1965.  62:2609g.

Mackay, D. and A.W. Wolkoff.  1973.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.   Environ. Sci. Technol.
  7(7):611-614.
                                  65-6

-------
Metcalf, R.L. and J.R. Sanborn.  1975.   Pesticides  and  environmental
  quality in Illinois.  111. Nat'l.  Hist.  Survey Bull.   31:381-436.

Patai, S.  1967.  The chemistry of the  ether  linkage.   Interscience
  Publishers, New York.  785p.

Shackelford, W.M. and L.H.  Keith.  1976.   Frequency of  organic  compounds
  identified in water.  U.  S. Environmental Protection  Agency,  (ERL),
  Athens, Ga.  617p.  (EPA 600/4-76-062).

Verschueren, K.  1977.  Handbook of  environmental data  on organic
  chemicals.  Van Nostrand/Reinhold, New York.   659p.

Weast, R.C.  1977.  CRC handbook of  chemistry and physics.   58th Edition.
  CRC Press, Inc., Cleveland, Ohio.   2398p.
                                    65-7

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                     66.   BIS(2-CHLOROISOPROPYL)ETHER
66.1  Statement of Probable Fate

    Based on available information,  it is not possible to determine the
most probable aquatic fate of bis(2-chloroisopropyl)ether.   The relative
importance of volatilization in comparison to other processes  is not known
for this compound.  In the event that a portion of  bis(2-chloroisopropyl)-
ether should enter the atmosphere, it will probably undergo photodestruc-
tion in the troposphere.   Slow hydrolysis of the carbon-chlorine bonds  may
provide the greatest contribution to the aquatic fate of this  pollutant.
No information was found from which  any conclusion  regarding biodegradation
in surface waters can be drawn.

66. 2  Identification

    Bis(2-chloroisopropyl)ether has  been detected in finished  drinking
water, surface waters, industrial effluents, and sea water (Shackelford and
Keith 1976).  The chemical structure of bis(2-chloroisopropyl)ether is
shown below.
Cl
                    H
                              CI
CH
                  CH
                              H
                                              Alternate Names

                                       Bis(2-chloro-l-methylethyl)ether
                                       2,2'-Oxybis(l-chloropropane)
                                       Dichlorodiisopropyl ether
                                       2, 2'-Dichloroisopropyl ether
    Bis(2-chloroisopropyl)ether

    CAS NO. 108-60-1
    TSL NO. KN 17500

66. 3  Physical Properties

    The general physical properties of bis(2-chloroisopropyl)ether are  as
follows.
Molecular weight
(Weast 1977)

Melting point
(Verschueren 1977)

Boiling point at 760 torr
(Verschueren 1977)

Vapor pressure at 20 C
(Verschueren 1977)
                                             171.07
                                             -97 C
                                             189C
                                             0.85 torr
                                    66-1

-------
    Solubility in water*                     1,700 mg/1
    (Verschueren 1977)

    Log octanol/water partition coefficient  2.58
    (Calc. by Leo et al. 1971)
*Experimental data generated at room temperature; no specific temperature
 reported.

66.4  Summary of Fate Data

    66.4.1  Photolysis

         Direct photolysis would not be expected to occur in surface waters
or the troposphere since bis(2-chloroisopropyl)ether does not possess any
chromophores that absorb radiation in the visible or near ultraviolet re-
gions of the electromagnetic spectrum (Jaffa and Orchin 1962).  No in-
formation was found that would suggest photolysis as an environmental fate
process.

    66.4.2  Oxidation

         Although water may have an inhibitory effect on the formation of
ether peroxides, it apparently does not prevent their formation (Patai
1967).  Since no information discussing the formation of peroxides from
ethers and molecular oxygen in dilute aqueous solutions was found, it is
uncertain whether ether peroxides form in the aquatic environment.  In-
direct photolysis, involving abstraction of alkyl hydrogens by the hydroxyl
radicals normally present in surface waters, is considered to be too slow
to be environmentally relevant (Dorfman and Adams 1973).

         The relative importance of volatilization in comparison to other
processes is unknown for this compound.  In the event that a portion of
bis(2-chloroisopropyl)ether should enter the atmosphere, it will probably
undergo photodestruction in the troposphere.  From the smog chamber studies
of Altshuller t a_l. (1962) and Laity _et _al. (1973), it can be inferred
that the half-life with respect to photodestruction in a smog chamber for
ethyl ether should be four hours.  Since oxidation reactions of alkyl
ethers involve carbon-hydrogen scission at the carbon atom adjacent to the
ether linkage, it can be inferred that bis(2-chloroisopropyl)ether will
also have a half-life of about four hours under similar conditions.  It
must be emphasized, however, that temporal stabilities based on smog cham-
ber data do not take into account all of the meteorological variables en-
countered in a natural environmental airshed.
                                    66-2

-------
    66.4.3  Hydrolysis

         A reaction medium of concentrated mineral acid is usually re-
quired for the solvolysis of dialkyl ethers to proceed at a measurable rate
(Fieser and Fieser 1956).  Hydrolytic cleavage of dialkyl ethers is thus
environmentally irrelevant.  The only other covalent bonds that are capable
of hydrolytic cleavage are the carbon-chlorine bonds.  Bohme and Sell
(1948) report a first order rate constant of chloride hydrolysis for bis-
(2-chloroethyl)ether in aqueous dioxane at 100C as 1.5 x 10"^ min~l.
No environmentally relevant kinetic data were found for the hydrolysis of
this compound.  Billing _ejt al. (1975) reported that the half-lives with
respect to hydrolysis for one and two carbon chloroaliphatic compounds are
six months to several years.  Bis(2-chloroisopropyl)ether may have a cor-
responding rate of hydrolysis.

    66,4.4  Volatilization

         The vapor pressure of bis(2-chloroisopropyl)ether (0.85 torr at
20C) suggests that it might be sufficiently volatile to be transported
into the atmosphere.  Using the approach of Mackay and Wolkoff (1973),
Durkin _e_t _al. (1975) calculated the half-life with respect to volatiliza-
tion for bis(2-chloroisopropyl)ether from a body of water to be 1.37 days.
(It should be noted that this method of calculating evaporative half-lives
is not universally accepted). In view of this haloether's solubility in
water (1,700 mg/1), it appears likely that it could be precipitated from
the atmosphere with rain and, in this manner, continuously recycle between
surface water and atmosphere until it is destroyed.  Because of its water
solubility, some migration through the soil may occur.

         Some information on the volatility of bis(2-chloroisopropyl)ether
in a natural aquatic environment has been provided by Kleopfer and Fairless
(1972).  These investigators monitored the concentration of bis(2-chloro-
isopropyl)ether in the Ohio River 150 miles from an industrial outfall, and
found that the pollutant was present at approximately the expected level
calculated from the dilution factors that would obtain during river trans-
port.  (The calculated concentration of pollutant was 1.8 yg/1 and the
measured concentrations were within the range of 0.5 to 5.0 yg/1).  This
observation suggests that neither sedimentary sorption, volatilization, nor
biodegradation were overtly operative during transport of this haloether
over that particular 150 miles of the Ohio River.

    66.4.5  Sorption

         With the exception of the field study of Kleopfer and Fairless
(1972) on the transport of bis(2-chloroisopropyl)ether in the Ohio River,
no information specifically pertaining to the relevance of sorption pro-
cesses for beta-haloalkyl ethers within the aquatic environment was found.
The results of this field study indicated that low molecular weight beta-
                                     66-3

-------
haloalkyl ethers do not readily become immobilized as part of the bed sedi-
ment in a river system.  The value of the log octanol/water partition co-
efficient of bis(2-chloroisopropyl)ether 2.58, (Leo et al. 1971) does indi-
cate, however, some potential for adsorption on suspended organic matter.

    66.4.6  Bioaccumulation

         No information indicating that bis(2-chloroisopropyl)ether will
bioaccumulate was found.  Moreover, Metcalf and Sanborn (1975) maintain
that compounds with solubilities of 50 mg/1 or more generally have little
potential for aquatic bioaccumulation.  The basis of this statement is the
strong dependence of bioaccumulation upon the partitioning processes in
aquatic systems; compounds which exhibit solubility in water of 50 mg/1 (or
greater) have been empirically observed not to partition preferentially to
biological systems or organic particulates.

    66.4.7  Biotransformation and Biodegradation

         No information was found from which any conclusion regarding
biodegradation in surface waters can be reached with any degree of confi-
dence.  Ludzack and Ettinger (1963) found that significant degradation of
bis(2-chloroethyl)ether, which had been added to Ohio River water  supple-
mented with settled sewage, occurred only after a 25-30 day period of
acclimation.  Kleopfer and Fairless (1972) reported no detectable degrada-
tion of bis(2-chloroisopropyl)ether five days after it had been added to a
sample of Ohio River water.

66.5  Data Summary

    Table 66-1 summarizes the preceding discussion.  Atmospheric photo-
oxidation contributes an indeterminate amount to the pollutant's destruc-
tion, and hydrolysis of the carbon-chlorine bonds is a slow process.  The
uncertain relationship of volatilization to the destructive fate processes
precludes a determination of the probable aquatic fate of bis(2-chloro-
isopropyl)ether.
                                    66-4

-------
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66.6  Literature Cited

Altshuller, A.P., I.R. Cohen,  S.F. Sleva,  and  S.L.  Kopczynski.   1962.   Air
  pollution:  photooxidation of  aromatic hydrocarbons.   Science.
  138(3538).-442-443.

Bb'hme, H. and K. Sell.  1948.   Die hydrolyse halogenierter ather  und
  thioather in dioxan-wasser-gemischen.  Chem.  Ber.   81(2):123-130.

Dilling, W.L., N.B. Tefertiller, and G.J.  Kallos.   1975.   Evaporation  rates
  of methylene chloride, chloroform, 1,1,1-trichloroethane,  trichloro-
  ethylene, tetrachloroethylene, and other chlorinated  compounds  in  dilute
  aqueous solutions.  Environ. Sci. Technol.  9(9):833-838.

Dorfman, L.M. and G.E. Adams.   1973.  Reactivity of  the hydroxyl  radical in
  aqueous solution.  NSRDS-NBS-46.  NTIS:COM-73-50623.   Springfield, Va.

Durkin, P.R., P.H. Howard, and J. Saxena.   1975.  Investigation of
  selected potential environmental contaminants:  haloethers.   U.S.
  Environmental Protection Agency, Office  of Toxic  Substances,  Washington,
  D.C.  168p.  (EPA 560/2-75-006).

Fieser, L.F. and M. Fieser.  1956.  Organic chemistry.   3rd  Edition.  D.C.
  Heath and Co., Boston, Mass.  1112p.

Jaffa  H.H. and M. Orchin.  1962.^  Theory  and  application of ultraviolet
  spectroscopy.  John Wiley and Sons, New  York.  253p.

Kleopfer, R.D. and B.J. Fairless.  1972.  Characterization of  organic
  compounds in a municipal water supply.  Environ.  Sci. Technol.
  6(12)1036-1037.

Laity, J.L., I.G. Burstain, and B.R. Appel.  1973.   Photochemical smog and
  the atmospheric reactions of solvents.  Chap. 7,  pp.  95-122.   Solvents
  Theory and Practice.  R.W. Tess (ed.).  Advances  in Chemistry Series 124.
  Am. Chem. Soc., Washington,  D.C.

Leo, A., C. Hansch and D. Elkins.  1971.  Partition  coefficients  and their
  uses.  Chem. Rev.  71:525-612.

Ludzack, F.J. and M.B. Ettinger.  1963.   Biodegradability of organic
  chemicals isolated from rivers.  Purdue  Univ., Eng. Bull., Ext. Ser. No.
  115:278-282.  (Abstract only).  CA.  1965.  62:2609g.

Mackay, D. and A.W. Wolkoff.  1973.  Rate  of evaporation of  low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci.  Technol.
  7(7):611-614.

Metcalf, R.L. and J.R. Sanborn.  1975.  Pesticides  and  environmental
  quality in Illinois.  111. Nat'l. Hist.  Survey Bull.   31:381-436.

                                     66-6

-------
Patai, S.  1967.   The chemistry of the ether linkage.   Interscience
  Publishers, New York.  785p.

Shackelford, W.M. and L.H. Keith.   1976.   Frequency of  organic  compounds
  identified in water.  U. S. Environmental  Protection  Agency,  (ERL),
  Athens, Ga.  617p.  (EPA 600/4-76-062).

Verschueren, K.  1977.  Handbook of environmental  data  on  organic
  chemicals.  Van Nostrand/Reinhold, New  York.   659p.

Weast, R.C.  1977.  CRC handbook of chemistry and  physics.  58th Edition.
  CRC Press, Inc., Cleveland, Ohio.  2398p.
                                      66-7

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                      67.   2-CHLOROETHYL VINYL ETHER
67. 1  Statement of Probable Fate

    Based on the available information, it appears that the predominant
process for removal of 2-chloroethyl vinyl ether from the aquatic environ-
ment is volatilization to the atmosphere.  That portion of the pollutant
which volatilizes to the troposphere is probably destroyed rapidly by
photooxidation of the oxygen- substituted double bond.  It should be noted,
however, that the solubility of 2-chloroethyl vinyl ether is relatively
high;  consequently, persistence of some 2-chloroethyl vinyl ether is to be
expected.  Although the role of sorption onto clays and humic materials
cannot be established based on the reviewed literature, such sorption
processes could provide heterogeneous sites for acid-catalyzed hydrolysis.
No information pertaining to biodegradation was found.

67.2  Identification

    2-Chloroethyl vinyl ether has been detected in industrial effluents
near Louisville, Kentucky (Shackelford and Keith 1976).  The chemical
structure of 2-chloroethyl vinyl ether is shown below.

                                             Alternate Names
       Cl     H           H    H
              ,
                             c _ H       (2-Chloroethoxy)-ethene
                                             Vinyl 2-chloroethyl ether
       H      H

      2-Chloroethyl vinyl ether

      CAS NO. 110-75-8
      TSL NO. KN 63000

67.3  Physical Properties

    The general physical properties of 2-chloroethyl vinyl ether are as
follows.

    Molecular weight                         106.55
    (Weast 1977)

    Melting point                            No data found

    Boiling point at 760 torr                108C
    (Weast 1977)
                                    67-1

-------
    Vapor pressure at 20C                   26.75 torr
    (Gale, from Dreisbach 1952)

    Solubility in water at 25C               15,000 mg/1
    (Calc. by method of Moriguchi 1975)

    Log octanol/water partition  coefficient  1.28
    (Calc. by method of Leo e_t _al.  1971)

67.4  Summary of Fate Data

    67.4.1  Photolysis

         2-Chloroethyl vinyl ether  would not be expected to undergo direct
photolysis in surface waters or  the troposphere since the compound does not
possess any chromophores that absorb radiation in the visible or near
ultraviolet regions of the electromagnetic spectrum (Jaffe and Orchin
1962).  No information was found that would suggest photolysis as an en-
vironmental fate process.

    67.4.2  Oxidation

         The estimated vapor pressure of 26.75 torr at 20C (Dreisbach
1952) indicates that volatilization will be a major transport process for
the removal of this compound from water.  Although it is very likely that
this pollutant will be reprecipitated with water during the formation of
rain, its expected rate of destruction in the troposphere probably makes
atmospheric photooxidation the major fate process.  Organic molecules with
unsaturated double bonds are the most reactive compounds that have been
studied in simulated smog chambers  (Altshuller et al. 1962;  Laity et al.
1973).  It has been extensively demonstrated that the mechanism for de-
composition involves electrophilic  attack by hydroxyl radical, ozone, or
other oxidants on the double bond.   Thus, reactivity generally increases
with substitution of electron-donating groups on the two carbon atoms of
the double bond.  Altshuller t  al. (1962) have reported that the half-
conversion time for the disappearance of ethylene and m-xylene under the
conditions employed for smog chamber studies is approximately four hours.
From this value and the table of relative reactivities given by Laity et
al. (1973), it can be inferred that the corresponding half-conversion times
for ethyl ether and 2-methyl-2-butene (two extensively studied compounds)
would be four hours and 30 minutes, respectively.  The decomposition of
2-chloroethyl vinyl ether will undoubtedly proceed more easily than ethyl
ether and may have a half-life closer to 30 minutes.  It must be emphasized
that the temporal stability of 2-chloroethyl vinyl ether under actual
atmospheric conditions is unknown.   Experiments performed in laboratory
                                    67-2

-------
irradiation chambers are usually conducted for relatively short periods and
cannot account for all of the meteorological variables within an
environmental airshed.

    67.4.3  Hydrolysis

         The first order rate constant for the hydrolysis of 2-chloroethyl
vinyl ether at 25C and pH 7 is 4.4 x 10~10 sec"* (Jones and Wood
1964).  This first order rate would correspond to a maximum half-life of
0.48 years.  Since this hydrolysis is second order with respect to hydrogen
ion concentration and, in addition, exhibits general acid catalysis
(Salomaa et ail. 1966;  Loudon and Ryono 1975), the hydrolytic half-life may
vary considerably within the limits of the ambient aquatic environment.
For example, the small amount of fully protonated phosphoric acid
(H3?04) still existing at neutral and slightly basic conditions can
contribute a measurable catalytic effect to the hydrolysis of vinyl ethers
(Loudon and Ryono 1975).  Even though the log octanol/water partition
coefficient is calculated as 1.28, some adsorbtion of this polar molecule
can be expected by suspended clays and humic materials.  Inasmuch as the
cations of the clay surface are Lewis acids (Gabel and Ponnamperuma 1967),
and the structure of humic materials apparently contains many phenolic acid
groups (Rook 1977), general acid catalysis may occur at these sites.  The
extent to which heterogeneous acid catalysis could contribute to the
hydrolysis of 2-chloroethyl vinyl ether is, however, unknown and must
remain, at this point, purely conjectural.

    67.4.4  Volatilization

         Although no information pertaining specifically to the volatiliza-
tion of 2-chloroethyl vinyl ether was found, the vapor pressure of 26.75
torr at 20C, calculated from data given by Dreisbach (1952), indicates
that volatilization will be important in the transport of this pollutant
from the aquatic environment into the atmosphere.  In view of this
haloether's solubility in water (15,000 mg/1), it appears likely that it
will be precipitated from the atmosphere with rain, so that a continuous
cycling of this compound will occur until it is destroyed.  Because of its
water solubility, some migration through the soil may occur.

    67.4.5  Sorption

         No information specifically pertaining to sorption processes with
environmental significance was found.  The log octanol/water partition
coefficient is calculated as 1.28 and the solubility as 15,000 mg/1 indi-
cating that there is little potential for partitioning of this aquatic
pollutant  into suspended lipophilic material;  however, it is possible that
this polar, unsaturated molecule could become transitorily sorbed by
suspended  clays and huinic materials.
                                   67-3

-------
    67.4.6  Bioaccumulation

         No information was found indicating that 2-chloroethyl vinyl ether
will bioaccumulate.   Moreover,  Metcalf and Sanborn (1975)  maintain that
compounds with solubilities of  50 mg/1 or more generally have little poten-
tial for aquatic bioaccumulation.  This statement is based on the empiri-
cally observed effect of aquatic partitioning processes upon bioaccumula-
tion; compounds with solubilities of 50 mg/1 (or greater)  do not partition
preferentially to biological systems or organic particulates.

    67.4.7  Biotransformation and Biodegradation

         No information.was found from which any conclusion regarding
biodegradation can be reached with any degree of confidence.

    67.4.8  Other Reactions

         Chlorination of 2-chloroethyl vinyl ether can introduce a chlorine
substituent at either carbon atom that is adjacent to the ether linkage
(Summers 1955).  As a consequence of this reaction, 2-chloroethyl vinyl
ether can be modified to an alpha-chloroether in a water treatment facil-
ity.  alpha-Chloroethers are very hydrolytically unstable and usually de-
compose within seconds (Summers 1955;   Tou and Kallos 1974).

67.5  Data Summary

    Table 67-1 summarizes the aquatic fate data discussed above.  The
predominant transport process for removal of 2-chloroethyl vinyl ether from
the aquatic environment appears to be volatilization to the atmosphere.
Photooxidation in the troposphere should be very rapid, and it is probably
the main fate process.  Although the extent of sorption onto clays and
humic materials cannot be established, such sorption processes could
provide heterogeneous sites for acid-catalyzed hydrolysis.
                                   67-4

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67.6  Literature Cited

Altshuller, A.P., I.R. Cohen,  S.F.  Sleva,  and  S.L.  Kopczynski.   1962.   Air
  pollution:  photooxidation of aromatic hydrocarbons.   Science
  138(3538)-.442-443.

Dreisbach, R.R.  1952.  Pressure-volume-temperature relationships of
  organic compounds.  3rd edition.   Handbook Publishers,  Inc.,  Cleveland,
  Ohio.  349p.

Gabel, N.W. and C. Ponnamperuma.  1967.   Model for  origin of
  monosaccharides.  Nature 216:453-455.

Jaffe, H.H. and M. Orchin.  1962.  Theory and  application of ultraviolet
  spectroscopy.  John Wiley and Sons,  New York.  253p.

Jones, D.M. and N.F. Wood.  1964.  The mechanism of vinyl ether hydrolysis.
  J. Chem. Soc.  5400-5403.

Laity, J.L., I.G. Burstain, and B.R. Appel.   1973.   Photochemical smog and
  the atmospheric reactions of solvents.  Chap. 7,  pp.  95-112.   Solvents
  Theory and Practice.  R.W. Tess (ed.).  Advances  in Chemistry Series 124.
  Am. Chem. Soc., Washington,  D.C.

Leo, A., C. Hansch and D. Elkins.  1971.  Partition coefficients and their
  uses.  Chem. Rev.  71:525-612.

Loudon, G.M. and D.E. Ryono,  1975.  An unusual rate law for vinyl ether
  hydrolysis.  Observation of H3?04 catalysis at high pH.  J. Org.
  Chem. 40(24):3574-3577.

Metcalf, R.L. and J.R. Sanborn.  1975.  Pesticides  and  environmental
  quality in Illinois.  111. Nat'l. Hist. Survey Bull.   31:381-436.

Moriguchi, I.  1975.  Quantitative structure-activity studies on parameters
  related to hydrophobicity.  Chem. Phann. Bull.  23:247-257.

Rook, J.J.  1977.  Chlorination reactions of fulvic acids in natural
  waters.  Environ. Sci. Technol.  11(5):477-482.

Shackelford, W.M. and L.H. Keith.  1976.  Frequency of  organic compounds
  identified in water.  U.S. Environmental Protection Agency, (ERL),
  Athens, GA.  617p.  (EPA 600/4-76-062).

Salomaa,  P., A. Kankaanpera, and M. Lajunen,  1966.  Protolytic cleavage
  of vinyl ethers.  Acta Chem. Scand. 20(7):1790-1801.
                                      67-6

-------
Summers, L.  1955.  The alpha-haloalkyl ethers.  Chem. Rev.  55:301-353.

Tou, J.C. and G.J. Kallos.  1974.  Study of aqueous HC1 and formaldehyde
  mixtures for formation of bis(chloromethyl)ether.  Am. Ind. Hyg. Assoc,
  J. 35(7):419-422.

Weast, R.E. (ed.).  1977.  Handbook of chemistry and physics.  58th ed.
  CRC Press, Inc., Cleveland, Ohio.  2398p.
                                    67-7

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                     68.   4-CHLOROPHENYL PHENYL ETHER
68.1  Statement of Probable Fate

    It is not possible to determine the most probable aquatic fate for
4-chlorophenyl phenyl ether from available data.  This pollutant is re-
ported to be rapidly degraded by acclimated sewage sludge, but biodegrada-
tion data from river-water die-away experiments indicate that this compound
has a potential for persistence in natural surface waters.  Sorption by
organic-rich sediments and bioaccumulation in fish have been demonstrated.
Although photolysis may make a minor contribution to the degradation of
this pollutant near the-air-water surface, oxidation and hydrolysis are
probably not important as fate processes.  The role of volatilization is
uncertain.

68.2  Identification

    The chemical structure of 4-chlorophenyl phenyl ether is presented be-
low.

                                             Alternate Names

                                             l-Chloro-4-phenoxybenzene
                                             p-Chlorophenyl phenyl ether
                                             4-Chlorodiphenyl ether
                                             4-Chlorophenyl ether
                                             Monochlorodiphenyl oxide
    4-Chlorophenyl phenyl ether

    CAS NO.  7005-72-3
    TSL NO.  None assigned

68.3  Physical Properties

    The general physical properties are as follows.

    Molecular weight                         203.66
    (Calc.  from Weast 1977)

    Melting  point                            -8C*
    (Dow Chemical Company 1979)

    Boiling  point at 760 torr                284C**
    (Mailhe  and Murat 1912)
                                      68-1

-------
    Vapor pressure at 25C                   0.0027 torr
    (Calc. by Branson 1977)

    Solubility in water at 25C              3.3 mg/1
    (Branson 1977)

    Log octanol/water partition coefficient  4.08
    (Branson 1977)
 *Brewster and Stevenson (1940) report a melting point of 46--47C for 2-
chlorophenyl phenyl ether, and they were apparently unable to prepare a
crystalline sample of 4-chlorophenyl phenyl ether.

**Dow Chemical Company (1979) has determined the boiling point at 760 torr
to be 293.03C.

68.4  Summary of Fate Data

    68.4.1  Photolysis

         4-Chlorophenyl phenyl ether has electromagnetic absorption maxima
at 272, 279 and 293 nm (Choudhry _et _al. 1977).  Irradiation of a methanolic
solution in Pyrex containers (X> 290 nm) results only in dechlorination and
the production of diphenyl ether (Choudhry _e_t al. 1977).  Although diphenyl
ether does not absorb electromagnetic radiation above 300 nm to an appre-
ciable extent (Ungnade 1953) , there is the possibility of a photochemically
induced rearrangement which yields ortho- and para-hydroxybiphenyl and a
trace of phenol (Ogata _et _al. 1970).  No information was found from which
rates of photodegradation in the aquatic environment could be estimated.

    68.4.2  Oxidation

         No information was found in the reviewed literature that would
support any role for oxidation of this compound as 'an aquatic fate.  In-
direct photolysis, involving interaction of hydroxyl radical with the
aromatic ring, is considered to be too slow in water to be environmentally
significant for this compound (Dorfman and Adams 1973).

         It is at present uncertain how much of this pollutant will vo-
latilize into the atmosphere from surface waters (see Section 68.4.4).  Any
4-chlorophenyl phenyl ether that enters into the troposphere will be sub-
ject  to photodegradation and reprecipitation with rain.  The atmospheric
half-life of unsubstituted benzene, proposed by Darnall _e _a_l. (1976), is
2.4 to 24 hours.  According to Laity _e_t _al. (1973), a chlorine substituent
on an aromatic ring should decrease its susceptibility  to photodegradation
in the troposphere.  Although an oxygen substituent should facilitate des-
truction, it is unknown what effect the presence of both groups would have
on the atmospheric destruction of 4-chlorophenyl phenyl ether.
                                   68-2

-------
    68.4.3  Hydrolysis

         No information was found in the reviewed literature that would
indicate hydrolysis as an aquatic fate for this compound.  It is, however,
considered to be unlikely that any of the covalent bonds of 4-chlorophenyl
phenyl ether will hydrolyze at ambient environmental conditions, since the
negative charge-density of the aromatic ring will impede the nucleophilic
attack of water or hydroxide ion.

    68.4.4  Volatilization

         The rate of evaporation from water of 4-chlorophenyl phenyl ether
has been calculated by Branson (1977), using the equations of Liss and
Slater (1974), to be 7.0 cm-hr"-*-.  The half-life, corresponding to this
evaporative rate constant for a pond one meter deep, is approximately seven
hours (Branson 1978).  Some assumptions made in the development of these
equations were:  1)  the pollutant is in solution, rather than in sus-
pended, sorbed, colloidal, or complexed form;  2)  the vapor is in equili-
brium with the liquid at the interface;  3)  water diffusion or mixing is
sufficiently rapid so that the concentration at the interface approaches
that of the bulk of the water; and  4)  the rate of evaporation of water is
negligibly affected by the presence of solutes and suspended matter.  This
calculated evaporative half-life should, therefore, be considered as a min-
imum half-life, since it is highly probable that 4-chlorophenyl phenyl
ether will be sorbed by both organic particulates and suspended clays (see
Section 68.4.5).

    68.4.5  Sorption

         The octanol/water partition coefficient for 4-chlorophenyl phenyl
ether, corresponding to log P =4.08 (Branson 1977), is indicative of a
marked preference for lipophilic organic materials over water.  In addi-
tion, the polar nature of the ether bond and the carbon-halogen bond,
coupled with the compound's miscibility with most lipophilic material en-
sures sorption of 4-chlorophenyl phenyl ether by organic detritus.  Adsorp-
tion by clay particles is also thought to be highly probable, inasmuch as
the polarity and planar geometry of this molecule should facilitate its
intercalation within the layered clay structure.  Evidence for sorption by
suspended organic material is given by a sediment to water distributional
ratio of 4:1 during biodegradation studies with sludge (Branson 1977).

    68.4.6  Bioaccumulation

         Limited information was found in the reviewed literature from
which bioaccumulation could be evaluated.  Several loading concentrations
                                    68-3

-------
of 4-chlorophenyl phenyl ether in water were tested for bioconcentration
inrai'nbow trout (Branson 1977).   Trout muscle reached a steady-state after
8.9 days and exhibited a bioconcentration factor of 736 + 89.   The log oc-
tanol/water partition coefficient also indicates a definite potential for
bioaccumulation in aquatic ecosystems.

    68.4.7  Biotransformation and Biodegradation

         Branson (1978) has reported that the half-life of 4-chlorophenyl
phenyl ether with respect to biodegradation in activated sludge is ap-
proximately four hours.  Data of greater relevance to natural  surface
waters were obtained in a river-water die-away experiment (Branson 1978).
Two water samples, taken from two different locations in the Tittabawassee
River and spiked with the pollutant at a concentration of 1 mg/1,  showed
two different rates of biodegradation.  One sample exhibited no detectable
degradation before 120 hours and the other, no degradation before  312
hours.  Greater than 90 percent  degradation required 264 hours in  the first
sample and 528 hours in the second.  The experiments illustrated that bio-
degradation rate constants are not reproducible when the microbial popula-
tions have a dissimilar history.  In these two experiments the half-lives
with respect to biodegradation would be rather meaningless since both
samples exhibited lag periods (during which loss of the pollutant  was un-
detectable) that were approximately one-half of the time required  for 90
percent degradation.  Thus, it is reasonable to assume that genetic induc-
tion levels for most degradative organisms would not be reached except in
the vicinity of discharges.  In a non-stagnant body of surface water, a
distinct microbial population would probably not be in contact with the
pollutant long enough to become  acclimated.

68.5  Data Summary

    Table 68-1 summarizes the aquatic fate data discussed above for
4-chlorophenyl phenyl ether.  Information on this compound is limited and a
statement of its most probable aquatic fate is, therefore, not feasible at
this time.  This ether is reported to be biodegradable by acclimated sewage
sludge, and there is also a possibility that photolysis might contribute to
its destruction near the air-water surface.  Biodegradation in most natural
surface waters, however, would appear not to be an important process,
whereas bioaccumulation in fish has been demonstrated and sorption by
organic sediments may be important.  The role of volatilization is un-
certain because of a lack of scientific concensus on methods of calculating
rates of pollutant volatilization from natural surface waters.  Oxidation
and hydrolysis are probably not  important as fate processes.
                                    68-4

-------
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                                 68-5

-------
68.6  Literature Cited

Branson, D.R.  1977.   A new capacitor fluid - a case study in product
  stewardship,  pp.  44-61.  Aquatic Toxicology and Hazard Evaluation.  ASTM
  Special Technical  Publication 634.  F.L. Mayer and J.L. Hamelink. (eds.).
  American Society for Testing and Materials, Philadelphia, Pa.

Branson, D.R.   1978.  Predicting the fate of chemicals in the aquatic
  environment  from laboratory data.  pp.  55-70.   Estimating the Hazard of
  Chemical Substances to Aquatic Life.  ASTM Special Technical Publication
  657. J. Cairns, Jr., K.L. Dickson, and A.W. Maki. (eds.).  American
  Society for  Testing and Materials, Philadelphia, Pa.

Brewster, R.Q. and  G. Stevenson.  1940.   The chlorination of phenyl ether
  and orientation in 4-chlorophenyl ether.  J. Am. Chem. Soc.
  62:3144-3146.

Choudhry, G.G., G. Sundstrom, L.O. Ruzo,  and 0.  Hutzinger.  1977.
  Photochemistry of  chlorinated diphenyl ethers.  J. Agric. Food Chem.
  25(6):1371-1376.

Darnall, K.R., A.C.  Lloyd, A.M. Winer, and J.N.  Pitts, Jr.  1976.
  Reactivity scale for atmospheric hydrocarbons  based on reaction with
  hydroxyl radical.   Environ. Sci. Technol. 10(7):692-696.

Dorfman, L.M.  and G.E. Adams.  1973.  Reactivity of the hydroxyl radical in
  aqueous solution.   NSRDS-NBS-46.  NTIS:COM-73-50623.  Springfield,   Va.

Dow Chemical Company.  1979.  Personal communication from M. Thomas (Dow)
  to N.W. Gabel, Versar, Inc.

Laity, J.L., I.G. Burstain, and B.R. Appel.  1973.  Photochemical smog
  and the atmospheric reactions of solvents.  Chap. 7, pp.95-112.  Solvents
  Theory and Practice.  R.W. Tess (ed.).   Advances in Chemistry Series 124.
  American Chemical  Society, Washington, D.C.

Liss, P.S. and P.G.  Slater.  1974.  Flux of gases across the air-sea
  interface.  Nature 247:181-184.

Mailhe, A. and M. Murat.  1912.  Derives halogenes de 1'oxyde de
  phenyle.  Bull. soc. chim. France.  11:328-332.

Ogata Y., K. Takagi  and I. Ishirao.  1970.  Photochemical rearrangement of
  diaryl ethers.  Tetrahedron.  26(11):2703-2709.
                                    68-6

-------
Ungnade, H.E.  1953,   The effects of solvents on the absorption spectra of
  aromatic compounds.   J. Am.  Chem.  Soc.  75:432-434.

Weast, R.C.  1977.  Handbook of chemistry and physics.   58th edition.
  CRC Press, Inc., Cleveland,  Ohio.   2398p.
                                    68-7

-------
                      69.   4-BROMOPHENYL PHENYL ETHER
69.1  Statement of Probable Fate

     Very little information pertaining to the environmental transport and
fate of 4-bromophenyl phenyl ether was found, and it is, therefore, not
possible to determine the most probable aquatic fate at this time.   Some
inferences can be drawn from experiments performed with this pollutant's
chloro analog.  4-Chlorophenyl phenyl ether is reported to be rapidly de-
graded by acclimated sewage sludge, but biodegradation data from river-
water die-away experiments indicate that this compound has a potential for
persistence in natural surface waters.  Sorption by organic-rich sediments
and bioaccumulation in fish may be important.  Although photolysis  may make
a minor contribution to degradation near the air-water surface, oxidation
and hydrolysis are probably not important as fate processes.  The role of
volatilization in the removal of halogenated aromatic ethers from aquatic
systems has not been demonstrated and remains uncertain.

69.2  Identification

     The chemical structure of 4-bromophenyl phenyl ether is shown  below.

                                             Alternate Names

                                             l-Bromo-4-phenoxybenzene
                                             p-Bromophenyl phenyl ether
                                             4-Bromodiphenyl ether
                                             4-Bromophenyl ether
     4-Broraophenyl phenyl ether

     CAS NO. 101-55-3
     TSL NO. None assigned

69.3  Physical Properties

     The general physical properties of 4-bromophenyl phenyl ether are
given below.
     Molecular weight
     (Weast 1977)
249.11
                                     69-1

-------
     Melting point                           18.72C
     (Weast 1977)

     Boiling point at 760 torr               310.14C
     (Weast 1977)

     Vapor pressure at 20C                  0.0015  torr
     (Calc. from Dreisbach 1952)

     Solubility in water                     No data found

     Log octanol/water partition             4.28
       coefficient (Calc. by method
       of Leo et al. 1971 using the
       data of Branson 1977)

69.4  Summary of Fate Data

     69.4.1  Photolysis

          It is reasonable to assume that 4-bromophenyl phenyl ether will
absorb electromagnetic radiation in the ultraviolet  region of the terres-
trial solar spectrum, since its analog, 4-chlorophenyl phenyl ether, has
electromagnetic absorption maxima at 272, 279 and 293 nm (Choudhry et al.
1977).  Irradiation of a methanolic solution of 4-chlorophenyl phenyl ether
in Pyrex containers (A.> 290 nm) results only in dechlorination and the
production of diphenyl ether (Choudhry et_ _al. 1977).  Although unsubsti-
tuted diphenyl ether apparently does not absorb electromagnetic radiation
above 300 nm to an appreciable extent (Ungnade 1953), there is the possi-
bility of a photochemically induced rearrangement which yields ortho- and
para-hydroxybiphenyl and a trace of phenol (Ogata et^ al. 1970).  No infor-
mation was found from which rates of photodegradation in the aquatic en-
vironment could be estimated.

     69.4.2  Oxidation

          No information was found in the reviewed literature that would
support any role for oxidation of this compound as an aquatic fate.  In-
direct photolysis, involving interaction of hydroxyl radical with the
aromatic ring, is considered to be too slow in water to be environmentally
significant for this compound (Dorfman and Adams 1973).

          It is at present uncertain how much of this pollutant will vola-
tilize into the atmosphere from surface waters (see  Section 69.4.4).  Any
4-bromophenyl phenyl ether that enters into  the troposphere will be subject
to photodegradation and reprecipitation with rain.  The atmospheric half-
                                     69-2

-------
life of unsubstituted benzene, proposed by Darnall ejt _al.  (1976),  is 2.4 to
24 hours.  According to Laity jit _al. (1973), a halogen substituent on an
aromatic ring should decrease its susceptibility to  photodegradation in
the troposphere.   Although the electron-donating resonance effect  of an
oxygen substituent should facilitate destruction, it is uncertain  how the
presence of both groups will affect the atmospheric destruction of 4-bromo-
phenyl phenyl ether.

     69.4.3  Hydrolysis

          No information was found in the reviewed literature that would
indicate hydrolysis as an aquatic fate for this compound.   It is,  however,
considered to be unlikely that any of the covalent bonds of 4-bromophenyl
phenyl ether will hydrolyze at ambient environmental conditions, since the
negative charge-density of the aromatic ring will impede the nucleophilic
attack of water or hydroxide ion.

     69.4.4  Volatilization

          The rate constant for evaporation from water of 4-chlorophenyl
phenyl ether has been calculated by Branson (1977), using the equations of
Liss and Slater (1974), to be 7.0 cm-hr"1.  The half-life, corresponding
to this evaporative rate constant for a pond one meter deep, is approxi-
mately seven hours (Branson 1978).  A similar calculated rate constant can
be expected for the structurally analgous 4-bromophenyl phenyl ether.  Some
assumptions made in the development of these equations were:  1)  the
pollutant is in solution, rather than in suspended, sorbed, colloidal, or
complexed form;  2)  the vapor is in equilibrium with the liquid at the
interface;  3)  water diffusion or mixing is sufficiently rapid so that the
concentration at the interface approaches that of the bulk of the  water;
and 4)  the rate of evaporation of water is negligibly affected by the pre-
sence of solutes and suspended matter.  This calculated evaporative half-
life should, therefore, be considered as a minimum half-life, since it is
highly probable that halogenated diphenyl ethers will be sorbed by both
organic particulates and suspended clays (see Section 69.4.5).

     69.4.5  Sorption

          The octanol/water partition coefficient for 4-bromophenyl phenyl
ether, corresponding to log P = 4.28 (Leo _et _al. 1971;  Branson 1977), is
indicative of a marked preference for lipophilic organic materials over
water.  In addition, the polar nature of the ether bond and the carbon-
halogen bond, coupled with the compound's raiseibility with most lipophilic
material ensures sorption of 4-bromophenyl phenyl ether by organic de-
tritus.  Adsorption by clay particles is also thought to be highly
                                     69-3

-------
probable, inasmuch as the polarity and  planar geometry of  this molecule
should facilitate its intercalation within the layered clay structure.
Evidence for sorption by suspended organic material  is given by a sediment
to water distributional ratio of 4:1 for 4-chlorophenyl phenyl ether during
biodegradation studies with sludge (Branson 1977).

     69.4.6  Bioaccumulation

          Limited information was found in the reviewed literature from
which bioaccumulation could be evaluated.   Several loading concentrations
in water of the structurally analogous  pollutant,  4-chlorophenyl phenyl
ether, were tested for bioconcentration in rainbow trout (Branson 1977).
Trout muscle reached a steady-state after 8,9 days and exhibited a biocon-
centration factor for 4-chlorophenyl phenyl ether  of 736 + 89.  It can  be
expected that 4-bromophenyl phenyl ether will behave similarly.  The log
octanol/water partition coefficient also indicates a definite potential for
bioaccumulation in aquatic ecosystems.

     69.4.7  Biotransformation and Biodegradation

          No specific information was found in the  reviewed literature  with
which to assess the biodegradation of 4-bromophenyl  phenyl ether.  Since
this pollutant is structurally analogous to 4-chlorophenyl phenyl ether,  it
may exhibit similar properties with respect to biodegradation.

          Branson (1978) has reported that the biodegradative half-life of
4-chlorophenyl phenyl ether in activated sludge is approximately four
hours.  Data of greater relevance to natural surface waters were obtained
in a river-water die-away experiment (Branson 1978).  Two  water samples,
taken from two different locations in the Tittabawassee River and spiked
with the pollutant at a concentration of 1 mg/1, showed two different rates
of biodegradation.  One sample exhibited no detectable degradation before
120 hours and the other, no degradation before 312 hours.   Greater than 90
percent degradation required 264 hours in the first  sample and 528 hours  in
the second.  The experiments illustrated that biodegradation rate constants
are not reproducible when the microbial populations  have a dissimilar his-
tory.  In these two experiments the half-lives with  respect to biodegrada-
tion are rather meaningless since both samples exhibited lag periods
(during which loss of the pollutant was undetectable) that were approxi-
mately one-half of the time required for 90 percent  degradation.  Thus, it
is reasonable to assume that genetic induction levels for most degradative
organisms will not be reached except in the vicinity of discharges.  In a
non-stagnant body of surface water, a distinct microbial population would
probably not be in contact with the pollutant long enough to become
acclimated.
                                     69-4

-------
69.5  Data Summary

     Table 69-1 summarizes the aquatic fate data discussed above for 4-
bromophenyl phenyl ether.   Information on this compound is very limited and
a statement of its most probable aquatic fate is, therefore,  not feasible
at this time.  The structurally analogous pollutant,  4-chlorophenyl phenyl
ether, is reported to be biodegradable by acclimated  sewage sludge, and
there is also a possibility that photolysis might contribute  to environ-
mental destruction near the air-water surface.  Based on data obtained from
experiments with 4-chlorophenyl phenyl ether, biodegradation  in most
natural surface waters does not appear to be an important process,  whereas
bioaccumulation in fish and sorption by organic sediments may be important.
The role of volatilization is uncertain because of a  lack of  scientific
concensus on methods of calculating rates of pollutant volatilization from
natural surface waters.  Oxidation and hydrolysis are probably not  impor-
tant as fate processes.
                                     69-5

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>s Ul
co e c co
4-1 -H CU rH
C  -1
CU CU O. 6
g 0 0 TH
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                                                                             u  o.
                                                                             s  

                                                                             ll
                                                                             ?  CJ
                                                                              CU u_,


                                                                              >H O
                                                                              15
                                                                              U n
                                                                              s
                                                                             ^y o
                                                                              C 4-1
                                                                              c 3
                                                                              OJ a.
                                                                              =  3
                                    69-6

-------
69.6  Literature Cited

Branson, D.R.  1977.   A new capacitor fluid - a case study in product
  stewardship,  pp.44-61.  Aquatic Toxicology and Hazard Evaluation.  ASTM
  Special Technical Publication 634.   F.L Mayer and J.L. Hamelink.  (eds.).
  American Society for Testing and Materials, Philadelphia, Pa.

Branson, D.R.  1978.   Predicting the fate of chemicals in the aquatic
  environment from laboratory data.  pp. 55-70.  Estimating the Hazard of
  Chemical Substances to Aquatic Life.  ASTM Special Technical Publication
  657.  J. Cairns, Jr., K.L. Dickson, and A.W. Maki.  (eds.).  American
  Society for Testing and Materials,  Philadelphia, Pa.

Choudhry, G.G., G. Sundstrom, L.O. Ruzo, and 0. Hutzinger.  1977.
  Photochemistry of chlorinated diphenyl ethers.  J. Agric. Food Chem.
  25(6):1371-1376.

Darnall, K.R., A.C. Lloyd, A.M. Winer, and J.N. Pitts, Jr.  1976.
  Reactivity scale for atmospheric hydrocarbons based on reaction with
  hydroxyl radical.  Environ. Sci. Technol.  10(7):692-696.

Dorfman, L.M. and G.E. Adams.  1973.   Reactivity of the hydroxyl radical in
  aqueous solution.  NSRDS-NBS-46.  NTIS:COM-73-50623.  Springfield, Va.

Dreisbach, R.R.  1952.  Pressure-volume-temperature relationships of
  organic compounds.   Handbook Publishers, Inc.  Sandusky, Ohio.  349p.

Laity, J.L., I.G. Burstain, and B.R.  Appel.  1973.  Photochemical smog and
  the atmospheric reactions of solvents.  Chap. 7, pp. 95-112.  Solvents
  Theory and Practice.  R.W Tess (ed.).  Advances in Chemistry Series 124.
  American Chemical Society, Washington, D.C.

Leo A., C. Hansch and D. Elkins.  1971.  Partition coefficients and their
  uses.  Chem. Rev.  71:525-612.

Liss, P.S. and P.G. Slater.  1974.  Flux of gases across the air-sea
  interface.  Nature 247:181-184.

Ogata Y., K. Takagi and I. Ishimo.  1970.  Photochemical rearrangement of
  diaryl ethers.  Tetrahedron 26(11):2703-2709.

Ungnade, H.E.  1953.   The effects of solvents on the adsorption spectra of
  aromatic compounds.  J. Am. Chem. Soc.  75:430-434.

Weast, R.C.  (ed.).  1977.  Handbook of chemistry and physics.  58th
  edition.  CRC Press, Inc., Cleveland, Ohio.  2398p.
                                    69-7

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                      70.   BIS(2-CHLOROETHOXY)METHANE


70.1  Statement of Probable Fate

    Based on the available data, it appears that the most probable fate of
bis(2-chloroethoxy)methane in the aquatic environment is slow hydrolysis.
Neither volatilization nor sorption processes would seem to be able to
affect the transport of this highly soluble, slightly volatile compound.
No information pertaining to biodegradation was found.

70.2  Identification

    Bis(2-chloroethoxy)methane has been detected in industrial effluents
(Shackelford and Keith 1976).  The chemical structure of bis(2-chloro-
ethoxy)methane is shown below.


             H           H           H     H

              I           !           II
 ci  c  c  o  c  o  c  c  a

       II           l           II
       H     H           H           H     H
    Bis(2-chloroethoxy)methane               Alternate Names
                                             Dichlorodiethyl methylal
    CAS NO. 111-91-1                         Bis(B-chloroethyl)formal
    TSL NO. PA 36750                         6,8,-Dichlorodiethyl formal

70.3  Physical Properties

    General physical properties of bis(2-chloroethoxy)tnethane are as fol-
lows.

    Molecular weight                         173.1
    (Webb et_ al.  1962)

    Melting point                            No data found

    Boiling point at 760 torr                218.1C*
    (Webb et al.  1962)
                                     70-1

-------
    Vapor pressure at 20C                   <0.1 torr
    (Calc. from Dreisbach 1952 based
    on the data of Webb _et al. 1962)

    Solubility in water at 25 C              81,000 mg/1
    (Calc. by method of Moriguchi 1975)

    Log octanol/water partition coefficient  1.26
    (Calc. based on method of Leo ^t al.
    1971)
*The boiling point at 760 torr has been reported as 105-106 by Durkin et
al. (1975).  Based on the detailed study of Webb _et al. (1962) on the
properties of this pollutant and other compounds in this series, the value
reported by Durkin _e_t al. (1975) is incorrect.

70.4  Summary of Fate Data

    70.4.1  Photolysis

         Bis(2-chloroethoxy)methane would not be expected to undergo direct
photolysis in surface waters or the troposphere since the compound does not
possess any chromophores that absorb radiation in the visible or near
ultraviolet regions of the electromagnetic spectrum (Jaffe and Orchin
1962).  No information was found that would suggest photolysis as an en-
vironmental fate process.

    70.4.2  Oxidation

         Even though water has an inhibitory effect on the formation of
ether and acetal peroxides, it apparently does not prevent their formation
(Patai 1967).  Since no information discussing the efficacy of molecular
oxygen in dilute aqueous solutions to form ether and acetal peroxides was
found, it is uncertain whether such peroxides exist in the aquatic environ-
ment.  Indirect photolysis, involving abstraction of alkyl hydrogens by the
hydroxyl radicals normally present in surface waters, is considered to be
too slow to be environmentally relevant (Dorfman and Adams 1973).

    70.4.3  Hydrolysis

         There are two sites within this molecule wherein hydrolysis could
take place:  (1) the carbon-oxygen bonds of the acetal linkage and (2) the
carbon-chlorine bonds.  The hydrolysis of acetal bonds is acid catalyzed.
Kankaanpera (1969) gives this acid-catalyzed rate constant for bis(2-
chloroethoxy)methane as 2.53 x 10~6 liter mole"'- sec"-'-.  Applying the
assumptions of Radding et al. (1977), the maximum hydrolytic half-life in
                                     70-2

-------
pure water at pH 7 and 25C would be within the range of thousands of
years.  What effect suspended clays and humic material would have on this
rate is purely conjectural.

         No environmentally relevant kinetic data were found for the
hydrolysis of the carbon-chlorine bond of this compound.  Dilling et al.
(1975) reported that the half-lives with respect to hydrolysis for one and
two carbon chloroaliphatic compounds are six months to several years.
Bis(2-chloroethoxy)methane may have a corresponding rate of hydrolysis.

    70.4.4  Volatilization

         Although no information pertaining specifically to the volatiliza-
tion of bis(2-chloroethoxy)methane was found, the estimated vapor pressure
of less than 0.1 torr based on the relationships given by Dreisbach (1952)
and the data reported by Webb _et _al. (1962) indicates that volatilization
would not be an important transport process.  Furthermore, the calculated
solubility of 81,000 mg/1 and the expected hydrogen bonding between the
acetal oxygen atoms and the water of solvation virtually precludes any role
for volatilization as a removal mechanism from water.

    70.4.5  Sorption

         No information specifically pertaining to sorption processes with
environmental significance was found.  The log octanol/water partition
coefficient, calculated as 1.26, projects little potential for adsorption
by lipophilic materials, and the aqueous solubility of 81,000 mg/1 indi-
cates that whatever sorption processes occur will be only of a transi-
tory nature. This summation, however, does not preclude a possible cata-
lytic role in hydrolysis for suspended particulates.

    70.4.6  Bioaccumulation

         No information was found indicating that bis(2-chloroethoxy)-
methane will bioaccumulate.  Based on experimental and empirical data,
Metcalf and Sanborn (1975) maintain that compounds with aqueous solubili-
ties of about 50 mg/1 or greater generally show little potential for bio-
accumulation.  The aqueous solubility of bis(2-chloroethoxy)methane, cal-
culated by the method of Moriguchi (1975) to be 81,000 mg/1, appears to be
much higher than the arbitrary 50 mg/1 limit suggested by Metcalf and
Sanborn (1975).

    70.4.7  Biotransformation and Biodegradation

         No information was found from which any conclusion regarding
biodegradation can be reached with any degree of confidence.
                                    70-3

-------
70.5  Data Summary

    Table 70-1 summarizes what is known about  the aquatic  fate of bis(2-
chloroethoxy)methane.   Neither volatilization  nor sorption appear to affect
the transport of this  highly soluble,  slightly volatile compound.  The most
probable fate for bis(2-chloroethoxy)methane is slow hydrolysis.
                                     70-4

-------
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70.6  Literature Cited

Billing, W.L., N.B. Tefertiller, and G.J.  Kallos.  1975.   Evaportion
  rates of methylene chloride, chloroform, 1,1,1,-trichloroethane,
  trichloroethylene, tetrachloroethylene,  and other chlorinated compounds
  in dilute aqueous solutions.  Environ. Sci. Technol. 9(9):833-838.

Dorfman, L.M. and G.E. Adams.  1973.  Reactivity of the hydroxyl radical
  in aqueous solution.  NSRDS-NBS-46.  NTIS:COM-73-50623.   Springfield, Va.

Dreisbach, R.R.  1952.  Pressure-volume-temperature relationships of
  organic compounds.  3rd Edition.  Handbook Publishers,  Inc., Cleveland,
  Ohio.  349p.

Durkin, P.R., P.H.  Howard, and J. Saxena.   1975.  Investigation of  selected
  potential environmental contaminants:  haloethers.  U.S. Environmental
  Protection Agency, Office of Toxic Substances, Washington, D.C.  168p.
  (EPA 560/2-75-006).

Jaffe, H.H. and M.  Orchin.  1962.  Theory and application of ultraviolet
  spectroscopy.  John Wiley and Sons, New York.  253p.

Kankaanpera, A.  1969.  Basicities of the oxygen atoms in symmetrical and
  unsymmetrical acetals.  Part II.  The base strengths and their relation
  to the rate coeffficients of the different partial fission reactions of
  acetal hydrolysis.  Acta Chem. Scand.  23(5):1728-1732.

Leo, A., C. Hansch and D. Elkins.  1971.  Partition coefficients and their
  uses.  Chem. Rev.  71:525-612.

Metcalf, R.L. and J.R. Sanborn.  1975.  Pesticides and environmental
  quality in Illinois.  111. Nat'l. Hist.  Survey Bull.  31:381-436.

Moriguchi,  I.  1975.  Quantitative structure-activity studies on parameters
  related to hydrophobicity.  Chem. Phara. Bull.  23:247-257.

Patai, S. (ed).  1967.  The chemistry of the ether linkage.  Interscience
  Publishers, New York.  785p.

Radding, S.B., D.H. Liu, H.L. Johnson, and T. Mill.  1977.  Review of the
  environmental fate of selected chemicals.  U.S. Environmental Protection
  Agency, Office of Toxic Substances, Washington, D.C.  147p.   (EPA
  560/5-77-003).
                                      70-6

-------
Shackelford,  W.M. and L.H.  Keith.  1976.   Frequency of organic compounds
  identified in water.  U.S. Environmental Protection Agency, (ERL,) ,
  Athens, Ga.   617p.  (EPA 600/4-76-062).

Webb, R.F., A.J. Duke, and L.S.A. Smith.   1962.   Acetals and oligoacetals,
  Part I.  Preparation and properties of reactive oligoformals.   J.  Chem.
  Soc. (Lond.)  4307-4319.
                                       70-7

-------
SECTION VII:  MONOCYCLIC AROMATICS
          Chapters 71-93

-------
                               71.  BENZENE


71.1  Statement of Probable Fate

    Based on the information found, it appears that the predominant process
for removal of benzene from the water column is volatilization to the atmos-
phere.  That portion of the benzene which volatilizes to the atmosphere is
probably depleted at a fairly rapid rate due to attack by hydroxyl radi-
cals.  It must be noted, however, that the solubility of benzene in water
is relatively high; consequently, persistence of some benzene in the water
column would be expected.  Although the role of benzene sorption onto sedi-
ments and suspended solids cannot be established based on the reviewed
literature, there is evidence of gradual biodegradation of benzene at low
concentrations by aquatic microorganisms.  The rate of benzene biodegrada-
tion is enhanced when other hydrocarbons are present.

71.2  Identification

    Benzene has been detected in finished drinking water (U.S. Environ-
mental Protection Agency 1975), in water and sediment samples from the
lower Tennessee River in ppb concentrations (Goodley and Gordon 1976) and
in the atmosphere (Howard and Durkin 1974).  The chemical structure of
benzene is shown below.

                                             Alternate Names

                                             Benzol
                                             Cyclohexatriene


     Benzene

     CAS NO. 71-43-2
     TSL NO. CY 14000

71.3  Physical Properties

    The general physical properties of benzene are as follows.

    Molecular weight                         78.12
    (Weast 1977)

    Melting point                            5.5C
    (Weast 1977)
                                    71-1

-------
    Boiling point at 760 torr                80.1C
    (Weast 1977)

    Vapor pressure at 25C*                  95.2 torr
    (Mackay and Leinonen 1975)

    Solubility in water                      **

    Log octanol/water partition coefficient  ***
*Vapor pressure values found in the literature include 45.5 torr at 10C
(Mackay and Leinonen 1975), 95.2 torr at 25C (Mackay and Leinonen 1975;
Mackay and Wolkoff 1973), and 100 torr at 26.075C (Howard and Durkin
1974).

**Several values for the solubility of benzene in water were found in the
literature.  Some of the reported values found in the literature include
1750 mg/1 at 10C (Mackay and Leinonen 1975), 820 mg/1 at 22C (Chiou_et
_al. 1977), 1780 mg/1 at 25C (Mackay and Wolkoff 1973; Mackay and Leinonen
1975), and 1800 mg/1 at 25C (Howard and Durkin 1974).

***The log octanol/water partition coefficient of benzene is reported to be
1.95 by Leo _et al. (1971) and 2.13 by Chiou _et _al. (1977).

71.4  Summary of_Fate Data

    71.4.1  Photolysis

         Since the ozone layer in the upper atmosphere effectively filters
out wavelengths of light less than 290 nm, and since the ultraviolet spec-
trum of benzene indicates that this compound does not absorb wavelengths of
light longer than 260 nm (Bryce-Smith and Gilbert 1976) direct: excitation
of benzene in the environment is unlikely unless a substantial wavelength
shift is caused by the media (Howard and Durkin 1974).  For instance, al-
though benzene does not absorb light directly in appreciable amounts at
wavelengths longer than 280 nm when dissolved in cyclohexane (Silverstein
and Bassler 1968) , slight shifts in wavelength absorption might be expected
in more representative environmental media such as water or the surface of
particulate organic matter (Howard and Durkin 1974).

    71.4.2  Oxidation

         No specific information pertaining to oxidation of benzene in the
aqueous environment under ambient conditions was found.  Howard and Durkin
(1974), however, report that catalysts, elevated temperature, elevated
pressure, or any of these conditions operating together may serve as ini-
tiators of benzene oxidation.  Based on this information, it can be
                                     71-2

-------
inferred that direct oxidation of benzene in environmental waters is un-
likely.

         Inasmuch as the main transport process that would account for re-
moval of benzene from water appears to be volatilization, the atmospheric
destruction of benzene probably is much more likely than any other fate
process.  These complex photochemical reactions have been studied in simu-
lated smog chambers (Altshuller et_ al. 1962; Laity _e_t _al. 1973) that
measured the rate of disappearance of the volatilized organic material.
The half-conversion time of m-xylene and 1,3,5-trimethylbenzene have been
reported to be somewhat less than four hours (Altshuller &t_ _al. 1962).
From this value and the table of relative reactivities given by Laity et
al. (1973), it can be inferred that the corresponding range for the half-
conversion time for benzene would be approximately 20 to 50 hours.  This
value for the estimated half-conversion time of benzene is in reasonable
agreement with the estimated half-life of benzene proposed by Darnall et
al. (1976) of 2.4 to 24 hours.  This half-life value is based on the
assumptions that benzene depletion is due solely to attack by hydroxyl radi-
cal (OH')) and that even high concentrations of ozone present in ambient
atmospheres will not contribute significantly to the photooxidation of
alkanes and aromatics, in general.  A second-order rate of reaction of ben-
zene with hydroxyl radicals of 0.85 x 10~9 i .mol'-'-sec"-'- has been ob-
tained by Darnall et al. (1976) by averaging rates from smog chamber data
by Hansen ^t al. (T9757 and Davis e al. (1975).  The temporal stability of
benzene under actual atmospheric conditions is, as yet, unknown.  Experi-
ments performed in laboratory irradiation chambers are usually conducted
for relatively short periods and cannot account for all of the meterologi-
cal variables within a natural airshed.

    71.4.3  Hydrolysis

         No specific information pertaining to the hydrolysis of benzene
under ambient conditions was found.  The hydrolysis of benzene is an un-
likely process under environmental conditions since nucleophilic attack of
the aromatic ring by water or hydroxide ion will be impeded by its negative
charge-density (Morrison and Boyd 1973).

    71.4.4  Volatilization

         The half-life with respect to volatilization from a water column
one meter thick has been estimated by Mackay and Leinonen (1975) to be 4.81
hours for benzene at 25C; at 10C the half-life with respect to volatili-
zation from the same depth of water has been estimated to be 5.03 hours.
Mackay and Leinonen (1975) point out that for benzene the rates and half-
lives of volatilization are insensitive to temperature and that tempera-
ture only affects the rate of volatilization significantly if the system
                                   71-3

-------
is vapor-phase controlled.   Some assumptions made in the estimation of
rates of volatilization and half-lives were as follows:   1)  the contaminant
concentration is in solution rather than in suspended,  colloidal, ionic,
complexed, or adsorbed form;  2) the vapor is in equilibrium  with the liquid
at the interface;  3) the water diffusion or mixing is sufficiently fast so
that the concentration at the interface approaches that of the bulk of the
water; and 4) the  rate of evaporation of water is negligibly affected by
the presence of the contaminants.

         Mackay and Leinonen (1975) point out that interpretation of the
significance of the evaporative rate of compounds such as benzene from en-
vironmental waters using values calculated with a one-meter  depth is de-
pendent upon the type of environmental situation encountered.  In situa-
tions where the water body is turbulent with frequent mixing between the
surface layer and  the bulk, as in a rapidly flowing shallow  river or during
white-capping on a lake or ocean, the evaporative rate would be more rapid
than for depths greater than one meter or in quiescent water, as evidenced
in a deep, slowly flowing river.

    71.4.5  Sorption

         Although no specific environmental sorption studies were found in
the reviewed literature, the values of the log octanol/water partition co-
efficient found for benzene (log P=1.95, Leo _et al. 1971; log P=2.13, Chiou
_e_t al. 1977) indicate that sorption processes may be significant for ben-
zene under conditions of constant exposure.  Presumably, benzene will be
adsorbed by sedimentary organic material; the extent to which this possible
adsorption will interfere with volatilization has not been considered.

    71.4.6  Bioaccumulation

         Neely et al. (1974) have shown that the bioaccumulation potential
of a compound is related to the log octanol/water partition coefficient
(log P) of the compound.  The log P values obtained'from the literature for
benzene of 1.95 (Leo _et al. 1971) and 2.13 (Chiou _et al. 1977) indicate
that the bioaccumulation potential of benzene by aquatic organisms at
pollutant concentrations anticipated in environmental waters would probably
be low.

    71.4.7  Biotransformation and Biodegradation

         Some species of soil bacteria have been demonstrated to be capable
of utilizing benzene as the sole source of carbon (Zobell 1950; Gibson
1976; Glaus and Walker 1964).  A study by Walker and Colwell (1975), how-
ever, showed that petroleum-degrading bacteria isolated from the oil-
polluted Colgate Creek of the Chesapeake Bay would only utilize benzene
                                     71-4

-------
when present in combination with dodecane, or with dodecane and naphtha-
lene.  This utilization was suggested by Walker and Colwell (1975) to most
likely occur as a result of co-oxidation or because of a lower concentra-
tion of benzene present than when petroleum-degrading bacteria were treated
with benzene alone.  Since measurable utilization of benzene at a 0.1%
concentration occurred for more than 30% (68 of 200) of the pure cultures
of hydrocarbon-utililizing bacteria, Walker and Colwell (1975) feel that
the latter explanation cannot be excluded.

         Gibson (1976) and Gibson et al. (1968) conducted experiments to
determine the metabolic pathway involved in the microbial, oxidative de-
gradation of benzene.  Although the microorganism used in these experi-
ments, Pseudomonas putida, could utilize benzene as the sole source of
carbon and energy for growth, toluene served as a better substrate, and
cells grown with toluene were used to investigate benzene metabolism.
Gibson (1976) found that the initial reactions in the bacterial oxidation
of aromatic hydrocarbons involved the formation of c is-dihydrodiols which
undergo further oxidation to yield catechols.  Gibson (1976) found that
mammals, on the other hand, oxidize benzene to arene oxides which are
hydrated to form trans-dihydrodiols prior to oxidation to yield catechols.

         For several reasons, the view that only a few genera of bacteria
such as Pseudomonas (Gibson 1976), and Achromobacter (Glaus and Walker
1964) can utilize benzene as a sole carbon source may not be valid.  The
usual enrichment procedures for isolating such bacteria tend to select only
those that grow rapidly.  Vigorous growth in pure culture is a great ad-
vantage in biochemical studies but may not encompass all of the more impor-
tant features of a natural habitat.  A specific compound may in fact be
readily metabolized in soil despite the failure to isolate single microbial
species capable of using that compound as a sole carbon source (National
Research Council 1977).  On the other hand, the isolation of single species
cabable of using a test compound as a sole carbon source must also be
viewed with caution.  An organism capable of using a test substrate as a
sole carbon source in pure culture may not be able to assimilate the com-
pound under natural conditions.  Generally, the concentration of substrate
used in pure culture studies is considerably higher than normally en-
countered in nature.  As a result, the enzymes essential for biodegradation
may not be induced under natural conditions.  Further, pure culture studies
rarely lead to useful degradation rate information (Howard and Durkin
1974).

         Helfgott e_t_ _aJL_. (1977) report the refractory index (often referred
to as the biorefractory index by other authors) of benzene to be 0.23
                                       71-5

-------
indicating that benzene is quite resistant to degradation.   The refractory
index of a compound as defined by Helfgott et al.  (1977)  is a parameter for
estimating the biodegradability and treatability of organic materials found
in and entering into the aquatic environment.  Values for the refractory
index normally range from 0.0 to 1.0.   Refractory materials are defined as
those materials having refractory index values (R.I.) of  less than 0.6.
According to Helfgott et^ al.  (1977) such refractory compounds are not
expected to degrade in an aerobic wastewater treatment system such as an
activated sludge process and  are expected to persist in natural water
systems such as aquifers, rivers, and  lakes.  They suggest  that the
aromatic symmetry and electron delocalization of benzene  are probably the
factors which impart persistence to this compound.  In a  study by Thorn and
Agg (1975) benzene is listed  as a synthetic organic chemical which should
be degradable by biological sewage treatment provided that  suitable
acclimatization can be achieved.

71.5  Data Summary

    Table 71-1 summarizes the aquatic  fate information found for benzene.
Volatilization appears to be  the major transport process  of benzene from
the water column to the atmosphere.  The atmospheric photooxidation of
volatilized benzene probably subordinates all other fate  processes.  The
reported rates of oxidation are atmospheric photooxidation rates based on
smog chamber data.  One half-life value and rate reported for the
atmospheric photooxidation of benzene  is based on the assumptions that ben-
zene depletion is due solely to attack by hydroxyl radicals and that even
high concentrations of ozone present in ambient atmospheres will not con-
tribute significantly to the photooxidation of alkanes and  aromatics, in
general.  Since benzene is relatively soluble in water, some benzene is ex-
pected to persist in the water column.  That portion of benzene which
persists in the water column would be  expected to eventually biodegrade at
a slow rate. The biodegradation of benzene would probably be enhanced by
the presence of other hydrocarbons.
                                    71-6

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71.6  Literature Cited

Altshuller, A.P., I.R. Cohen, S.F. Sleva, and S.L. Kopczynski.  1962.  Air
  pollution:  photooxidation of aromatic hydrocarbons.  Science
  138:442-443.

Bryce-Smith, D. and A. Gilbert.  1976.  The organic photochemistry of
  benzene - I.  Tetrahedron.  32:1309-1326.

Chiou, C.T., V.H. Freed, D.W. Schmedding, and R.L. Kohnert.  1977.
  Partition coefficients and bioaccumulation of selected organic chemicals.
  Environ. Sci. Technol. 11(5):475-478.

Glaus, D. and N. Walker.  1964.  The decomposition of toluene by soil
  bacteria.  J. Gen. Microbiol. 36:107-122.

Da mail, K.R, A.C. Lloyd, A.M. Winer, and J.N. Pitts, Jr.  1976.
  Reactivity scale for atmospheric hydrocarbons based on reaction with
  hydroxyl radical.  Environ. Sci. Technol. 10(7):692-696.

Davis, D.D., W. Bellinger, and S. Fischer.  1975.  Kinetics study of the
  reaction of the OH free radical with aromatic compounds.   I.  Absolute
  rate constants for reaction with benzene and toluene at 300K.  J. Phys.
  Chem. 79:293-294.

Gibson, D.T., J.R. Koch, and R.E. Kallio.  1968.   Oxidative degradation of
  aromatic hydrocarbons by microorganisms.  I.  Enzymatic formation of
  catechol from benzene.  Biochemistry 7:2653-2662.

Gibson, D.T. 1976.  Initial reactions in the bacterial degradation of
  aromatic hydrocarbons.  Zbl. Bakt. Hyg. I. Abt. Orig. B.  162:157-168.

Goodley, P.C. and M. Gordon.  1976. Characterization of industrial organic
  compounds in water.  Trans. Kentucky Academy of Science 37(1-2):11-15.

Hansen, D.A. , R. Atkinson, and J.N. Pitts, Jr.  1975.  Rate constants for
  the reaction of OH radicals with a series of aromatic hydrocarbons. J.
  Phys. Chem.   79:1763-1766.

Helfgott, T.B., F.L. Hart, and R.G. Bedard.  1977.  An index of refractory
  organics.  U.S. Environmental Protection Agency, (Office of Research and
  Development), Ada, Oklahoma. 131p. EPA 600/2-77-174.

Howard, P.H. and P.R. Durkin.  1974.  Sources of  contamination, ambient
  levels, and fate of benzene in  the environment.  U.S. Environmental
  Protection Agency,  (Office of Toxic Substances), Washington, D.C. 65p.
  EPA 560/5-75-005.
                                     71-J

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Laity, J.L., I.G. Burstain, and B.R. Appel.  1973.  Photochemical smog and
  the atmospheric reactions of solvents.  Chap. 7, pp. 95-112. Solvents
  Theory and Practice.  R.W. less (ed.)  Advances in Chemistry Series 124.
  Am. Chem. Soc., Washington, D.C.

Leo, A., C. Hansch, and D. Elkins.  1971.  Partition coefficients and their
  uses.  Chem. Rev. 71:525-616.

Mackay, D. and A.W. Wolkoff.  1973.  Rate of evaporation of low-solubility
  contaminants from water bodies  to atmosphere.  Environ. Sci. Technol.
  7(7):611-614.

Mackay, D. and P.J. Leinonen.  1975.  Rate of evaporation of low-solubility
  contaminants from water bodies  to atmosphere.  Environ. Sci. Technol.
  9(13):1178-1180.

Morrison, R.T. and R.N. Boyd.  1973.  Organic chemistry.  3rd Edition.
  Allyn and Bacon, Inc., Boston.  1258p.

National Research Council,  1977.  Fates of pollutants.  Research and
  development needs.  National Academy of Sciences, Washington, D.C. 144p.

Neely, W.B., D.R. Branson, and G.E. Blau.  1974.  Partition coefficient to
  measure bioconcentration potential of organic chemicals in fish. Environ.
  Sci. Technol. 8:1113-1115.

Silverstein, R.M. and G.C. Sassier.  1968.  Spectrometric identification of
  organic compounds.  2nd Edition.  John Wiley and Sons, New York.  256 p.

Thorn, N.S. and A.R. Agg.  1975.   The breakdown of synthetic organic
  compounds in biological processes. Proc. Roy. Soc. Lond. B 189:347-357.

U.S. Environmental Protection Agency.  1975.  Preliminary assessment of
  suspected carcinogens in drinking water.  U.S. Environmental Protection
  Agency, (Office of Toxic Substances), Washington, D.C. 33p. EPA
  560/4-75-003.

Walker, J.D. and R.R. Colwell.  1975.  Degradation of hydrocarbons and
  mixed hydrocarbon substrate by  microorganisms from Chesapeake Bay.
  Prog. Water Technol. 7(3-4):783-791.

Weast, R.C. (ed). 1977.  Handbook of chemistry and physics.  58th Edition.
  CRC Press, Inc., Cleveland, Ohio.  2398p.

Zobell, C.E. 1950.  Assimilation  of hydrocarbons by microorganisms.
  Advances in Enzymology and Related Subjects of Biochemistry.  F.F. Nord
  (ed.) Volume 10:443-486.  Interscience Publishers, Inc. New York.
                                    71-9

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                            72.   CHLOROBENZENE


72.1  Statement of Probable Fate

    Based on the information found,  it is not  possible to determine the
predominant aquatic fate of chlorobenzene.  There is some experimental
evidence indicating that volatilized chlorobenzene will undergo atmospheric
photooxidation in the presence of nitric oxide.   The extent to which such
photooxidation will occur at chlorobenzene concentrations prevalent in the
atmospheric environment in the presence of nitric oxide, as well as other
free radical initiators, is unknown.  Information found concerning the
biodegradation potential of chlorobenzene indicates that this compound,
being very persistent, will probably eventually  biodegrade, but not at a
substantial rate unless the microorganisms present are already growing on
another hydrocarbon source.

    Chlorobenzene has a high affinity for lipophilic materials, and it also
is reported to have a relatively low solubility  at temperatures anticipated
to be prevalent in most ambient waters.  Consequently, sorption, bioaccumu-
lation, and volatilization are expected to be  competing processes.  The rate
at which each of these competing processes occur will determine which fate
is predominant for chlorobenzene in the aquatic  environment.  Should vol-
atilization occur at a more rapid rate than sorption or bioaccumulation,
then atmospheric processes would be expected to  regulate the fate of chloro-
benzene.  On the other hand, should sorption and bioaccumulation occur more
rapidly than volatilization, biodegradation of chlorobenzene by aquatic
microorganisms would be anticipated to regulate  the fate of this compound.

72.2  Identification

    Chlorobenzene has been detected in finished  drinking water, in ground
water, in uncontaminated upland water, in waters contaminated by either
industrial, municipal, or agricultural wastes  (U.S. Environmental Protec-
tion Agency 1975), and in the atmosphere (Ware and West 1977).  The
chemical structure of chlorobenzene is shown below.
                                             Alternate Names

                                             Monochlorobenzene
    Chlorobenzene                            Benzene Chloride

    CAS NO. 108-90-7
    TSL NO. CZ 01750
                                     72-1

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72.3  Physical Properties

    The general physical properties of chlorobenzene are given below.

    Molecular weight                         112.56
    (Verschueren 1977)

    Melting point                            -45C
    (Verschueren 1977)

    Boiling point at 760 torr                132C
    (Verschueren 1977)

    Vapor pressure at 20C                   *

    Solubility in water                      **

    Log octanol/water partition coefficient  2.84
    (Leo et al. 1971;  Chiou et al. 1977)
*Values for the vapor pressure of chlorobenzene in water were reported to
be 8.8 torr at 20C (Verschueren 1977) and were calculated to be 11.72 torr
at 20C from the table of vapor pressures, critical temperatures and criti-
cal pressures of organic compounds in-Weast (1973).

**Values for the solubility of chlorobenzene in water were reported to be
500 mg/1 at 20C (Verschueren 1977),  488 mg/1 at 25C (Mardsen and Marr
1963), and 448 mg/1 at 30C (Chiou _etal. 1977).

72.4  Summary of Fate Data

    72.4.1  Photolysis

         No specific information pertaining to the direct photolysis of
chlorobenzene in the aqueous or atmospheric environments under ambient con-
ditions was found.

    72.4.2  Oxidation

         No specific information pertaining to the oxidation of chloro-
benzene in the aqueous environment under ambient conditions was found.

         Billing e_t al. (1976) studied the photodecomposition rate of
chlorobenzene under simulated atmospheric conditions.  The reactor was a
waterjacketed Pyrex cylinder maintained at 27 + 1C and 35% relative
                                     72-2

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humidity.  The ultraviolet light sources were two General Electric
275-Wreflector sunlamps, each of which had a short wavelength cutoff of 290
nm.  Chlorobenzene decomposed relatively slowly when compared to other
chlorinated compounds studied.  In the presence of nitric oxide, a probable
reaction initiator, chlorobenzene, at a concentration of 10 mg/1, underwent
43% reaction in 7.5 hours;  this corresponds to an estimated half-life of
8.7 hours.  No photodecomposition products were reported.  For comparison,
the estimated half-life of benzene proposed by Darnall et_ aL. (1976) was
2.4 to 24 hours.  Using the half-conversion time of m-xylene and
1,3,5-trimethylbenzene, which was reported to be somewhat less than four
hours (Altshuller ^t_ a_l. 1962), and the table of relative reactivities
given by Laity et_ al_. (1973), the corresponding range for the half-
conversion time for benzene was inferred to be approximately 20 to 50
hours.  Due to decreased susceptibility to electrophilic attack, chloro-
benzene would be anticipated to have a longer half-life than benzene under
similar conditions of exposure.  Because the experimental procedure of
Dilling et_ al. (1976) is different from the hypothetical conditions used
for the theoretical calculation of Darnall e_t_ a^L. (1976) the results cannot
be quantitatively compared.  However, they do show that chlorobenzene will
undergo photolysis over a defined range of times.  In addition, experiments
of the type carried out by Dilling e_t_ aJ. (1976) should, for accuracy, be
carried out for at least 2 or 3 half-times.

    72.4.3  Hydrolysis

         Chlorobenzene, being an aryl halide, will undergo nucleophilic
substitution only with extreme difficulty.  For example, Morrison and Boyd
(1973) report that chlorobenzene is converted into phenol by aqueous sodium
hydroxide only at temperatures over 300C.  Consequently, hydrolysis of
chlorobenzene under environmental conditions would not be expected to
occur.

    72.4.4  Volatilization

         Available data on chlorobenzene indicate that this compound prob-
ably volatilizes from the water column to the atmosphere at a relatively
rapid rate.  Garrison and Hill (1972) reported that at 300 mg/1 of chloro-
benzene volatilized almost completely (less than 1 mg/1 chlorobenzene re-
mained) from aerated distilled water in less than four hours.  The same
concentration of chlorobenzene volatilized almost completely (less than 1
mg/1 chlorobenzene remained) from unaerated distilled water in less than 3
days.  No further details of this experiment were reported.  Assuming a
first order process and also assuming that volatilization is the only re-
moval process operating (i.e., no gas-stripping and negligible adsorption
onto container walls), the data of Garrison and Hill (1972) yield estimates
                                 72-3

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of about 0.5 hours and 9 hours for evaporative half-lives of chlorobenzene
in water under conditions of aeration and quiescence,  respectively.
Because of the limited data on experimental  conditions,  it is difficult to
interpret the effect of aeration on the rate of evaporation.   Furthermore,
because of the possibility of air-stripping  during aeration,  it is
concluded that the 0.5 hour value is not a valid evaporative  half-life for
chlorobenzene in agitated water.  The value  of 9 hours appears to  be valid
for the non-agitated case.

         According to Mackay and Wolkoff (1973) the rate of evaporation of
pollutants having a low solubility in water  can be quite rapid even though
these compounds often have a high molecular  weight and a low vapor
pressure, and should, on these bases, evaporate slowly.   Mackay and  Wolkoff
(1973) contend that these compounds often have high activity coefficients
in water which cause unexpectedly high partial vapor pressures at  equili-
brium and thus high rates of evaporation. Although chlorobenzene, which
has a reasonably low solubility in water of  488 mg/1 at  25C (Mardsen and
Marr 1963), was not mentioned specifically,  other chlorinated hydrocarbons
having a low solubility in water were predicted to have  relatively rapid
rates of evaporation.  Presumably, chlorobenzene would have a rate of
evaporation similar to compounds having vapor pressure and solubility
values approximately equal to those values for chlorobenzene.

         The published values for the vapor  pressure (Weast 1973)  and
solubility (Mardsen and Marr 1963) of chlorobenzene in water at 25C were
used to compute the Henry constant as being  approximately 3.56 x 10"^
atmos. m-Ymole. This constant for chlorobenzene may be compared to the
value of 3.51 x 10"3 atmos m3/mole predicted for Aroclor 1248 at 25C
(Mackay and Leinonen 1975).  In the case of  Aroclor 1248, the half-life for
evaporation from a water column one meter thick was estimated by Mackay and
Leinonen (1975) to be 9.53 hours at 25C. An estimate of the corresponding
half-life for evaporation of chlorobenzene under the same conditions would
presumably be of the same order, approximately 10 to 11  hours, or  close to
the value of about 9 hours discussed previously.

         The interpretation of the environmental significance of the rate
of evaporation of chlorobenzene based on calculations involving an assumed
depth of one meter is dependent upon the type of environmental situation
encountered.  In situations where the water  body is turbulent with frequent
mixing between the surface layer and the bulk, as in a rapidly flowing
shallow river, or during white-capping on a  lake or ocean, the rate of
evaporation would be more rapid than for depths greater than one meter or
in quiescent water, as exists in a deep, slowly flowing  river.
                                      72-4

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    72.4.5  Sorption

         Although no specific environmental sorption studies were found in
the literature, the values of the log octanol/water partition coefficient
found for chlorobenzene (log P = 2.84, Leo e_t_ al_. 1971;  Chiou _e_t_ _al_. 1977)
indicate that sorption processes may be substantial for chlorobenzene at
pollutant concentrations anticipated in environmental waters.  Presumably,
chlorobenzene will be adsorbed by sedimentary organic material;  the extent
to which this possible adsorption will interfere with volatilization has
not been considered.

    72.4.6  Bioaccumulation

         Chlorobenzene was studied for its bioaccumulation potential in a
model aquatic ecosystem developed by Lu and Metcalf (1975).  This aquatic
ecosystem was devised for studying relatively volatile organic compounds
and simulating direct discharge of chemical wastes.  A 3-liter flask
containing members of an aquatic food chain, including daphnia, mosquito
larvae, snails, mosquito fish, and green filamentous algae, was maintained
at 80F with 12 hours daylight exposure.  Radiolabelled derivatives were
added to the flask in concentrations of 0.01 to 0.1 ppm.   After 48 hours,
the experiment was terminated.

         The contaminative efficacy of chlorobenzene was evaluated by de-
termining the quantitative distribution of the radioactivity in the
organisms, water, and air of the model aquatic ecosystem.  Chlorobenzene
was found to be a highly persistent compound, as demonstrated by the eco-
logical magnification (EM) values (often referred to as the bioconcentra-
tion or bioaccumulation factors by other authors) of 645 in fish, 1292 in
mosquito larva, 1313 in snail, 2789 in daphnia, and 4185 in alga.  Lu and
Metcalf (1975) define ecological magnification as the ratio of the con-
centration of the parent compound in an organism to the concentration in
the water.  Hydroxylation served as the predominant detoxification mecha-
nism of chlorobenzene by all of the organisms in the model aquatic food
chain except filamentous green alga, Oedogonium cardiacum.  Hydroxylation
of chlorobenzene to o- and p-chlorophenol and to 4-chlorocatechol was found
in mosquito larva and in water extracts.

         In addition to experimental evidence, there is empirical evidence
that chlorobenzene has a high potential for bioaccumulation in organisms.
Neely et al. (1974) and Lu and Metcalf (1975) have shown that the log
octanoTTwater partition coefficient (log P) correlates well with the
ability of a compound to accumulate in the lipids of tissues of living
organisms.  The log P value of 2.84 (Leo et al. 1971;  Chiou et al. 1977)
                                     72-5

-------
indicates that bioaccumulation of chlorobenzene by aquatic organisms at
pollutant concentrations anticipated in environmental waters will probably
occur.

    72.4.7  Biotransformation and Biodegradation

         Some species of soil bacteria have been demonstrated to be cap-
able of obtaining their energy and carbon requirements from chlorobenzene
(Matthews 1924).  A study by Gibson e_t_ _al_. (1968), however, indicated that
the microorganism Pseudomonas putida could oxidize chlorobenzene only when
it was already growing on an aromatic hydrocarbon source.  In this study,
cultures of Pseudomonas putida were initially grown on toluene as the sole
source of carbon for 15 hours and were subsequently grown on chlorobenzene
for up to 20 hours.  This oxidative degradation of chlorobenzene by
Pseudomonas putida resulted in the formation of 3-chlorocatechol (Gibson et
_al. 1968).

         In a study by Garrison and Hill (1972), chlorobenzene was exposed
to aerated mixed cultures of aerobic microorganisms to test their resis-
tance to microbial action.  Garrison and Hill (1972) reported that chloro-
benzene volatilized completely from these aerated cultures in less than one
day.  No other information on this particular study was given.

         In a report by Heukelekian and Rand (1955) BOD (Biochemical Oxygen
Demand) data on pure compounds were compiled and assembled into broad
groups on the basis of certain chemical similarities.  The BOD values were
expressed as grains per gram of chemical tested at 20C.  Most of the data
reported have BOD values based on a 5-day incubation period, although there
were  some BOD values based on an incubation period of 10 days or based on
ultimate values.  From the reported data, it appears that the presence of a
chlorine atom on a benzene ring results in a BOD value lower than that of
benzene itself.  For comparison, the BOD of benzene was reported as 1.20
g/g after a 10-day incubation period whereas the BOD of chlorobenzene was
reported to be 0.03 g/g after a 5-day incubation period.  This decrease in
BOD value as a result of the presence of a chlorine atom on the benzene
ring  indicates an increased resistance to microbial attack.  Although the
difference in the incubation time for benzene and chlorobenzene could
account to some degree for the difference in BOD values for the two
compounds, Alexander and Lustigman (1966) also found that the presence of a
chlorine atom on the benzene ring retarded the rate of biodegradation.

         Lu and Metcalf (1975) reported values for the biodegradability
index of chlorobenzene ranging from 0.014 to 0.063 in organisms found in a
model aquatic ecosystem.  The biodegradability index is defined as the
ratio of polar products of degradation to the non-polar products.  A  low
value for the biodegradability index indicates that a compound resists
biodegradation.  Comparison of the biodegradability index of chlorobenzene
with  that of some more widely studied persistent pollutants such as DDT and
                                      72-6

-------
aldrin, give a better idea of the significance of these values.  Lu and
Metcalf (1975) reported a biodegradability index for DDT of 0.012 in
mosquito fish compared to a value of 0.015 for aldrin, and 0.014 for
chlorobenzene .

    72.4.8  Other Reactions

         In an investigation by Carlson et_ _a_l_. (1975), monosubstituted
aromatics were exposed to low concentrations of aqueous chlorine (7 x
10~^M) for 20 minutes at varying pH conditions to determine the extent of
chlorine incorporation into aromatic compounds.  Chlorine, being an elec-
trophile, was found to be more slowly incorporated into aromatic compounds
containing deactivating groups such as chloro, nitro, nitrile, and carbonyl
than into aromatic compounds containing activating groups such as hydroxyl,
ether, amine derivatives, or alkyl.  For comparison, when chlorobenzene, a
compound containing a ring-deactivating group, and phenol, a compound
containing a ring-activating group, were exposed at pH 3 to 7 x 10~^M
aqueous chlorine for 20 minutes, the chlorine uptake was 1.8 +_ 0.1% and
97.8 + 0.1%, respectively.  Chlorobenzene was not chlorinated at higher
pH. Since the typical range of pH found during the course of most water
treatment processes is 5 to 9, there is a low probability of forming higher
chlorinated benzenes from reactions of chlorine and chlorobenzene.

72.5  Data Summary

    There is not enough environmentally significant information on photo-
lysis, oxidation, or sorption processes to be able to predict the aquatic
fate of chlorobenzene.  Chlorobenzene has a high affinity for lipophilic
materials, and it is also reported to have a relatively low solubility at
temperatures expected to prevail in most ambient waters.  Consequently,
sorption, bio accumulation, and volatilization are expected to be competing
processes.  The rate at which each of these competing processes occur will
dictate which fate is predominant for chlorobenzene in the aquatic
environment.  These data are summarized in Table 72-1.
                                     72-7

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                                                   Table  72-1

                                   Summarv of Aauatic Fate of Chlorobenzene
Environmental
   Process

Photolysis

Oxidation
Hydrolysis
Volatilization
Sorption
Bioaccumulation
 Biotransformation/
   Biodegradation
           Summary
          Statement                       Rate

No information found.

No information was found concerning
the oxidation of Chlorobenzene
in ambient waters.  Experimental
evidence indicates that volatilized
Chlorobenzene will undergo atmos-
pheric photooxidation in the pre-
sence of nitric oxide.  The extent
to which photooxidation will occur
at Chlorobenzene concentrations
prevalent in the atmospheric en-
vironment in the presence of nitric
oxide or other initiators in unknown.

Chlorobenzene probably will not
hydrolyze in ambient waters due to
the extreme difficulty with which
aryl halides undergo nucleophilic
substitution.

This compound probably volatilizes
from the water column to the atmos-
phere at a relatively rapid rate.

The high log P value found for             -
Chlorobenzene indicates that sorp-
tion processes may \>e substantial
for Chlorobenzene at pollutant
concentrations anticipated in
environmental waters.

Experimental and empirical evidence
indicates that Chlorobenzene has an inter-
mediate potential for bioaccumulation in
the lipids of tissues of living organisms.

Information  found concerning  the           -
biodegradation potential of Chloro-
benzene  indicates that  this compound
will probably eventually biodegrade,
but not  at a substantial rate unless
the microorganisms present are already
growing  on another hydrocarbon source.
                                                                           Half-Life
  8.7 hours
Confidence
 of Data

   Low

   Low
   9 hours
1C or 11 hours
                                                                                                Medium
                                                                                                Medium
                                                                                                Low
                                                                                                Medium
                                                                                                Medium
 a.   There  is insufficient  information  in  the  reviewed literature to permit assessment of a most probable fate.
 b.   This half-life is the  estimated half-life based on the experimental results and conditions of Billing et_ &._
     (1976)  and  probably  does  not  reflect  the  photooxidation half-life of Chlorobenzene under environmental
     conditions.
 c.   This half-life is tha  estimated half-life based on the experimental results and conditions of Garrison and
     Hill  (1972)  in unaerated  distilled water.
 d.   This half-life is based on  the calculated Henry constant of Chlorobenzene which is of the same order as the
     Henry  constant predicted  for  Aroclor  1248 at  25C by Mackay and Leinonen (1975).   Since the half-life for
     evaporation of Aroclor 1248 from a water  column one meter thick was estimated by Mackay and Leinonen (1975)
     to  be  9.53  hours  at  25 C,  the half-life for evaporation of Chlorobenzene under the same conditions was
     assumed to  be  only slightly longer.
                                                    72-8

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72.6  Literature Cited

Alexander, M. and B.K. Lustigman.  1966.  Effect of chemical structure on
  microbial degradation of substituted benzenes.  J. Agr. Food Chera.
  14(4):410-413.

Altshuller, A.P., I.R. Cohen, S.F. Sleva, and S.L. Kopcaynski.    1962.
  Air pollution: photooxidation of aromatic hydrocarbons.    Science
  138:442-443.

Carlson, R.M., R.E. Carlson, H.L. Kopperman, and R. Caple.  1975.  Facile
  incorporation of chlorine into aromatic systems during aqueous
  chlorination processes.  Environ. Sci. Technol.  9(7):674-675.

Chiou, C.T., V.H. Freed, D.W. Schmedding, and R.L. Kohnert.  1977.
  Partition coefficient and bioaccumulation of selected organic chemicals.
  Environ. Sci. Technol.  11(5) :475-478.

Darnall, K.R., A.C. Lloyd, A.M. Winer, and J.N. Pitts, Jr.  1976.
  Reactivity scale for atmospheric hydrocarbons based on reaction with
  hydroxyl radical.  Environ. Sci. Technol. 10(7):692-696.

Dilling, W.L., C.J. Bredeweg, and N.B. Tefertiller.  1976.  Simulated
  atmospheric photodecomposition rates of methylene chloride,
  1,1,1-trichloroethane, trichloroethylene, and other compounds.  Environ.
  Sci. Technol.  10(4):351-356.

Garrison, A.W. and D.W. Hill.  1972.  Organic pollutants from mill persist
  in downstream waters.  Am. Dyest. Rep. 21-25.

Gibson, D.T., J.R. Koch, C.L. Schuld, and R.E. Kallio.  1968.  Oxidative
  degradation of aromatic hydrocarbons by microorganisms.  II.  Metabolism
  of halogenated aromatic hydrocarbons.  Biochemistry 7(11):3795-3802.

Heukelekian, H. and M.C. Rand.  1955.  Biochemical oxygen demand of pure
  organic compounds.  Sewage and Industrial Wastes 27(9):1040-1053.

Laity, J.L., I.G. Burstain, and B.R. Appel.  1973.  Photochemical smog and
  the atmospheric reactions of solvents.  Chap. 7.  pp. 95-112.  Solvents
  Theory and Practice.  R.W. Tess (ed.)  Advances in Chemistry Series 124.
  Am. Chem. Soc., Washington, D.C.

Leo, A., C. Hansch, and D. Elkins.  1971.  Partition coefficients and
  their uses.  Chem. Rev. 71:525-616.

Lu,  P. and R.L. Metcalf.  1975.  Environmental fate and biodegradability of
  benzene derivatives as studied in a model aquatic ecosystem.   Environ.
  Health Perspect. 10:269-284.
                                    72-9

-------
Mackay, D. and A.W. Wolkoff.  1973.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. Technol.
   7(7):611-614.

Mackay, D. and P.J. Leinonen.  1975.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. Technol.
  9(13):1178-1180.

Mardsen, C. and S. Marr.  1963.  Solvents guide.  Cleaver-Hume   Press
  Ltd., London.  239p.

Matthews, A.   1924.  Partial sterilization of soil by antiseptics.  J. Agr.
  Sci.  14:1-57.

Morrison, R.T. and R.N. Boyd.  1973.  Organic chemistry.  3rd Edition.
  Allyn and Bacon, Inc., Boston.   1258p.

Neely, W.B., D.R. Branson, and G.E. Blau.  1974.  Partition coefficient  to
  measure bioconcentration potential of organic chemicals in fish.
  Environ. Sci. Technol. 8:1113-1115.

U.S. Environmental Protection Agency.   1975.  Preliminary assessment  of
  suspected carcinogens in drinking water.  U.S. Environmental Protection
  Agency, Office of Toxic Substances, Washington, D.C.  33p.   (EPA
  560/4-75-003).

Verschueren, K.  1977.  Handbook of environmental data  on organic
  chemicals.   Van Nostrand/Reinhold Press, New York.  659p.

Ware,  S.A. and W.L. West.  1977.   Investigation of selected potential
  environmental contaminants:  halogenated benzenes.  U.S. Environmental
  Protection Agency, Office of Toxic Substances, Washington, D.C.   283p.
  EPA  560/2-77-004.

Weast, R.C.   1973.  (ed).  Handbook of  chemistry and physics.  54th
  Edition.  CRC Press,  Inc., Cleveland, Ohio.  2452p.
                                    72-10

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                73.   1,2-DICHLOROBENZENE  (o^DICHLOROBENZENE )


 73.1   Statement  of Probable  Fate

    Based  on the information found,  it is  not  possible  to  determine  the
 predominant aquatic  fate  of  1,2-dichlorobenzene.   There is some  evidence
 that  dichlorobenzenes  in  general  are reactive  toward  hydroxyl  radicals in
 air with a half-life of approximately three  days.   Products  and  further de-
 tails of such  photooxidation reactions,  however, were not  indicated.   In-
 formation  concerning the  biodegradation  potential  of  1,2-dichlorobenzene
 indicates  that  this  compound is very persistent and will probably, at  best,
 be biodegraded  very  slowly by microorganisms already  growing on  another
 hydrocarbon source.

     1,2-Dichlorobenzene has  a high affinity  for lipophilic materials and it
 is reported to  have  a  relatively  low vapor pressure and low  aqueous  solu-
 bility at  ambient temperatures.   Consequently, sorption, bioaccumulation,
 and volatilization are expected to be competing processes.   The  rate at
 which each of  these  competing processes  occur  will determine which fate is
 predominant for 1,2-dichlorobenzene  in the aquatic environment.  Should
 volatilization  occur at a more rapid rate  than sorption or bioaccumulation,
 then  atmospheric processes would  be  expected to regulate the fate of
 1,2-dichlorobenzene.   On  the other hand, should sorption and bioaccumula-
 tion  occur more rapidly than volatilization, biodegradation  of 1,2-di-
 chlorobenzene  by aquatic  microorganisms  would  be anticipated to  regulate
 the fate of this compound.

 73.2   Identification

     1,2-Dichlorobenzene has  been  detected  in finished drinking water, in
 superchlorinated municipal wastewaters,  in ground  water (U.S.  Environmental
 Protection Agency 1975),  in  wastewater effluent (Glaze  and Henderson 1975),
 and in the atmosphere  (Ware  and West  1977).  The chemical structure of
 1,2-dichlorobenzene  is shown below.
                                             Alternate Names

                                             o-Dichlorobenzene
                                             Orthodichlorobenzene
     1,2-Dichlorobenzene                      Dowtherm E

     CAS NO. 95-50-1
     TSL NO. CZ 45000
                                     73-1

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73.3  Physical Properties

    The general physical properties  of 1,2-dichlorobenzene are given below.

    Molecular weight                        147.01
    (Weast 1977)

    Melting point                            -17.0C
    (Weast 1977)

    Boiling point at 760 torr                180.5C
    (Weast 1977)

    Vapor pressure at 25 C                   1.5 torr*

    Solubility in water at 25C              145 mg/1
    (Verschueren 1977)

    Log octanol/water partition coefficient  3.38
    (Leo et al. 1971)
*Values for the vapor pressure of 1,2-dichlorobenzene in water were re-
ported to be 1.5 torr at 25C (Verschueren 1977)  and were calculated to be
1.45 torr from the table of vapor pressures,  critical temperatures and
critical pressures of organic compounds in Weast  (1973).

73.4  Summary of Fate Data

    73.4.1  Photolysis

         No specific information pertaining to the direct photolysis of
1,2-dichlorobenzene in the aquatic or atmospheric environments was found.

    73.4.2  Oxidation

         According to Ware and West (1977) 1,2-dichlorobenzene is resistant
to autooxidation by the peroxy radical (R02*) in  water.  No more de-
tails of this phenomenon were reported.  Dichlorobenzenes in general were
reported by Ware and West (1977) to be reactive toward hydroxyl radicals
(OH') in air with a half-life of approximately three days.  Products and
further details of such photooxidation reactions  were not indicated.
1,2-Dichlorobenzene specifically was reported by  Ware and West (1977) to be
resistant to autooxidation by ozone in air.
                                    73-2

-------
    73.4,3  Hydrolysis

         No specific information pertaining to the hydrolysis of 1,2-di-
chlorobenzene has been found.  Although Ware and West (1977) report that
the inductive electronegative effect of halogen substitutents on an aromat-
ic ring facilitates attack by nucleophiles such as OH~, Morrison and Boyd
(1973) report that aryl halides are characterized by very low re-
activity toward nucleophilic reagents such as OH~.

         As an example of the difficulty with which aryl halides undergo
nucleophilic substitution, the conditions necessary for the nucleophilic
substitution of hexachlorobenzene to a pentachlorophenyl derivative were re-
ported by Patai (1973) to be the presence of aqueous ammonia at a tempera-
ture of at least 250C.  On this basis, 1,2-dichlorobenzene, being less
chlorinated and, consequently, less easily attacked by nucleophiles than
hexachlorobenzene, would not be expected to undergo hydrolysis at an
appreciable rate under environmental conditions.

    73.4.4  Volatilization

         Available data on 1,2-dichlorobenzene indicate that this compound
probably volatilizes from the water column to the atmosphere at a rela-
tively rapid rate.  Garrison and Hill (1972) reported that a 100 mg/1 con-
centration of 1,2-dichlorobenzene volatilized almost completely (less than
1 mg/1 of 1,2-dichlorobenzene remained) from aerated distilled water in
less than 4 hours.  The same concentration of 1,2-dichlorobenzene volatil-
ized almost completely (less than 1 mg/1 of 1,2-dichlorobenzene remained)
from unaerated distilled water in less than 3 days.  No further details of
this experiment were reported.  The data of Garrison and Hill (1972) can be
used to calculate approximate values for evaporative half-lives.  For the
aerated solution, the calculated half-life is less than 30 minutes;
however, since the aeration probably caused air-stripping of the 1,2-di-
chlorobenzene, this value is not representative of an evaporative half-life
under conditions of agitation.  The data for unaerated conditions,
apparently close to quiescence, correspond to a half-life of less than
about nine hours.

         According to Mackay and Wolkoff (1973) the rate of evaporation of
pollutants having a low solubility in water can be quite rapid even though
these compounds often have a high molecular weight and a low vapor
pressure, and should, on these bases, evaporate slowly.  Mackay and Wolkoff
(1973) contend that these compounds often have high activity coefficients
in water which cause unexpectedly high equilibrium partial vapor pressures
and thus high rates of evaporation.  Although 1,2-dichlorobenzene  was not
mentioned specifically, other chlorinated hydrocarbons having relatively
low solubility in water were predicted to have somewhat rapid rates of
evaporation.
                                       73-3

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         The published value for the vapor pressure (Verschueren 1977) and
solubility in water (Verschueren 1977) of 1,2-dichlorobenzene at 25C were
used to compute the Henry constant as being approximately 1.99 x 10~3
atmos. m-Vmole which may be compared to the value of 1.55 x 10"^ atmos.
m-Vmole predicted for biphenyl at 25C (Mackay and Leinonen 1975).  In
the case of biphenyl, the half-life for evaporation from a water column one
meter thick was estimated by Mackay and Leinonen (1975) to be 7.52 hours at
25C.  An estimate of the corresponding half-life for evaporation of
1,2-dichlorobenzene under the same conditions would presumably be of the
same order,  at most, approximately 8 or 9 hours.  Mackay and Leinonen
(1975) point out that interpretation of the environmental significance of
the rate of evaporation of compounds such as 1,2-dichlorobenzene from en-
vironmental waters using values calculated at 1 meter depth is dependent
upon the type of environmental situation encountered.  In sitviations where
the water body is turbulent with frequent mixing between the surface layer
and the bulk, as in a rapidly flowing shallow river or during white-capping
on a lake or ocean, the rate of evaporation would be more rapid than for
depths greater than 1 meter or in quiescent water, as exists in a deep,
slowly flowing river.

    73.4.5  Sorption

         Although no specific environmental sorption studies were found in
the literature, the value of the log octanol/water partition coefficient
for 1,2-dichlorobenzene (log P = 3.38, Leo _e_t _al. 1971) indicates that
sorption processes may be substantial for 1,2-dichlorobenzene at pollutant
concentrations anticipated in environmental waters.  Presumably, 1,2-di-
chlorobenzene will be adsorbed by sedimentary organic material;  the extent
to which this possible adsorption will interfere with volatilization has
not been considered.

    73.4.6  Bioaccumulation

         Although no experimental evidence of the bioaccumulation potential
of 1,2-dichlorobenzene was found, there is indirect evidence that 1,2-di-
chlorobenzene has a high potential for bioaccumulation in aquatic organ-
isms.  Neely _e_t al. (1974) and Lu and Metcalf (1975) have shown that the
log octanol/water partition coefficient (log P) correlates well with the
ability of a compound to accumulate in the lipids of tissues of living
organisms.  Furthermore, the incorporation of chlorine into an organic
molecule increases its lipophilic character resulting in an increased
bioaccumulation potential (Kopperman _e_t_ _al_. 1976).  For comparison, chloro-
benzene, a compound containing only one chlorine atom, and 1,2-dichloro-
benzene, a compound having two chlorine atoms, have log P values of 2.84
(Leo _et _al. 1971;  Chiou _e_t al. 1977) and 3.38 (Leo _et _al. 1971), re-
spectively.  Since it has been established experimentally that chloro-
                                    73-4

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benzene bioaccumulates in aquatic organisms (Lu and Metcalf 1975), 1,2-di-
chlorobenzene,  would be expected to be bioaccumulated by aquatic organisms
at least as much as chlorobenzene.

    73.4.7  Biotransformation and Biodegradation

         According to Ware and West (1977), the more highly halogenated a
compound becomes, the more resistant it is to biodegradation.   Experimental
(Lu and Metcalf 1975) as well as empirical evidence (Leo et_ l. 1971; Chiou
et_ a_l.  1977) has been found indicating that chlorobenzene is a persistent
chemical and is not readily biodegraded unless the microorganisms present
are already growing on another hydrocarbon source. Furthermore, Alexander
and Lustigman (1966) found that the presence of a chlorine atom on the ben-
zene ring retarded the rate of biodegradation.  On these bases, 1,2-di-
chlorobenzene,  being more highly chlorinated than chlorobenzene, would
presumably biodegrade at least as slowly as chlorobenzene under the same
conditions of exposure to microorganisms.

         Evidence that 1,2-dichlorobenzene is a persistent chemical is
presented in a study by Thorn and Agg (1975), which lists 1,2-dichloroben-
zene as a synthetic organic chemical which is unlikely to be removed during
biological sewage treatment even after prolonged exposure of the biota. In
contrast, Thorn and Agg (1975) list chlorobenzene, a compound with one less
chlorine atom than 1,2-dichlorobenzene, as a synthetic organic chemical
which should be degradable by biological sewage treatment provided that
suitable acclimatization can be achieved.  Based upon inference 1,2-di-
chlo.robenzene should be resistent to biodegradation; but in the absence of
solid experimental evidence, no definitive conclusion can be reached.

73.5  Data Summary

    There is not enough environmentally significant information on photoly-
sis, oxidation, sorption, or biodegradation processes to be able to predict
the aquatic fate of 1,2-dichlorobenzene.   1,2-Dichlorobenzene has a high
affinity for lipophilic materials and it is reported to have a relatively
low vapor pressure and low solubility at temperatures expected to prevail
in most ambient waters.  Consequently, sorption, biaccumulation, and
volatilization are expected to be competing processes. The rate at which
each of these competing processes occur will dictate which fate is pre-
dominant for 1,2-dichlorobenzene in the aquatic environment.  These data
are summarized in Table 73-1.
                                     73-5

-------
                                                     Table  73-1

                            Summary  of  Aqtmtlc Fnle of 1,2-Dichlorobenzene
Environmental
   Process a
              .Summary
             SLatempDt
                                                                  Rate
                                                                                 Half-Life
                                                                           Conf ttlnncp
                                                                             of Data
Photolysis           No information found

Oxidation            1,2-Dichlorobenzene is reported to be
                     resistant to autooxidation by the peroxy
                     radical (ROs-) in water.   Dichlorobenzenes
                     were reported to be reactive toward hydroxyl
                     radicals (OH1) in air.

Hydrolysis           1,2-Dichlorobenzene will  probably not hydro-  -
                     lyze in ambient waters due to the extreme
                     difficulty with which aryl halides undergo
                     nucleophilic substitution.

Volatilization       This compound probably volatilizes from the
                     water column to the atmosphere at a relatively
                     rapid rate.

Sorption             The high log P value found for 1,2-dichloro-  -
                     benzene indicates that sorption processes
                     may be sustantial for this compound at pollu-
                     tant concentrations anticipated in environ-
                     mental waters.
                                                                           Low

                                                           ~3 daysb        Low
                                                                           Medium
                                                      ^8  or  9  hours    a   Medium
                                                      less than 9 hours
                                                                           Low
Bioaccumulation
Empirical evidence indicates that 1,2-di-
chlorobenzene will bioaccumulate in the
lipids of tissues of living organisms.
                                                                                                Low
Biotransformat ion/
  Biodegradation
Will biodegrade, at best,
rate.
                          at a very slow
                                                                           Low
a.   There is insufficient information in the reviewed literature to permit assessment
     of a most probable fate.

b.   This half-life is the reported half-life of dichlorobenzene toward hydroxyl radicals in air
     cited in Ware and West (1977).

c.   This half-life is based on the calculated Henry constant of 1,2-dichlorobenzene which is of
     the same order as the Henry constant predicted for biphenyl at 25C by Mackay and Leinonen
     (1975) .   Since the half-life for evaporation of biphenyl from a water column one meter thick
     was estimated by Mackay and Leinonen to be 7.52 hours at 25C, the half-life for evaporation
     of 1,2-dichlorobenzene under the same conditions was assumed to be of the same order, at most,
     approximately 8 or 9 hours.
 d.
     This half-life is the estimated half-life based on the experimental results and conditions of
     Garrison and Hill (1972) in unaerated distilled water.
                                                   73-6

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73.6  Literature Cited

Alexander, M. and B.K. Lustigman.  1966.  Effect of chemical structure on
  microbial degradation of substituted benzenes.  J. Agr. Food Chem.
  14(4):410-413.

Chiou, C.T.,  V.H. Freed, D.W. Schmedding, and R.L. Kohnert.  1977.
  Partition coefficient and bioaccumulation of selected organic chemicals.
  Environ. Sci. Technol.  11(5 ):475-478.

Garrison, A.W. and D.W. Hill.  1972.   Organic pollutants from mill persist
  in downstream waters.  Am. Dyest. Rep. 21-25.

Glaze, W.H. and J.E. Henderson.   1975.  Formation of organochlorine
  compounds from the chlorination of a municipal secondary effluent.   J.
  Water Poll. Cont. Fed.  47(10 ):2511-2515.

Kopperman, H.L., D.W. Kuehl, and G.E. Glass.  1976.  Chlorinated compounds
  found in waste treatment effluents and their capacity to bioaccumulate.
  Proceedings of the conference on the environmental impact of water
  chlorination.  Oak Ridge, Tennessee, October 22-24.  (Preprint only.)

Leo, A., C. Hansch, and D. Elkins.  1971.  Partition coefficients and their
  uses.  Chem. Rev. 71:525-616.

Lu, P. and R.L. Metcalf.  1975.   Environmental fate and biodegradability of
  benzene derivatives as studied in a model aquatic ecosystem.   Environ.
  Health Perspect.  10:269-284.

Mackay, D. and A.W. Wolkoff.  1973.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. Technol.
  7(7):611-614.

Mackay, D. and P.J. Leinonen.  1975.   Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. Technol.
  9(13):1178-1180.

Morrison, R.T. and R.N. Boyd.  1973.   Organic chemistry.  3rd Edition.
  Allyn and Bacon, Inc., Boston.  1258p.

Neely, W.B.,  D.R. Branson, and G.E. Blau.  1974.  Partition coefficient to
  measure bioconcentration potential of organic chemicals in fish.
  Environ. Sci. Technol. 8:1113-1115.

Patai, S.  (ed).  1973.  The chemistry of the carbon-halogen bond:  Part 2.
  John Wiley Interscience, New York.   1215p.
                                     73-7

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Thorn, N.S. and A.R. Agg.  1975.  The breakdown of synthetic organic
  compounds in biological processes.  Proc. Roy. Soc. Lond.  B 189:347-357.

U.S. Environmental Protection Agency.  1975.  Preliminary assessment of
  suspected carcinogens in drinking water.  U.S. Environmental Protection
  Agency, Office of Toxic Substances, Washington, D.C.  33p.  EPA
  560/4-75-003.

Verschueren, K.  1977.  Handbook of environmental data on organic
  chemicals.  Van Nostrand/Reinhold Press, New York.  659p.

Ware, S.A. and W.L. West.  1977.  Investigation of selected potential
  environmental contaminants:  halogenated benzenes.  U.S. Environmental
  Protection Agency, Office of Toxic Substances, Washington, D.C.  283p.
  EPA 560/2-77-004.

Weast, R.C. (ed).  1973.  Handbook of chemistry and physics.  54th Edition.
  CRC Press, Inc., Cleveland, Ohio.  2452p.

Weast, R.C. (ed).  1977.  Handbook of chemistry and physics.  58th Edition.
  CRC Press, Inc., Cleveland, Ohio.  2398p.
                                     73-8

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               74.  1.3-DICHLOROBENZENE (m-DICHLOROBENZENE )


74.1  Statement of Probable Fate

    Based on the information found, it is not possible to determine the
predominant aquatic fate of 1,3-dichlorobenzene.  The probable aquatic fate
of this compound can only be inferred from information available on its
more thoroughly studied structural isomers, 1,2- and 1,4-dichlorobenzene,
and from information available on dichlorobenzenes in general.  There is
some evidence that dichlorobenzenes are reactive toward hydroxyl radicals
in air with a half-life of approximately three days.  Products and further
details of such photooxidation reactions, however, were not indicated.
Information concerning the biodegradation potential of 1,3-dichlorobenzene
indicates that this compound is very persistent and will probably, at best,
be biodegraded only very slowly by microorganisms already growing on
another hydrocarbon source.

    1,3-Dichlorobenzene has a high affinity for lipophilic materials, a
relatively low vapor pressure, and low aqueous solubility at ambient
temperatures.  Consequently, sorption, bioaccumulation, and volatilization
are expected to be competing processes.  The rate at which each of these
competing processes occurs will determine which fate is predominant for
1,3-dichlorobenzene in the aquatic environment.  Should volatilization
occur at a more rapid rate than sorption or bioaccumulation, then
atmospheric processes would be expected to regulate the fate of 1,3-
dichlorobenzene.  On the other hand, should sorption and bioaccumulation
occur more rapidly than volatilization, biodegradation of 1,3-dichloro-
benzene by aquatic microorganisms would be anticipated to control the fate
of this compound.

74.2  Identification

    1,3-Dichlorobenzene has been detected in drinking water and ground
water (U.S. Environmental Protection Agency 1975).  The chemical structure
of 1,3-dichlorobenzene is shown below.
                  C\                         Alternate Names
    1,3-Dichlorobenzene                      m-Dichlorobenzene
                                             Metadichlorobenzene
    CAS NO. 541-73-1
    TSL NO. None Assigned
                                    74-1

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74.3  Physical Properties

    The general physical properties of 1,3-dichlorobenzene are given below.

    Molecular weight                         147.01
    (Weast 1977)

    Melting point                            -24.7C
    (Weast 1977)

    Boiling point at 760 torr                173C
    (Weast 1977)

    Vapor pressure at 25C                   2.28 torr*
    (Weast 1973)

    Solubility in water at 25C              123 mg/1
    (Verschueren 1977)

    Log octanol/water partition coefficient  3.38
    (Leo et al. 1971)
*This vapor pressure was calculated from the table of vapor pressures,
critical temperatures and critical pressures of organic compounds in Weast
(1973).

74.4  Summary of Fate Data

    74.4.1  Photolysis

         No specific information pertaining to the direct photolysis of
1,3-dichlorobenzene in the aquatic or atmospheric environments was found.

    74.4.2  Oxidation

         No specific information pertaining to the oxidation of 1,3-di-
chlorobenzene in ambient waters was found.

         Dichlorobenzenes in general were reported by Ware and West (1977)
to be reactive toward hydroxyl radicals (OH*) in air with a half-life of
approximately three days.  Products and further details of such photooxida-
tion reactions were not indicated.

    74.4.3  Hydrolysis

         No specific information pertaining to the hydrolysis of 1,3-di-
chlorobenzene in ambient waters was found.   Although Ware and West (1977)
report that the inductive electronegative effect of halogen substitu-
                                     74-2

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ents on the benzene ring facilitates attack by nucleophiles such as OH~~,
Morrison and Boyd (1973) report that aryl halides will undergo nucleophilic
substitution only with extreme difficulty.  For example, Patai (1973) re-
ported that the minimum conditions necessary for the nucleophilic substitu-
tion of hexachlorobenzene to a pentachlorophenyl derivative were the
presence of aqueous ammonia and a temperature of at least 250C.  On these
bases, 1,3-dichlorobenzene, being less chlorinated and, consequently, less
easily attacked by nucleophiles than hexachlorobenzene, would not be anti-
cipated to undergo hydrolysis at an appreciable rate under environmental
conditions.

    74.4.4  Volatilization

         Garrison and Hill (1972) found that chlorobenzene, 1,2-dichloro-
benzene, 1,4-dichlorobenzene, and 1,2,4-trichlorobenzene volatilized almost
completely (less than one mg/1 of the compounds remained) from aerated
distilled water in less than four hours.  Initial concentrations of the
compounds were 300 mg/1, 100 mg/1, 300 mg/1, and 100 mg/1, respectively.
The same concentrations of the four chlorinated benzenes volatilized from
unaerated distilled water in less than three days.  1,2,4-Trichlorobenzene
disappeared from unaerated distilled water in less than two days.  Pre-
sumably, 1,3-dichlorobenzene under similar conditions would volatilize at a
similar rate.

         According to Mackay and Wolkoff (1973) the rate of evaporation of
pollutants having a low solubility in water can be quite rapid even though
these compounds often have a high molecular weight and a low vapor pres-
sure, and should, on these bases, evaporate slowly.  Mackay and Wolkoff
(1973) contend that these compounds often have high activity coefficients
in water which cause unexpectedly high equilibrium partial vapor pressures
and thus high rates of evaporation.  Although 1,3-dichlorobenzene, which
has a reasonably low vapor pressure of 2.28 torr (calculated from Weast
1973) and a low solubility in water of 123 mg/1 at 25C (Verschueren 1977)
was not mentioned specifically, other chlorinated hydrocarbons having a low
solubility in water and a low vapor pressure were predicted to have rela-
tively rapid rates of evaporation.  Presumably, 1,3-dichlorobenzene would
have a rate of evaporation similar to these chlorinated compounds having
vapor pressure and solubility values approximately equal to those values
for 1,3-dichlorobenzene.

         The calculated value for the vapor pressure (Weast 1973) and the
published value for the solubility in water (Verschueren 1977) of 1,3-di-
chlorobenzene at 25C were used to compute the Henry constant as being
approximately 3.58 x 10~3 atmos. ra-Vraole, which may be compared to the
value of 3.51 x 10"^ atmos. m-^/mole predicted for Aroclor 1248
                                    74-3

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at 25C (Mackay and Leinonen 1975).  In the case of Aroclor 1248, the
half-life for evaporation from a water column one meter thick was estimated
by Mackay and Leinonen (1975) to be 9.53 hours at 25C.  The corresponding
half-life for evaporation of 1,3-dichlorobenzene under the same conditions
would presumably be of the same order, at most, approximately 10 or 11
hours.   Mackay and Leinonen (1975) point out that interpretation of the en-
vironmental significance of the rate of evaporation of compounds such as
1,3-dichlorobenzene from environmental waters using values calculated at 1
meter depth is dependent upon the type of environmental situation en-
countered.  In situations where the water body is turbulent with frequent
mixing between the surface layer and the bulk, as in a rapidly flowing
shallow river or during white-capping on a lake or ocean, the rate of
evaporation would be more rapid than for depths greater than 1 meter or in
quiescent water, as evidenced in a deep, slowly flowing river.

    74.4.5  Sorption

         Although no specific environmental sorption studies were found in
the literature, the value of the log octanol/water partition coefficient
found for 1,3-dichlorobenzene (log P = 3.38, Leo et_jl. 1971) indicates
that sorption processes may be substantial for 1,3-dichlorobenzene at
pollutant concentrations anticipated in environmental waters.   Presumably,
1,3-dichlorobenzene will be adsorbed by sedimentary organic material;  the
extent to which this possible adsorption will interfere with volatilization
has not been considered.

    74.4.6  Bioaccumulation

         Although no experimental evidence of the bioaccumulation potential
of 1,3-dichlorobenzene was found, there is empirical evidence that 1,3-di-
chlorobenzene has a high potential for bioaccumulation in aquatic organ-
isms.  Neely et_ jJ.. (1974) and Lu and Metcalf (1975) have shown that the
log octanol/water partition coefficient (log P) correlates well with the
ability of a compound to accumulate in the lipids of tissues of living
organisms.  Furthermore, the incorporation of chlorine into an organic
molecule increases its lipophilic character resulting in an increased
bioaccumulation potential (Kopperman e_t_ _a_l. 1976).  For comparison,
chlorobenzene, a compound containing only one chlorine atom, and 1,3-di-
chlorobenzene, a compound having two chlorine atoms, have log P values of
2.84 (Leo et_ a_l. 1971;  Chiou et_ _al.  1977) and 3.38 (Leo et_ al. 1971), re-
spectively.  Since it has been established experimentally that chloro-
benzene bioaccumulates in aquatic organisms (Lu and Metcalf 1975), 1,3-di-
chlorobenzene would be expected to be bioaccumulated by aquatic organisms
at least as much as chlorobenzene.
                                     74-4

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    74.4.7  Biotrans format ion and Biodegradation

         According to Ware and West (1977), the more highly halogenated a-
compound becomes,  the more resistant it is to biodegradation.   Experimental
(Lu and Metcalf 1975) as well as empirical evidence (Leo et al.  1971; Chiou
et_ al_.  1977) has been found indicating that chlorobenzene is a persistent
chemical and is not readily biodegraded unless the microorganisms present
are already growing on another hydrocarbon source. Furthermore,  Alexander
and Lustigman (1966) found that the presence of a chlorine atom on the ben-
zene ring retarded the rate of biodegradation.  On these bases,  1,3-di-
chlorobenzene,  being more highly chlorinated than chlorobenzene, would
presumably biodegrade at least as slowly as chlorobenzene under the same
conditions of exposure to microorganisms.

         Evidence that 1,3-dichlorobenzene is a persistent chemical is
presented in a study by Thorn and Agg (1975), which lists 1,3-dichloro-
benzene as a synthetic organic chemical which is unlikely to be removed
during biological sewage treatment, even after prolonged exposure of the
biota.   In contrast, Thorn and Agg (1975) list chlorobenzene, a compound
with one less chlorine atom than 1,3-dichlorobenzene, as a synthetic
organic chemical which should be degradable by biological sewage treatment
provided that suitable acclimatization can be achieved.  Based upon
inference 1,3-dichlorobenzene should be resistant to biodegradation but in
the absence of solid experimental evidence, no definitive conclusions can
be reached or its biodegradability.

74.5  Data Summary

    There is not enough environmentally significant information on pho-
tolysis, oxidation, sorption, or biodegradation processes to be able to
predict the aquatic fate of 1,3-dichlorobenzene.  1,3-Dichlorobenzene has a
high affinity for lipophilic materials and is reported to have a relatively
low vapor pressure and low solubility at temperatures expected to prevail
in most ambient waters.  Consequently, sorption, bioaccumulation, and
volatilization are expected to be competing processes. The rate at which
each of these competing processes occur will dictate which fate is
predominant for 1 , 3-dichlorobenzene in the aquatic environment.   These data
are summarized in Table 74-1.
                                       74-5

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                                                   Table 74-1

                            Summary of Aquatic Fate of 1,3-Dichlorobenzene
Environmental               Summary
   Processa               Statement

Photolysis           No information found.

Oxidation            Dichlorobenzenes in
                     general were reported to
                     be reactive toward hydroxyl
                     radicals (OH-) in air.

Hydrolysis           1,3-Dichlorobenzene will
                     probably not hydrolyze
                     in ambient waters due
                     to the extreme difficulty
                     with which aryl  halides
                     undergo nucleophilic sub-
                     stitution.

Volatilization       This compound probably
                     volatilizes from the
                     water column to  the
                     atmosphere at a  relatively
                     rapid rate.

Sorption             The high log P value found
                     for 1,3-dichlorobenzene in-
                     dicates that sorption  pro-
                     cesses may be substantial
                     for this compound at pollutant
                     concentrations anticipated in
                     environmental waters.
                                                          Rate
                                                                     Half-Life
                                                                        daysn
                                                          Confidence
                                                            of Data
                                                                                     Medium
                                                                     MO  or  more
                                                                      hours0
                                                          Medium
                                                                                     Medium
      Bioaccumulation
                           Empirical evidence indicates
                           that 1,3-dichlorobenzene will
                           bioaccumulate in the lipids of
                           tissues of living organisms.
                                                                                     Medium
      Bio trans formation/
        Biodegradation
Will biodegrade, at best, at
a very slow rate.
                                                                                     Low
a.  There is insufficient  information in  the reviewed  literature  to  permit  assessment
    of a most probable fate.

b.  This half-life is the  reported half-life of  dichlorobenzenes  toward  hydroxyl  radicals in
    air cited in Ware and  West  (1977).

c.  This half-life is based  on  the calculated Henry constant  of 1,3-dichlorobenzene which is of
    the same order as the  Henry constant  predicted  for Aroclor  1248  at 25C by  Mackay  and Leinonen
    (1975).   Since the half-life for-evaporation of Aroclor 1248  from a  water column one  meter thick
    was estimated by Mackay  and Leinonen  to  be 9.53 hours  at  25C, the half-life  for evaporation of
    1,3-dichlorobenzene under  the same  conditions was  assumed to  be  of the  same order, or approxi-
    mately 10 or more hours.
                                                    74-6

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74.6  Literature Cited

Alexander, M. and B.K. Lustigman.  1966.  Effect of chemical structure on
  microbial degradation of substituted benzenes.  J. Agr. Food Chem.
   14(4):410-413.

Chiou, C.T.,  V.H. Freed, D.W. Schmedding, and R.L. Kohnert.  1977.
  Partition coefficient and bioaccunmlation of selected organic chemicals.
  Environ. Sci. Technol.  11(5):475-478.

Garrison, A.W. and D.W. Hill.  1972.  Organic pollutants from mill persist
  in downstream waters.  Am. Dyest. Rep. 21-25.

Kopperman, H.L., D.W. Kuehl, and G.E. Glass.  1976.  Chlorinated compounds
  found in waste treatment effluents and their capacity to bioaccumulate.
  Proceedings of the conference on the environmental impact of water
  chlorination.  Oak Ridge, Tennessee, October 22-24.  (Preprint only).

Leo, A., C. Hansch, and D. Elkins.  1971.  Partition coefficients and their
  uses.  Chem. Rev. 71:525-616.

Lu, P. and R.L. Metcalf.  1975.  Environmental fate and biodegradability of
  benzene derivatives as studied in a model aquatic ecosystem.   Environ.
 Health Perspect.  10:269-284.

Mackay, D. and A.W. Wolkoff.  1973.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. Technol.
  7(7):611-614.

Mackay, D. and P.J. Leinonen.  1975.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. Technol.
  9(13):1178-1180.

Morrison, R.T. and R.N. Boyd.  1973.  Organic chemistry.  3rd Edition.
  Allyn and Bacon, Inc., Boston.   1258p.

Neely, W.B.,  D.R. Branson, and G.E. Blau.  1974.  Partition coefficient to
  measure bioconcentration potential of organic chemicals in fish.
  Environ. Sci. Technol. 8:1113-1115.

Patai, S.  (ed).  1973.  The chemistry of the carbon-halogen bond:  Part 2.
  John Wiley Interscience, New York.  1215p.

Thorn, N.S. and A.R. Agg.  1975.  The breakdown of synthetic organic
  compounds in biological processes.  Proc. Roy. Soc. Lond. B 189:347-357.
                                     74-7

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U.S. Environmental Protection Agency.  1975.  Preliminary assessment of
  suspected carcinogens in drinking water.  U.S. Environmental Protection
  Agency, Office of Toxic Substances, Washington, B.C.  33p.  EPA
  560/4-75-003.

Verschueren, K.  1977.  Handbook of environmental data on organic
  chemicals.  Van Nostrand/Reinhold Press, New York.  659p.

Ware, S.A. and W.L. West.  1977.  Investigation of selected potential
  environmental contaminants:  halogenated benzenes.  U.S. Environmental
  Protection Agency, Office of Toxic Substances, Washington, D.C.  283p.
  EPA 560/2-77-004.

Weast, R.C. (ed).  1973.  Handbook of chemistry and physics.  54th Edition.
  CRC Press, Inc., Cleveland, Ohio.  2452p.

Weast, R.C. (ed).  1977.  Handbook of chemistry and physics.  58th Edition.
  CRC Press, Inc., Cleveland, Ohio.  2398p.
                                     74-8

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               75.  1,4-DICHLQRQBENZENE (_p_-DICHLOROBENZENE )


75.1  Statement of Probable Fate

    Based on the information found, it is not possible to determine the
predominant aquatic fate of 1,4-dichlorobenzene.  There is some evidence
that dichlorobenzenes in general are reactive toward hydroxyl radicals in
air with a half-life of approximately three days. Products and further de-
tails of such photooxidation reactions, however, were not indicated.  In-
formation concerning the biodegradation potential of 1,4-dichlorobenzene
indicates that this compound is very persistent and will probably, at best,
be biodegraded very slowly by microorganisms already growing on another
hydrocarbon source.

    1,4-Dichlorobenzene has a high affinity for lipophilic materials, have
a relatively low aqueous solubility and low vapor pressure at ambient
temperatures.  Consequently, sorption, bioaccumulation,  and volatilization
are expected to be competing processes. The rate at which each of these
competing processes occur will determine which fate is predominant for
1,4-dichlorobenzene in  the aquatic environment.  Should volatilization
occur at a more rapid rate than sorption or bioaccumulation, then atmo-
spheric processes would be expected to regulate the fate of 1,4-dichloro-
benzene.  On the other  hand, should sorption and bioaccumulation occur more
rapidly than volatilization, biodegradation of 1,4-dichlorobenzene by
aquatic microorganisms  would be anticipated to regulate  the fate of this
compound.

75.2  Identification

    1,4-Dichlorobenzene has been detected in drinking water, in superchlo-
rinated municipal wastewaters,  in ground water (U.S. Environmental Protec-
tion Agency 1975), in wastewater effluent (Glaze and Henderson 1975), and
in the atmosphere (Ware and West 1977).  The chemical structure of
1,4-dichlorobenzene is  shown below.
                                             Alternate Names

                                             p-Dichlorobenzene
                                             Paradichlorobenzene
              Cl
     1,4-Dichlorobenzene

     CAS NO.  106-46-7
     TSL NO.  CZ 45500
                                      75-1

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75.3  Physical Properties

    The general physical properties of 1,4-dichlorobenzene are given below.

    Molecular weight                        147.01
    (Weast 1977)

    Melting point                            53.1C
    (Weast 1977)

    Boiling point at 760 torr                174C
    (Weast 1977)

    Vapor pressure at 25C                   *

    Solubility in water at 25C              79 mg/1
    (Verschueren 1977)

    Log octanol/water partition              3.39
    coefficient (Leo et. al. 1971;
    Chiou et al. 1977)
*T,he value of the vapor pressure of 1,4-dichlorobenzene at 25C was ob-
tained by interpolation from the values at 20C (0.6 torr) and 30C (1.8
torr) from Gray (1957). The value at 25C was calculated to be 1.18 torr.
For further explanation, see Methods section.

75.4  Summary of Fate Data

    75.4.1  Photolysis

         No specific information pertaining to the direct photolysis of
1,4-dichlorobenzene in the aquatic or atmospheric environments was found.

    75.4.2  Oxidation

         According to Ware and West (1977) 1,4-dichlorobenzene is resistant
to autooxidation by the peroxy radical (RC>2') in water.  No more de-
tails of this phenomenon were reported.

         Dichlorobenzenes in general were reported by Ware and West (1977)
to be reactive toward hydroxyl radicals (OH*) in air with a half-life of
approximately three days.  Products and further details of such photooxida-
tion reactions were not indicated.  1,4-Dichlorobenzene, specifically, was
reported by Ware and West (1977) to be resistant to autooxidacion by ozone
in air.
                                     75-2

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    75.4.3  Hydrolysis

         No specific information pertaining to the hydrolysis of 1,4-di-
chlorobenzene has been found.  Although Ware and West (1977) report that
the inductive electronegative effect of halogen substituents activates the
ring making such compounds more easily attacked by nucleophiles such as
OH~, Morrison and Boyd (1973) report that aryl halides are characterized
by very low reactivity toward nucleophilic reagents such as OH~.  As an
example of the difficulty with which aryl halides undergo nucleophilic sub-
stitution, the conditions necessary for the nucleophilic substitution of
hexachlorobenzene to a pentachlorophenyl derivative were reported by Patai
(1973) to be the presence of aqueous ammonia at a temperature of at least
250C.  On these bases,-1,4-dichlorobenzene, being less chlorinated and,
consequently, less easily attacked by nucleophiles than hexachlorobenzene,
would not be expected to undergo hydrolysis at an appreciable rate under
environmental conditions.

    75.4.4  Volatilization

         Available data on 1,4-dichlorobenzene indicate that this compound
probably volatilizes from the water column to the atmosphere at a rela-
tively rapid rate.  Garrison and Hill (1972) reported that a 300 mg/1
concentration of 1,4-dichlorobenzene volatilized almost completely (less
than 1 mg/1 of 1,4-dichlorobenzene remained) from  aerated distilled water
in less than 4 hours.  The same concentration of 1,4-dichlorobenzene
volatilized almost completely (less than 1 mg/1 of 1,4-dichlorobenzene re-
mained) from unaerated distilled water in less than 3 days.  No further de-
tails of this experiment were reported.  The data of Garrison and Hill
(1972) can be used to calculate approximate values for evaporative half-
lives.  For the aerated solution, the calculated half-life is less than 30
minutes;  however, since the aeration probably caused air-stripping of the
1,4-dichlorobenzene, this value is not recommended as an evaporative half-
life under conditions of agitation.  The data for unaerated conditions,
apparently close to quiescence, correspond to a half-life of less than
about nine hours.

         According to Mackay and Wolkoff (1973) the rate of evaporation of
pollutants having a low solubility in water can be quite rapid even though
these compounds often have a high molecular weight and a low vapor pres-
sure, and should, on these bases, evaporate slowly.  Mackay and Wolkoff
(1973) contend that these compounds often have high activity coefficients
in water which cause unexpectedly high equilibrium partial vapor pressures
and thus high rates of evaporation.  Although 1,4-dichlorobenzene, which
has a moderately low vapor pressure (1.18 torr) and a low solubility in
water of 79 mg/1 at 25C (Verschueren 1977) was not mentioned specifically,
other chlorinated hydrocarbons having a low solubility in water were pre-
dicted to have relatively rapid rates of evaporation.  Presumably, 1,4-
                                     75-3

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dichlorobenzene would have a rate of evaporation similar to those chlor-
inated compounds having vapor pressure and solubility values approximately
equal to those values for 1,4-dichlorobenzene.

         The calculated value for the vapor pressure (Gray 1957) and
published value for the solubility in water (Verschueren 1977) of 1,4-di-
chlorobenzene were used to compute the Henry constant as being approxi-
mately 2.88 x 10~3 atmos. m-Vmole which may be compared to the value of
2.76 x 10~~3 atmos. m^/mole predicted for Aroclor 1254 at 25C (Mackay
and Leinonen 1975).  In the case of Aroclor 1254, the half-life for
evaporation from a water column one meter thick was estimated by Mackay and
Leinonen (1975) to be 10.3 hours at 25C.  An estimate of the corresponding
half-life for evaporation of 1,4-dichlorobenzene under the same conditions
would presumably be of the same order, approximately 11 or more hours.
Mackay and Leinonen (1975) point out that interpretation of the environ-
mental significance of the rate of evaporation of compounds such as 1,4-
dichlorobenzene from environmental waters using values calculated at 1
meter depth is dependent upon the type of environmental situation en-
countered.  In situations where the water body is turbulent with frequent
mixing between the surface layer and the bulk, as in a rapidly flowing
shallow river or during white-capping on a lake or ocean, the rate of
evaporation would be more rapid than for depths greater than 1 meter or in
quiescent water, as exists in a deep, slowly flowing river.

    75.4.5  Sorption

         Although no specific environmental sorption studies were found in
the literature, the value of the log octanol/water partition coefficient
for 1,4-dichlorobenzene (log P = 3.39, Leo _e_t a_l. 1971) indicates that
sorption processes may be substantial for 1,4-dichlorobenzene at pollutant
concentrations anticipated in environmental waters.  Presumably, 1,4-di-
chlorobenzene will be adsorbed by sedimentary organic material; the ex-
tent to which this possible adsorption will interfere with volatilization
has not been considered.

    75.4.6  Bioaccumulation

         Although no experimental evidence of the bioaccumulation potential
of 1,4-dichlorobenzene was found, there  is empirical evidence that 1,4-
dichlorobenzene has a high potential for bioaccumulation in aquatic organ-
isms.  Neely _e_t _al, (1974) and Lu and Metcalf (1975) have shown that the
log octanol/water partition coefficient  (log P) correlates well with the
ability of a compound to accumulate in the lipids of tissues of living
organisms.   Furthermore, the incorporation of chlorine into an organic
molecule increases its lipophilic character resulting in an increased
bioaccumulation potential (Kopperman et  al. 1976).  For comparison, chloro-
benzene, a compound containing only one  chlorine atom, and 1,4-dichloro-
                                    75-4

-------
benzene, a compound having two chlorine atoms, have log P values of 2.84
(Leo _et _al. 1971; Chiou ejt _al. 1977) and 3.39 (Leo _et _al. 1971), respec-
tively.  Since it has been established experimentally that chlorobenzene
bioaccumulates in aquatic organisms (Lu and Metcalf 1975), 1,4-dichloro-
benzene, would be expected to be bioaccumulated by aquatic organisms at
least as much as chlorobenzene.

    75.4.7  Biotransformation and Biodegradation

         According to Ware and West (1977), the more highly halogenated a
compound becomes, the more resistant it is to biodegradation.  Experimental
(Lu and Metcalf 1975) as well as empirical evidence (Leo et al. 1971; Chiou
_e_t jj. 1977) has been found indicating that chlorobenzene is a persistent
chemical and is not readily biodegraded unless the microorganisms present
are already growing on another hydrocarbon source.  Furthermore, Alexander
and Lustigman (1966) found that the presence of a chlorine atom on the ben-
zene ring retarded the rate of biodegradation.  On these bases, 1,4-di-
chlorobenzene, being more highly chlorinated than chlorobenzene, would
presumably biodegrade at least as slowly as chlorobenzene under the same
conditions of exposure to microorganisms.

         Evidence that 1,4-dichlorobenzene is a persistent chemical is
presented in a study by Thorn and Agg (1975) which lists 1,4-dichlorobenzene
as a synthetic organic chemical which is unlikely to be removed during
biological sewage treatment, e-ven after prolonged exposure of the biota.
In contrast, Thorn and Agg (1975) list chlorobenzene, a compound with one
less chlorine atom than 1,4-dichlorobenzene, as a synthetic organic
chemical which should be degradable by biological sewage treatment provided
that suitable acclimatization can be achieved.  Based upon inference
1,4-dichlorobenzene should be resistant to biodegradation but in the
absence of solid experimental evidence no definite conclusions can be
reached on its biodegradability.

75.5  Data Summary

    There is not enough environmentally significant information on photoly-
sis, oxidation, sorption, or biodegradation processes to be able to predict
the fate of 1,4-dichlorobenzene.  1,4-Dichlorobenzene has a high affinity
for lipophilic materials, yet it is reported to have a relatively low vapor
pressure and low solubility at temperatures expected to prevail in most
ambient waters.  Consequently, sorption, bioaccumulation, and volatiliza-
tion are expected to be competing processes. The rate at which each of
these competing processes occur will dictate which fate is predominant for
1,4-dichlorobenzene in the aquatic environment.  These data are summarized
in Table 75-1.
                                    75-5

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-------
75.6  Literature Cited

Alexander, M. and B.K. Lustigman.  1966.  Effect of chemical structure on
  microbial degradation of substituted benzenes.  J. Agr. Food Chem.
  14(4):410-413.

Chiou, C.T., V.H. Freed, D.W. Schmedding, and R.L. Kohnert.  1977.
  Partition coefficient and bioaccumulation of selected organic chemicals.
  Environ. Sci. Technol.  11(5) :475-478.

Garrison, A.W. and D.W. Hill.  1972.  Organic pollutants from mill persist
  in downstream waters.  Am. Dyes. Rep. 21-25.

Glaze, W.H. and J.E. Henderson.  1975.  Formation of organochlorine
  compounds from the chlorination of a municipal secondary effluent.   J.
  Water Poll. Cont. Fed.  47(10) :2511-2515.

Gray, D.E. (ed.).  1957.  American institute of physics handbook.
  McGraw-Hill, New York. 2052p.

Kopperman, H.L., D.W. Kuehl, and G.E. Glass.  1976.  Chlorinated compounds
  found in waste treatment effluents and their capacity to bioaccumulate.
  Proceedings of the conference on the environmental impact of water
  chlorination.  Oak Ridge, Tennessee, October 22-24.  (Preprint only).

Leo, A., C. Hansch, and D. Elkins.  1971.  Partition coefficients and their
  uses.  Chemical Reviews  71:525-616.

Lu, P. and R.L. Metcalf.  1975.  Environmental fate and biodegradability of
  benzene derivatives as studied in a model aquatic ecosystem.   Environ.
  Health Perspect.  10:269-284.

Mackay, D. and A.W. Wolkoff.  1973.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. Technol.
  7(7):1178-1180.
Mackay, D. and P.J. Leinonen.  1975.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. Technol.
  9(13):1178-1180.

Morrison, R.T. and R.N. Boyd.  1973.  Organic chemistry.  3rd Edition.
  Allyn and Bacon, Inc., Boston.  1258p.

Neely, W.B., D.R. Branson, and G.E. Blau.  1974.  Partition coefficient to
  measure bioconcentration potential of organic chemicals in fish.
  Environ. Sci. Technol. 8:1113-1115.
                                    75-7

-------
Patai, S.  (ed).  1973.  The chemistry of the carbon-halogen bond:  part 2.
  John Wiley Interscience, New York.  1215p.

Thorn, N.S. and A.R.  Agg.  1975.  The breakdown of synthetic organic
  compounds in biological processes.  Proc.  Roy. Soc. Lond. B 189:347-357.

U.S. Environmental Protection Agency.  1975.   Preliminary assessment of
  suspected carcinogens in drinking water.  U.S. Environmental Protection
  Agency, Office of Toxic Substances, Washington, D.C.  33p.  EPA
  560/4-75-003.

Verschueren, K.  1977.  Handbook of environmental data on organic
  chemicals.  Van Nostrand/Reinhold Press, New York.  659p.

Ware, S.A. and W.L.  West.  1977.  Investigation of selected potential
  environmental contaminants:  halogenated benzenes.  U.S. Environmental
  Protection Agency, Office of Toxic Substances, Washington, D.C.  283p.
  EPA 560/2-77-004.

Weast, R.C. (ed).  1977.  Handbook of chemistry and physics,.  58th Edition.
  CRC Press Inc., Cleveland, Ohio.  2398p.
                                    75-8

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                        76.   1,2,4-TRICHLOROBENZENE
76.1  Statement of Probable Fate

    1,2,4-Trichlorobenzene has a high affinity for lipophilic materials and
is reported to exhibit a relatively high rate of volatilization from aque-
ous systems.   Consequently, sorption, bioaccumulation,  and volatilization
are anticipated to be competing processes.   The rates at which these com-
peting processes occur will determine which fate is predominant for 1,2,4-
trichlorobenzene in the aquatic environment, but the data found were
insufficient  to determine the most rapid process.   Should volatilization
occur at a more rapid rate than sorption or bioaccumulation,  then atmo-
spheric processes would be expected to regulate the fate of 1,2,4-tri-
chlorobenzene.  On the other hand, should sorption and  bioaccumulation
occur more rapidly than volatilization,  biodegradation  of 1,2,4-tri-
chlorobenzene by aquatic microorganisms  would be anticipated  to regulate
the fate of this compound.

76.2  Identification

    1,2,4-Trichlorobenzene has been detected in drinking water (U.S. En-
vironmental Protection Agency 1975), in  wastewater effluent (Glaze and
Henderson 1975), and in the atmosphere (Ware and West 1977).   The chemical
structure of  1,2,4-trichlorobenzene is shown below.
                                             Alternate Name
                                             unsym-Trichlorobenzene
    1,2,4-Trichlorobenzene

    CAS NO. 120-82-1
    TSL NO. DC 21000

76.3  Physical Properties

    The general physical properties of 1,2,4-trichlorobenzene are as
follows.
    Molecular weight
    (Weast 1977)

    Melting point
    (Weast 1977)
181.45
16.95C
                                     76-1

-------
    Boiling point at 760 torr
    (Weast 1977)

    Vapor pressure at 25C

    Solubility in water at 25C
    (Dow 1978)
213.5C


0.42 torr (calculated)

30 mg/1
    Log octanol/water partition coefficient  4.26
    (Calculated from Tute 1971;  See
    Methods section on Bioaccumulation)

76.4  Summary of Fate Data

    76.4.1  Photolysis

         No specific information pertaining to the direct photolysis of
1,2,4-trichlorobenzene in the aquatic or atmospheric environments was
fo und .

    76.4.2  Oxidation

         No specific information pertaining to the oxidation of 1,2,4-tri-
chlorobenzene in the aquatic environment was found.  1,2,4-Trichloroben-
zene, however, was reported by Simmons et al.  (1976) to  be susceptible to
attack by hydroxyl radicals in the atmosphere.  The rate of this react'ion
is  not known, but the half-life was estimated  by Simmons _e_t _a_l_. (1976) to
be  one to several days.

    76.4.3  Hydrolysis

         No specific information pertaining to the hydrolysis of 1,2,4-tri-
chlorobenzene in ambient waters was found.  Although Ware and West (1977)
report that the inductive electronegative effect of halogen substitutents
activates the ring making such compounds more easily attacked by nucle-
ophiles such as OH~, Morrison and Boyd (1973)  report that aryl halides
will undergo nucleophilic substitution only with extreme difficulty.  For
example, Patai (1973) reported that the minimum conditions necessary for
the nucleophilic subtitution of hexachlorobenzene to a pentachlorophenyl
derivative were the presence of aqueous ammonia and a temperature of 250C.
On these bases, 1,2,4-trichlorobenzene, being less chlorinated and, con-
sequently, less easily attacked by nucleophiles than hexachlorobenzene,
would not be expected to undergo hydrolysis at an appreciable rate under
environmental conditions.

    76.4.4  Volatilization

         Available data on 1,2,4-trichlorobenzene indicate that this com-
pound probably volatilizes from the water column to the atmosphere at a
                                       76-2

-------
relatively rapid rate.  Ware and West (1977) cited information in which the
half-life for evaporation of 1,2,4-trichlorobenzene from water is 45
minutes at standard temperature and pressure.  Garrison and Hill (1972) re-
ported that at 100 mg/1 1,2,4-trichlorobenzene volatilized almost
completely (less than 1 mg/1 of 1,2,4-trichlorobenzene remained) from
aerated distilled water in less than 4 hours.  The same concentration of
1,2,4-trichlorobenzene volatilized almost completely (less than 1 mg/1 of
1,2,4-trichlorobenzene remained) from unaerated distilled water in less
than 2 days.   No further details of this experiment were reported.

         Assuming that volatilization as reported by Garrison and Hill
(1972) is a first order process, then the evaporative half-lives corres-
ponding to the above data are 36 minutes for the aerated condition and 7.2
hours for the unaerated (quiescent) condition.  The evaporation half-life
of 45 minutes cited by Ware and West (1977) is very close to the 36 minutes
resulting from aeration;  aeration is assumed to have involved some gas
stripping of  the 1,2,4-trichlorobenzene from the solution.

         In a separate experiment by Garrison and Hill (1972) an aqueous
solution of 1,2,4-trichlorobenzene (50 mg/1) was exposed to mixed cultures
of aerobic microorganisms and aerated.  Based on the graphical
representation of the data by Garrison and Hill (1972) and assuming a
negligible rate of biodegradation in comparison to volatilization (no
degradation products were found), the evaporative half-life for the
1,2,4-trichlorobenzene in this system appears to be on the order of 4 to 5
hours.  A measurable concentration of the compound was detected after nine
days of exposure.  In this experiment, the half-life with respect to
volatilization is reasonably close to the 7.2 hours cited above for the
quiescent condition in distilled water.  It is believed that this result
illustrates the effect of partitioning of 1,2,4-trichlorobenzene between
the water and suspended biological material.

         In summary, the rate of volatilization of 1,2,4-trichlorobenzene
from distilled water is relatively high, but, when suspended organic
materials are present, the partitioning between water and" suspended solids
can greatly reduce the volatilization rate.

    76.4.5  Sorption

         Although no specific environmental sorption studies were found in
the literature, the calculated value of the log octanol/water partition
coefficient found for 1,2,4-trichlorobenzene (log P = 4.26, Tute 1971)
indicates that sorption processes, may be substantial for this compound at
pollutant concentrations anticipated in environmental waters.  Presumably,
1,2,4-trichlorobenzene will be adsorbed by sedimentary organic material;
the extent to which this possible adsorption will interfere with
volatilization has not been studied.
                                    76-3

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    76.4.6  Bioaccumulation

         Macek _e_t _al. (1977) investigated the relative significance of
aqueous and dietary exposure of bluegill sunfish (Lepomis macrochirus) to
1,2,4-trichlorobenzene using an aquatic food chain which consisted of a
food organism, the water flea,  Daphnia magna, and a consumer organism, the
bluegill sunfish, Lepomis macrochirus.  Prior to initiation of the food
chain experiment, the duration of exposure required to "load" daphnids and
fish to an equilibrium concentration of radiolabelled 1,2,4-trichloroben-
zene was determined.  Two flow-through experimental units (A,B) were dosed
to maintain continuous aqueous exposure of bluegill sunfish to a concentra-
tion of 3.0 ug/1 of 1,2,4-trichlorobenzene.  Two other flow-through
experimental units (C,D) contained no aqueous concentrations of 1,2,4-tri-
chlorobenzene.  Fish in units B and C were fed daphnids allowed to reach
equilibrium concentrations of 1,2,4-trichlorobenzene, while fish in A and D
were fed uncontaminated daphnids.  As a result, fish in unit A experienced
only aqueous exposure to the chemical, fish in unit B experienced both
dietary and aqueous exposure, fish in unit C experienced only dietary
exposure, and fish in unit D experienced no exposure and served as a con-
trol group.  Quantitative measures of the accumulation of -^C-residues of
1,2,4-trichlorobenzene in bluegill sunfish were determined by removing five
fish from each experimental unit on day 1,3,7,10,14 and each succeeding 7
days during exposure and radiometrically analyzing each individual fish.

         Daphnids continuously exposed to a mean measured aqueous concen-
tration of 3.1 + 0.4 ug/1 of 1,2,4-trichlorobenzene had a mean measured
equilibrium 14-C residue body burden of 0.44 j- 0.16 mg/1.  Therefore, the
estimated equilibrium bioconcentration factor for daphnids is 142X.  Fish
continuously exposed to a mean measured aqueous concentration of 2.9 + 1.1
ug/1 of 1,2,4-trichlorobenzene had a mean measured equilibrium 14-C residue
body burden of 0.53 + 0.25 mg/1.  Consequently, the estimated equilibrium
bioconcentration factor for fish is 182X.  Macek _e_t al. (1977) define bio-
concentration as that process wher.eby chemical substances enter aquatic
organisms through gills or epithelial tissues directly from water.  Bio-
accumulation, on the other hand, is defined by Macek _e_t _al. (1977) as a
broader term referring to a process which includes bioconcentration as well
as any uptake of chemical residues from dietary sources.  The data from
investigating the bioaccumulation of 14-C 1,2,4-trichlorobenzene by blue-
gill sunfish from food and water yielded a mean equilibrium 14-C residue
body burden in sunfish of 0.57 +_ 0.5 mg/1 when exposed to 2.9 ug/1 of
1,2,4-trichlorobenzene and fed daphnids having a mean concentration of 0.44
mg/1 of 14-C residue.

         The drawback of the experiments of Macek et al. 1977 is that
either results are based upon total c-^-carbon analysis and hence do not
take into account metabolic by-products and bound residues.  If biodegrada-
tion does not occur, their bioconcentration factors will be maximal.  Under
the conditions of the experiment, however, bioconcentration factors very
likely will be lower.
                                    76-4

-------
          Macek et al. (1977) concluded that the contribution of the
dietary source of 1,2,4-trichlorobenzene to the ultimate equilibrium body
burden is probably insignificant compared to the aqueous source of this
chemical even if the  two sources are additive.   The lack of significant
contribution of dietary 1,2,4-trichlorobenzene  to the equilibrium 14-C
residue body burden of bluegill sunfish was confirmed by results of a study
where bluegill sunfish were held in uncontaminated water but fed daphnids
containing 0.44 mg/1  of 14-C residues as 1,2,4-trichlorobenzene.  The mean
14-C residue body burden in bluegill sunfish during dietary exposure to
approximately 3 yg/1  of 1,2,4-trichlorobenzene  was 0.03 + 0.01 mg/1.  This
value represents only about 5 percent of the 1,2,4-trichlorobenzene 14-C
residue body burden in bluegill sunfish exposed for a comparable period to
either aqueous or combined  aqueous and dietary  1,2,4-trichlorobenzene.
Macek et_ al. (1977) conclude that the percent of residue burden due to
dietary intake alone  of 1,2,4-trichlorobenzene  by bluegill sunfish fed
contaminated daphnids is less than the error associated with the ability to
estimate residues due to bioconcentration alone.

         Neely _et al. (1974) and Lu and Metcalf (1975) have shown that the
log octanol/water partition coefficient (log P) correlates well with the
ability of a compound to accumulate in the lipids of tissues of living
organisms.  Furthermore, the incorporation of chlorine into an organic
molecule increases its lipophilic character resulting in an increased po-
tential for bioaccumulation (Kopperman _et _al. 1976).  The rather high log P
value of 4.26 as calculated from the method of  Tute (1971) indicates that
the bioaccumulation potential of 1,2,4-trichlorobenzene by aquatic organ-
isms at pollutant concentrations anticipated in environmental waters would
probably be relatively high.  Based upon the minimal available evidence,
however, 1,2,4-trichlorobenzene appears not to  be bioaccumulated as ex-
tensively as chlorobenzene.  Whether this is a  function of insufficient
sampling or an exception to the connection between log P and bioaccumula-
tion cannot yet be determined.

    76.4.7  Biotransformation and Biodegradation

         Experimental evidence indicates that chlorobenzene, which is
structurally similar to 1,2,4-trichlorobenzene  but less highly chlorinated,
is not readily biodegraded unless the microorganisms present are already
utilizing another carbon source (Gibson et_ _al.  1968).  According to Ware
and West (1977), resistance to biodegradation generally increases as the
degree of halogenation of a compound increases, and Alexander and Lustigman
(1966) found that the presence of a chlorine atom on the benzene ring re-
tarded the rate of biodegradation.  From the above, it can be inferred that
1,2,4-trichlorobenzene, being more highly chlorinated than chlorobenzene,
would presumably be biodegraded less rapidly than chlorobenzene under the
same conditions of exposure to microorganisms.
                                       76-5

-------
          Ware and West (1977) report an experiment in which 1,2,4-tri-
chlorobenzene at a concentration of 50 mg/1 was exposed to preconditioned
treatment plant cultures.   1,2,4-Trichlorobenzene was very persistent as
evidenced by the slight decrease in concentration probably due primarily to
atmospheric loss after a 50-hour period. The microorganisms absorbed
approximately 40 percent of 1,2,4-trichlorobenzene but did not substan-
tially biodegrade the absorbed portion.

         Ware ana West (1977)  also report another experiment in which the
biodegradation of 1,2,4-trichlorobenzene was studied in order to assess the
extent and rate of degradation and the role of acclimatization.  When un-
acclimatized industrial wastewater microorganisms were used, 99 percent of
the 1,2,4-trichlorobenzene at  an initial concentration of 1.7 mg/1 dis-
appeared from the experimental mixture after 10 days.  The BOD (Biochemical
Oxygen Demand) value  after 10  days, however, indicated that only 55 percent
of the theoretical value was removed.  The 45 percent theoretical oxygen
demand remaining was  assumed to be due to incompletely oxidized metabolites
of 1,2,4-trichlorobenzene which were probably incorporated in the cell
wall.  The experiments found no apparent degradation as measured by BOD in
the first few days of monitoring.  Analysis, however, indicated a 14 per-
cent reduction in the concentration of 1,2,4-trichlorobenzene after 24
hours, a 36 percent reduction  at 72 hours, and a 43 percent reduction at 7
days.  At a concentration of 2.6 mg/1 the rate of degradation was appar-
ently lower for the first few days.  After seven days of exposure of the
microorganisms to a concentration of 2.6 mg/1 of 1,2,4-trichlorobenzene,
analysis indicated 28 percent  decrease in the concentration of 1,2,4-tri-
chlorobenzene.  This  is quite  different from the 43 percent decrease of
1,2,4-trichlorobenzene after seven days of exposure of the microorganisms
to a concentration of 1.7 mg/1 of this compound.  After ten days, however,
all of the 1,2,4-trichlorobenzene at the higher concentration had
disappeared from the  BOD bottles.

         Ware and West (1977)  reported another phase of the previous ex-
periment in which the formation rate of carbon-14 labeled C02 through
biodegradation of 1,2,4-trichlorobenzene by activated sludge was examined.
After 5 days, 13 percent of the  1,2,4-trichlorobenzene remained, while 56
percent was converted to carbon dioxide, 23 percent to polar metabolites,
and 7 percent was volatilized.  Approximately 80 percent of the 1,2,4-tri-
chlorobenzene was adsorbed on solids, accounting for the low volatility
from the system.  In summary,  biodegradation of 1,2,4-trichlorobenzene has
been demonstrated under conditions where microorganisms are acclimated and
another carbon source is utilized, as in the activated sludge process.
Under environmental conditions, biodegradation would be expected but at a
very much lower rate than in the waste treatment studies reported above.
                                       76-6

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76.5  Data Summary

    Information found concerning photolysis, oxidation, or sorption pro-
cesses was insufficient to allow prediction of the aquatic fate of
1,2,4-trichlorobenzene.  1,2,4-Trichlorobenzene has a high affinity for
lipophilic materials and is  predicted to have a relatively low vapor
pressure and low solubility  at ambient temperatures.  Consequently,  sorp-
tion and bioaccumulation are expected to be competing with volatilization
in removing the compound from the water column. The rate at which each of
these competing processes occur will dictate which fate is predominant for
1,2,4-trichlorobenzene in ambient waters.  These data are summarized in
Table 76-1.
                                       76-7

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76.6  Literature Cited

Alexander, M. and B.K. Lustigman.  1966.  Effect of chemical structure on
  microbial degradation of substituted benzenes.  J. Agr. Food Chem.
  14(4):410-413.

Dow Chemicals U.S.A.  1978.  Technical data bulletin for 1,2,4-di-
  chlorobenzene organic chemicals development.  Midland, Michigan

Garrison, A.W. and D.W. Hill.  1972.  Organic pollutants from mill persist
  in downstream waters.  Amer. Dyest. Rep.    :21-25.

Gibson, D.T., J.R. Koch, C.L. Schlud, and R.E. Kallio.  1968.  Oxidative
  degradation of aromatic hydrocarbons by microorganisms.  II.
  Metabolism of halogenated aromatic hydrocarbons.  Biochemistry
  7(11):3795-3802.

Glaze, W.H. and J.E. Henderson.  1975.  Formation of organochlorine
  compounds  from the chlorination of a municipal secondary effluent.   J.
  Wat. Poll. Cont. Fed.  47(10):2511-2515.

Kopperman, H.L., D.W. Kuehl, and G.E. Glass.  1976.  Chlorinated compounds
  found in waste treatment effluents and their capacity to bioaccumulate.
  Proceedings of the conference on the environmental impact of water
  chlorination.  Oak Ridge, Tennessee, October 22-24.  (Preprint only).

Lu, P. and R.L. Metcalf.  1975.  Environmental fate and biodegradability of
  benzene derivatives as studied in a model aquatic ecosystem.   Environ.
  Health Perspect.  10:269-284.

Macek, K.J., S.R. Petrocelli, and B.H. Sleight, III.  1977.  Considerations
  in assessing the potential for, and significance of biomagnification of
  chemical residues in aquatic food chains.  Presented at:  ASTM Second
  Symposium on Aquatic Toxicology, October 31-November 1, Cleveland, Ohio.

Morrison, R.T. and R.N. Boyd.  1973.  Organic chemistry.  3rd Edition.
  Allyn and Bacon, Inc. , Boston.   1258p.

Neely, W.B., D.R. Branson, and G.E. Blau.  1974.  Partition coefficient  to
  measure bioconcentration potential of organic chemicals in fish.
  Environ. Sci. Technol. 8:1113-1115.

Patai, S.  (ed).  1973.  The chemistry of the carbon-halogen bond:  part 2.
  John Wiley Interscience, New York.  1215p.
                                     76-9

-------
Simmons, P., D. Branson and R.  Bailey.   1976.   1,2,4-Trichlorobenzene:
  biodegradable or not?  Canadian Association of Textile Colorists and
  Chemists, Intern. Technical Conf., Quebec.   (Preprint only).

Tute, M.S.  1971.   Principles and practice of Hansch analysis:  a guide
  to structure-activity correlation for the medicinal chemist.   Adv.
  Drug Res.  6:1-77.

U.S. Environmental Protection Agency.  1975.   Preliminary assessment of
  suspected carcinogens in drinking water.  U.S. Environmental Protection
  Agency, Office of Toxic Substances, Washington, D.C.  33p. (EPA
  560/4-75-003).

Ware, S.A. and W.L. West.  1977.  Investigation of selected potential
  environmental contaminants:  halogenated benzenes.  U.S. Environmental
  Protection Agency, Office of Toxic Substances, Washington, D.C.  283p.
  (EPA 560/2-77-004).

Weast, R.C. (ed).   1977.  Handbook of chemistry and physics.  58th Edition.
  CRC Press Inc.,  Cleveland, Ohio.  2398p.
                                      76-10

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                          77.  HEXACHLOROBENZENE
77.1  Statement of Probable Fate

    Based on the information found, it appears that hexachlorobenzene is a
very persistent compound.  None of the destructive processes studied, which
include photolysis, oxidation, hydrolysis, and biodegradation, appear to
exert an appreciable effect on the fate of hexachlorobenzene in the aquatic
environment.  Some investigators believe that naturally occurring complex
organic compounds present in rivers and streams may serve as photosensiti-
zers and thus enhance the degradation of organic pollutants, such as hexa-
chlorobenzene, by sunlight.  The photolysis of hexachlorobenzene in the
presence of such possible photosensitizers, however, has not been studied.
The relative volatility of hexachlorobenzene is not known.  Should hexa-
chlorobenzene prove to be volatile and enter into the upper atmosphere,
there is a possibility that short wavelength light may eventually convert
this pollutant into other compounds.  No evidence that this does occur,
however, has been found.

    Hexachlorobenzene has a high affinity for lipophilic materials.  Con-
sequently, sorption and bioaccumulation are anticipated to occur quite
readily.  It appears that the majority of hexachlorobenzene found in
aquatic organisms is from aqueous rather than dietary sources.  Further-
more, rates of depuration of hexachlorobenzene from exposed aquatic or-
ganisms are substantially more rapid than rates of depuration reported for
such highly persistent compounds as DDT.  Therefore, biomagnification of
hexachlorobenzene through aquatic food chains probably does not occur.

77.2  Identification

    Hexachlorobenzene has been found in water and sediment samples, and in
aquatic biota from a known site of contamination in the lower Mississippi
River (Laseter et_ _al. 1976;  Laska &t_ _a_l. 1976), in many species of fish
collected throughout the United States (Johnson _et_ &L_. 1974), in drinking
water supplies and finished drinking water of several cities in the United
States (U.S. Environmental Protection Agency 1975), in food products and in
samples of human fatty tissue from individuals in the population of Italy
(Leoni and D'Arca 1976), and as an impurity in several agricultural
pesticides (Leoni and D'Arca 1976).
                                      77-]

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    The chemical structure of hexachlorobenzene is shown below.
                                             Alternate Names

                                             Perchlorobenzene
    Hexachlorobenzene                        RGB

    CAS NO. 118-74-1
    TSL NO. DA 29750

77.3  Physical Properties

    The general physical properties of hexachlorobenzene are given below.

    Molecular weight                         284.79
    (Weast 1977)

    Melting point                            230C
    (Weast 1977)

    .Boiling point                            322C
    (Weast 1977)

    Vapor pressure at 20C                   1.089 x 10~5 torr
    (Leoni and D'Arca 1976;   Isenee _et: al.
    1976)

    Solubility in water                      *

    Log octanol/water partition coefficient  6.18
    (Neely et al. 1974)
^According to Laseter et_ a.1. (1976) the upper limit of solubility of hexa-
chlorobenzene in water is 20yg/l.  Although no temperature or pH was given
at which this upper limit of solubility was determined, it is assumed that
these parameters would be within the limits found in environmental waters.
Metcalf _et _al. (1973) measured the solubility of HCB at 25C by radioassay
and obtained a value of 6 pg/1.

77.4  Summary of Fate Data

    77.4.1  Photolysis

         Plimmer and Klingebiel (1976) investigated photodecomposition as a
possible route for the environmental degradation of hexachlorobenzene.  In
                                  77-2

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one experiment, a layer of crystalline hexachlorobenzene was placed on a
glass plate under a quartz cover and exposed for five months to a sunlamp
or to ambient laboratory illumination.  In another experiment, a hexane  .
solution of hexachlorobenzene was spotted on silica gel coated thin layer
chromatography plates which were subsequently exposed for 4.5 hours to out-
door sunlight, to a 40-W sunlamp (maximum wavelength about 310 nm), or to
laboratory illumination (fluorescent lighting).  These experiments indi-
cated that photodecomposition of hexachlorobenzene was extremely slow. No
photodecomposition products were identified.  According to Plimmer and
Klingebiel (1976), this photodecomposition is unlikely to be sensitized by
triplet-state energy transfer since two triplet sensitizers of similar tri-
plet state energies, diphenylamine and benzophenone, did and did not sen-
sitize hexachlorobenzene photolysis, respectively.  It was suggested that
other related amines might function in a manner analogous to diphenyl-
amine through a charge-transfer mechanism.  From these studies, Plimmer and
Klingebiel (1976) conclude that photodecomposition of hexachlorobenzene in
the solid state is expected to be slow.  They indicate, however, that com-
plex naturally occurring organic compounds present in rivers and streams
(such as humic acids) may serve as photosensitizers and thus enhance the
degradation of organic pollutants by sunlight.  The photolysis of hexa-
chlorobenzene in the presence of such possible photosensitizers has not
been studied.

         Laseter _e_t _al. (1976) exposed solutions of hexachlorobenzene in
both hexane and benzene to irradiation for periods of 30, 65, and 120
minutes.  The "Laboratory Methods and Materials" section of their report
states that irradiations were conducted at a wavelength of 253.7 nm in
capped quartz test tubes, but the "Results" section of the paper states
that irradiations were conducted at a wavelength of 273.5 nm.  Less than 10
percent of the original hexachlorobenzene was reported to remain after 60
minutes of irradiation in benzene.  Concurrently, a gradual increase of
lower molecular weight products commenced after 30 minutes of irradiation.
Based on the fact that both of the aforementioned wavelengths are below the
atmospheric cutoff, it appears that direct photolysis would probably not be
an important fate process for hexachlorobenzene in the aquatic environment.

    77.4.2  Oxidation

         The stability of hexachlorobenzene is such that it is resistant to
oxidation except under the most extreme conditions (Fieser and Fieser
1956).  Consequently, oxidation of hexachlorobenzene would not be expected
to be an important fate under ambient conditions.

    77.4.3  Hydrolysis

         Although Ware and West (1977) report that the electronegative
effect of aromatic halogen substituents activates the benzene ring and
                                      77-3

-------
renders it more easily attacked by nucleophiles such as OH", Morrison and
Boyd (1973) report that aryl halides will undergo nucleophilic substitution
only with extreme difficulty.  For example, Patai (1973) reported that the
minimum conditions necessary for the nucleophilic substitution of hexa-
chlorobenzene to a pentachlorophenyl derivative were the presence of aque-
ous ammonia and a temperature of 250C.   Furthermore, Leoni and D'Arca
(1976) report that hexachlorobenzene is chemically very inert at room tem-
perature and reacts with caustic alkalis to form the corresponding pen-
tachlorophenolates only at 130-200C.  Thus, hexachlorobenzene would not
be expected to undergo hydrolysis at an appreciable rate under environ-
mental conditions.

    77.4.4  Volatilization

         Because of its relatively high boiling point and correspondingly
low vapor pressure at ambient temperature, hexachlorobenzene might be ex-
pected to exhibit very slow rates of volatilization from water.  Mackay and
Wolkoff (1973), however, contend that sparingly soluble organic compounds
often have high activity coefficients in aqueous solution which cause un-
expectedly high equilibrium partial vapor pressures and high rates of evap-
oration.  Hexachlorobenzene, with an aqueous solubility on the order of 10
Ug/1 (Laseter e_t_ aJL. 1976;  Metcalf ^t_ a_l. 1973) may be such a compound,
but experimental information on this point was not found.  Calculations
according to Mackay and Leinonen (1975) predict a rate constant of 0.08/min
or a tj/2 of - 8 hours for evaporation from a water column 1 m deep.
Although such an estimate does not include environmentally important
parameters such as adsorption by sediment and variable mixing rates, it
does suggest that transport by volatilization might be important in the
absence of other processes.

    77.4.5  Sorption

         Laseter e_t_ a_l. (1976) conducted an experiment in which a soil
sample was exposed to a regular flow of water with a concentration of 8.3
Mg/1 hexachlorobenzene.  After one day of exposure to this concentration,
the soil sample had a concentration of 332 Mg/1 of hexachlorobenzene, a con-
centration factor of 40X.  Four days after initiation of the test, the con-
centration of hexachlorobenzene in the soil was 269 yg/1, a concentration
factor of 32X.  Depuration was initiated on the fourth day of exposure, and
after four days of depuration the sample of sediment had a concentration of
303 ug/1 of hexachlorobenzene, a concentration almost as high as that
measured on the first day of exposure.  Laseter e_t_ ajL. (1976) concluded
that bottom sediment accumulated proportionately less hexachlorobenzene
than did organisms, but retained it longer.

         Laska e_t_ aJ. (1976) studied the distribution of hexachlorobenzene
in the vicinity of an industrialized region bordering the Mississippi River
between Baton Rouge and New Orleans, Louisiana.  The mean concentration of
                                     77-4

-------
hexachlorobenzene in water from the Mississippi River near Baton Bouge was
2.2 yg/1 as compared to a concentration of 167.0 yg/1 (dry weight) in the
soil of the levee adjacent to the river.  Laska _e_t al. (1976) attributed
the high levels of hexachlorobenzene in the soil on the river side of the
levee to accumulation of hexachlorobenzene from the load carried in solu-
tion and suspension in river water.

    77.4.6  Bioaccumulation

         The bioaccumulation potential of hexachlorobenzene has been
studied using radiotracer techniques in three model aquatic ecosystems
(Metcalf _e_t _al. 1973;  Lu and Melcalf 1975;  Isensee _ejt _al. 1976).  De-
scriptions and pertinent results for all three ecosystems are presented be-
low.  In all studies, hexachlorobenzene was found to be a highly persist-
ent compound as demonstrated by the ecological magnification values (EM)
and bioaccumulation ratios (BR; equivalent to EM).

         Metcalf _et _al. (1973) and Lu and Metcalf (1975) define ecological
magnification as the ratio of the parent compound in an organism to the
concentration in the water;  the term bioaccumulation ratio used by Isensee
_e_t al. (1976) has the same meaning as ecological magnification.  From this
definition and from the three experimental designs, it is unclear as to
whether ecological magnification and bioaccumulation ratio are measures of
bioconcentration or measures of bioaccumulation.  According to Macek et al.
(1977), bioconcentration is defined as that process whereby chemical sub-
stances enter aquatic organisms through gills or epithelial tissue directly
from water.  Bioaccumulation, on the other hand, is defined by Macek et al.
(1977) as a broader term referring to a process which includes bioconcen-
tration as well as any uptake of chemical residues from dietary souces.

         Since the relative contributions of aqueous and dietary sources of
hexachlorobenzene were not quantitatively evaluated at equilibrium condi-
tions, it is not possible to determine whether ecological magnification and
bioaccumulation ratio are measures of bioaccumulation or bioconcentration.
It is, in fact, very possible that the body burden due to dietary and
aqueous exposure of hexachlorobenzene would be statistically indistinguish-
able from the body burden due to aqueous exposure alone of hexachloroben-
zene.  In a critical analysis of selected chemicals by Macek _e_t _al. (1977)
which included Kepone, PCB's, endrin, cadmium, DDT, and 1,2,4-trichloroben-
zene, the equilibrium body burden due to dietary exposure of these chemi-
cals was statistically indistinguishable from the equilibrium body burden
due  to aqueous exposure except in the case of DDT.

         Reported values of ecological magnification or bioaccumulation
ratio for the various species studied in each model ecosystem are presented
in Table 77-1.  As can be seen, the agreement between results for a given
species varies from poor to excellent;  in general, the ratios reported by
Metcalf _e_t _a_l. (1973) were somewhat lower than those from the other two
sets of studies, but the reason for the variation is unclear.
                                       77-5

-------
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                         77-6

-------
         The model aquatic ecosystem of Lu and Metcalf (1975) was devised
for studying relatively volatile organic compounds and simulating direct
discharge of chemical wastes.  A 3-liter flask which contained members of
an aquatic food chain, including daphnia, mosquito larvae, snails, mosquito
fish, and green filamentous algae, was maintained at 80F with 12 hours
daylight exposure.  Radiolabeled pollutants were added to the flask in con-
centrations of 0.01 to 0.1 mg/1.  After 48 hours, the experiment was ter-
minated and samples were taken and analyzed.  The data for hexachloroben-
zene are shown in Table 77-1.

         From a different set of experiments Metcalf _e_t al. (1973) esti-
mated the comparative environmental properties of DDT, methoxyclor, and
other DDT analogs including hexachlorobenzene.  A small glass aquarium that
had a sloping terrestrial-aquatic interface of pure sand was used to carry
out the aquatic model ecosystem evaluation.  A 5-mg portion (equivalent to
1.1 kg/he) of radiolabeled pesticide was applied to sorghum seedlings grown
on the terrestrial portion after which fourth instar salt marsh cater-
pillars were fed on the leaves until these were consumed.  Thus, the fecal
products of the larvae and the larvae themselves contaminated the aquatic
portion of the system.  The radiolabeled products were transferred through
several trophic levels in the aquatic food chain, e.g., alga (Oedogonium
cardiacum)     snail (Physa); plankton	^-water flea (Daphnia magna)    
mosquito (Culex pipiens)     fish (Gambusia affinis).  After 33 days in
this environmental plant growth chamber at 80F with a 12-hour photoperiod,
the experiment was terminated and samples taken and analyzed, with the
results for hexachlorobenzene as shown on Table 77-1.

         Isenee _e_t _al. (1976) developed their model aquatic ecosystem
specifically to determine the bioaccumulation potential of hexachloro-
benzene.  In this study, three replicates each of control soil and soils
treated with hexachlorobenzene at concentrations of 0.1, 1, and 10 mg/1
were placed in tanks to which was added aqueous solutions with concentra-
tions of 0, 10, 100, and 1000 yg/1 of hexachlorobenzene, respectively.
Twenty-four hours later, approximately 100 daphnids (Daphnia magna), eight
snails (Helisoma sp.), a few strands of an alga (Oedogonium cardiacum), and
10 ml of old aquarium water which contained various diatoms, protozoa, and
rotifers were added.  Water lost by evaporation was replenished as needed.
At 30 days, daphnids were sampled for analysis (20 to 30 mixed age organ-
isms per sample) and two 0.15 - 0.25g mosquito fish (Gambusia affinis) were
added.  Three days later all organisms were harvested and two 2.0 - 2.5g
fingerling channel catfish (Ictalurus punctatus) were added to each tank
and exposed for 8 days.

         Based on their results, Isensee _et al_, (1976) concluded that
hexachlorobenzene is very persistent in aquatic systems.  They also found
that the amounts of hexachlorobenzene accumulated by all studied aquatic
                                     77-7

-------
model food chain organisms increased as the treatment concentrations
increased.  In addition, these researchers found that, for any given
treatment concentration, higher food chain organisms (such as snails and
mosquito fish) always contained 1.5 to 2 times more hexachlorobenzene than
lower food chain organisms such as algae and daphnids.  Furthermore, cat-
fish, being the highest food chain organism in the model aquatic ecosystem,
accumulated 10 times more hexachlorobenzene than did any other organisms.
Isensee j|t. _al. (1976) theorize that either biomagnification within the food
chain and/or a species -specific response was important in contributing to
the aforementioned accumulation patterns.

         Due to the experimental design used by Isensee _e_t _al (1976) it is
difficult to assess whether biomagnification through food chains was, in
fact, occurring, since there is no evidence that the bioaccumulation ratios
measured were quantitatively evaluated at equilibrium concentrations.

         In a study by Laseter t JL. (1976), an experiment was carried out
for a duration of one week to determine the difference in uptake of hexa-
chlorobenzene between sunfish (Lepomis macrochirus) held in water free of
hexachlorobenzene 16 ug/g and fed sailfin mollies (Poecilia latipinna)
contaminated with hexachlorobenzene and sunfish held in water having a
concentration of 2.7 yg/1 of hexachlorobenzene and fed sailfin mollies
contaminated with hexachlorobenzene 16 yg/g.  The mean concentration of
hexachlorohenzene in the two groups of tish at the end of the experiment
was 594 and 3578 yg/1 hexachlorobenzene, respectively.  From these results,
it appeared that aqueous sources contributed far more to the sunfish body
burden of hexachlorobenzene than did dietary sources.

         Other experiments conducted by Laseter _e_t _aJ. (1976) on uptake and
depuration of hexachlorobenzene by various aquatic organisms showed that
this chemical was depurated quite rapidly.  For example, after 15 days ex-
posure of fingerling bass (6 to 10 cm in length) to concentrations of 2 to
10 yg/1 hexachlorobenzene, only 8.6 to 26.9% (depending upon initial ex-
posure concentration) of the residue remained after exposure to water free
of hexachlorobenzene for 13 days.  Laseter j^t a_l. (1976) note that fish
initially exposed to lower levels of hexachlorobenzene had succeeded in
eliminating a greater proportion of the substance than had those exposed to
higher levels.

         The findings of Laseter _e_t _al. (1976) support the contention that
hexachlorobenzene bioaccumulates but does not biomagnify in aquatic food
chains.  It must be noted, however, that there is no evidence that the rel-
ative contributions of aqueous and dietary sources of hexachlorobenzene
were measured at equilibrium conditions.  It is also uncertain whether
predators of fish such as aquatic birds will ultimately biomagnify this
compound.
                                       77-8

-------
         In addition to experimental evidence, there is empirical evidence
that hexachlorobenzene has a high potential for accumulation in organisms.
Neely et_ al.. (1974) and Metcalf et_ al. (1973)  have  shown  that  the log  oc-
tanol/water partition coefficient (log P) correlates well with the ability
of a compound to accumulate in the lipids of tissues of living organisms.
The high log P value of 6.18 (Neely et_ _al. 1974) indicates that the bio-
accumulation potential of hexachlorobenzene by aquatic organisms at pollu-
tant concentrations anticipated in environmental waters is very high.

    77.4.7  Biotransformation and Biodegradation

         In a study on the uptake and excretion of some halogenated aroma-
tic hydrocarbons by juvenile Atlantic salmon (Salmo salar), Zitko e t a1.
(1977) stated that it is generally believed that fish excrete halogenated
aromatic hydrocarbons mainly unchanged, at a rate determined primarily by
the lipophilicity and/or water solubility of the compounds.  In the aquatic
model ecosystem of Metcalf et_ aJL. (1973 ) radiolabeled hexachlorobenzene was
found in substantial quantities in the various organisms with little evi-
dence of degradation products except "highly polar materials and conju-
gates".  For instance, hexachlorobenzene comprised 85.1 percent of the
total radioactivity in alga (Oedogonium cardiacum), 90.8 percent in the
snail (Physa sp. ), 87.2 percent in the water flea (Daphnia magna), 58.3
percent in mosquito larva (Culex pipiens quinquefasciatus), and 27.2 per-
cent in fish (Gambusia affinis).  The aqueous phase of the ecosystem,  how-
ever, contained an appreciable quantity of pentachlorophenol.  This compound
was not found in free form in any of the organisms of the system.  In the
aquatic model ecosystem of Lu and Metcalf (1975) hexachlorobenzene com-
prised 84 percent of the total radioactivity in the snail, 67 percent  in
the water flea, 65 percent in mosquito larva, and 64 percent in fish.   Pen-
tachlorophenol, however, was found in alga, mosquito larva, and the aqueous
phase as the only identified degradation product. Zitko et_ l. (1977) re-
port that hydroxylated metabolites of halogenated aromatic hydrocarbons
appear to play a relatively minor role in the excretion of halogenated
aromatic hydrocarbons since the mixed function oxidase system is much less
active in fish than in mammals.  Zitko e_t_ a_l. (1977) contend that, inasmuch
as the mixed function oxidase system has been demonstrated to be induced in
trout by PCBs (Lidman et_ _a_l. 1976), it is likely that most of the polar
metabolites of halogenated aromatic hydrocarbons, described in mammals, are
also formed in fish but in small amounts which are difficult to detect.

         Lu and Metcalf (1975) reported a value for the biodegradability
index of hexachlorobenzene of 0.377 in mosquito fish (Gambusia affinis) of
the aquatic model ecosystem.  The biodegradability index is defined by Lu
and Metcalf (1975) as the ratio of polar products of degradation to the
non-polar products.  A low value for the biodegradability index indicates
that a compound resists biodegradation.  Comparison of the biodegradability
                                    77-9

-------
index of hexachlorobenzene with that of some widely studied  persistent
pollutants, such as DDT and aldrin,  give a better idea of the significance
of this value.  Lu and Metcalf (1975) reported a biodegradability index for
DDT of 0.012 in mosquito fish (Gambusia affinis) compared to values of
0.015 for aldrin and 0.377 for hexachlorobenzene.

         According to Ware and West  (1977), the more highly  halogenated a
compound becomes, the more resistant it is to biodegradation.  Experimental
evidence has been found indicating that chlorobenzene is a persistent com-
pound (Lu and Metcalf 1975) and is not readily biodegraded unless the
microorganisms present are already growing on another hydrocarbon source
(Gibson _e_t_ al_. 1968).   Furthermore, Alexander and Lustigman (1966) found
that the presence of a chlorine atom on the benzene ring retarded the rate
of biodegradation.  On these bases,  hexachlorobenzene, being more highly
chlorinated than chlorobenzene, would presumably biodegrade  more slowly
than chlorobenzene under similar conditions of exposure to microorganisms.

77.5  Data Summary

    Data found concerning the processes for removal of hexachlorobenzene
are insufficient to allow designation of a most probable fate pathway for
this compound.  Hexachlorobenzene has a high affinity for lipophilic
materials; consequently, sorption and bioaccumulation are anticipated to
occur quite readily.  It appears that the major portion of hexachloroben-
zene found in aquatic organisms is from aqueous rather than  dietary
sources.  Furthermore, experimentally determined rates of depuration of
hexachlorobenzene appear to be substantially more rapid than rates of de-
puration for other persistent chemicals, such as DDT,  Consequently, bio-
magnification of hexachlorobenzene through aquatic food chains probably
does not occur.  Not enough data however is available to determine if this
statement is true where birds are the predators of fish.  These data are
summarized in Table 77-2.
                                       77-10

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                                           77-11

-------
77.6  Literature Cited

Alexander, M. and B.K. Lustigman.  1966.  Effect of chemical structure on
  microbial degradation of substituted benzenes.  J. Agr. Food Chem.
  14 (4):410-413.

Fieser, L.F. and M. Fieser.  1956.  Organic chemistry.  3rd Edition.  D.C.
  Heath and Co., Boston, Mass.   1112p.

Gibson, D.T., J. R. Koch, C.L. Schuld, and R.E. Kallio.  1968.  Oxidative
  degradation of aromatic hydrocarbons by microorganisms.  II.  Metabolism
  of halogenated aromatic hydrocarbons.  Biochemistry  7(11):3795-3802.

Isensee, A.R., E.R. Holden, E.A. Woolson, and G.E. Jones.  1976.  Soil
  persistence and aquatic bioaccumulation potential of hexachlorobenzene.
  J. Agric. Food Chem.  24(6 ):1210-1214.

Johnson, J.L., D.L. Stalling, and J.W. Hogan.  1974.  Hexachlorobenzene
  residues in fish.  Bull. Environ. Contam. Toxicol.  11(5):393-406.

Laseter, J.L., C.K. Bartell, A.L. Laska, D.G. HoLnquist, D.B. Condie, J.W.
  Brown, and R.L. Evans.  1976.  An ecological study of hexachlorobenzene.
  U.S. Environmental Protection Agency, Office of Toxic Substances,
  Washington, D.C.  62p.  EPA 560/6-76-009.

Laska, A.L. , C.K. Bartell, and J.L. Laseter.  l"976.  Distribution of
  hexachlorobenzene and hexachlorobutadiene in water, soil, and selected
  aquatic organisms along the lower Mississippi River, Louisiana.   Bull.
  Environ. Contam. Toxicol.  15(5 ):535-542.

Leoni, V. and S.U. D'Arca.   1976.  Experimental data and critical review of
  the occurrence of hexachlorobenzene in the  Italian environment.   Sci.
  Total Environ.  5:253-272.

Lidman, U., L. Forlin, 0. Molander, and G. Axelson.   1976.  Induction of
  the drug metabolizing system in rainbow trout (Salmo gairdneri) liver in
  polychlorinated biphenyls  (PCBs).  Acta Pharmacol. Toxicol.  39:262-272.

Lu, P. and R.L. Metcalf.  1975.  Environmental fate and biodegradability of
  benzene derivatives as studied in a model aquatic ecosystem.   Environ.
  Health Perspect.  10:269-284.

Macek, K.J., S.R. Petrocelli, and B.H. Sleight, III.  1977.  Considerations
  in assessing the potential for, and significance of biomagnification of
  chemical  residues in aquatic food chains.   Presented at:  ASTM Second
  Symposium on Aquatic Toxicology, October 31 - November 1, Cleveland,
  Ohio.
                                     77-12

-------
Mackay, D. and P.J. Leinonen.  1975.  Rate of evaportion of
  low-solubility contaminants from water bodies to atmosphere.  Environ.
  Sci. Technol.  9(13):1178-1180.

Mackay, D. and A.W. Wolkoff.  1973.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. Technol.
  7(7):611-614.

Metcalf, R.L., I.P. Kapoor, P. Lu, C.K. Schuth, and P. Sherman.  1973.
  Model ecosystem studies of the environmental fate of six organochlorine
  pesticides.  Environ.  Health Perspect.  (4):35-44.

Morrison, R.T. and R.N.'Boyd.  1973.  Organic chemistry.  3rd Edition.
  Allyn and Bacon, Inc.,  Boston.  1258p.

Neely, W.B., D.R. Branson, and G.E. Blau.  1974.  Partition coefficient to
  measure bioconcentation potential of organic chemicals in fish.  Environ.
  Sci. Technol.  8:1113-1115.

Patai, S.  (ed).  1973.   The chemistry of the carbon-halogen bond:  Part 2.
  John Wiley Interscience, New York.  1215p.

Plimmer, J.R. and U.I. Klingebiel.  1976.  Photolysis of hexachlorobenzene.
  J. Agric. Food Chem.  24(4):  721-723.

U.S. Environmental Protection Agency.  1975.  Preliminary assessment of
  suspected carcinogens  in drinking water.  U.S. Environmental Protection
  Agency, Office of Toxic Substances, Washington, B.C.  33p.  (EPA
  560/4-75-003).

Ware, S.A. and W.L. West.  1977.  Investigation of selected potential
  environmental contaminants:  halogenated benzenes.  U.S. Environmental
  Protection Agency, Office of Toxic Substances, Washington, D.C.  283p.
  (EPA 560/2-77-004).

Weast, R.C. (ed).  1977.   Handbook of chemistry and physics.  58th Edition,
  CRC Press, Inc., Cleveland, Ohio.  2398p.

Zitko, V.  1977.  Uptake and excretion of chlorinated and brominated
  hydrocarbons by fish.   Fisheries and Marine Service Technical Report No.
  737.  Fisheries and Environmental Sciences Resource Branch, St. Andrews,
  New Brunswick.  14p.
                                    77-13

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                         78.  ETHYLBENZENE
78.1  Statement of Probable Fate

    From the available data it would appear that the principal mechanism
for removal of ethylbenzene from the aquatic environment is volatilization.
The atmospheric photodestruction of ethylbenzene probably overshadows all
other fates.  Adsorption on sediments and suspended solids appears to play
a role that cannot be established quantitatively at this time.  It is not
possible to estimate the relative importance of biodegradation in the de-
termination of the fate of ethylbenzene in the aquatic environment.

78.2  Identification

    Ethylbenzene has been detected at several geographical locations in
finished drinking water, effluents, and ambient surface waters (Shackelford
and Keith 1976).  The chemical structure of ethylbenzene is shown below.
                                             Alternate Names

                                             Phenylethane
                                             Ethylbenzol
    Ethylbenzene

    CAS NO.  100-41-4
    TSL NO.  DA 07000

78.3  Physical Properties

    The general physical properties of ethylbenzene are as follows.

    Molecular weight                         106.16
    (Verschueren 1977)

    Melting  point                            -94.9C
    (Verschueren 1977)

    Boiling  point                            136.2C
    (Verschueren 1977)

    Vapor pressure at 20C                   7 torr
    (Verschueren 1977)
                                    78-1

-------
    Solubility in water at 20C              152 mg/1
    (Verschueren 1977)

    Log octanol/water partition coefficient  3.15
    (Tute 1971)

78.4  Summary of Fate Data

    78.4.1  Photolysis

         Inasmuch as the main transport process for ethylbenzene appears to
be volatilization, which removes it from water, the atmospheric destruction
of ethylbenzene probably subordinates all other fate processes.  These com-
plex photochemical reactions have been studied in simulated smog chambers
(Altshuller _e_t _al. 1962;  Laity _et al. 1973) that measured the rate of dis-
appearance of the volatilized organic material.  The half-conversion time
of m-xylene and 1,3,5-trimethylbenzene have been reported to be somewhat
less than four hours (Altshuller et_ a_l. 1962).  From this value and the
table of relative reactivities given by Laity ejt _al_. (1973), it can be in-
ferred that the corresponding half-conversion time for ethylbenzene would
be approximately 15 hours.  The temporal stability of ethylbenzene under
actual atmospheric conditions is, as yet, unknown. Experiments performed in
laboratory irradiation chambers are usually conducted for relatively short
periods and cannot account for all of the meteorological variables within a
natural airshed.

    78.4.2  Oxidation

         Ethylbenzene is readily oxidized in the liquid phase by molecular
oxygen, but this oxidation is effectively inhibited by the presence of
water (Stephens and Roduta 1935).  No data were found from which a relevant
rate of oxidation of ethylbenzene in environmental waters could be deter-
mined with confidence.

    78.4.3  Hydrolysis

         No data have been found that would support any role for hydrolysis
at ambient environmental conditions.

    78.4.4  Volatilization

         From the vapor pressure data (Gray 1957) and the aqueous solubil-
ity (Verschueren  1977), the computed Henry constant (H = 6.44 atmos. m-*
mole"*) lies in the same range as that for toluene, (Hto^ = 6.68 atmos.
m^ mole"*) (Mackay and Leinonen 1975).  Therefore, the half-life with
respect to volatilization (from a water layer one meter thick) could
                                      78-2

-------
approximate the 5 to 6 hour half-life estimated by Mackay and Leinonen
(1975) for toluene.  Thus, volatilization is probably an important removal
process for ethylbenzene from the aquatic environment.  No experimental
data on rates of evaporation of ethylbenzene from water were found.

    78.4.5  Sorption

         Although no specific environmental sorption studies were found in
the reviewed literature, the log octanol/water partition coefficient (log
P=3.15, Tute 1971) indicates that sorption processes may be significant for
ethylbenzene.  Presumably, ethylbenzene will be adsorbed by sedimentary
organic material;  however, the extent to which this adsorption will inter-
fere with volatilization, has not been considered.

    78.4.6  Bioaccumulation

         No information was found indicating that ethylbenzene would bio-
accumulate.  Moreover, Metcalf and Sanborn (1975) maintain that compounds
with solubilities of 50 mg/1 or more generally have little potential for
aquatic bioaccumulation.

    78.4.7  Biotransformation and Biodegradation

         Some species of soil bacteria have been demonstrated to be capable
of using ethylbenzene as a sole carbon source (Glaus and Walker 1964;
Gibson et_ _al. 1966).  This microbial oxidative degradation proceeds via
hydroxylation of the aromatic ring to 2,3-dihydroxy-l-ethylbenzene (Gibson
t al. 1973).

78.5  Data Summary

    The data obtained for ethylbenzene are summarized in Table 78-1.
Volatilization appears to be the major route of removal of this chemical
from aquatic environments.  The atmospheric reactions of ethylbenzene
probably overshadow all other fate processes.  The estimated photooxidative
half-life in Table 78-1 is based on smog chamber data and is, therefore, an
approximation applicable only to a metropolitan environment.
                                    78-3

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78.6  Literature Cited

Altshuller, A.P., I.R. Cohen, S.F. Sleva,  and S.L. Kopczynski.  1962.  Air
  pollution:  photooxidation of aromatic hydrocarbons.  Science.
  138(3538):442-443.

Glaus, D. and N. Walker.  1964.  The decomposition of toluene by soil
  bacteria.  J. Gen. Microbiol. 36:107-122.

Gibson, D.T., J. R. Koch, and R.E. Kallio.  1966.  Oxidative degradation of
  aromatic hydrocarbons by microorganisms.  Enzymatic formation of catechol
  from benzene.  Biochemistry.  7(7):2653-2662.

Gibson, D.T., B. Gschwendt, W.K. Yeh,  and V.M. Kobal.  1973.  Initial
  reactions in the oxidation of ethylbenzene by Pseudomonas putida.
  Biochemistry.  12(8):1520-1528.

Gray, D.E. (ed.)  1957.  American Institute of Physics handbook,  McGraw
  Hill, New York. 4224p.

Laity, J.L., I.G. Burstain, and B.R. Appel.  1973.  Photochemical smog and
  the atmospheric reactions of solvents.  Chap. 7, pp. 95-112.  Solvents
  Theory and Practice.  R.W. Tess (ed.)  Advances in Chemistry Series 124.
  Am. Chem. Soc., Washington, B.C.

Mackay, D. and P.J. Leinonen.  1975.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. Technol.
  9(13):1178-1180.

Metcalf, R.L. and J.R. Sanborn.  1975.  Pesticides and environmental
  quality in Illinois.  111. Nat'l. Hist.  Survey Bull.  31:381-436.

Shackelford, W.M. and L.H. Keith.  1976.  Frequency of organic compounds
  identified in water.  U.S. Environmental Protection Agency, (ERL) ,
  Athens, GA. 617p. (EPA 600/4-76-062).

Stephens, H.N. and F.L. Roduta.  1935.  Oxidation in the benzene series by
  gaseous oxygen.  The oxidation of tertiary hydrocarbons.  J. Am. Chem.
  Soc.  57:2380-2381.

Tute, M.S.  1971.  Principles and practice of Hansch analysis:  a guide to
  structure-activity correlation for the medicinal chemist.  Adv. in
  Drug Res.  5:1-77.  Academic Press,  New York.

Verschueren, K.  1977.  Handbook of environmental data on organic
  chemicals.  Van Nostrand/Reinhold, New York.  659p.
                                     78-5

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                             79.   NITROBENZENE


79.1  Statement of ProbableJFate

    The aquatic fate of nitrobenzene might involve both photoreduction of
the nitro group and biodegradation.  It is not possible with the available
data, to determine which of these two fates predominates.   It should be
noted that both photochemical and biological degradation can lead potenti-
ally to a large variety of organic nitrogen compounds, two of which,
namely, diphenylhydrazine and benzidine, are presently listed as priority
pollutants.  The persistence of these compounds, as well as nitrobenzene
itself, cannot be ascertained from existing data.

79.2  Identification

    Nitrobenzene has been found in unfinished drinking water and ambient
water (Shackelford and Keith 1976).  The chemical structure of nitrobenzene
is shown below.

                                             Alternate Names

                                             Nitrobenzol


     Nitrobenzene

     CAS NO. 98-95-3
     TSL NO. DA 64750

79.3  Physical Properties

    The general physical properties of nitrobenzene are as follows.

    Molecular weight                         123.11
    (Windholz 1976)

    Melting point                            5.6C
    (Windholz 1976)

    Boiling point at 760 torr                211C
    (Windholz 1976)

    Vapor pressure at 20C                   0.15 torr
    (Vershueren 1977)

    Solubility in water at 20C              1900 mg/1
    (Verschueren 1977)

    Log octanol/water partition coefficient  1.85
    (Neely et al. 1974)
                                     79-1

-------
79.4  Summary of Fate Data

    79.4.1  Photolysis

         Although dissociation of an N-0 nitro bond by light,  of  wavelengths
longer than 190 nm is energetically improbable,  photoreduction of  aromatic
nitrocompounds occurs at least to 436 nm (Leighton and Lucy 1934;  Morrison
1969).  The vapor phase irradiation products of  nitrobenzene  are nitroso-
benzene and 4-nitrophenol (Hastings and Matsen 1948).
                                                     NO,
         Inasmuch as the quantum yield of this vapor-phase reaction can be
expected to be quite low (Morrison 1969) and evaporation of nitrobenzene
from water is not thought to be a major aquatic transport process (see Sec-
tion 79.4.4,  Volatilization),  there is little likelihood that  vapor-phase
photolysis or photooxidation will make a significant contribution to the
aquatic fate of this compound.

         In the liquid phase,  nitrobenzene has been demonstrated to photo-
oxidize suitable hydrogen donors (Morrison 1969).   For example,  exposure of
nitrobenzene and toluene to sunlight leads to the  formation of a complex
mixture, the principal compounds of which are aniline, 4-aminophenol,
azoxybenzene and benzoic acid (Vecchiotti and Zanetti 1931; Morrison
1969).
                            O(-)
                                                        COOH

-------
         Adsorption of nitrobenzene by humus could allow reactions of this
type to play a significant role in the abiotic, aquatic fate of this
compound.  The most suitable hydrogen donors appear to be alkylaromatics
that can form benzyl-type radicals, although aromatic hydrogens are also
reported to be abstracted (Morrison 1969).  In this respect, suspended
humus should have an abundance of reaction sites (Rook. 1977) for the
photoreduction of nitrobenzene.

    79.4.2  Oxidation

         Although no specific environmental data regarding oxidation were
found, it is probable that oxidation does not operate as an initial aquatic
fate process for nitrobenzene.  Substitution by nitro groups decreases the
electron density of aromatic rings (Fieser and Fieser 1956) resulting in a
decreased susceptibility to attack by molecular oxygen and hydroxyl radi-
cals.  Nitrobenzene itself is a strong chemical oxidizing agent, but intra-
molecular oxidation-reduction reactions involving the aromatic hydrogens
are unknown;  an intermolecular oxidation-reduction reaction, however, has
been postulated as the mechanism of its vapor-phase photolysis  (Morrison
1969).  Oxidative processes undoubtedly contribute, via quinone formation
(Cason 1948), to the ultimate destruction of nitrobenzene's photoreduction
products.
                       NH,
                            dissolved
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and
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    79.4.3  Hydrolysis

         No specific information was found regarding the hydrolysis of
nitrobenzene.  Nonetheless, the hydrolytic scission of any covalent bond of
nitrobenzene under ambient environmental conditions is highly improbable
(Fieser and Fieser 1956).

    79.4.4  Volatilization

         The vapor pressure data for nitrobenzene  (Gray 1957) may be ex-
tended to ambient temperature using the equation:
                                     79-3

-------
    In VP = a + b/T (where VP is in torr;   T in K)

from which Che value included in Section 79.3 was derived using a = 20.2
and b = 6415.  Further, the solubility of  nitrobenzene in water is given as
1900 mg/1 (Verschueren 1977).  Combining these data,  the calculated Henry
constant for nitrobenzene, H = 1.53 x 10~-> atmos. wr  mole"-'-,  is com-
parable to that of aldrin (Mackay and Leinonen 1975)  for which the half-
life with respect to volatilization from water is approximately 185 hours.
Under these circumstances, it would appear that volatilization, while con-
tributing to the transport of nitrobenzene, will probably not be of over-
bearing consequence.  In the model aquatic ecosystem, studied: by Lu and
Metcalf (1975), 2.22 percent of the nitrobenzene introduced into the water
was recovered in the air at the end of the experiment.

    79.4.5  Sorption

         The octanol/water partition coefficient for  nitrobenzene, corres-
ponding to log p = 1.85 (Neely e_ al_. 1974), suggests only a limited pre-
ference for lipophilic organic materials over water.   No data on adsorptive
transport or removal of nitrobenzene from water were  found. The polar
nature of the nitro group, however, coupled with nitrobenzene's miscibility
with most organic material, insures the adsorption of nitrobenzene by
humus.  Adsorption onto clay is also thought to be very highly probable.
Photoreduction on humus, and acid-base catalyzed reactions of the reduction
products on clay, could be a major fate pathway for nitrobenzene.

    79.4.6  Bioaccumulation

         Radio-labelled nitrobenzene was studied for  its effect in a model
aquatic ecosystem developed by Lu and Metcalf (1975).  This aquatic ec-
osystem was devised for studying relatively volatile  organic compounds and
simulating direct discharge of chemical wastes.  The  food chain members
consisted of green filamentous algae, snails, water fleas, mosquito larvae,
and fish.

         The contaminative efficacy of nitrobenzene was evaluated by de-
termining the quantitative distribution of the radioactivity in the organ-
isms, water, and air of the model aquatic ecosystem.   Nitrobenzene was
neither stored nor ecologically magnified, (ecological magnification of 29)
and it was found to be resistant to degradation (biodegradability index of
0.023).  Compared to halogenated organic compounds, such as hexachloroben-
zenes and DDT, nitrobenzene was found less likely to  be stored in the fatty
tissue of aquatic organisms.

    79.4.7  Biotransformation and Biodegradation

         A nitro substituent generally decreases the  biodegradability of
aromatic rings  (Venulet and Van Etten 1970;  Lu and Metcalf 1975), and it
has been reported that  the decomposition of nitrobenzene by soil microflora
requires at least 64 days  (Verschueren 1977).


                                      79-4

-------
         The metabolism of C14 labeled nitrobenzene by rabbits has been
extensively investigated (Robinson et_ al. 1951; Parke 1956).  Fifty-
eight percent of the label was eliminated in the urine, with 31 percent
being present as aminophenol and its detoxification conjugates.  The con-
jugates were largely the glucuronide or the N-acetyl derivative of the
glucuronide.  In addition, detoxification conjugates of 3-nitrophenol,
4-nitrophenol, and aniline were also isolated from the urine.

79.5  Data Summary

    Table 79-1 summarizes the aquatic fate data discussed above for
nitrobenzene.  Photoreduction of the nitro group and/or hydroxylation of
the ring while adsorbed-on humus could be an important abiotic fate.
Biodegradation to the same type of reduction products could also be an
important fate process.  There is no way, with the available data, to
ascertain which of these two fates predominates.  It should be noted that
both photochemical and biological degradation can lead potentially to a
large variety of organic nitrogen compounds, two of which, namely,
diphenylhydrazine and benzidine, are presently listed as priority pollu-
tants.
                                      79-5

-------
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                                79-6

-------
79.6  Literature Cited

Cason, J.  1948.  Synthesis of benzoquinones by oxidation. Organic
  Reactions.   4:305-361.

Fieser, L.F. and M.  Fieser.  1956.  Organic chemistry.  3rd edition D.C.
  Heath and Co., Boston.  1112p.

Gray, D.E.  1957.  American institute of physics handbook.  McGraw-Hill,
  New York.  4224p.

Hastings, S.H. and F.A. Matsen.  1948.  The photodecomposition of
  nitrobenzene.  J.  Am. Chem. Soc.  70:3514-3515.

Leighton, P.A. and F.A. Lucy.  1934.  The photoisomerization of the
  o-nitrobenzaldehydes.  J. Chem. Phys.  2:756-759.

Lu, P.Y. and R. Metcalf.  1975.  Environmental fate and biodegradability of
  benzene derivatives as studied in a model aquatic ecosystem.  Environ.
  Health Perspect.  19:269-273.

Mackay, D. and P.J.  Leinonen.  1975.  Rates of evaporation of
  low-solubility contaminants from water bodies to atmosphere.  Environ.
  Sci. Technol. 9(1):1178-1180.

Morrison, H.A.  1969.   The photochemistry of the nitro and nitroso groups.
  H. Feuer (ed).  The chemistry of the nitro and nitroso groups.  Part 1.
  Chap. 4.  pp. 165-212.  Interscience Publishers, New York.

Neely, W.B., D.R. Branson and G.E. Blau.  1974.  Partition coefficient to
  measure bioconcentration potential of organic chemicals in fish.
  Environ. Sci. Technol. 8:1113-1115.

Parke, D.V.  1956.  Studies in detoxication.  The metabolism of
  nitrobenzene  in the rabbit and guinea pig.  Biochem. J. 62:339-346.

Robinson, D., J.N. Smith, and R.T. Williams.  1951.  Studies in
  detoxication.  The metabolism of nitrobenzene in the rabbit.  Biochem. J.
  50:228-235.

Rook, J.J.  1977.  Chlorination reactions of fulvic acids in natural
  waters.  Environ.  Sci. Technol. 11(5):477-482.

Shackelford, W.M. and L.H. Keith.  1976.  Frequency of organic compounds
  identified in water.  U.S. Environmental Protection Agency, (ERL),
  Athens, GA. 617p.  (EPA-600/4-76-062).
                                       79-7

-------
Vecchiotti, L. and G. Zanetti.  1931.  Chemical reactions promoted by
  light. Gazz. Chim. Ital. 61:789-802.

Venulet, J. and R.L. Van Etten.  1970.  Biochemistry and pharmacology of
  the nitro and nitroso groups.  H.  Feuer (ed). The chemistry of the nitro
  and nitroso groups.  Part 2.  Chap. 4. pp.201-287. Interscience
  Publishers, New York.

Verschueren, K.  1977.  Handbook of environmental data on organic
  compounds.  Van Nostrand/Reinhold, New York.  659p.

Windholz, M. (ed).  1976.  The Merck index, vol. 9.  Merck and Co. Rahway,
  N.J. , 1313p.
                                     79-8

-------
                               80.  TOLUENE


80.1  Statement of Probable Fate

    From the available data it appears that the principal mechanism for re-
moval of toluene from the aquatic environment is volatilization.  The atmo-
spheric photodestruction of toluene probably subordinates all other fates.
Adsorption on sediments and suspended solids probably plays a role in the
fate of toluene, but it cannot be established quantitatively at this time.
The data do not allow the estimation of the relative importance of bio-
degradation in the determination of the fate of toluene in the aquatic
environment.

80.2  Identification

    Toluene has been detected at several geographical locations in finished
drinking water, industrial effluents, and ambient surface waters
(Shackelford and Keith 1976).  The structure of toluene is shown below.


                                             Alternate Names

                                             Toluol
                                             Phenylmethane
                                             Methylbenzene
                                             Methyl benzol
                                             Methacide
      Toluene

      CAS NO. 108-88-3
      TSL NO. XS 52500

80.3  Physical Properties

    The general physical properties of toluene are as follows.


    Molecular weight                         92.13
    (Weast 1977)

    Melting point                            -95C
    (Weast 1977)

    Boiling point                            110.6C
    (Weast 1977)
                                   80-1

-------
    Vapor pressure at 25C                   28.7 torr
    (Weast 1977)

    Solubility in water at 25C              534.8 mg/1
    (Sutton and Calder 1975)

    Log octanol/water partition              2.69
    coefficient (Tute 1971)

80.4  Summary of Fate Data

    80.4.1  Photolysis

         The predominant photochemical reaction of toluene is generally re-
garded as a dissociation with formation of benzyl radical (Porter and
Norman 1954).  Reaction of this benzyl radical with molecular oxygen is re-
ported to be extremely fast (k~10^ 1 mole sec"-'-) resulting in the
production of benzyl hydroperoxide.  Although benzyl hydroperoxide is
thermally stable, it can be photochemically dissociated to benzyl alcohol
and benzaldehyde.

         Toluene itself does not absorb light at wavelengths longer than
286 nm, but a charge-transfer complex between toluene and molecular oxygen
absorbs electromagnetic radiation to at least 350 nm.  According to Wei and
Adelman (1969), it is the photolysis of this charge-transfer complex that
is responsible for the observed oxidation products (benzyl alcohol and
benzaldehyde) at ambient conditions.  No information was found on the
extent of formation of this complex in water.  Furthermore, no information
was found  from which the rate of benzylic hydrogen abstraction in water
could be determined.

         Inasmuch as the main transport process removing toluene from water
appears to be volatilization, the atmospheric reactions of toluene probably
subordinate all other fate processes.  These complex photochemical re-
actions have been studied in simulated smog chambers (Altshuller et al.
1962; Laity et al. 1973) that measured the rate of disappearance of the
volatilized organic material.  The half-conversion time of m-xylene and
1,3,5-trimethylbenzene  have been reported to be somewhat less than four
hours (Altshuller _e_t al. 1962).  From this value and the table of relative
reactivities given by Laity e_t _a_l. (1973), it can be inferred that the
corresponding half-conversion time for toluene would be approximately
fifteen hours.  Benzaldehyde is reported to be the principal organic
product from the photochemical reaction of toluene (Laity et al. 1973).
The temporal stability of toluene under actual atmospheric conditions is,
as yet, unknown.  Experiments performed in laboratory irradiation chambers
are usually conducted for relatively short periods and cannot account for
all of the meteorological variables within a natural airshed.
                                   80-2

-------
    80.4.2  Oxidation

         Toluene is readily oxidized in the liquid phase by molecular oxy-
gen, but this oxidation is effectively inhibited by the presence of water
(Stephens and Roduta 1935). Reaction of toluene in water with hydroxyl
radicals from the irradiation of hydrogen peroxide produces benzaldehyde,
benzyl alcohol, and an isomeric mixture of  cresols (Jefcoate _e_t ai. 1969).
No data were found from which a relevant rate of oxidation of toluene in
environmental waters could be determined.
                          CHO
CH2 OH
    80.4.3  Hydrolysis

         No data have been found that would support any role for hydrolysis
at ambient environmental conditions.

    80.4.4  Volatilization

         The half-life with respect to volatilization from a water column
one meter thick has been estimated by Mackay and Leinonen (1975) to be 5.18
hr for toluene.  Some assumptions made in this estimation were:  1) the con-
taminant concentration is in solution, rather than in suspended, colloidal,
ionic, complexed, or adsorbed form; 2) the vapor is in equilibrium with the
liquid at the interface; 3) water diffusion or mixing is sufficiently rapid
so that the concentration at the interface approaches that of the bulk of
the water; and 4) the rate of evaporation of water is negligibly affected
by the presence of the contaminants.

    80.4.5  Sorption

         Although no specific environmental sorption studies were found in
the reviewed literature, the log octanol/water partition coefficient (log
P= 2.69; Tute 1971) indicates that sorption processes may be significant
for toluene.  Presumably, toluene will be adsorbed by sedimentary organic
material, but the extent to which this absorption will interfere with
volatilization has not been considered.
                                    80-3

-------
    80.4.6  Bioaccumulation

         No information was found indicating that toluene would bio-
accumulate. Moreover, Metcalf and Sanborn (1975) maintain that, in general,
compounds with solubilities of 50 mg/1 or more have little potential for
aquatic bioaccumulation.

    80.4.7  Biotransformation and Biodegradation

         Some species of soil bacteria have been demonstrated to be capable
of using toluene as a sole carbon source (Glaus and Walker 1964; Gibson _e_t
al. 1966).  This oxidative microbial degradation proceeds via hydroxylation
of the aromatic ring to-a mixture of cresols and catechols, which are meta-
bolized further to acetic acid and pyruvic acid.  In mammals,, toluene is
detoxified by oxidation to benzoic acid, which then reacts with glycine to
form hippuric acid (Ogata _!  1970).  Hippuric acid is rapidly excreted
in the urine.

80.5  Data Summary

    The data obtained for toluene are summarized in Table 80-1.  Volatili-
zation appears to be the major route of removal of this chemical from
aquatic environments.  The atmospheric reactions of toluene probably sub-
ordinate all other fate processes. The precipitation of the atmospheric
oxidation products could introduce benzaldehyde into the water.  The es-
timated photooxidative half-life  in Table 80-1 is based on smog chamber
data and is, therefore, an approximation applicable only to & metropolitan
environment.
                                     80-4

-------




















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80.6  Literature Cited

Altshuller, A.P., I.R. Cohen, S.F. Sleva, and S.L. Kopczynski.  1962.  Air
  pollution:  photooxidation of aromatic hydrocarbons.  Science.
  138(3538):442-443.

Glaus, D. and N. Walker.  1964.  The decomposition of toluene by soil
  bacteria.  J. Gen. Microbiol. 36:107-122.

Gibson, D.T., J.R. Koch, and R.E. Kallio.  1966.  Oxidative degradation of
  aromatic hydrocarbons by microorganisms.  Enzymatic formation of catechol
  from benzene.  Biochemistry.  7(7):2653-2662.

Jefcoate, C.R.E., J.R. Lindsay-Smith, and R.O.C. Norman. 1969.  Oxidation
  of some benzenoid compounds by Fenton's reagent and the ultraviolet
  irradiation of hydrogen peroxide.  J. Chem. Soc. B. 1013-1018.

Laity, J.L., I.G. Burstain, and B.R. Appel.  1973.  Photochemical smog and
  the atmospheric reactions of solvents.  Chap. 7.  pp. 95-112.  Solvents
  Theory and Practice.  R.W. Tess (ed.)  Advances in Chemistry Series 124.
  Am. Chem. Soc., Washington, D.C.

Mackay, D. and P.J. Leinonen.  1975.  Rate of evaporation of low-solubility
  contaminants from water bodies to atmosphere.  Environ. Sci. and Technol.
  9(13):1178-1180.

Metcalf, R.L. and J.R. Sanborn.  1975.  Pesticides and environmental
  quality in Illinois.  111. Natl. Survey Bull. 31:381-436.

Ogata, M., K. Toraokuni and Y. Takatsuka.  1970.  Urinary excretion of
  hippuric acid in  the urine of persons exposed to toluene.  Brit. J.
  Industr. Med.   27:43-50.

Porter, G. and I. Norman. 1954.  Trapped atoms and radicals in a glass
  cage.  Nature.  174(4428):508-509.

Shackelford, W.M. and L.H. Keith.  1976.  Frequency of organic compounds
  identified in water.  U.S. Environmental Protection Agency, (ERL),
  Athens, Ga. 6l7p.   (EPA-600/4-76-062).

Stephens, H.N. and  F.L. Roduta.  1935.  Oxidation in the benzene series by
  gaseous oxygen.   The oxidation of tertiary hydrocarbons.  J. Am. Chem.
  Soc. 57:2380-2381.

Sutton, C. and J.A. Calder.  1975.  Solubility of alkylbenzenes in
  distilled water and seawater at 25C.  J. Chem. Eng. Data
  20(3):320-322.
                                      80-6

-------
Tute, M.S. 1971.  Principles and practice of Hansch analysis: a guide  to
  structure-activity correlation for the medicinal chemist.  Adv. Drug Res.
  5:1-77.

Weast, R.C. (ed.)  1977.  CRC handbook of cheiaisty and physics. CRC Press,
  Inc., Cleveland, Ohio. 2398p.

Wei, K.S. and A.H. Adelman.  1969.  ,The photooxidation of  toluene.  The
  role of an excited charge-transfer complex.  Tetrahedron Lett.
  (38):3297-3300.
                                    80-7

-------
                          81.  2,4-DINITROTOLUENE
81.1  Statement of Probable Fate

    The aquatic fate of 2,4-dinitrotoluene might involve photodestruction,
oxidation, and biodegradation.  It is not possible, with the available
data, to determine which of these fates predominates.  Adsorption onto
sediment probably plays a major role in transport and may also provide re-
action sites for destruction.  It should be noted that both abiotic and
biotic degradation can lead potentially to a large variety of organic ni-
trogen compounds.  The persistence of these compounds, as well as 2,4-di-
nitrotoluene itself, cannot be ascertained from existing data.

81.2  Identification

    2,4-Dinitrotoluene has been detected in ambient waters and industrial
effluents (Shackelford and Keith 1976).  The chemical structure of
2,4-dinitrotoluene is shown below.

                                             Alternate Names
                                             Dinitrotoluol
                                             DNT
                                             l-Methyl-2,4-dinitrobenzene
     2,4-Dinitrotoluene

     CAS NO. 121-14-2
     TSL NO. XT 15750

81.3  Physical Properties

    The general physical properties of 2,4-dinitrotoluene are as follows.

                                             182.14
Molecular weight
(Verschueren 1977)

Melting point
(Verschueren 1977)

Boiling point at 760 torr
(Verschueren 1977)

Vapor pressure at 59C
(Lenchitz and Velicky 1970)
                                             70C


                                             300C


                                             0.0013 torr
                                     81-1

-------
   Solubility in water at 22C
   (Verschueren 1977)
270 mg/1
   Log octanol/water partition coefficient  2 01
   (Calc. by method of Tute 1971)

81.4  Summary of Fate Data

    81.4.1  Photolysis

         Although dissociation of an N-0 nitro bond by light: of wavelengths
longer than 190 nm is energetically improbable, photoreduction of aromatic
nitro compounds occurs at least to 436 nm (Leighton and Lucy 1934;  Morrison
1969).  It is not expected that vapor phase photolysis of 2,4-dinitro-
toluene will have any significant effect on its aquatic fate in view of the
compound's low volatility and moderate solubility.   Photolysis in solution,
however, may be a highly probable fate process.

         Wettermark (1962) demonstrated that 2,4-dinitrotoluene is photo-
chromic, i.e., it has the property of becoming colored on exposure to light
and then  becoming colorless in the dark.  When dilute aqueous solutions
(10~^M) of 2-nitrotoluene, or any of several, similar compounds including
2,4-dinitrotoluene, are irradiated with ultraviolet light they become in-
tensely colored and then slowly fade with the cessation of irradiation.
Although the absorption of light of wavelengths longer than 330 nm is
greatly diminished in the case of 2-nitrotoluene and 2,4-dinitrotoluene,
absorption still does occur.  The photochromism of  these compounds appears
to be dependent on the ease of formation of a structural isomer analogous
to an aci-nitroparaffin.
                              l_
                              2  light
                                 dark
                                               =N
                                                    OH
                                           NO,
         The environmental consequence of this phenomenon should be the re-
duction, in sunlight, of the nitro group to a hydroxylamino,  nitroso, or
amino group with concomitant oxidation of the methyl group to an alcohol,
aldehyde, or carboxylic acid group.  Specific substantiation for this
supposition has not been found, thus far, in the reviewed literature.  The
exposure to sunlight, however, of a mixture of nitrobenzene arid toluene has
been shown to lead to a complex mixture, the principal compounds of which
                                    81-2

-------
are aniline,  4-aminophenol,  azoxybenzene,  and benzoic acid (Vecchiotti and
Zanetti 1931; Morrison 1969).
                                                           NH.,
                              O(-)
                                                           COOH
         Adsorption of 2,4-dinitrotoluene on a suspended clay particle,  or
its incorporation into an acidic micelle, could provide an additional de-
gradative pathway for the initial photoproduct.  Since aci-nitroparaffins
undergo acid catalyzed hydrolysis to olefinic carbonyl compounds  (Johnson
and Degering 1943), the following sequence might take place.
                              NOn
                                               CH2OH
    81.4.2  Oxidation

         Oxidation of the methyl group of 2,4-dinitrotoluene by aqueous
hydroxyl radical or dissolved oxygen is a distinct  environmental possibil-
ity.  In very strongly basic hydroxylic solvents,  the benzyl anion of
2,4-dinitrotoluene transfers an electron to molecular oxygen or any other
available electron acceptor to form the 2,4-dinitrobenzyl radical
                                    81-3

-------
(Russell et_ al.  1967).   This organic radical can then couple to form 2,2',
4,4'-tetranitrobibenzyl or react further with oxygen to form a hydroperox-
ide.   Further sequential oxidation of these two compounds  leads to a mix-
ture  of oxidation products.
                                                           CH2OOH
         Since it is highly probable that 2,4-dinitrotoluene will be ad-
sorbed by suspended clay particles, these reactions are environmentally
feasible on the clay surface, inasmuch as this surface provides basic re-
action sites for ionization.

    81.4.3  Hydrolysis

         There is little likelihood that direct hydrolysis of this compound
will occur.  No data were found suggesting that dinitrotoluenes are subject
to hydrolysis under environmental conditions.

    81.4.4  Volatilization

         Since the vapor pressure of 2,4-dinitrotoluene is quite low at
ambient conditions and the solubility is moderate (270 mg/1 at 22C)
(Verschueren 1977), it is anticipated that the Henry constant will be with-
in the range of 1 x 10~^ to 1 x 10~^ atmos.  m-^ mole""*.  These
values indicate that the half-life with respect to volatilization will be
approximately hundreds of days (Mackay and Leinonen 1975).  Volatilization,
therefore, is probably not an important transport process for 2,4-dinitro-
toluene.

    81.4.5  Sorption

         The log octanol/water partition coefficient calculated by the
method of Tute (1971) (log P =2.01) is sufficiently large to indicate that
adsorption by humus may be significant for 2,4-dinitrotoluene. Although no
data were found specifically pertaining to adsorption of this compound, the
ability of polynitroaroniatic compounds to form very
                                     81-4

-------
stable charge-transfer complexes with more highly electronegative aromatic
compounds (Hall and Poranski 1970) indicates that 2,4-dinitrotoluene should
be strongly adsorbed by both humus and clay.  In addition,  the basic sites
on the clay surface might form addition-type complexes with this compound
(Hall and Poranski 1970).

    81.4.6  Bioaccumulation

         No environmentally relevant data on the bioaccumulation of 2,4-di-
nitrotoluene have been found.   By analogy with nitrobenzene, bioaccumula-
tion may not be an important process, and this contention is supported by
the relatively low octanol/water partition coefficient (log P = 2.01) for
2,4-dinitrotoluene compared to highly bioaccumulated compounds (PCBs, for
example) which exhibit log P value of 5 to 6 or more.

    81.4.7  Biotransformation and Biodegradation

         Detoxification of 2-nitrotoluene by dogs leads to  urinary excre-
tion of 2-nitrobenzyl alcohol and 2-nitrobenzoic acid (Venulet and VanEtten
1970); 2,4-Dinitrotoluene might be metabolized in an analogous fashion.
Alexander and Lustigman (1966) report that dinitrobenzenes  are resistant to
biodegradation by soil microorganisms.  It has been reported that
dinitrotoluenes are decomposed very slowly in a reservoir (Galuzova 1963).
Biodegradation by Azotobacter has also been reported to be  slow (Bringmann
and Kuehn 1972).  Reduction of the nitro groups by the fungus Mucrosporium
has been reported to result in the formation of 2-amino-4-nitrotoluene,
4-amino-2-nitrotoluene, 2,2'-dinitro-4,4'-azoxytoluene, 4,4'-
dinitro-2,2'-azoxytoluene, and 4-acetamido-2-nitrotoluene (McCormick et al.
1978).

81.5  Data Summary

    Table 81-1 summarizes the aquatic fate data discussed above for 2,4-di-
nitrotoluene.  Intramolecular photolysis, oxidation by dissolved oxygen,
and, to a small extent, biodegradation may all be aquatic fates for this
compound.  Adsorption onto sediment probably plays a major  role in trans-
port and may also provide reaction sites for destruction.  With the
available data, it is not possible to determine which fate  predominates.
                                      81-5

-------
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                                      81-6

-------
81.6  Literature Cited

Alexander, M. and B.K. Lustigman.  1966.  Effect of chemical structure  on
  microbial degradation of substituted  benzenes.  J. Agr. Food Chem.
  14(4):410-413.

Bringmann, G. and R. Kuehn.  1972.  Biological decomposition of
  nitrotoluenes and nitrobenzenes by azotobacter agilis.  Gesundh. Ing.,
  92(9) 273-276; C.A., 76:49516f. (Abstract only)

Galuzova, L.V. 1963.  Maximum permissible concentration of dinitrotoluene
  in the water or reservoirs.  Gig. Sanit. 28(2):14-19.

Hall, T.N. and C.F. Poranski, Jr.  1970.  Polynitroaromatic addition
  compounds.  H. Feuer (ed. ) The chemistry of the nitro and nitroso groups,
  Part 2. Chap. 6. pp 329-384.  Interscience Publishers, New York.

Johnson, K. and E.F. Degering.  1943.   Production of aldehydes and ketones
  from nitroparaffins.  J. Org. Chem. 8(1):10-11.

Leighton, P.A. and F.A. Lucy.  1934.  The photoisomerization of the
  o-nitrobenzaldehydes.  J. Chem. Phys. 2:756-759.

Lenchitz, C. and R.W. Velicky.  1970.   Vapor pressure and heat of
  sublimation of three nitrotoluenes. J. Chem. Eng. Data.  15(3 ):401-403.

Mackay, D. and P.J. Leinonen.  1975.  Rate of evaporation of
  low-solubility  contaminants from water bodies to atmosphere.  Environ.
  Sci. Technol. 9(13):1178-1180.

McCormick, N.G., J.H. Cornell, and A.M. Kaplan.  1978.  Identification  of
  biotransformation products from 2,4-dinitrotoluene.  Appl. Environ.
  Microbiol.  35(5 ):945-948.

Morrison, H.A.  1969.  The photochemistry of the nitro and nitroso groups.
  H. Feuer (ed. ) The chemistry of the nitro and nitroso groups.  Part I.
  Chap. 4. p.165-212.  Interscience Publishers, New York.

Russell, G.A., A.J. Move, E.G. Janzen,  S. Mak, and E.R. Talaty.  1967.
  Oxidation of carbanions.  II. Oxidation of p-nitrotoluene and
  derivatives in basic solution.  J. Org. Chem. 32(1):137-146.

Shackelford, W.M. and L.H. Keith.  1976.  Frequency of organic compounds
  identified in water.  U.S. Environmental Protection Agency (ERL),
  Athens,  Ga. 617p.  (EPA-600/4-76-062).
                                    81-7

-------
Tute, M.S.   1971.  Principles and practice of Hansch analysis: a guide to
  structure-activity correlation for the medicinal chemist.  Adv. Drug Res.
  6:1-77.

Vecchiotti, L. and G. Zanetti.  1931.  Chemical reactions promoted by
  light.  Gazz. Chim. Ital. 61:798-802.

Venulet, J. and R.L. VanEtten.  1970.  Biochemistry and pharmacology of the
  nitro and nitroso groups. H. Feuer (ed. ) The chemistry of the nitro and
  nitroso groups.  Part 2. Chap. 4. pp 201-287.  Interscience Publishers,
  New York.

Verschueren, K.  1977.  Handbook of environmental data on organic
  compounds.  Van Nostrand/Reinhold, New York. 659p.

Wettennark, G. 1962.  Photochromism of o-nitrotoluenes.  Nature
  194(4829):677.
                                    81-8

-------
                          82.  2,6-DINITROTOLUENE


82.1  Statement of Probable Fate

    The aquatic fate of 2,6-dinitrotoluene may involve both photodestruc-
tion and biodegradation.  It is not possible, with the available data, to
determine which of these fates predominates.  Adsorption onto sediment
probably plays a major role in transport and may also provide reaction
sites for destruction.  It should be noted that both abiotic and biotic de
gradation can lead potentially to a large number of organic nitrogen com-
pounds.  The persistence of these compounds, as well as 2,6-dinitrotoluene
itself, cannot be ascertained from existing data.

82.2  Identification                                                   ,
      	                                                   /

    2,6-Dinitrotoluene, a relatively minor intermediate in the manufacture
of trinitrotoluene (Fieser and Fieser 1956), has been detected in finished
drinking water, industrial effluents, and ambient waters (Shackelford and
Keith 1976).  The chemical structure of 2,6-dinitrotoluene is shown below.
                                             Alternate Names
                                             Dinitrotoluol
     2,6-Dini tro toluene

     CAS NO. 606-20-2
     TSL NO. XT 19250

82.3  Physical Properties

     The general physical properties of 2,6-dinitrotoluene are as follows.

     Molecular weight                        182.14
     (Weast 1977)

     Melting point                           65C
     (Weast 1977)

     Boiling point                           285C
     (Maksimov 1968)

     Vapor pressure                          No data found
                                     82-1

-------
     Solubility in water

     Log octanol/water partition
     coefficient (Calc. by method
     of Tute 1971)
No data found

2.05
 82.4  Summary of Fate Data

    82.4.1  Photolysis

         Although dissociation of an N-0 nitro bond by light of wavelengths
longer than 190 nm is energetically improbable, photoreduction of aromatic
nitro compounds occurs at least to 436 nm (Leighton and Lucy 1934; Morrison
1969).  It is not expected that vapor-phase photolysis of 2,6-dinitro-
toluene will have any significant effect on its aquatic fate in view of the
compound's expected low volatility and moderate solubility.  Photolysis in
\solution, however, may be a highly probable fate process.
 \
         Wettermark (1962) demonstrated that 2,6-dinitrotoluene is photo-
chromic, i.e., it has the property of becoming colored on exposure to  light
and then becoming colorless in the dark.  When dilute aqueous solutions
(10~^M) of 2-nitrotoluene, or any of several similar compounds, are
irradiated with ultraviolet light they become intensely colored and then
slowly fade with the cessation of irradiation.  Although the absorption of
light of wavelengths longer than 330 nm is greatly diminished in the case
of 2,6-dinitrotoluene, absorption still does occur.  The photochromism of
these compounds appears to be dependent on the ease of formation of a
structural isomer analogous to an aci-nitroparaffin.
                                   light
                                   dark
                                                       N.
                                                         "OH
          The  environmental consequence of this phenomenon should be the  re-
 duction,  in sunlight,  of the nitro  group to a hydroxylamino, nitroso, or
 amino  group with concomitant oxidation of the methyl group to an alcohol,
 aldehyde,  or  carboxylic acid group.   Specific substantiation for this
 supposition has not  been found, thus  far, in the  reviewed literature.  The
 exposure  to sunlight,  however, of a mixture of nitrobenzene and toluene  has
                                      82-2

-------
been shown to lead to a complex mixture, the principal compounds of which
are aniline, 4-aminophenol, azoxybenzene,  and benzoic acid (Vecchiotti and
Zanetti 1931;  Morrison 1969).
           CH,
                                                               COOH
                                    82-3

-------
          Adsorption of 2,6-dinitrotoluene  on  a  suspended clay particle, or
its incorporation into an acidic micelle, could  provide an additional de-
gradative pathway for the initial photoproduct.   Since aci-nitroparaffins
undergo acid  catalyzed hydrolysis to olefinic  carbonyl compounds (Johnson
and Degering  1943), the following sequence  might take place.
                                                  CH-OH
=0    0.
                                             HQT
    82.4.2  Oxidation

         Oxidation of  the methyl group of 2,6-dinitrotoluene by aqueous
hydroxyl radical or dissolved oxygen is not environmentally feasible.  In
very strongly basic solutions, the benzyl anion of  2,6-dinitrotoluene has
been observed to be very stable to oxidation by dissolved oxygen  (Russell
et al. 1967).

    82.4.3  Hydrolysis

         No data were  found suggesting that the dinitrotoluenes are subject
to hydrolysis under environmental conditions.

   82.4.4  Volatilization

         The vapor pressure of the dinitrotoluenes  is  probably low at ambi-
ent conditions and the solubility is expected to be moderate, perhaps about
300 ppm (by inference  from  the reported solubility  of  the 2,4-isomer in
Verschueren 1977).  From this it is anticipated that the Henry constant for
                                    82-4

-------
the 2,6-isomer will be in the range of 1 x 10"^ to 1 x 10~ atmos. nr
mole"^-.  This would suggest that the half-life with respect to
volatilization would be-measured in hundreds of days (Mackay and Leinonen
1975).  Volatilization, therefore, may not be an important transport pro-
cess for 2,6-dinitrotoluene.  No experimental data, however, were found to
support this supposition.

    82.4.5  Sorption

         The log/octanol water partition coefficient calculated by the
method of Tute (1971) (log P = 2.05) is sufficiently large to indicate that
adsorption by humus may be significant for 2,6-dinitrotoluene.  Although no
data were found specifically pertaining to adsorption of this compound, the
ability of polynitroaromatic compounds to form very stable charge-transfer
complexes with more highly electronegative aromatic compounds (Hall and
Poranski 1970) indicates that 2,6-dinitrotoluene should be strongly
adsorbed by both humus and clay.  In addition, the basic sites on the clay
surface might form addition-type complexes with this compound (Hall and
Poranski 1970).

    82,4.6  B ioaccumulat ion

         No environmentally relevant data on the bioaccumulation of 2,6-
dinitrotoluene have been found.  By analogy with nitrobenzene, bioaccumu-
lation may not be an important process; this contention is supported by the
relatively low value of the log octanol/water partition coefficient for
2,6-dinitrotoluene (log P = 2.05) compared to log P values of 5 to 6 or
more for strongly bioaccumulated compounds (PCBs, for example).

    82.4.7  Biotransformation andBiodegradation

         Detoxification of 2-nitrotoluene by dogs leads to urinary ex-
cretion of 2-nitrobenzyl alcohol and 2-nitrobenzoic acid (Venulet and
VanEtten 1970).  Similarly, 2,6-dinitrotoluene might be metabolized in an
analogous fashion.  Alexander and Lustigman (1966) report that dinitro-
benzenes are resistant to biodegradation by soil microorganisms.  It has
been reported that dinitrotoluenes are decomposed very slowly in a re-
servoir (Galuzova 1963).  Biodegradation by Azqtobacter has also been re-
ported to be slow (Bringmann and Kuehn 1972).

82.5  Data Summary

    Table 82-1 summarizes the aquatic fate discussed above for
2,6-dinitrotoluene.  Intramolecular photolysis and, to a small extent,
biodegradation may both be aquatic fates for this compound.  Adsorption
onto sediment probably plays a major role in transport and may also provide
reaction sites for destruction.  With the available data, however, it is
not possible to determine which fate predominates.
                                    82-5

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                      82-6

-------
82.6  Literature Cited

Alexander, M. and B.K. Lustigman.   1966.  Effect of chemical structure  on
  microbial degradation of substituted  benzenes.  J. Agr.  Food  Chem.
  14 (4) -.410-413.

Bringmann, G. and R. Kuehn.   1972.  Biological decomposition of
  nitrotoluenes and nitrobenzenes by azotobacter agilis.   Gesundh.  Ing.,
  92(9):273-276;  CA. 76:495l6f.  (Abstract only).

Fieser, L.F. and M. Fieser.   1956,  Organic chemistry,  3rd edition. D.C.
  Heath and Co., Boston, Mass.  1112p.

Galuzova, L.V.  1963.  Maximum  permissible concentration  of dinitrotoluene
  in the water  of reservoirs.   Gig. Sanit. 28(2):14-19.   (Abstract  only).

Hall, T.N. and  C.F. Poranski, Jr.   1970.  Polynitroaromatic addition
  compounds.  H. Feuer (ed).  The chemistry of the nitro  and nitroso
  groups.  Part 2.  Chap. 6.  pp.329-384.  Interscience Publishers, New
  York.

Johnson, K. and E.F. Degering.   1943.   Production of aldehydes  and  ketones
  from nitroparaffins.  J. Org.  Chem. 8(1):10-11.

Leighton, P.A.  and F.A. Lucy.   1934.  The photoisomerization of  the
  o-nitrobenzaldehydes.  J. Chem. Phys.  2:756-759.

Mackay, D. and  P.J. Leinonen.   1975.  Rate of evaporation of low-solubility
  contaminants  from water bodies to atmosphere.  Environ.  Sci.  Technol.
  9(13):1178-1180.

Maksimov, Y.Y.  1968.  Vapor  pressures  of aromatic nitro  compounds  at
  various temperatures.  Zh.  Fiz. Khim.  42(11):2921-2925;  CA.  1969.
  70:61315y.  (Abstract only).

Morrison, H.A.  1969.  The photochemistry of the nitro  and nitroso  groups.
  H. Feuer (ed).  The chemistry  of  the  nitro and nitroso  groups.  Part  1.
  Chap. 4.  pp.165-212.  Interscience Publishers, New York.

Russell, G.A., A.J. Moye, E.G.  Janzen,  S. Mak, and E.R. Talaty.   1967.
  Oxidation of carbanions.  II.  Oxidation of p-nitrotoluene and
  derivatives in basic solution.  J. Org. Chem. 32(1):137-146.

Shackelford,W.M.,  and L.H. Keith.   1976. Frequency of organic compounds
  identified in water.  U.S.  Environmental Protection Agency, (ERL),
  Athens, GA. 617p.   (EPA-600/4-76-062).
                                   82-7

-------
Tute, M.S.  1971.  Principles and practice of Hansch analysis:  a guide to
  structure-activity correlation for the medicinal chemist.  Adv. Drug
  Res. 6:1-77.

Vecchiotti, L.  and G. Zanetti.  1931.  Chemical reactions promoted by
  light.  Gazz. Chim. Ital.  61:789-802.

Venulet, J. and R..  Van Etten.  1970.  Biochemistry and pharmacology of the
  nitro and nitroso  groups.  H. Feuer (ed).  The chemistry of the nitro and
  nitroso groups.  Part 2.  Chap. 4. pp.201-287.  Interscience Publishers,
__ New York.

Verschueren, K.  1977.  Handbook of environmental data on organic
  compounds.  Van Nostrand/Reinhold, New York. 659p.

Weast, R.E. (ed).  1977.  Handbook of chemistry and physics.  58th Edition.
  CRC Press, Inc., Cleveland, Ohio. 2398p.

Wettermark, G.   1962.  Photochromism of o-nitrotoluenes.  Nature.
  194(4829):677.
                                    82-8

-------
                               83.   PHENOL
83.1  Statement of Probable Fate

    Photooxidation, metal-catalyzed oxidation,  and biodegradation probably
all contribute to the fate of phenol in the aquatic environment.   The dom-
inance of any of these destructive pathways depends upon the particular en-
vironmental conditions of the aqueous medium but the degradation products
are very similar for all fate pathways.  The first step usually involves
further hydroxylation of the aromatic ring followed by oxidation to a ben-
zoquinone and cleavage of the ring structure.  There is a possibility that
some of the phenol that is present in surface waters volatilizes  into the
atmosphere and is rapidly destroyed by oxidation in the troposphere.
Neither sorption nor bioaccumulation appear to be important processes in
the aquatic fate of phenol.

83.2  Identification

    Phenol has been detected in finished drinking water, surface waters,
and industrial effluents (Shackelford and Keith 1976).  The chemical struc-
ture of phenol is shown below.

                                             Alternate Names

                                             Carbolic acid
                                             Hydroxybenzene
                                             Phenyl hydroxide
                                             Phenic acid
                                             Phenyl hydrate
    Phenol

    CAS NO. 108-95-2
    TSL NO. SJ 33250

83.3  Physical Properties

    The general physical properties of phenol are as follows.

    Molecular weight                         94.11
    (Weast 1977)

    Melting point                            40.90C
    (Andon et_ al. 1960)

    Boiling point at 760 torr                181.75C
    (Weast 1977)
                                    83-1

-------
    Vapor pressure at 20C                   0.5293  torr*
    (Andon e al.  1960)

    Solubility in water  at 25C              93,000  mg/1
    (Morrison and Boyd 1973)

    Log octanol/water partition coefficient  1.46
    (MeCall 1975)

    pKa                                     10.02
    (Herington and Kynaston 1957)
*Vapor pressure of phenol as a supercooled liquid.

83.4  Summary of Fate Data

    83.4.1  Photolysis

         Phenol is a very weak acid, pKa = 10.02 (Herington and Kynaston
1957), and exists principally as its protonated, non-ionized form in en-
vironmental surface waters.  Coordination of the phenolic oxygen atom with
dissolved or suspended di- and trivalent metal cations,  however, can
markedly increase the ionization of the phenolic proton.  In the near ul-
traviolet spectral region the absorption maximum of undissociated phenol
occurs at 270 nm and does not extend beyond 290 nm.  The anion of phenol
has an absorption maximum at 287 nm which extends to 310 nm (Herington and
Kynaston 1957).  Any environmental photolysis of phenol  that could occur
would, therefore, probably involve the phenolic anion or photosensitization
of the undissociated form as an adsorbate.  It should be noted that com-
plexes of phenol with metal cations, such as iron (III), absorb light
strongly at about 600 nm (Ackermann and Hesse 1970).

         Solid or liquid phenol has long been known to form reddish high
molecular weight material when exposed to sunlight and air (Joschek and
Miller 1966).  A possible explanation for these observations could be the
formation and photolysis of an oxygen-phenol charge-transfer complex.
Joschek and Miller (1966) report that the steady irradiation of aqueous
solutions of phenol at 254 nm in the presence of oxygen yields isolable
amounts of 4,4'-dihydroxybiphenyl, 2,4'-dihydroxybiphenyl, 2,2'-dihydroxy-
biphenyl, hydroquinone, and catechol as well as many uncharacterized com-
pounds.  The latter two compounds predominate as products in the more
dilute solutions.  The intermediate that is postulated to explain this dis-
tribution of products is the phenoxyl radical.  If the phenoxyl radical can
also be produced by irradiation of phenol under environmental conditions,
it is expected that hydroquinone and catechol will be its main degradation
products.  In  the presence of photosensitizers that can transfer their
electronic excitation energy to molecular oxygen, the phenoxyl radical can
                                      83-2

-------
react at the C-2 and C-4 positions with an excited triplet oxygen molecule
to produce a phenoxylperoxy radical which decomposes directly to a 2- or
4-benzoquinone (Pfoertner and Bose 1970).

         Environmentally relevant substantiation for photooxidative degra-
dation has been reported by Perelshtein and Kaplin (1968) and Kinney and
Ivanuski (1969).  The former investigators used natural sunlight as a
source of radiation, whereas the latter investigators used commercially
available sun lamps.  Both found a gradual reduction of aqueous phenol that
could not be attributed to microbial degradation.  Perelshtein and Kaplin
(1968) proposed, on the basis of the ultraviolet absorption spectra of
their solution samples, that hydroquinone was being synthesized from the
phenol. The possible role of direct oxidation or volatilization in their
experiments was not taken into account.

         There is a possibility that volatilization of phenol will con-
tribute to its removal from water (see Section 83.4.4).  In the event that
some phenol does evaporate with water into the atmosphere, it can be ex-
pected that rapid photooxidation will take place in the troposphere.  Based
upon the smog chamber studies of Altshuller et al. (1962) and Laity et al.
(1973), the half-conversion times of m-xylene and toluene in a metropolitan
airshed are about four hours and twelve hours, respectively.  According to
Laity _ejt ad. (1973) aromatic substituents that increase a molecule's sus-
ceptibility to electrophilic attack will increase its rate of photodestruc-
tion in the atmosphere.  Inasmuch as the effect of a hydroxy group, in this
regard, is much greater than that of a methyl group (Morrison and Boyd
1973), it can be assumed that any phenol which does get into the tropo-
sphere will be destroyed within a few hours.

    83.4.2  Oxidation

         Hydroxylation of aqueous phenol at the C-2 position in the pres-
ence of air and iron(III) or copper(II) ions has been reported but at tem-
peratures and pressures far above what would be normally encountered in en-
vironmental surface waters (Makalets and Ivanova 1969).  In addition,
phenol has been oxidized by passing molecular oxygen into an aqueous solu-
tion at 25C and pH 9.5-13 (Kirso et al. 1967).  These observations, al-
though not environmentally relevant in themselves, raise the possibility
that phenol could be non-photolytically oxidized in highly aerated waters
that also contained iron and copper in solution or as part of the suspended
particulates.

    83.4.3  Hydrolysis

         There are no data to suggest that hydrolysis of phenol is an en-
vironmentally significant process.  The covalent bond of a substituent
attached to an aromatic ring is usually resistant to hydrolysis because of
                                    3j-3

-------
the high negative charge-density of the aromatic nucleus (Morrison and Boyd
1973).

    83.4.4  Volatilization

         The vapor pressure of supercooled liquid phenol at 20C is 0.5293
torr at 20C (Andon e_t al. 1960) and the solubility has been given as
93,000 mg/1 (Morrison and Boyd 1973).   A moderately low vapor pressure and
a high solubility usually imply that there is little tendency for volati-
lization from water.  Furthermore, it  can be expected that aqueous phenol
will be highly solvated which will increase its persistence in water at low
levels of concentration.  Nonetheless, it has been reported by Hakuta
(1975) that the vapor-liquid distribution ratio of phenol in water at a
concentration of 1 mg/1 is 1.8 at atmospheric pressure, thus making volati-
lization from surface waters a distinct possibility.  Moreover, when a thin
layer of montmorillonite, that had been previously saturated with gaseous
phenol, is exposed for one week to an atmosphere with a 40% relative humid-
ity, the phenol becomes almost completely desorbed from the clay (Saltzman
and Yariv 1975).

    83.4.5  Sorption

         Phenol has a log octanol/water partition coefficient of 1.46
(McCall 1975) and should, therefore, have only a slight tendency to become
sorbed onto the organic detrius.  Furthermore, phenol apparently would have
very little affinity for microcrystalline clay particulates in the aquatic
environment inasmuch as it can be almost completely desorbed from a thin
layer of montmorillonite that has been exposed for one week to the atmo-
sphere at 40% relative humidity (Saltzman and Yariv 1975).  From the data
of Chang and Anderson 1968, phenol also appears to be ineffective as a
flocculant of clays and soils.  This latter observation implies that phenol
does not form stable organic-inorganic aggregates in an aqueous medium.

    83.4.6  Bioaccumulation

         The log octanol/water partition coefficient indicates that phenol
should not be bioaccumulated to any extent in the aquatic environment.  A
review of the current literature revealed no information concerning the
bioaccumulation of phenol by aquatic microorganisms or by aquatic inverte-
brates or vertebrates.

    83.4.7  Biotransformation and Biodegradation

         The microbial degradation of phenol has been observed in many
laboratory studies in which phenol represented the primary carbon source
provided for isolated and adapted microorganisms.  Alexander and Lustigman
(1966) observed that phenol was degraded rapidly by a mixed population of
                                     83-4

-------
soil microorganisms.  Their data suggested that the hydroxy group, compared
to other benzene ring substituents, facilitated microbial degradation.
Bayly et al. (1966) reported that Pseudomonas putida converted phenol  to
catechol.  Buswell and Twomey (1975) and Buswell (1975) demonstrated the
oxidation of phenol by thermophilic bacteria, Bacillus stearothermo-
philus.  Other reports have documented the decomposition of phenol by
yeasts, including strains belonging to the genera Oospora, Sacaromycetes,
Candida, and Debaryomyces (Buswell 1975).  Neujahr and Varga (1970) re-
ported  the oxidation of phenol by both intact cells and cell extracts  of
the yeast, Trichosporon cutaneum.  The metabolic pathways involved in  the
microbial degradation of phenols have been well established (Neujahr and
Varga 1970;  Buswell 1975;  and Buswell and Twomey 1975).  The compound is
first converted to catechol and the aromatic ring is subsequently cleaved
to form 2-hydroxymuconic semialdehyde.

         Happold and Key (1932) were among the first to demonstrate the
bacterial degradation of phenol in phenolic wastes.  Baird et al. (1974)
examined the biodegradation of phenol under conditions simulating an aero-
bic biological sewage treatment system.  At concentrations of 1 mg/1 to 10
mg/1, phenol was biodegraded beyond the limits of detection.  At a concen-
tration of 100 mg/1, only 20% of the phenol was removed.  At concentrations
as low as 10 mg/1, however, phenol began to inhibit the oxygen uptake  of
the unacclimated sludge. These effects were greatly accentuated at 100
mg/1.  With long acclimation periods, activated sludge can be conditioned
to metabolize up to 500 mg/1 phenol without exhibiting toxic effects
(McKinney e_t al_. 1956).

         Biodegradation has been suggested as the mechanism for the decom-
position of phenol in natural waters (Streeter 1929;  Mischonsniky 1934;
Krombach and Barthel 1964;  Polisois e_t al_. 1975;  Wuhrraann 1972), and re-
cent studies have examined the importance of microorganisms in this
process.  Visser st_ &!._ (1977) conducted an in situ investigation of the
phenol-degrading activity of bacteria in river water.  Phenol (125 yg/1)
was added to containers holding large quantities of river water.  The
containers were incubated in the river along with sterilized controls.  The
removal rate of phenol was 30 JJg/1 per hour from the natural samples com-
pared to < 1 yg/1 per hour from the sterilized controls.

         Increasing the aeration of a natural system appears to enhance the
removal of phenol by microorganisms (Borighem and Vereecken 1978).
Borighem and Vereecken (1978) found that the presence or absence of light
had no influence on the biodegradation of phenol, but decreasing the con-
centration of phenol significantly reduced the lag time necessary to initi-
ate degradation and increased the rates of removal.

         In addition to aquatic microorganisms, the goldfish, Carrassius
auratus, also is reported to be able to biotransform phenol.  Kobayashi et
al. (1976a) found phenol sulfate to be present in all organs, particularly
the gallbladder.  Biliary excretion was suggested as the mechanism of
                                    83-5

-------
elimination in fish.  This supposition appears to be supported by data ob-
tained by Kobayashi e_t al. (1976b) who reported the _in situ sulfate conju-
gation of phenol by liver tissue.

    83.4.8  Other Reactions

         Chlorophenols may be produced inadvertently by chlorination re-
actions which take place during the disinfection of wastewater effluents or
drinking water sources.  Phenol has been reported to be highly reactive to
chlorine in dilute aqueous solutions over a considerable pH range (Carlson
et al. 1975).  The formation of 2- and 4-chlorophenol as well as more high-
ly chlorinated phenols such as 2,4- and 2,6-dichlorophenol and 2,4,6-tri-
chlorophenol has been reported under conditions similar to those employed
during the disinfection of wastewater effluents (Aly 1968, Barnhart and
Campbell 1972).  The synthesis of 2-chlorophenol took place in 1 hour in
aqueous solutions containing as little as 10 mg/1 and 20 mg/1 of phenol and
chlorine, respectively (Barnhart and Campbell 1972).

83.5  Data Summary

    Table 83-1 summarizes the aquatic fate data for phenol.  Photooxida-
tion, metal-catalyzed oxidation and biodegradation probably all contri-
bute to the aquatic destruction of this pollutant.  There is a possibility
that some volatilization into the atmosphere can occur.  Any phenol that
passes into the atmosphere would be rapidly destroyed by oxidation in the
troposphere.  Neither sorption nor bioaccumulation appear to be important
processes in the aquatic fate of phenol.
                                       83-6

-------
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83.6  Literature Cited

Ackermann, G. and D. Hesse.  1970.   tiber eisen(HI)-komplexe mit phenolen.
  III.  Die absorptionsspektren und deren auswertung.   Z.  Anorg. Allg.
  Chem.  375(l):77-86.

Alexander, M. and B.K. Lustigman.  1966.  Effect  of chemical structure  on
  raicrobial degradation of substituted benzenes.   J. Agr.  Food  Chem.
  U(4):440-413.

Altshuller, A.P., I.R. Cohen, S.F.  Sleva, and  S.L. Kopczynski,   1962.   Air
  pollution:  photooxidation of aromatic hydrocarbons.   Science
  138(3538):442-443.

Aly, O.M.  1968.  Separation of phenols in waters by thin-layer
  chromatography.  Water Res.  2:587-590.

Andon, R.J.L., D.P. Biddiscombe, J. D. Cox, R. Handley, D. Harrop,  E.F.G.
  Herington, and J.F. Martin.  1960.  Thermodynamic properties  of organic
  oxygen compounds.  Part 1.  Preparation and  physical  properties of  pure
  phenol, cresols, and xylenols.  J. Chera. Soc. (Lond.)  5246-5254.

Baird, R.B., C.L. Kuo, J.S. Shapiro, and W.A.  Yanko.  1974.  The fate  of
 .phenolics in wastewater-determination by direct injection GLC and Warburg
  respirometry.  Arch. Environ. Contam. Toxicol.   6:165-178.

Barnhart, E.L. and G.R. Campbell.  1972.  The  effect of chlorination    on
  selected organic chemicals.  Government Printing Office.  Water
  Pollution Control Research Series, 12020 EXG 03/72.   Washington,  D.C.
  103p.

Bayly, R.C., S. Dagley, and D.T. Gibson.  1966.  The metabolism of  cresols
  by a species of Pseudomonas.  Biochem. Jour. 101:293-301.

Borighem, G. and J. Vereecken.  1978.  Study of the biodegradation of
  phenol in river water.  Ecol. Model.  4:51-59.

Buswell, J.A.  1975.  Metabolism of phenols and cresols by Bacillus
  stearothermophilus.  J. Bact.  124(3):1077-1083 .

Buswell, J.A. and D.G. Twomey.  1975.  Utilization of phenol and cresols by
  Bacillus stearothermophilus.  Strain PH24.  J.  Gen.  Microbiol.
  87:377-379.

Carlson, R.M., R.E. Carlson, H.L. Kopperman, and  R. Caple.  1975.  Facile
  incorporation of chlorine into aromatic systems during aqueous
  chlorination processes.  Environ. Sci. Technol.  9(7):674-675.
                                      83-8

-------
Chang, C.W. and J.U Anderson.  1968.   Flocculation of  clays and  soils  by
  organic compounds.  Proc. Soil Sci.  Soc.  Amer.   32(1):23-27.

Hakuta, T.  1975.  Vapor-liquid equilibriums of pollutants  in sea  water.
  II.  Vapor-liquid equilibriums of phenolic substance water systems.
  Nippon Kaisui Gakkai-Shi   28(156):379-385.   (Abstract  only).  CA
  1976.  84:126558r.

Happold, F.C. and A. Key.   1932.  The  bacterial purification of  gas works
  liquors.  The action of  the liquors  on the bacterial flora of  sewage.
  J. Hyg.  32:573-577.

Herington, E.F.G. and W. Kynaston.  1957.  The ultraviolet  absorption
  spectra and dissociation constants  of certain phenols in  aqueous
  solution.  Trans. Faraday Soc.  53:138-142.

Joschek, H.I. and S.I. Miller.  1966.   Photooxidation  of  phenol, cresols,
  and dihydroxybenzenes.  J. Am. Chem. Soc.  88(14):3273-3281.

Kinney, L.C. and V.R. Ivanuski.  1969.  Photolysis mechanisms for  pollution
  abatement.  Robert A.  Taft Water Res. Cent. Rep., No.  TWRC-13.   41p.
  (Abstract only).  CA 1971.  75:67258g.

Kirso, U., K. Kuiv, and M. Gubergrits.  1967.   Kinetics of  phenol  and
  m-cresol oxidation by molecular oxygen in an aqueous medium.   Zh. Prikl.
  Khim.  40(7) =1583-1589.   (Abstract  only).  CA1968.   '68:12174b.

Kobayashi, K. , H. Akitake, and S. Kimura.  1976a.   Studies  on the
  metabolism of chlorophenols in fish:  VI.  Turnover  of  absorbed  phenol in
  goldfish.  Bull. Jap. Soc. Sci. Fish.  42(1):45-50.

Kobayashi, K., S. Kimura,  and H. Akitake.  1976b.   Studies  on the
  metabolism of chlorophenols in fish:  VII.  Sulfate  conjugation  of
  phenol and PCP in fish livers.  Bull. Jap. Soc.  Sci. Fish.
  42(2):171-176.

Krombach, H and J. Barthel.  1964.  Investigation  of a small watercourse
  accidentally polluted by phenol compounds.  Advan. Water  Pollut. Res.
  1:191-224.

Laity, J.L., I.G. Burstain, and B.R. Appel.  1973. Photochemical  smog and
  the atmospheric reactions of solvents.  Chap. 7, pp. 95-112.   Solvents
  Theory and Practice.  R.W. Tess (ed.)  Advances  in Chemistry Series  124.
  Am. Chem. Soc., Washington, D.C.

Makalets, B.I. and L.G. Ivanova.  1969.  Oxidation of  phenol by  atmospheric
  oxygen in aqueous solutions.  Neftekhimiya 9(2):280-285.   (Abstract
  only).  CA 1969.  71:29831y.
                                      83-9

-------
McCall, J.C.  1975.  Liquid-liquid partition coefficients by high-pressure
  liquid chromatography.  J. Med. Chem.  18(6):549-552.

McKinney, R.E., H.D. Tomlinson, and R.L.  Wilcox.   1956.   Metabolism of
  aromatic compounds by activated sludge.  Sewage Ind.  Wastes 28:547.

Mischonsniky, S.  1934.  A study of the pollution of fish-containing waters
  by waste phenolic waters.  14th Cong. Chem. Ind. (Paris, October 1934,
  Abstract only.)  J. Am. Water Works Assoc.  1937.  29:304.

Morrison, R.T. and R.N. Boyd.  1973.  Organic Chemistry,  3rd edition.
  Allyn and Bacon, Inc., Boston.  1258p.

Neujahr, H.Y. and J.M. Varga.  1970.  Degradation of phenols by intact
  cells and cell-free preparations of Trichosporon cutaneum.  Eur. J.
  Biochem. 13:37-44.

Perelshtein, E.I. and V.T. Kaplin.  1968.  Mechanism of  the
  self-purification of inland surface waters by the removal of phenol
  compounds.  II.  Effect of natural uv rays on aqueous  solutions of phenol
  compounds.  Gidrokhim.  Mater.  48:139-144.  (Abstract  only).  CA1969.
  71:73795p.

Pfoertner, K. and D. Bose.  1970.  Die photosensibilisierte oxydation
  einwertiger phenole zu chinonen.  Helv. Chim.  Acta 53(7):1553-1566.

Polisois, G., A. Tessier, P.G.C. Campbell, and J.P. Villeneuve.  1975.
  Degradation of phenolic compounds downstream from a petroleum refinery
  complex.  J. Fish. Res. Bd. Can.  32(11):2125-2131.

Saltzman, S. and S. Yariv.  1975.  Infrared study of the sorption of phenol
  and  p-nitrophenol by montmorillonite.  Proc. Soil Sci.  Soc. Amer.
  39(3):474-479.

Shackelford, W.M. and L.H. Keith.  1976.  Frequency of organic   compounds
  identified in water.  U.S. Environmental Protection Agency,   (ERL),
  Athens, Ga.  617p.  (EPA 600/4-76-062).

Streeter, H.W.  1929.  Chlorophenol tastes and odors in water supplies of
  Ohio River cities.  Am. J. Pub. Health 19(8):929-934.

Visser,  S.A., G. Lamontagne, V. Zoulalian, and A. Tessier.  1977.  Bacteria
  active  in  the degradation of phenols in polluted waters of the St.
  Lawrence River.  Arch. Environ. Contam. Toxicol.  6:455-469.
                                      83-10

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Weast, R.C. (ed.)  1977.  Handbook of chemistry and physics.  CRC Press,
  Inc., Cleveland, Ohio.  2398p.

Wuhnnann, K.  1972.  Stream purification.  In Mitchell, R. (ed.):  Water
  pollution microbiology.  Wiley Interscience, New York.  119p.
                                     83-11

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                            84.   2-CHLOROPHENOL
84. 1  Statement of Probable Fate

    From the information found in the reviewed literature,  it is not
possible to determine the most probable aquatic fate of this compound.   Mi-
crobial degradation and photolysis have been demonstrated,  but their en-
vironmental importance cannot be assessed from the available data.   One or
both of these processes probably would account for most of  the degradation
of 2-chlorophenol in an aqueous discharge.  Other fate processes are prob-
ably not important for this compound.  Similarly, the transport processes
of volatilization and sorption do not appear to have an overbearing effect
on removal of 2-chlorophenol.

84.2  Identification

    2-Chlorophenol has been detected in finished drinking water and in
industrial effluents (Shackelford and Keith 1976).
                                             Alternate Names
                                             o-Chlorophenol
     2-Chlorophenol

    CAS NO.  95-57-8
    TSL NO.  SK 26250

84.3  Physical Properties

    The general physical properties of 2-chlorophenol are as  follows.

    Molecular weight                         125.56
    (Verschueren 1977)

    Melting  point                            8.4C
    (Drahonovsky and Vacek 1971)

    Boiling  point at 760 torr                175.6C
    (Verschueren 1977)

    Vapor pressure at 20C                   2.2 torr (calculated)
                                    84-1

-------
    Solubility in water at 20C              28,500 mg/1
    (Verschueren 1977)

    Log octanol/water partition coefficient  2.17
    (Leo _e _al. 1971)

    PKa                                      8.52
    (Drahonovsky and Vacek 1971)

84.4  Summary of Fate Data

    84.4.1  Photolysis

         2-Chlorophenol is a weak acid, pka = 8.52 (Drahonovsky and Vacek
1971), and will exist both as an anion and as an undissociated phenolic
compound in environmental surface waters.   The undissociated compound in
cyclohexane does not absorb electromagnetic radiation above 290 nm (Sadtler
Standard Spectra 1975).  In aqueous solution, however, the artion has an
absorption maximum at 293 nm that extends  beyond 310 nm (Drahonovsky and
Vacek 1971).  It also should be noted that the aqueous iron(HI) complex of
2-chlorophenol has an absorption maximum at 550 nm (Ackermann and Hesse
1970).

         It is uncertain whether the reported photolytic experiments on
2-chlorophenol can be extrapolated for interpretation within an environ-
mental context.  Grabowski (1961) has reported that the photolysis of
2-chlorophenol in aqueous alkali at 313 nm results in the replacement of
chlorine by a hydroxy group.  Omura and Matsuura (1971) found that irra-
diation of 2-chlorophenol in aqueous alkali produced intractable tars both
at 254 nm and above 290 nm.  Although it is obvious that it is the anion of
2-chlorophenol that is undergoing photolysis, it is not clear whether the
resulting intermediate reacts with water or hydroxide ion.  Oraura and
Matsuura (1971) were able to demonstrate with 4-chlorophenol that when
cyanide ion was present in the photolysis solutions, some of the substi-
tuent chlorine was replaced by cyano groups.  This latter observation
supports a reaction mechanism involving anion interaction with the pho-
tolyzing carbon-chlorine bond.  Based on the results of these experiments,
the anion of 2-chlorophenol should be capable of undergoing photolysis
within the environment of ambient surface waters but it is not clear what
the resulting products would be.

         It is not known whether 2-chlorophenol can volatilize from water
into  the atmosphere.  In the event that some of this compound should
evaporate with water into the troposphere, it probably will undergo pho-
todegradation.  The atmospheric half-life of benzene, proposed by Darnall
et al. (1976), is 2.4 to 24 hours.  According to Laity et al. (1973), a
                                    84-2

-------
chlorine substituent on benzene should decrease its susceptibility to
photodegradation in the troposphere while a hydroxy group should facilitate
destruction.  What effect the presence of both groups would have on the '
atmospheric destruction of the aromatic ring is unknown.

    84.4.2  Oxidation

         Gunther e al. (1971) have demonstrated that hydroxyl radicals
attack 2-chlorophenol at the C-2 and C-4 positions resulting in the forma-
tion of a complex mixture.  No information was found, however, from which
an environmentally relevant rate could be estimated for this reaction.

    84.4.3  Hydrolysis

         There are no data to suggest that hydrolysis of 2-chlorophenol is
an environmentally significant process.  The covalent bond of a substituent
attached to an aromatic ring is usually resistant to hydrolysis because of
the high negative charge-density of the aromatic nucleus (Morrison and Boyd
1973).

    84.4.4  Volatilization

         The calculated vapor pressure of 2-chlorophenol, 2.2 torr at 20C,
is moderate and thus volatilization could be a potential transport process
for removal of this pollutant from surface waters.  The high solubility,
28,500 mg/1, of 2-chlorophenol, however, would increase its resistance to
volatilization at low concentrations in water.  Furthermore, acidic sub-
stances are usually highly solvated.  No specific data on the rate of
volatilization of 2-chlorophenol was found although it is surmised that
volatilization is not a competing removal process.

    84.4.5  Sorption

         The value of the log octanol/water partition coefficient (log
P=2.17) indicates a slight potential for sorption by lipophilic materials
in the sediment and particulates.  The only specific study of adsorption to
soils of a chlorinated phenol appears in Aly and Faust (1964).  Sorption of
2,4-dichlorophenol to kaolinite, Wyoming bentonite (a montmorillonitic
clay), and Fithian illite was measured.  Sorption to illite and bentonite,
which are more negatively charged than kaolinite, was considerably greater
than to the latter.  Bentonite has a larger specific surface area than the
other clays, and the order of decreasing sorption (bentonite, illite,
kaolinite) correlates with decreasing specific surface area.  The essential
conclusion is that sorption to sedimentary clays or suspended clays in sur-
face waters will not remove significant amounts of chlorinated phenols.
                                      84-3

-------
    84.4.6  Bioaccumulation

         No information concerning the bioaccumulation of 2-chlorophenol in
aquatic plants or animals was found.   Spencer and Williams (1950) investi-
gated the fate of oral doses of 2-chlorophenol administered to rabbits. The
results indicated that urinary excretion was rapid and probably rep-
resents the major route of elimination from mammals.   As a group, the
chlorophenols (2-chlorophenol, 2,4-dichlorophenol, 2,4,6-trichlorophenol,
and pentachlorophenol) are more likely to bioaccumulate as the number of
attached chlorine groups increases (Alexander and Aleem 1961).

    84.4.7  Biotransformation and Biodegradation

         In general, chlorophenols are more stable to biodegradation than
phenol, and resistance to microbial catabolism is greatest among the more
highly chlorinated phenols (Alexander and Aleem 1961).  Information found
in the reviewed literature on the biodegradability of 2-chlorophenol is
limited to laboratory studies.  At low concentrations (1-10 ppm) 2-chloro-
phenol is completely degraded by pure and mixed cultures of bacteria after
3-6 hours (Baird e^ _al. 1974;  Loos jet_ _al. 1967).  At 100 ppm, however,
only 20% of it is degraded (Baird e al. 1974).  2-Chlorophenol is probably
degraded by co-metabolism, although it does support the growth of common
aquatic bacteria (Knackmuss and Hellwig, 1978).  Natchtigall and Butler
(1974) report the microbial degradation of 2-chlorophenol by several
species of Pseudomonas and a species of Arthrobacter, all isolated from
various soils. The extrapolation of these laboratory conditions to
environmental situations may have very limited value.  Verschueren (1977)
reports:  (a)  a decomposition rate of 14 days for complete disappearance
of 2-chlorophenol in a soil suspension, and (b)  a decomposition period by
soil microflora of over 64 days.

         No studies were found that discussed what minimum levels of the
pollutant are required to induce catabolic pathways.  It is reasonable to 
suppose that genetic induction levels for most degradative organisms will
not be reached except in the vicinity of discharges.  Thus, stagnant re-
ceiving waters with stable, high levels of chlorinated phenols may estab-
lish and maintain a degradative microflora able quickly to degrade the
pollutant, provided oxygen is not limiting.  Rapid dilution in fast-flowing
receiving waters, on the contrary, will make establishment of an adapted
microflora much less likely and biodegradation should be slower, even with
greater aeration.

    84.4.8  Other Reactions

         Chlorophenols may be produced inadvertently by chlorination re-
actions which take place during the disinfection of wastewater effluents or
                                      84-4

-------
drinking water sources.  Ungubstituted phenol has been reported to be
highly reactive to chlorine in dilute aqueous solutions over a considerable
pH range (Carlson et al. 1975).  The formation of 2- and 4-chlorophenol as
well as more highly chlorinated phenols such as 2,4- and 2,6-dichlorophenol
and 2,4,6-trichlorophenol has been reported under conditions similar to
those employed during the disinfection of wastewater effluents (Aly 1968;
Barnhart and Campbell 1972).  2-Chlorophenol can serve as an intermediate
in the synthesis of the more highly chlorinated phenols (Morrison and Boyd
1973).

84.5  Data Summary

    Table 84-1 summarizes the aquatic fate data for 2-chlorophenol.  Micro-
bial degradation has been" well substantiated but the rates are dependent on
numerous variables.  Photolysis also has been demonstrated but its signifi-
cance in the environment is uncertain.  Other fate processes probably do
not compete to any signficant extent.  The transport processes of volatili-
zation and adsorption probably do not have an overbearing effect on the
removal of this pollutant.
                                      84-5

-------
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                                         84-6

-------
84.6  Literature Cited

Ackermann, G. and D. Hesse.  1970.  Uber eisen(lll)-komplexe mit phenolen.
  III.  Die absorptionsspektren und deren auswertung.   Z.   Anorg,  Allg.
  Chem. 375(l):77-86.

Alexander, M. and M.J.H. Aleem.  1961.  Effect of chemical structure on
  microbial decomposition of aromatic herbicides.  J.  Agr. Food Chem.
  9:44-47.

Aly, O.M.  1968.  Separation of phenols in waters by thin layer
  chromatography.  Water Res. 2:587-590.

Aly, O.M. and S.D. Faust.-  1964.  Studies on the fate  of 2,4-D and ester
  derivatives in natural surface waters.  J. Agr. Food Chem.
  12(6):541-546.

Baird, R.B., C.L. Kuo, J.S. Shapiro and W.A. Yanko.  1974.  The fate of
  phenolics in wastewater-determination by direct injection GLC and Warburg
  respirometry.  Arch. Environ. Contain. Toxicol.  6:165-178.

Barnhart, E.L. and G.R. Campbell.  1972.  The effect of chlorination on
  selected organic chemicals.  Government Printing Office.  Water  Pollution
  Control Research Series, 12020 EXG 03/72.  Washington, D.C.  103p.

Carlson, R.M., R.E. Carlson, H.L. Koppennan, and R. Caple.  1975.   Facile
  incorporation of chlorine into aromatic systems during aqueous
  chlorination processes.  Environ. Sci. Technol.  9(7) :674-675.

Darnall, K.R., A.C. Lloyd, A.M. Winer and J.N. Pitts,  Jr.   1976.
  Reactivity scale for atmospheric hydrocarbons based  on reaction  with
  hydroxyl radical.  Environ. Sci. Technol.  10(7):692-696.

Drahonovsky, J. and Z. Vacek.  1971.  Dissoziationskonstanten und
  austauscherchromatographie chlorierter phenole.  Coll. Czech. Chem.
  Commun.  36(10)=3431-3440.

Grabowski, Z.R.  1961.  Photochemical reactions of some aromatic halogen
  compounds.  Z. Physik. Chem.  27:239-243.

Gilnther, K. , W.G. Filby and K. Eiben.  1971.  Hydroxylation of substituted
  phenols:  an ESR-study in the Ti3+/H202 system.  Tetrahedron Lett.
  (3):251-254.

Knackmuss, H.J. and M. Hellwig.  1978.  Utilization and cooxidation of
  chlorinated phenols by Pseudomonas sp. B13.  Arch. Microbiol. 117:1-7.
                                      84-7

-------
Laity, J.L., I.G. Burstain, and B.R.  Appel.   1973.   Photochemical smog and
  the atmospheric reactions of solvents.   Chap.  7,  pp.  95-112.   Solvents
  Theory and Practice.  R.W. less (ed.).   Advances  in Chemistry Series 124.
  Am. Chem. Soc., Washington, D.C.

Leo, A., C. Hansch and D.  Elkins.  1971.   Partition coefficients and their
  uses.  Chem. Rev. 71:525-616.

Loos, M.A., J.M. Bollag, and M. Alexander.  1967.   Phenoxyacetate herbicide
  detoxication by bacterial enzymes.   J.  Agr.  Food  Chem.   15:(5):858-860.

Morrison, R.T. and R.N. Boyd.  1973.   Organic  Chemistry,  3rd edition.
  Allyn and Bacon, Inc., Boston.  1258p.

Nachtigall, H. and R.G. Butler.  1974.  Metabolism  of phenols  and
  chlorophenols by activated sludge microorganisms.  Abstr.  Annual Meet.
  Am. Soc. Microbiol.  74:184.

Omura, K. and T. Matsuura.  1971.  Photolysis  of halogenophenols in aqueous
  alkali and in aqueous cyanide.  Tetrahedron  27:3101-3109.

Sadtler Standard Spectra.   1975.  2-Chlorophenol.   Sadtler Research
  Laboratories, Inc., a Subsidiary of Block Engineering,  Inc.

Shackelford, W.M. and L.H. Keith.  1976.   Frequency of  organic  compounds
  identified in water.  U.S. Environmental Protection Agency,  (ERL),
  Athens, Ga.  617p.  (EPA 600/4-76-062) .

Spencer, B. and R.T. Williams.  1950.  Studies in detoxication.  The
  metabolism of halogenobenzenes.  A comparison  of  the  glucuronic acid,
  ethereal sulphate and mercapturic acid  conjugations of  chloro-, bromo-,
  and iodo-benzenes and of the o-,  m- and p-chlorophenols.  Biosynthesis of
  o-, m- and p-chlorophenylglucuronides.   Biochera.   J.   47:279-284.

Verschueren, K.  1977.  Handbook of environmental data  on organic
  chemicals.  Van Nostrand/Reinhold Co.,  New York.   659p.
                                      84-8

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                          85.   2,4-DICHLOROPHENOL
85.1  Statement of Probable Fate

    2,4-Dichlorophenol is readily biodegraded and probably does not bio-
accumulate in the aquatic environment.  Although photolysis of 2,4-di-
chlorophenol in the presence of a photosensitizer has been reported to
occur, the rate of photolysis has been observed to be slower than the rate
of microbial degradation.  Neither oxidation nor hydrolysis contributes
significantly to the aquatic fate of this pollutant, and neither volatili-
zation nor sorption is considered to be important in its transport.

85.2  Identification

    2,4-Dichlorophenol (2,4-DCP) has been detected in municipal and indus-
trial effluents, finished drinking water, and surface waters (Shackelford
and Keith 1976).  The chemical structure is shown below.
                                             Alternate Name
                                             2,4-DCP
    2,4-Dichlorophenol

    CAS NO.  120-83-2
    TSL NO.  SK 85750

85.3  Physical Properties

    The general physical properties of 2,4-dichlorophenol are as follows.

    Molecular weight                        163.0
    (Verschueren 1977)

    Melting  point                            45C
    (Verschueren 1977)

    Boiling  point at 760 torr                210C
    (Verschueren 1977)

    Vapor pressure at 20C                   0.12 torr (calculated)
                                     85-1

-------
    Solubility in water at 20C              4500 mg/1
    (Verschueren 1977)

    Log octanol water/partition coefficient  2.75
    (Tute 1971)

    pKa                                      7.85
    (Pearce and Simkins 1968)

85.4  Summary of Fate Data

    85.4.1  Photolysis

         Crosby and Tutass (1966) have reported the photodegradation of
aqueous 2,4-dichlorophenol when it is exposed to near ultraviolet or solar
radiation under conditions of good aeration.   After 10 days of solar irra-
diation, 2,4-dichlorophenol could no longer be detected in the solution and
most of it appeared to have been converted to an insoluble dark acidic
tarry substance.  Crosby and Tutass (1966) proposed that the tarry sub-
stance was formed by the polymerization of 2-hydroxybenzoquinone, an oxida-
tion product of 1,2,4-benzenetriol which results from the photolysis of the
carbon-chlorine bonds.

         Plimmer et al. (1971), however, found that irradation of 2,4-di-
chlorophenol with light at wavelengths above 280 nm induced a negligible
amount of photolysis.  No photoreaction was observed without the presence
of a photosensitizer, such as riboflavin.  The reaction products from
2,4-dichlorophenol, in the presence of riboflavin, were a mixture of tetra-
chlorophenoxyphenols and tetrachlorodihydroxybiphenyls.  None of the highly
toxic chlorinated dibenzodioxins could be detected.  Despite extensive
destruction of the 2,4-dichlorophenol, its conversion to dimeric products
was less than 5%.

         In a study of microbial degradation of this pollutant in samples
of natural lake water, Aly and Faust (1964) observed that photolysis was
insignificant compared to microbial catabolism.

    85.4.2  Oxidation

         Gunther et_ _a].. (1971) have demonstrated that hydroxyl radicals
attack 2-chlorophenol at the C-2 and C-4 positions resulting in the forma-
tion of a complex mixture.  No information was found, however, from which
an environmentally relevant rate could be estimated for this reaction.  It
can be inferred that 2,4-dichlorophenol could also undergo reactions of
this type but at a much slower rate (Morrison and Boyd 1973).
                                     85-2

-------
    85.4.3  Hydrolysis

         There are no data to suggest that hydrolysis of 2,4-dichlorophenol
is an environmentally significant process.  The covalent bond of a substi-
tuent attached to an aromatic ring is usually resistant to hydrolysis be-
cause of the high negative charge-density of the aromatic nucleus (Morrison
and Boyd 1973).

    85.4.4  Volatilization

         Compounds with a moderate solubility (4500 mg/1) and a low vapor
pressure (0.12 torr at 20C) generally do not volatilize from water.
Furthermore, 2,4-dichlorophenol is a weak acid (pKa = 7.85; Pearce and
Sirakins 1968) and will be about 50 percent ionized and very surely solvated
in environmental surface waters.

    85.4.5  Sorption

         The value of the log octanol/water partition coefficient (P=2.75)
indicates a slight potential for sorption by lipophilic materials in the
sediment and particulates.  The only specific study of 2,4-dichlorophenol
adsorption by soils appears in Aly and Faust (1964).  Sorption to kaoli-
nite, Wyoming bentonite (a montmorillonitic clay), and Fithian illite was
measured.  Sorption to illite and bentonite, which are more negatively
charged than kaolinite, was considerably greater than to the latter.  Ben-
tonite has a larger specific surface area than the other clays, and the
order of decreasing sorption (bentonite, illite, kaolinite) correlates with
decreasing specific surface area.  The essential conclusion is that sorp-
tion to sedimentary clays or suspended clays in surface waters will not
remove significant amounts of 2,4-dichlorophenol.

    85.4.6  Bioaccumulation

         Little information exists concerning the bioaccumulation of 2,4-
dichlorophenol.  Isensee and Jones (1971), using -^C-labelled 2,4-di-
chlorophenol, demonstrated that oats and soybean seedlings concentrated
2,4-dichlorophenol from dilute solutions (0.2 mg/1) by factors of 9.2X for
oats and 0.65X for soybean.  No further concentration took place during the
remaining 13 days of the experiment.  As a group, the chlorophenols are
more likely to bioaccumulate as the number of attached chlorine groups
increase.

    85.4.7  Biotransformation and Biodegradation

         Microbial decomposition of 2,4-dichlorophenol has been studied ex-
tensively in connection with work on the herbicide 2,4-D (2,4-dichloro-
                                    85-3

-------
phenoxyacetic acid).  Alexander and Aleem (1961) claimed complete dis-
appearance of 2,4-dichlorophenol in 5 or 9 days in two different silt loam
suspensions when the initial concentration was 50 mg/1.  Aly and Faust
(196A) studied 2,4-dichlorophenol oxidative degradation in samples of
natural lake water under laboratory conditions (pH 7, aeration, 25C).  An
initial concentration of 100 Ug/1 was completely eliminated in  9 days.
Concentrations of 500 and 1000 Ug/1 were 97.5 percent eliminated in 30
days.  The half-life of 2,4-dichlorophenol in their cultures was 6 days.
Adaptation of the cultures was apparently unnecessary and photolysis was
not rapid enough to be competitive.  (It should be noted that, these
observations could also be explained by the volatilization of this
pollutant.)  In a lake water culture simulating eutrophic conditions,  the
2,4-dichlorophenol was much more persistent with significant levels
remaining after 43 days.  The eutrophic cultures showed hydrogen sulfide
generation and a drop in pH.  In a non-aquatic rate study, Macrae _e_t al.
(1963) found that a soil bacterium (F la vo bag t e r i um sjp.) adapted to 2,4-D
butyl ester was able to completely oxidize 163 y.g/1 of 2,4-dichlorophenol
in 60 minutes.

         Soil bacteria that have been shown to carry out 2,4-dichlorophenol
oxidation include Pseudomonas, Achromobacter, Arthrobacter, Flavobacterium,
and mixed soil cultures (Alexander and Aleem 1961; Macrae et al. 1963;
Paris and Lewis 1973; Bollag _et _al. 1968; Chu and Kirsch 1972Ingols _e_t
al. 1966; Loos et_ al_. 1967; Tiedje et_ al. 1969).  Loos e_t al_. (1967) found
that Arthrobacter could methylate 2,4-dichlorophenol to form 2,4-dichloro-
anisole.  The significance of this chlorinated anisole as a biotransforma-
tion product has not been explored.  Chu and Kirsch (1972) showed that an
unidentified bacillus soil culture adapted to pentachlorophenol was also
readily able to degrade 2,4-dichlorophenol.  The dichlorophenol was 67 per-
cent oxidized by a suspension of the bacillus in 150 rain.  Ingols et al.
(1966) also observed 2,4-dichlorophenol degradation by pentachlorophenol-
adapted sludge.

         No studies of induction kinetics or of metabolic interactions in
mixed cultures were found in the reviewed literature.  Similarly, no
studies were found that discussed what minimum levels of the pollutant are
required to induce catabolic pathways.  It is reasonable to suppose that
genetic induction levels for most degradative organisms will not be reached
except in the vicinity of discharges.  Thus, stagnant receiving waters with
stable, high levels of 2,4-dichlorophenol will establish and maintain  a de-
gradative microflora able quickly to degrade the pollutant, provided oxygen
is not limiting.  Rapid dilution in fast-flowing receiving waters, on  the
contrary, will make establishment of an adapted microflora much less likely
and biodegradation should be slower, even with greater aeration.
                                      85-4

-------
    85.4.8  Other Reactions

         Chlorophenols may be produced inadvertently by chlorination re-
actions which take place during the disinfection of wastewater effluents or
drinking water sources.  Unsubstituted phenol has been reported to be
highly reactive to chlorine in dilute aqueous solutions over a considerable
pH range (Carlson _e_t _al. 1975).  The formation of 2- and 4-chlorophenol as
well as more highly chlorinated phenols such as 2,4- and 2,6-dichlorophenol
and 2,4,6-trichlorophenol has been reported under conditions similar to
those employed during the disinfection of wastewater effluents (Aly 1968;
Barnhart and Campbell 1972).  2,4-Dichlorophenol can serve as an interme-
diate in the synthesis of 2,4,6-trichlorophenol (Morrison and Boyd 1973).

85.5  Data Summary

    Table 85-1 summarizes the aquatic fate data for 2,4-dichlorophenol.
Microbial degradation has been well substantiated but the rates are depend-
ent on numerous variables.  Other fate processes probably do not compete to
any significant extent.  The transport processes of volatilization and
adsorption probably do not have an overbearing effect on the removal of
this pollutant.
                                   85-5


-------
85.6  Literature Cited

Alexander, M. and M.J.H. Aleem.  1961.  Effect of chemical structure on
  microbial decomposition of aromatic herbicides. J. Agr. Food Chem.
  9:44-47.

Aly, O.M.  1968.  Separation of phenols in waters by thin layer
  chromatography.  Water Res. 2:587-590.

Aly, O.M. and S.D. Faust.  1964.  Studies on the fate of 2,4-D and ester
  derivatives in natural surface waters.  J. Agr. Food Chem. 12(6):541-546.

Barnhart, E.L. and G.R."Campbell.  1972.  The effect of chlorination on
  selected organic chemicals.  Government Printing Office.  Water
  Pollution Control Research Series, 12020 EXG 03/72.  Washington, D.C.
  103p.

Bollag, J.M., C.S. Helling, and M. Alexander.  1968.  2,4-D metabolism:
  enzymatic hydroxylation of chlorinated phenols.  J. Agr. Food Chem.
  16(5):826-828.

Carlson, R.M., R.E. Carlson, H.L. Kopperman, and R.  Caple.  1975.   Facile
  incorporation of chlorine into aromatic systems during aqueous
  chlorination processes.  Environ. Sci. Technol.  9(7):674-675.

Chu, J.P. and E.J. Kirsch.  1972.  Metabolism of PCP by axenic bacterial
  culture.  Appl. Microbiol.  23(5):1033-1035.

Crosby, D.G. and H.O. Tutass.  1966.  Photodecomposition of 2,4-dichloro-
  phenoxyacetic acid.  J. Agr.  Food Chem. 14(6):596-599.

Giinther, K. , W.G. Filby and K.  Eiben.  1971.  Hydroxylation of substituted
  phenols:  an ESR-study in the Ti^+/H202 system.  Tetrahedron Lett.
  (3):251-254.

Ingols, R.S., P.E. Gaffney, and P.C. Stevenson.  1966.  Biological
  activity of halophenols.  J.  Water Pollut. Control Fed.  38:629-635.

Isensee, A.R., and G.E. Jones.   1971.  Absorption and translocation of root
  and foliage applied 2,4-dichlorophenol, 2,7-dichlorodibenzo-p-dioxin, and
  2,3,7,8-tetrachlorodibenzo-p-dioxin.  J. Agr. Food Chem. 19(6):1210-1214.

Loos, M.A., J.M. Bollag, and M. Alexander.  1967.  Phenoxyacetate herbicide
  detoxication by bacterial enzymes.  J. Agr. Food Chem. 15:(5):858-860.

Macrae, I.C., M. Alexander, and A.D. Rovira.  1963.   The decomposition of
  4-(2,4-dichlorophenoxy)butryic acid by Flavobacterium sp. J. Gen.
  Microbiol. 32:69-76.
                                     85-7

-------
Morrison, R.T. and R.N. Boyd.  1973.   Organic  Chemistry,  3rd edition.
  Allyn and Bacon, Inc.,  Boston.   1258 p.

Paris, D.F., and D.L. Lewis.  1973.   Chemical  and microbial  degradation of
  ten selected pesticides in aquatic  systems.   Residue  Rev.  45:95-123.

Pearce, P.J. and R.J.J. Simkins.   1968.  Acid  strengths of some substituted
  picric acids.  Can. J.  Chem.  46(2):241-248.

Plimmer, J.R., U.I. Klingebiel, D.G.  Crosby, and A.S. Wong.   1971.   Photo-
  chemistry of dibenzo-p-dioxins.   Chap.  6,  pp. 45-54.   Chlorodioxins-
  Origin and Fate.  E.H.  Blair (ed.)  Advances  in Chemistry Series 120.  Am.
  Chem. Soc., Washington, D.C.

Shackelford, W.M. and L.H. Keith.   1976.   Frequency of  organic compounds
  identified in water.  U.S. Environmental Protection Agency, (ERL),
  Athens, Ga. 617p. (EPA 600/4-76-062).

Tiedje, J.M., J. Duxbury, M. Alexander, and J.E. Dawson.   1969.  2,4-D
  metabolism:  pathway of degradation of  chlorocatechols  by Arthrobacter
  sp. J. Agr. Food Chem.   17:1021-1025.

Tute, M.S.  1971.  Principles and  practice of  Hansch analysis:  a guide to
  structure activity correlation for  the  medicinal chemist.   Adv. Drug
  Res . 6:1-77.

Verschueren, K.  1977.  Handbook of environmental data  on organic
  chemicals.  Van Nostrand/Reinhold Co.,  New York.  659p.
                                       85-8

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                        86.  2,4 ,6-TRICHLORQPHENOL


86.1  Statement of Probable Fate

    From the information that was found in the reviewed literature, it is
not possible to determine the most probable aquatic fate of this pollutant.
Microbial degradation and photolysis have been demonstrated, and, although
their significance in surface waters is uncertain, one or both of these
processes probably would account for most of the degradation of environ-
mental 2,4,6-trichlorophenol.  Other fate processes are probably not rele-
vant to the aquatic environment.  Similarly, the transport process of
volatilization probably does not have an overbearing effect on the removal
of 2,4,6-trichlorophenol.  Based on the log octanol/water partition co-
efficient, there is a definite potential for sorption by organic matter.

86.2  Identification

    2,4,6-Trichlorophenol has been detected in finished drinking water
(Shackelford and Keith 1976).  The chemical structure is shown below.

                                             Alternate Names

                                             No data found
    2,4,6-Trichlorophenol

   CAS NO.  88-06-2
   TSL NO.  SN 15750

86.3  Physical Properties

    The general physical properties of 2,4,6-trichlorophenol are as
follows.

    Molecular weight                         197.45
    (Verschueren 1977)

    Melting point                            68C
    (Verschueren 1977)

    Boiling point at 760 torr                244.5C
    (Verschueren 1977)

    Vapor pressure at 76.5C                 1 torr
    (Verschueren 1977)
                                   86-1

-------
    Solubility in water at 25C              800 mg/1
    (Verschueren 1977)

    Log octanol/water partition coefficient  3.38
    (Leo _et al. 1971)

    pKa                                      5.99
    (Drahonovsky and Vacek 1971)

86.4  Summary of Fate Data

    86.4.1  Photolysis

         2,4,6-Trichlorophenol is a moderately acidic substance, pKa =
5.99 (Drahonovsky and Vacek 1971), and will exist substantially as an anion
in environmental surface waters.  The ultraviolet absorption spectrum of
the anion exhibits a maximum absorption peak at 311 nm while the undisso-
ciated form has an absorption maximum at 286 nm (Drahonovsky and Vacek
1971).  It also should be noted that the iron(III) complex of 2,4,6-tri-
chlorophenol has an absorption maximum at 570 nm (Ackermann and Hesse
1970).

         In the presence of an electron acceptor, 2,4,6-trichlorophenol can
be photooxidized to 2,6-dichlorophenoxyl semiquinone radical anion (Leaver
1971).  The semiquinones thus formed disproportionate rapidly to 2,6-di-
chlorobenzoquinone and 2,6-dichlorohydroquinone.  No further information
which specifically pertained to the photolysis of 2,4,6-trichlorophenol was
found in the reviewed literature.

         The experimental photolysis of 2,4-dichlorophenol, however, has
been described.  For example, Crosby and Tutass (1966) have reported the
photodegradation of aqueous 2,4-dichlorophenol when it is exposed to near
ultraviolet or solar radiation under conditions of good aeration.  After 10
days of solar irradiation, 2,4-dichlorophenol could no longer be detected
in the solution and most of it appeared to have been converted to an in-
soluble dark acidic tarry substance.  Crosby and Tutass (1966) proposed
that the tarry substance was formed by the polymerization of 2-hydroxyben-
zoquinone, an oxidation product of 1,2,4-benzenetriol which results from
the photolysis of the carbon-chlorine bonds.

         In contrast, Plimmer _ejt al. (1971) found that irradiation of 2,4-
dichlorophenol with light at wavelengths above 280 nm induced a negligible
amount of photolysis.  No photo reaction was observed without the presence
of a photosensitizer, such as riboflavin.  The reaction products from
2,4-dichlorophenol, in the presence of riboflavin, were a mixture of tetra-
                                  86-2

-------
chlorophenoxyphenols and tetrachlorodihydroxybiphenyls.  In a study of the
microbial degradation of 2,4-dichlorophenol in samples of natural lake
water, Aly and Faust (1964) reported that photolysis was insignificant com-
pared to microbial catabolism.  It is uncertain to what extent these re-
ported photolytic experiments on 2,4-dichlorophenol can be extrapolated to
2,4,6-trichlorophenol for interpretation within an environmental context.

    86.4.2  Oxidation

         Giinther _e_t _al. (1971) have demonstrated that hydroxyl radicals
attack 2-chlorophenol at the C-2 and C-4 positions resulting in the forma-
tion of a complex mixture.  No information was found, however, from which
an environmentally relevant rate could be estimated for this reaction.  It
can be inferred that 2,4,6-trichlorophenol might also undergo reactions of
this type but at a much slower rate (Morrison and Boyd 1973).

    86.4.3  Hydrolysis

         There are no data to suggest that the hydrolysis of 2,4,6-tri-
chlorophenol is an environmentally relevant process.  The covalent bond of
a substituent attached to an aromatic ring is resistant to hydrolysis be-
cause of the high negative charge-density of the aromatic nucleus (Morrison
and Boyd 1973).

    86.4.4  Volatilization

         Compounds with an appreciable solubility (800 mg/1) and a low
vapor pressure (1 torr at 76.5C) generally do not volatilize from water.
Furthermore, 2,4,6-trichlorophenol is a moderately acidic substance
(pka = 5.99; Drahonovsky and Vacek 1971) and will be substantially
ionized and very surely solvated in environmental surface waters.

    86.4.5  Sorption

         The value of the log octanol/water partition coefficient (3.38)
indicates a definite potential for sorption by lipophilic materials in the
sediment and particulates.  The only specific study of adsorption to soils
of a chlorinated phenol appears in Aly and Faust (1964).  Sorption of
2,4-dichlorophenol to kaolinite, Wyoming bentonite ( a montmorillonitic
clay), and Fithian illite was measured.  Sorption to illite and bentonite,
which are more negatively charged than kaolinite, was considerably greater
than to the latter.  Bentonite has a larger specific surface area than the
other clays, and the order of decreasing sorption (bentonite, illite, and
kaolinite) correlates with decreasing specific surface area.  The essential
conclusion is that sorption to sedimentary clays in surface waters will not
move significant amounts of chlorinated phenols.  Sorption by the organic
detritus, however, may be important.
                                    86-3

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    86.4.6  Bioaccumulation

         No information was found concerning the bioaccumulation of
2,4,6-trichlorophenol.  Isensee and Jones (1971), using -^C-labelled
2,4-dichlorophenol demonstrated that oats and soybean seedlings were able
to concentrate it from dilute solutions (0.2 mg/1) by factors of 9.2x for
oats and 0.65x for soybean.  No further concentration took place during the
remaining 13 days of the experiment.  As a group, the chlorophenols are
more likely to become bioaccumulated as the number of attached chlorine
atoms increases.

    86.4.7  Biotransformation and Biodegradation

         Limited published information was found in the reviewed literature
on biodegradation of 2,4,6-trichlorophenol.  In flask cultures inocculated
with sludge bacteria, 7-10 days were required to remove 95% of the 2,4,6-
trichlorophenol at an initial concentration of 300 ppm (Tabak et al. 1964).
At lower concentrations (100 ppm) respirometry experiments indicated that
70% of the 2,4,6-trichlorophenol can be removed in 3 hours (Tabak et al.
1964).  In soil cultures 5-13 days were required for complete removal of
2,4,6-trichlorophenol.  Alexander and Aleem (1961) found in their experi-
ments that the time required for the complete disappearance of some
chlorophenols, including 2,4,6-trichlorophenol from various soil samples
ranged from 1 to 9 days.  Ingols _et _al. (1966) reported complete aromatic
ring degradation of 2,4,6-trichlorophenol within 5 days by microbial action
in an acclimated sludge.  Thus, it appears that 2,4,6-trichlorophenol is
susceptible to biodegradation but its fate in situ in aquatic systems re-
mains uncertain.

         No studies of induction kinetics or of metabolic interactions in
mixed cultures were found in the reviewed literature.  Similarly, no
studies were found that discussed what minimum levels of the pollutant are
required to induce catabolic pathways.  It is reasonable to suppose that
genetic induction levels for most degradative organisms will not be reached
except in the vicinity of discharges.  Thus, stagnant receiving waters with
stable, high levels of 2,4,6-trichlorophenol will establish and maintain a
degradative microflora able quickly to degrade the pollutant, provided oxy-
gen is not limiting.  Rapid dilution in fast-flowing receiving waters, on
the contrary, will make establishment of an adapted microflora much less
likely and biodegradation could, under these circumstances, be insignifi-
cant.

86.5  Data Summary

    Table 86-1 summarizes the aquatic fate for 2,4,6-trichlorophenol.
Microbial degradation has been demonstrated but the rates are dependent on
numerous variables.  Photolysis also has been demonstrated but its signi-
                                     86-4

-------
ficance in the environment is uncertain.  Other fate processes probably do
not compete to any relevant extent.  The transport process of volatiliza-
tion probably does not have an overbearing effect on the removal of this
pollutant, but sorption by organic matter may be important.
                                       86-5

-------
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                                        86-6

-------
86,6  Literature Cited

Ackermann, G. and D. Hesse.  1970.  Uber eisen (Ill)-komplexe mit
  phenolen.  III.  Die absorptionsspektren und deren auswertung.  Z.
  Anorg.  Allg. Chem.  375(1):77-86.

Alexander, M. and M.J.H. Aleem.  1961.  Effect of chemical structure on
  microbial decomposition of aromatic herbicides.  J. Agr. Food Chem.
  9:44-47.

Aly, O.M. and S.D. Faust.  1964.  Studies on the fate of 2,4-D and ester
  derivatives in natural surface waters.  J. Agr. Food Chem. 12(6):541-546.

Crosby, D.G. and H.O. Tutass.  1966.  Photodecomposition of 2,4-dichloro-
  phenoxyacetic acid.  J. Agr.  Food Chem. 14(6):596-599.

Drahonovsky, J. and Z. Vacek.  1971.  Dissoziationskonstanten und
  austauscherchromatographie chlorierter phenole.  Coll. Czech. Chem.
  Commun.  36(10):3431-3440.

Gunther, K. , W.G. Filby and K.  Eiben.  1971.  Hydroxylation of substituted
  phenols:  an ESR-study in the Ti3+/H202 system.  Tetrahedron Lett.
  (3):251-254.

Ingols, R.S., P.E. Gaffney, and P.C. Stevenson.  1966.  Biological activity
  of halophenols.  J. Water Pollut. Control Fed.  38:629-635.

Isensee, A.R. and G.E. Jones.  1971.  Absorption and translocation of root
  and foliage applied 2,4-dichlorophenol, 2,7-dichlorodibenzo-p-dioxin, and
  2,3,7,8-tetrachlorodibenzo-p-dioxin.  J. Agr. Food Chem. 19(6):1210-1214.

Leaver, I.H.  1971.  Semiquinone radical intermediates in the eosin-sensi-
  tized photooxidation of phenols.  Aust. J. Chem.   24(4):891-894.

Leo, A., C. Hansch and D. Elkins.  1971.  Partition coefficients and their
  uses.  Chem. Rev. 7:525-616.

Morrison, R.T. and R.N. Boyd.  1973.  Organic Chemistry, 3rd Edition.
  Allyn and Bacon, Inc., Boston.  1258p.

Plimmer, J.R., U.I. Klingebiel, D.G. Crosby and A.S. Wong.  1971.  Photo-
  chemistry of dibenzo-p-dioxins.  Chap. 6, pp. 45-54.  Chlorodioxins-
  Origin and Fate.  E.H. Blair (ed.) Advances in Chemistry Series 120.
  Am. Chem. Soc., Washington, D.C.
                                    86-7

-------
Shackelford, W.M. and L.H.  Keith.  1976.  Frequency of organic compounds
  identified in water.  U.S.  Environmental Protection Agency, (ERL),
  Athens, Ga.  617p. (EPA 600/4-76-062).

Tabak, H.H., C.W. Chambers, and P.W. Kabler.  1964.   Microbial metabolism
  of aromatic compounds.   I.   Decomposition of phenolic compounds and
  aromatic hydrocarbons by phenol-adapted bacteria.   J. Bacteriol.
  87(4):910-919.

Verschueren, K.  1977.  Handbook of environmental data on organic
  chemicals.  Van Nostrand/Reinhold Co.,  New York.  659p.
                                     86-8

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                           87.  PENTACHLOROPHENOL
87.1  Statement of Probable Fate

    Both photolysis and biodegradation appear to be important fate pro-
cesses in the aquatic environment for pentachlorophenol.  Photolytic des-
truction is probably rapid near the water-surface but as depth increases
microbial metabolism assumes greater importance.  Photolysis leads to a
mixture of chloranilic acid, chlorinated phenols, chlorinated dihydroxyben-
zenes and smaller non-aromatic compounds.  Microbial metabolism produces
pentachloroanisole and a mixture of chlorinated phenols.  Pentachlorophenol
is accumulated by a. large number of aquatic organisms and, in some cases,
appears to be bioconcentrated.  Although detoxification and depuration
involving formation of a sulfate ester conjugate has been demonstrated in
fish, the accumulation of pentachloroanisole in the fish of a natural aqua-
tic environment has also been documented.  Sorption by the organic matter
of sediments and soil definitely plays a role in the storage and transport
of this pollutant.  In a study of a natural freshwater lake, leaf litter
and other organic matter in the soil and sediments of the lake's watershed
retained relatively high concentrations of pentachlorophenol and its
degradation products, and this served as a source for continual pollution
of the aquatic ecosystem.  Hydrolysis, oxidation, and volatilization do not
appear to affect the environmental fate of pentachlorophenol.

87.2  Identification

    Pentachlorophenol has been detected in drinking water, rainwater, sur-
face waters and effluents (Bevenue _e_t al. 1972;  Fountaine ^t _al. 1976;
Pierce and Victor 1978; Rudling 1970; Shackelford and Keith 1976).  The
chemical structure is shown below.

                                            Alternate Names

                                             PCP
                                             Chlorophen
                                             Penchlorol
     Pentachlorophenol

     CAS NO. 87-86-5
     TSL NO. SM 63000
                                    87-1

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87.3  Physical Properties

    The general physical properties of pentachlorophenol are given below.

    Molecular weight                       266.35
    (Verschueren 1977)

    Melting point                            190C
    (Verschueren 1977)

    Boiling point at 760 torr                310C
    (Verschueren 1977)

    Vapor pressure at 20C                   0.00011 torr
    (Bevenue and Beckman 1967)

    Solubility in water at 20C              14 mg/1*
    (Verschueren 1977)

    Log octanol/water partition coefficient  5.01
    (Leo et al. 1971)

    PKa                                      4.74
    (Drahonovsky and Vacek 1971)
*Windholz (1976) lists the solubility as 80 mg/1.  In a natural water sys-
tem this pollutant will be present primarily as an anion and. its solubil-
ity, therefore, will be dependent on the cationic composition of the water.
The anion, however, will be much more soluble than the undissociated com-
pound under any circumstances.

87.4  Summary of Fate Data

    87.4.1  Photolysis

         Pentachlorophenol (PCP) is a moderately acidic substance, pka =
4.74 (Drahonovsky and Vacek 1971), and will exist primarily as an anion in
environmental surface waters.  The ultraviolet absorption spectrum of pen-
tachlorophenol in dilute aqueous solutions at pH 7 has major absorption
peaks at 245 run and 318 nm (Hiatt et al. 1960).  This is essentially the
absorption spectrum of the anion (Drahonovsky and Vacek 1971).

         Kuwahara e_t _al. (1966 ab) found that a 2% solution of the sodium
salt of pentachlorophenol was degraded to  the extent of 50% when exposed to
sunlight for ten days.  The major products were identified as chloranilic
                                     87-2

-------
acid (2,5-dichloro-3,6-dihydroxybenzoquinone), tetrachlororesorcinol, and
several complex chlorinated benzoquinones.  Hiatt _ejt _al.  (1960) calculated
the rate constant for  the aqueous  photochemical degradation of sodium pen-
tachlorophenolate at an initial concentration of 10 mg/1  and with a light
intensity of approximately 0.04 watts/cm^ between 290 and 330 nm.  A
first-order plot was obtained with k =  3.4 x 10"^ sec~^.  Using this
value  to estimate the  rates of destruction of sodium pentachlorophenolate
at various depths at noontime in a clear body of water on a midsummer day
at the latitude of Cleveland, Ohio, the approximate half-lives at 10 cm and
300 cm were 0.2 hr and 4.75 hr, respectively.

         Wong and Crosby (1978) found that the rate of photolysis of penta-
chlorophenolate anion  was much faster than that of the undissociated com-
pound.  At pH 7.3 degradation of pentachlorophenol at a concentration of
100 mg/1 was achieved  within five  to seven days by sunlight during the sum-
mer months in Davis, California whereas at pH 3.3 photolysis was much
slower.  From the data of Wong and Crosby (1978) an approximate half-life
for aqueous photolysis at Davis, California can be estimated as 1.5 days
during the summer months.  When the irradiation was interrupted at a point
where  50 to 75% of the pentachlorophenol had been degraded, three types of
degradation products were isolated:  chlorinated phenols, chlorinated
dihydroxybenzenes, and non-aromatic fragments.  The chlorinated phenols
were reported to consist primarily of 2,3,4,6- and 2,3,5,6-tetrachloro-
phenol together with a mixture of  trichlorophenols.  Tetrachlororesorcinol
and tetrachlorocatechol were isolated but tetrachlorohydroquinone was con-
sidered to be too unstable for isolation and probably was the main pre-
cursor for the principal non-aromatic degradation product, dichloromaleic
acid.  Plimmer _e_t _al.  (1971) and Stehl _e_t _al. (1971) report that octa-
chlorodibenzo-p-dioxin is produced in trace amounts during the photolysis
of pentachlorophenolate anion.

         Pierce and Victor (1978)  monitored the fate of pentachlorophenol
and its degradation products that  had been introduced into a freshwater
lake from an accidental release of wood-treating wastes which had been held
in a settling pond.  The origin of two  of the major degradation products,
2,3,5,6- and 2,3,4,5-tetrachlorophenol  was attributed to  photolytic de-
chlorination of pentachlorophenol  while it was still being held in the
settling pond along with hydrocarbon wastes.

    87.4.2  Oxidation

         Although no specific information was found pertaining to the
oxidation of pentachlorophenol, highly  chlorinated organic compounds are
usually resistant to oxidation even at  temperature extremes that could not
be reached in the aquatic environment (Morrison and Boyd  1973).  Therefore,
oxidation would not be expected to be an important fate under ambient con-
ditions.
                                    87-3

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    87.4.3  Hydrolysis

         The covalent bond of a substituent attached to an aromatic ring is
usually resistant to hydrolysis because of the high negative charge-density
of the aromatic nucleus.  For example,  the synthesis of pentachlorophenol-
ate anion from hexachlorobenzene requires treatment of the hexachloroben-
zene with concentrated alkali at 130-200C (Leoni and D'Arca 1976).  It can
be assumed that more extreme conditions would be necessary for further hy-
drolysis of the pentachlorophenolate anion.  Therefore, hydrolysis is not a
relevant fate for this pollutant.

    87.4.4  Volatilization

         Compounds with a moderate solubility in water and a very low vapor
pressure generally do not volatilize from water.  Furthermore, pentachloro-
phenol is & moderately acidic substance and will be substantially ionized
and very surely solvated.  Volatilization, therefore, is not considered to
be an operative transport process.

    87.4.5  Sorption

         The data of Hiatt _e_t _al. (1960) indicate that sorption of penta-
chlorophenol occurs principally on acidic soil systems with little or no
adsorption occurring on neutral soils.   Choi and Aomine (1974) have re-
ported that pH is the most important variable in controlling the adsorption
of pentachlorophenol onto soils and that the actual amount of sorption is
directly related to the organic content of the soil.  Based on these data
and the values of the pka and the log octanol/water partition coefficient
(pka = 4.74; log P = 5.01), sorption of dissolved pentachlorophenol by
suspended organic particulates in circumneutral water would not be expected
to occur.  Organic rich sediments that become somewhat acidic due to an-r
aerobic microbial digestion products could, however, be capable of sorbing
substantial amounts of pentachlorophenol.

         The observations of Pierce (1978) and Pierce and Victor (1978) on
the fate of two accidental spills of pentachlorophenol into a freshwater
lake have borne out these expectations.  Sediments and leaf litter retained
high concentrations of pentachlorophenol and its degradation products
throughout the period that the lake was monitored, and thus provided a
source for continuing pollution of the aquatic ecosystem.  Furthermore,
rainfall produced a chronic influx of pentachlorophenol by either leaching
of the contaminated soil  in the watershed area or by transporting leaf
litter into the lake.  Thus, sorption by the organic material of soil and
sediments apparently plays an important role in the storage and transport
of this  pollutant.
                                      87-4

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    87.4.6  Bioaccumulation

         The log octanol/water partition coefficient indicates that penta-
chlorophenol (PCP) should be bioaccumulated significantly in the aquatic
environment.  This supposition seems to be supported by data obtained in a
number of laboratory studies.  Weinbach and Nolan (1956) observed a 33-fold
and 50-fold concentration of pentachlorophenol in a species of snail after
24 and 30 hours, respectively, in a solution that contained 2 mg/1 of the
compound.  Guppies, (Lebistes reticulatus), killed by lethal doses of pen-
tachlorophenol within 18 hours, were found to contain approximately 100 u g
PCP/g wet weight of tissue (Stark 1969).

         The Japanese littleneck clam (Tapes phillipinarum) has also been
reported to accumulate pentachlorophenol (Kobayashi _et _al. 1970).  Seidel
(1974) observed that the compound can be absorbed from water by two species
of marsh plants, soft rush (Juncus effusus L.) and sea rush (J. maritimis
L.).

         Akitake and Kobayashi (1975) demonstrated that a 72-hour exposure
to a sublethal level of pentachlorophenol of 100 ]jg/l resulted in a 900-
fold concentration of the compound in the goldfish, Carassius auratus.   In
another study (Kobayashi and Akitake 1975), these authors investigated the
absorption at three concentrations (0.1, 0.2 and 0.4 mg/1) of pentachloro-
phenol by goldfish and found that the amount accumulated by the fish in-
creased with time.  They reported a concentration factor of 1000 after ex-
posure for 120 hr. in 0.1 mg/1 pentachlorophenol and observed a maximum
concentration of 116 yg PCP/g body weight in fish exposed to 0.2 mg/1 in
the water.  Lu and Metcalf (1975) studied the fate of radiolabeled penta-
chlorophenol in a model aquatic ecosystem with a six element food chain.
Pentachlorophenol was reported to be ecologically magnified
(bioconcentrated) in the mosquito fish, Gambusia affinis.

         Several investigations have documented the distribution of penta-
chlorophenol in the aquatic environment.  Rudling (1970) observed a 1000-
fold concentration of the compound in the eel, Anguilla anguilla, living in
lake water that had been contaminated with a 3 yg/1 concentration of pen-
tachlorophenol from pulp mill discharges.  Zitko _e_t al. (1974) carried out
a survey of levels of pentachlorophenol in estuarine fauna in New
Brunswick, Canada, and found low, but easily detectable amounts in most of
the samples (Table 87-1).

         Pierce and Victor (1978) studied the fate of pentachlorophenol in
a freshwater lake near Hattiesburg, Mississippi, after accidental spills of
wood-treating PCP-containing wastes in fuel oil.  Pentachlorophenol was
found to persist in the water and in fish for over six months following the
                                     87-5

-------
                                Table 87-1
                   Concentration of Pentachlorophenol in
                   estuarine fauna (Zitko et al.  1974).
                                  Sample weight       Pentachlorophenol
Sample                            	(g)	       (ng/g wet weight)

Cod                                  486.8                   0.82
Winter flounder                      157.8                   1.77
                                     128.2                   3.99
Sea raven                            875.9                  <0.5
Silver hake                          214.0                   1.75
Atlantic salmon                       50.3                   1.26
                                       8.5                   0.54
White shark liver                      -                    10.83
Double-crested cormorant egg          46.2                   0.36
Herring gull egg                      96.1                   0.51
                                     87-6

-------
spills.  Fish were also observed Co accumulate several of the degradation
products of PCP, namely, pentachloroanisole and the 2,3,5,6- and 2,3,4,5-
tetrachlorophenol isomers.  The bioaccumulation of pentachloroanisole is to
be expected, based upon its log octanol/water partition coefficient (log P
= 5.66; calculated by the method of Tute 1971).

         Pierce and Victor (1978) found that fish liver developed the high-
est concentration of pentachlorophenol followed by gill and muscle tissue.
Zitko e_t al. (1974) and Holmberg et_ al. (1972) likewise observed higher
levels of the compound in the liver compared to the muscle of marine fish
and the yellow eel.  Contrary to these observations,  Statham et al_. (1976)
found the highest bioconcentration of radiolabeled pentachlorophenol in the
rainbow trout (Salmo gairdneri) to occur in the gill. Kobayashi and Akitake
(1975), however, reported that the highest concentration of pentachloro-
phenol was found in the gall bladder of fish.

         Although a rapid uptake of pentachlorophenol has been observed in
many organisms, depuration also appears to be significant.  Kobayashi and
Akitake (1975) reported that goldfish excreted 50 percent of absorbed pen-
tachlorophenol within 10 hours after transfer to fresh running water.  The
concentration decreased to approximately 20 percent of initial levels after
20 hours.  Holmberg e_t aJ._. (1972) observed that pentachlorophenol in the
muscle of the yellow eel, (Anguilla anguilla) decreased from 9.4 ug/g of
tissue to 3.6 yg/g after 8 days in clean water.

    87.4.7  Biotransformation and Biodegradation

         The microbial degradation of chlorophenols under laboratory condi-
tions has been discussed by several investigators (Alexander and Aleem
1961; Watanabe 1973; Nachtigall and Butler 1974; Gee and Peel 1974).  In
general, chlorophenols are more resistant to biological degradation than
phenols.  Highly chlorinated phenols, particularly those with halogens at
the meta position on the ring, are the most resistant to degradation
(Alexander and Aleera 1961).

         The microbial degradation of pentachlorophenol, however, has been
observed in several laboratory investigations in which the compound was
used as the sole carbon source provided to microorganisms,  Chu and Kirsch
(1972) observed the oxidation of pentachlorophenol by bacteria obtained
from a continuous flow enrichment culture.  Watanabe (1973) reported on the
ability of Pseudomonas sp., isolated from PCP-saturated soil cultures, to
grow in the presence of 40 ppm pentachlorophenol as the sole carbon source.
The data presented by Kirsch and Etzel (1973) also indicates that penta-
chlorophenol is amenable to microbial degradation under laboratory con-
ditions.  These investigators found that the rate of decomposition is de-
pendent on a number of variables, particularly the bacterial concentration
and the amount of aeration.
                                      87-7

-------
         Microbial decomposition of pentachlorophenol in soil and axenic
cultures has been studied by several investigators (Alexander and Aleem
1961; Kuwatsuka 1972; and Kirsch and Etzel 1973) who have reported the
presence of two decomposition products, a tri- and a tetrachlorophenol that
were relatively stable to further decomposition.  Pierce and Victor (1978),
in one of the few studies to investigate the fate of pentachlorophenol in a
natural aquatic environment, documented the presence of pentachloroanisole
and the 2,3,5,6- and 2,3,4,5-tetrachlorophenol  isomers as major degrada-
tion products.  Cserjesi and Johnson (1972) observed the methylation of
pentachlorophenol to pentachloroanisole' by the fungi, Trichoderma virgatum.
They noted, however, that the formation of pentachloroanisole did not
account for the total reduction in the concentration of pentachlorophenol
in the growth medium and concluded that methylation is either the first
step in the metabolism of this compound or a reaction parallel to degra-
dation.

          Pentachlorophenol has been shown to be detoxified by fish and
other aquatic organisms in several laboratory investigations.  Akitake and
Kobayashi (1975) found that the goldfish, Carassius auratus, transformed
pentachlorophenol to a pentachlorophenyl sulfate, which was identical to
that found in the littleneck clam (Kobayashi _et _al. 1970).  Lu and Metcalf
(1975) also noted that conjugation at the phenolic hydroxy group was the
most important detoxification mechanism among the organisms they studied in
their model aquatic ecosystem.  Studies in man indicate that the primary
route for removal of pentachlorophenol in mammals, including man, is urin-
ary excretion of a conjugate form (Jacobson and Yelner 1971; Braun and
Sauerhoff 1976).

87.5  Data Summary

    Table 87-2 summarizes the aquatic fate data for this pollutant.  Pho-
tolysis and degradation appear to be highly effective fate processes for
pentachlorophenol.  This pollutant does not appear to be persistent in the
aqueous medium itself but it is definitely sorbed by the organic matter of
soils and sediments in a freshwater ecosystem.  Moreover, it has been shown
to be bioaccumulated by numerous aquatic organisms.  The persistence of de-
gradation products may be important and one of them, pentachloroanisole,
probably has a greater tendency for bioaccumulation than the pollutant it-
self.  Oxidation, hydrolysis, and volatilization are probably of very
little environmental consequence.
                                    87-8

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87.6  Literature Cited

Akitake, H. and K. Kobayashi.  1975.  Mecabolism of chlorophenols in fish.
  III.  Isolation and identification of a conjugated PCP (pentachloro-
  phenolate) excreted by goldfish.  Bull. Jap. Soc. Sci. Fish.
  41(3)-.321-327.

Alexander, M. and M.I.H. Aleem.  1961.  Effect of chemical structure on
  microbial decomposition of aromatic herbicides.  J. Agr. Food Chem.
  9:44-47.

Bevenue, A. and H. Beckman.  1967.  Pentachlorophenol: a discussion of its
  properties and its occurrence as a residue in human and animal tissues.
  Residue Rev. 19:83-134.

Bevenue, A., J.N. Ogata, and J.W. Hylin.  1972.  Organochlorine pesticides
  in rainwater, Oahu, Hawaii, 1971-1972.  Bull. Environ. Contain. Toxicol.
  8(4):238-241.

Braun, W.H. and M.W. Sauerhoff.  1976.  The pharmacokinetic profile of
  pentachlorophenol in monkeys.  Toxicol. Appl. Pharmacol. 38:525-533.

Choi, J. and A. Aomine.  1974.  Adsorption of pentachlorophenol by soils.
  Soil Sci. Plant Nutr. (Tokyo). 20(2):134-144.

Chu, J.P. and E.J. Kirsch.  1972.  Metabolism of pentachlorophenol by an
  axenic bacterial culture.  Appl. Microbiol. 23(5): 1033-1035.

Cserjesi, A.J. and E.L. Johnson.  1972.  Methylation of pentachlorophenol
  by Trichoderma virgatum.  Can. J. Microbiol. 18:45-49.

Drahonovsky, J. and Z. Vacek.  1971.   Dissoziationskonstanten und
  austauscherchromatographie chlorierter phenole.  Coll. Czech. Chem.
  Commun. 36(10):3431-3440.

Fountaine, J.E., P.B. Joshipura, and P.N. Keliher.  1976.  Some
  observations regarding pentachlorophenol levels in Haverford Township,
  Pennsylvania.  Water Res. 10(3) .-185-188.

Gee, J.M. and J.L. Peel.  1974.  Metabolism of 2,3,4,6-tetrachlorophenol by
  microorganisms from  broiler house litter.  J. Gen. Microbiol.
  85:237-243.

Hiatt, C.W., W.T. Haskins, and L. Olivier.  I960.  The action of sunlight
  on sodium  pentachlorophenate.  Am. J. Trop. Med. Hyg. 9:527-531.
                                   87-10

-------
Holmberg, B.S., S. Jensen, A. Larsson, K. Lewander,  and M.  Olsson.   1972.
  Metabolic effects of technical PGP on the eel Anguilla anguilla L.  Comp.
  Biochem. Physiol. 43(B):171-183.

Jacobson, I. And S. Yelner.  1971.   Metabolism of -^C-pentachlorophenol
  in the mouse.  Acta. Pharmacol. Toxicol, 29:513-521.

Kirsch, E.J. and J.E. Etzel.  1973.  Microbial decomposition of penta-
  chlorophenol.  J. Water Pollut. Control Fed. 45(2)-.359-364.

Kobayashi, K. and H. Akitake.  1975.  Studies on the metabolism of
  chlorophenols in fish - I.  Absorption and excretion of PCP by goldfish.
  Bull. Jap. Soc. Sci. Fish.  41(l):87-92.

Kobayashi, K., H. Akitake, and T. Tomiyama.  1970.  Studies on the
  metabolism of PCP, a herbicide in aquatic organisms - III.  Isolation and
  identification of a. conjugated PCP yielded by a shellfish, Tapes
  phillippinarum.  Bull. Jap. Soc.  Sci. Fish.  36:103-108.

Kuwahara, M., N. Kato, and K. Munakata.  1966a.  The photochemical  reaction
  of pentachlorophenol.  Part I.  The structure of the yellow compound.
  Agr. Biol. Chem. (Tokyo). 30(3):232-238.

Kuwahara, M., N. Kato, and K. Munakata.  196bb.  The photochemical  reaction
  of pentachlorophenol.  Part II.  The chemical structures  of minor
  products.  Agr. Biol. Chem. (Tokyo). 30(3):239-244.

Kuwatsuka, S.  1972.  Degradation of several herbicides in soils under
  different  conditions.  pp.385-400.  In:  Environmental Toxicology of
  Pesticides.  Academic Press, New York.

Leo, A., C.  Hansch and D. Elkins.  1971.  Partition coefficients and their
  uses.  Chem. Rev.  7:525-616.

Leoni, V. and S.U. D'Arca.  1976.  Experimental data and critical review of
  the occurrence of hexachlorobenzene in the Italian environment.   Sci.
  Total Environ.  5:253-272.

Lu, P. Y. and R.L. Metcalf.  1975.   Environmental fate and  biodegradability
  of benzene derivatives as studied in a model aquatic system.  Environ.
  Health Perspect.  10:269-284.

Morrison, R.T. and R.N. Boyd.  1973.  Organic chemistry, 3rd edition.
  Allyn and  Bacon, Boston, Mass.  1258p.

Nachtigall,  M.H. and R.G. Butler.  1974.  Metabolism of phenols and
  chlorophenols by activated sludge microorganisms.  Abstr.  Annual Meet.
  Amer. Soc. Microbiol. 74:184.
                                   87-11

-------
Pierce, R.H.,  Jr.  1978.   Fate and impact of pentachlorophenol in a
  freshwater ecosystem.  U.S.  Environmental Protection Agency, (ERL),
  Athens, Ga.   61p.  (EPA 600/3-78-063).

Pierce, R.H. Jr., and D.M. Victor.  1978.  The fate of pentachlorophenol in
  an aquatic ecosystem,  pp.41-52.  In K.R. Rao (ed.).  Pentachlorophenol:
  Chemistry, Pharmacology and  Environmental Toxicology.   Plenum Press,
  New York. 402p.

Plimmer, J.R., U.I. Klingebiel, D.G. Crosby, and A.S.  Wong.   1971.
  Photochemistry of dibenzo-p-dioxins.  Chap. 6, pp.  45-54.  Chloro-
  dioxins-Origin and Fate.  E.H. Blair (ed.) Advances  in Chemistry Series
  120.  American Chemical Society, Washington, D.C.

Rudling, L.  1970.  Determination of pentachlorophenol in organic tissues
  and water.  Water Res.  4:533-537.

Shackelford, W.M. and L.H. Keith.  1976.   Frequency of organic compounds
  identified in water.  U.S. Environmental Protection Agency,  (ERL),
  Athens, Ga.  617p. (EPA 600/4-76-062).

Seidel, K.  1974.  Elimination of PCP from water bodies by plants.
  Naturwissenschaften 61(2) :81.

Stark, A.  1969.  Analysis of  pentachlorophenol residues in soil, water,
  and fish.  J. Agr. Food Chem. 17(4):871~873.

Statham, C.N., M.J. Melancon,  and J.J. Lech. 1976.   Bioconcentration  of
  xenobiotics  in trout bile:   a proposed  monitoring aid for some waterborne
  chemicals.  Science 193(4254).-680-681.

Stehl, R.H., R.R. Papenfuss,  R.A. Bredeweg, and R.W.  Roberts.   1971.   The
  stability of pentachlorophenol and chlorinated dioxins to sunlight,  heat,
  and combustion.  Chap.  13.   pp. 119-125.  Chlorodioxins-Origin and  Fate.
  E.H. Blair (ed.).  Advances  in Chemistry Series 120.  American Chemical
  Society, Washington, D.C.

Tute, M.S. 1971.  Principles  and practice of Hansch analysis:   a guide to
  structure - activity correlation for the medicinal chemist.   Adv. Drug
  Res. 6:1-77.

Verschueren, K.  1977.  Handbook of environmental data on organic
  compounds.  Van Nostrand/Reinhold, New York.  659p.

Watanabe, I.  1973.  Decomposition of pesticides by soil microorganisms.
  Jap. Agr. Res. Quart.  7:15-19.
                                   87-12

-------
Weinbach.  E.G. and M.O. Nolan.  1956.   Effect of pentachlorophenol on the
  metabolism of the snail, Australarbis glabratus.  Exper.  Parasitol.
  5:276-284.
Windholz, M.
  1313p.
(ed.).   1976.   The Merck Index.   Merck and Co.  Rahway,  N.J
Wong, A.S. and D.B. Crosby.  1978.  Photolysis of pentachlorophenol in
  water.  pp.19-25.  _In K.R. Rao (ed.).  Pentachlorophenol:   Chemistry,
  Pharmacology, and Environmental Toxicology.  Plenum Press,  New York.
  402p.

Zitko, V., 0. Hutzinger, and P.M.K. Choi.  1974.   Determination of
  pentachlorophenol and chlorobiphenylols in biological samples.  Bull.
  Environ. Contam. Toxicol.  12(6) :649-653.
                                   87-13

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                            88.   2-NITROPHENOL
88.1  Statement of Probable Fate

    Based on the information gathered for 2-nitrophenol,  4-nitrophenol,  and
2,4-dinitrophenol, 2-nitrophenol will probably undergo slow photooxidation
in an aerated aquatic environment.  There is a possibility for photoreduc-
tion of the nitro group if the 2-nitrophenol becomes absorbed by organic
particulates.  2-Nitrophenol should be strongly sorbed by clay minerals  and
may even undergo hydrolysis within the clay structure.  Volatilization and
bioaccumulation appear to be unlikely processes, and although biotransfor-
mation of 2-nitrophenol has been demonstrated, the nitrophenols are very
persistent in aqueous mixed cultures and will inhibit microbial growth in
natural aquatic systems into which they are introduced.

88.2  Identification

    2-Nitrophenol has been detected in river water and in industrial efflu-
ents (Shackelford and Keijth 1976).  The chemical structure of 2-nitrophenol
is shown below.
                                             Alternate Names

                                             o-Nitrophenol
                                             2-Hydroxynitrobenzene
    2-Nitrophenol

    CAS NO. 88-75-5
    TSL NO. SM 21000

88.3  Physical Properties

    The general physical properties of 2-nitrophenol are as follows.

    Molecular weight                        139.11
    (Weast 1977)

    Melting point                            45.3
    (Weast 1977)

    Boiling point at 760 torr                216C
    (Weast 1977)
                                    88-1

-------
    Vapor pressure at  49.3C                 1.0  torr
    (Weast 1977)

    Solubility in water at  20C              2,100  mg/1
    (Verschueren 1977)

    Log octanol/water  partition coefficient   1.76
    (Leo et_ al.  1971)

    PKa                                      7.21
    (Pearce and Simkins 1968)

88.4  Summary of Fate  Data

    88.4.1  Photolysis

         2-Nitrophenol is a somewhat acidic  substance, pKa = 7.21 (Pearce
and Simkins 1968), and will exist  to an appreciable extent as an anion in
environmental surface  waters.   The ultraviolet absorption spectrum of
2-nitrophenol in methanol exhibits a maximum at about 345 run which extends
out beyond 400 nm (Sadtler Standard Spectra  1975).   The  intensity of
absorption increases markedly  near 410 nm in basic  solutions.

         Nakagawa and  Crosby (1974) report that 4-nitrophenol was degraded
in aqueous solutions within a  period of 1-2  months  when  it was exposed to
sunlight at a concentration of 200 mg/1.   The principal  products were hy-
droquinone and 4-nitrocatechol.  A dark,  acidic intractable polymer was
also produced.  Although no specific information was found in the reviewed
literature demonstrating that  2-nitrophenol  would also be photochemically
hydroxylated in an analogous manner, the following  mechanistic considera-
tions for this photochemical reaction indicate that, under similar condi-
tions, 2-nitrophenol would give rise to catechol and 2-nitrohydroquinone.

         Nakagawa and  Crosby (1974) proposed that the observed reaction
products arising from  4-nitrophenol were the result of photonucleophilic
displacement reactions involving water or hydroxide ion.   This type of re-
action mechanism could be rationalized for the formation of hydroquinone,
but it is probably not valid in the case of  4-nitrocatechol since it is
difficult to envision  the development of a center of electron deficiency on
the unsubstituted carbon atom  adjacent to the phenolic group.  Suarez et_
al. (1970) and Giinther et_ a_l.  (1971) have demonstrated that hydroxyl
radicals preferentially attack 4-nitrophenol in water at  the C-2 and C-4
positions, and that the latter intermediate  results in the displacement of
the nitro group by a hydroxy group.  Although attack by  hydroxyl radical
can more easily explain the observed products, it is uncertain whether the
concentration of hydroxyl radicals in the experiment of  Nakagawa and Crosby
                                      3-2

-------
(1974) would have been sufficient to be responsible for the formation of
hydroquinone and 4-nitrocatechol.

         It has been reported that the ultraviolet irradiation of indivi-
dual nitrophenolic compounds at 254 nm was without effect (Mitchell 1961).
This experiment, however, was conducted by irradiating the nitro sub-
stituted phenol for 30 minutes after it had been applied in solution to a.
sheet of chromatographic paper which was then dried.  The experiment was
one of very short duration and the conditions were anhydrous or hypo-
hydrous.  For these reasons it is probably not valid to draw any conclu-
sions from this experiment with respect to either the aquatic or terres-
trial environment.

         A further photochemical reaction of 2-nitrophenol that must be
considered is photoreduction of the nitro group.  As an example, Nakagawa
and Crosby (1974) found that the nitro group of nitrofen, a nitroaromatic
pesticide, underwent photochemical reduction to amino and azo groups in an
aqueous solution that contained 10% raethanol.  Aromatic nitro groups are,
in general, reduced photochemically in the presence of suitable hydrogen
donors (Morrison 1969).  Since the log octanol/water partition coefficient
of 2-nitrophenol indicates a slight potential for absorption by aquatic or-
ganic matter,.this organic material could serve as a reducing agent in the
photoreduction of 2-nitrophenol to 2-aminophenol and 2,2'-dihydroxyazo-
benzene.  No specific information was found pertaining to photoreduction of
2-nitrophenol in the reviewed literature.

    88.4.2  Oxidation

          Suarez _e_t ad. (1970) and Gu'nther e_t a^. (1971) have demonstrated
that hydroxyl radicals attack. 2-nitrophenol at the C-2 and C-4 positions
resulting in the formation of a complex mixture that follows from the
generation of 1,2-benzosemiquinone.  No information was found, however,
from which an environmentally relevant rate could be estimated for this
reaction.

    88.4.3  Hydrolysis

         The covalent bond of a substituent attached to an aromatic ring is
usually resistant to hydrolysis because of the high negative charge-density
of the aromatic nucleus.  Nonetheless, there is a possibility, albeit not
specifically substantiated in the reviewed literature, that nitrophenols
could undergo hydrolysis while sorbed by clay minerals.  Specifically,
4-nitrophenol is not only adsorbed by montmorillonite but also becomes
intercalated within the clay structure (Saltzman and Yariv 1975), and there
is no apparent reason to believe that 2-nitrophenol would behave differ-
ently.  The water molecules within the clay structure can be highly acidic,
                                      3-3

-------
and the phenolic group of the intercalated 2-nitrophenol would become more
acidic than when the molecule is in aqueous solution due to coordination of
the phenolic oxygen atom with the metal cations of the clay.  This acidic
environment could conceivably lead to the hydrolysis of the 2-nitrophenol
anion in a manner similar to the hydrolysis of aci-nitroparaffins (Johnson
and Degering 1943).  The resulting compound would be 1,2-benzoquinone.
                                                          = 0
    88.4.4  Volatilization

         Compounds which exert high vapor pressures are generally con-
sidered volatile and thus have a greater tendency to enter the atmosphere
from an aqueous system than do those compounds having a low vapor pressure.
The vapor pressure of 2-nitrophenol at somewhat elevated temperatures
(49C) is only 1.0 torr.  The vapor pressure of 2-nitrophenol under ambi-
ent conditions would be even less and, as a result, volatilization is an
unlikely transport process.  In addition, the high solubility of 2-nitro-
phenol in water (2,100 mg/1 at 20C) favors a partitioning tendency toward
water rather than air.  A further factor which must be taken into con-
sideration is the ionization of the molecule in an aqueous medium.  The
value of the pKa (7.21; Pearce and Simkins 1968) indicates that about
one-half of the molecules of this pollutant will be present as non-volatile
anions in circumneutral water.

    88.4.5  Sorption

         The moderate value of the octanol/water partition coefficient
indicated by log P = 1.76 (Leo et al. 1971) suggests only limited potential
for sorption of 2-nitrophenol by organic particulates.  In contrast, the
stability of the clay complexes of 4-nitrophenol, an isomer of 2-nitro-
phenol, appears to be considerable, e.g., 4-nitrophenol cannot be desorbed
from montmorillonite even when the clay complex is heated under reduced
pressure (Saltzman and Yariv 1975).  It is uncertain whether 2-nitrophenol
will behave similarly since there is a sizable amount of intramolecular
                                     88-4

-------
interaction between the nitro group and the adjacent hydroxy group
(Morrison and Boyd 1973) which could interfere with the formation of such
unusually stable organic-inorganic complexes.  From the data of Chang and
Anderson (1968), however, 2-nitrophenol appears to be an effective floc-
culating agent for clays and soils in aqueous suspension.  These latter
observations imply that 2-nitrophenol does form stable organic-inorganic
aggregates in an aqueous medium.
    88.4.6  Bioaccumulation

         Based upon the observed relationship of the octanol/water parti-
tion coefficient and a compound's tendency to bioaccumulate in aquatic sys-
tems (Neely e_t al. 1974), nitrophenols in general are not expected to bio-
accumulate in aquatic organisms.  Further support for this contention is
given by the study of Lu and Metcalf (1975) on the biomagnification of
nitroaromatic compounds.  Nitrobenzene itself was not found to be biomagni-
fied, and the fish in the model aquatic ecosystem excreted the nitroben-
zene primarily in the form of 4-nitrophenol.

    88.4.7  Biotransformation and Biodegradation

         All nitrophenols inhibit the microbial growth of natural aquatic
systems because they uncouple the metabolic process of oxidative phos-
phorylation (Howard et_ a.1. 1976;  Makhinya 1966).  Most of the studies of
microbial- degradation of nitrophenols have concentrated on the isolation of
pure cultures which can utilize these pollutants as sources of energy, car-
bon, or nitrogen (Howard _e_t _a_l. 1976).  It is questionable whether these
studies can be extrapolated to the environment of ambient surface waters
since the concentration of the test chemical employed for enrichment of an
organism and for obtaining a reasonable amount of cell growth is far above
the concentrations generally found in nature.

         There are only a few published studies regarding the breakdown of
nitrophenolic compounds by natural communities of microorganisms (Howard et
al. 1976).  Brebion e_t _al. (1967) examined the ability of the microorgan-
isms taken from soil, water, and mud to degrade 4-nitrophenol.  The cells
were cultivated on a mineral nutrient solution in which nitrophenols were
added as the sole source of carbon.  The experimental findings  revealed no
significant removal of the compounds under these conditions.  The fate of
2-, 3-, and 4-nitrophenol in the presence of natural communities of soil
organisms has been studied also by Alexander and Lustigman (1966).  A small
amount of soil was suspended in water and the concentration of the nitro-
phenol was assayed by ultraviolet absorbance.  The 2-nitrophenol was the
most resistant to degradation and persisted unchanged for more than 64
days.  The 3-nitrophenol and the 4-nitrophenol remained unchanged for 4
days and 16 days, respectively, before being gradually biodegraded.  The
                                     88-5

-------
three biotransformation processes that have been observed for 2-nitrophenol
under optimal conditions of pure culture are reduction of the nitro group
(HcCormick. _e_t al. 1976), hydroxylation of the aromatic ring (Raymond and
Alexander 1971), and displacement of the nitro group by a hydroxy group
(Raymond and Alexander 1971;   Siddaramappa et al. 1973;  Munnecke and Hsieh
1974).

88.5  Data Summary

    Table 88-1 summarizes the aquatic fate data for 2-nitrophenol.  Nitro-
phenols interfere with biological oxidative phosphorylation and thereby can
greatly inhibit the microbial growth of aquatic systems into which they are
introduced.  2-Nitrophenol will probably undergo slow photolytic des-
truction in ambient surface waters.  Volatilization is not thought to be an
important transport process but 2-nitrophenol should be strongly sorbed by
clay minerals.  There is a possibility that 2-nitrophenol could undergo
hydrolysis within the clay structure.  Although the biotransformation of
2-nitrophenol has been demonstrated, it appears to be very persistent in
aqueous mixed cultures.
                                       3-6

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88.6  Literature Cited

Alexander, M. and B.K. Lustigman.  1966.   Effect of chemical structure on
  microbial degradation of substituted benzenes.  J. Agr.  Food Chem.
  14(4):410-413.

Brebion, G., R. Cabridenc, and B. Huriet.  1967.  Studying the
  biodegradation possibilities of industrial effluents.   Application  to the
  biodegradation of phenols.  Rev. Inst.  Fr. Petrole Ann.   Combust.
  Liquides.  22(6) :1029-1052.   Quoted by  Howard _e_t _al. (1976).

Chang, C.W. and J.U. Anderson.  1968.  Flocculation of clays and soils by
  organic compounds.  Proc. Soil. Sci. Soc. Amer.  32(l):23-27.

Giinther, K., W.G. Filby and K. Eiben.  1971.  Hydroxylation of
  substituted phenols:  an ESR-study in the Ti^+/H202 system.
  Tetrahedron Lett.  (3):251-254.

Howard, P.H., J. Santodonato,  J. Saxena,  J.E. Mailing, and D, Greninger.
  1976.  Investigation of selected potential environmental contaminants:
  nitroaromatics.  U.S. Environmental Protection Agency (Office of Toxic
  Substances), Washington, B.C., 600p. (EPA-560/2-76-010).

Johnson, K. and E.F. Degering.  1943.  Production of aldehydes and ketones
  from nitroparaffins.  J. Org. Chem.  8(1).-10-11.

Leo, A., C. Hansch and D. Elkins.  1971.   Partition coefficients and  their
  uses. Chem. Rev. 71:525-616.

Lu, P-Y. and R. Metcalf.  1975.  Environmental fate and biodegradability of
  benzene derivatives as studied in a model aquatic ecosystem.  Environ.
  Health Perspect.  19:269-273.

Makhinya, A.P.  -1966.  Effect of o-, m-,  and p-nitrophenols on the natural
  self-purification processes of reservoirs.  Vop. Koramunal. Gig.  6:76-79.
  (Abstract only).  CA1968.  68:98508y.

McCormick,  N.G., F.E. Feeherry, and H.S.  Levinson.  1976.   Microbial
  transformation of 2,4,6-trinitrotoluene and other nitroarotncitic
  compounds. Appl. Environ. Microbiol.  31(6):949-958 .

Mitchell, L.C.   1961.  The  effect of ultraviolet light (2537) on 141
  pesticide  chemicals by paper chromatography.  J. Assoc. Off. Agr. Chem.
  44:643-712.
                                     88-8

-------
Morrison, H.A.  1969.  The photochemistry of the nitro and nitroso groups.
  H. Feuer (ed).  The chemistry of the nitro and nitroso groups.  Part 1.
  Chap. 4.  pp. 165-212.  Interscience Publishers, New York.

Morrison, R.T. and R.N. Boyd.  1973.  Organic Chemistry, 3rd edition.
  Allyn and Bacon, Inc., Boston.  1258p.

Munnecke, D.E. and D.P.H. Hsieh.  1974.  Microbial decontamination of
  parathion and p-nitrophenol in aqueous media.  Appl. Microbiol.
  28(2):212-217.

Nakagawa, M.  and D.G. Crosby.  1974.  Photodecomposition of nitrofen.  J.
  Agr. Food Chem.  22(5):849-853.

Neely, W.B.,  D.R. Branson, and G.E.  Blau.  1974.  Partition coefficient  to
  measure bioconcentration potential of organic chemicals in fish.
  Environ. Sci. Technol.  8:1113-1115.

Pearce, P.J.  and R.J.J. Simkins.  1968.  Acid strengths of some substituted
  picric acids.  Can, J. Chem.  46(2):241-248.

Raymond, D.G.M. and M. Alexander.  1971.  Microbial metabolism and
  cometabolism of nitrophenols.  Pestic. Biochem.  Physio1.   1(2):123-130.

Sadtler Standard Spectra.  1975.  2-nitrophenol.  Sadtler Research
  Laboratories, Inc., a Subsidiary of Block Engineering, Inc.

Saltzman, S.  and S.  Yariv.  1975.  Intrared study of the sorption of phenol
  and p-nitrophenol by montmorillonite.  Proc. Soil Sci. Soc. Amer.
  39(3):474-479.

Shackelford,  W.M. and L.H. Keith.  1976.  Frequency of organic compounds
  identified  in water.  U.S.  Environmental Protection Agency, (ERL),
  Athens, Ga.  6l7p.  (EPA-600/4-76-062).

Siddaramappa, R., K.P. Rajaram, and N. Sethunathan.  1973.  Degradation  of
  parathion by bacteria isolated from flooded soil.  Appl. Microbiol.
  26(6):846-847.

Suarez, C. , F. Louys , K. Giinther and K. Eiben.  1970.   Hydroxyl radical
  induced denitration of nitrophenols.  Tetrahedron Lett.  (8) -.575-578.

Verschueren,  K.  1977.  Handbook of environmental data on organic
  chemicals.   Elsevier/Van Nostrand, New York.  659p.

Weast, R.C. (ed).  1977.  Handbook of chemistry and physics.  CRC Press,
  Inc.  Cleveland, Ohio.  2398p.
                                      88-9

-------
                             89.  4-NITROPHENOL


89.1  Statement of Probable Fate

    4-Nitrophenol will probably undergo slow photoxidation in an aerated
aquatic environment.  There is, however, a possibility for photoreduction
of the nitro group if the 4-nitrophenol becomes absorbed by organic parti-
culates.  4-Nitrophenol will be strongly sorbed by clay minerals and may
even undergo hydrolysis within the clay structure.  Volatilization and
bioaccumulation appear to be unlikely processes, and although biotransfor-
mation of 4-nitrophenol has been demonstrated, the nitrophenols are very
persistent in aqueous mixed cultures and will inhibit microbial growth in
natural aquatic systems into which they are introduced.

89.2  Identification

    4-Nitrophenol has been detected in industrial effluents (Shackelford
and Keith 1976).  The chemical structure of 4-nitrophenol is shown below.
                                             Alternate Names

                                             p-Nitrophenol
                                             4-Hydroxynitrobenzene
    4-Nitrophenol

    CAS NO. 100-07-7
    TSL NO. SM 22750

89.3  Physical Properties

    The general physical properties of 4-nitrophenol are as follows.

    Molecular weight                         139.11
    (Weast 1977)

    Melting point                            114.9C
    (Weast 1977)

    Boiling point at 760 torr                279C
    (Weast 1977)

    Vapor pressure at 146C                  2.2 torr
    (Verschueren 1977)
                                   89-1

-------
    Solubility in water at 25C              16,000 mg/1
    (Verschueren 1977)

    Log octanol/water partition coefficient  1.91
    (Leo et al. 1971)
    (Pearce and Simkins 1968)                 7.15

89.4  Summary of Fate Data

    89.4.1  Photolysis

         4-Nitrophenol is a somewhat acidic substance,  pKa = 7.15 (Pearce
and Siinklns 1968), and will exist to an appreciable extent as an anion in
environmental surface waters.   The ultraviolet absorption spectrum of 4-ni-
trophenol in methanol exhibits a maximum at about 310 nm which extends out
to 400 nm (Verzilina and Belotsvetov 1969).  The intensity of absorption
near 400 nm is enhanced in basic solutions.

         Nakagawa and Crosby (1974) report that 4-nitropheriol was degraded
in aqueous solution within a period of 1-2 months when it was exposed to
sunlight at a concentration of 200 mg/1.  The principal products were
hydroquinone and 4-nitrocatechol.  A dark, acidic intractable polymer was
also produced.  Nakagawa and Crosby (1974) propose that the observed  re-
action products arose via photonucleophilic displacement reactions in-
volving water or hydroxide ion.  This type of reaction mechanism might be
easily rationalized for the formation of hydroquinone,  but it is probably
not valid in the case of 4-nitrocatechol since it is difficult to envision
the development of a center of electron deficiency on the unsubstituted
carbon atom adjacent to the phenolic group,

         Suarez _e_t al. (1970)  and Gunther e al. (1971) have demonstrated
that hydroxyl radicals preferentially attack 4-nitrophenol in water at the
C-2 and C-4 positions, and that the latter intermediate results in the dis-
placement of the nitro group by a hydroxy group.  Although attack by hy-
droxyl radical can more easily explain the observed products, it is uncer-
tain whether the concentration of hydroxyl radicals in the experiment of
Nakagawa and Crosby (1974) would have been sufficient to be responsible for
the formation of hydroquinone  and 4-nitrocatechol.

         It has been reported  that the ultraviolet irradiation of individ-
ual nitrophenols at 254 nm was without effect (Mitchell 1961).  This ex-
perient, however, was conducted by irradiating the nitro substituted phenol
for 30 minutes after it had been applied in solution to a sheet of chroma-
tographic paper which was then dried.  The experiment was one of very short
duration and the conditions were anhydrous or hypohydrous.  It is, there-
                                    89-2

-------
         A further photochemical reaction of 4-nitrophenol that must be
considered is photoreduction of the nitro group.  As an example, Nakagawa
and Crosby (1974) found that the nitro group of nitrofen, a nitroaromatic
pesticide, underwent photochemical reduction to amino and azo groups in an
aqueous solution that contained 10% methanol.  Since the log octanol/water
partition coefficient of 4-nitrophenol indicates a slight potential for
absorption by suspended organic matter, this organic material could serve
as a reducing agent in the photoreduction of 4-nitrophenol to 4-aminophenol
and 4,4'-dihydroxyazobenzene.  No specific information was found pertaining
to photoreduction of 4-nitrophenol in the reviewed literature.

    89.4.2  Oxidation

         Suarez _e_t al. (1970) and Gunther _e_t al. (1971) have demonstrated
that hydroxyl radicals attack 4-nitrophenol at the C-2 and C-4 positions
resulting in the formation of a complex mixture, the principal products of
which are hydroquinone and 1,4-benzoquinone.  No information was found,
however,  from which an environmentally relevant rate could be estimated for
this reaction.

    89.4.3  Hydrolysis

         The covalent bond of a substituent attached to an aromatic ring is
usually resistant to hydrolysis because of the high negative charge-density
of the aromatic nucleus.  Nonetheless, there is a possibility, albeit not
specifically substantiated in the reviewed'literature, that nitrophenols
could undergo hydrolysis while sorbed by clay minerals.  4-Nitrophenol is
not only adsorbed by montmorillonite but also becomes intercalated within
the clay structure (Saltzman and Yariv 1975).  The water molecules within
the clay structure can be highly acidic, and the phenolic group of the
intercalated 4-nitrophenol itself is more acidic than when the molecule is
in aqueous solution due to coordination of the phenolic oxygen atom with
the metal cations of the clay.  This acidic environment could conceivably
lead to the hydrolysis of the 4-nitrophenol anion in a manner similar to
the hydrolysis of aci-nitroparaffins (Johnson and Deqering 1943).  The re-
sulting compound would be 1,4-benzoquinone.
                                      89-3

-------
    89.4.4  Volatilization

         The vapor pressure of 4-nitrophenol at elevated temperatures
(146C) is only 2.2 torr.  The vapor pressure of 4-nitrophenol under ambi-
ent conditions would be even less and, as a result, volatilization is a
highly unlikely transport process.  In addition, the high solubility of
4-nitrophenol in water (16,000 mg/1 at 25C) favors a partitioning tendency
toward water rather than air.  4-Nitrophenol, in fact, does not even
volatilize from boiling water due to intermolecular hydrogen bonding
(Morrison and Boyd 1973).

    89.4.5  Sorption

         The moderate value of the octanol/water partition coefficient in-
dicated by log P = 1.91 (Leo et_ al_. 1971) suggests only slight potential
for sorption by organic particulates.  The stability of the clay complexes
of 4-nitrophenol, however, appears to be considerable.  Intercalation of
4-nitrophenol within the structure of montmorillonite involves both ionic
attraction and orbital overlap.  4-Nitrophenol cannot be desorbed from
montmorillonite even when the clay complex is under reduced pressure
(Saltzman and Yariv 1975).

    89.4.6  Bio ace urn ula t i on

         Based upon the observed relationship of the octanol/water parti-
tion coefficient and a compound's tendency to bioaccumulate in aquatic sys-
tems (Neely _e_t al. 1974), nitrophenols in general are riot expected to bio-
accumulate in aquatic organisms.  Further support for this contention is
given by the study of Lu and Metcalf (1975) on the biomagnification of
nitroaromatic compounds.  Nitrobenzene itself was not found to be biomagni-
fied, and the fish in the model aquatic ecosystem excreted the nitrobenzene
primarily in the form of 4-nitrophenol.

    89.4.7. Biotransformation and Biodegradatlon

         All nitrophenols inhibit the microbial growth of natural aquatic
systems because they uncouple the metabolic process of oxidative phos-
phorylation (Howard _e_t a_l. 1976; Makhinya 1966).  Most of the studies of
microbial degradation of nitrophenols have concentrated on the isolation of
pure cultures which can utilize these pollutants as sources of energy, car-
bon, or nitrogen (Howard et^ _al. 1976).  It is questionable whether these
studies can be extrapolated to the environment of ambient surface waters
since the concentration of the test chemical employed for enrichment of an
organism and for obtaining a reasonable amount of cell growth is far above
the concentrations generally found in nature.
                                      89-4

-------
         There are only a few published studies regarding the breakdown of
nitrophenolic compounds by natural communities of microorganisms (Howard et
al. 1976).  Brebion _e_t _al_. (1967) examined the ability of the microorgan- .
isms taken from soil, water, and mud to degrade 4-nitrophenol.  The cells
were cultivated on a mineral nutrient solution in which nitrophenols were
added as the sole source of carbon.  The experimental findings  revealed no
significant removal of the compounds under these conditions.  The fate of
2-, 3-, and 4-nitrophenol in the presence of natural communities of soil
organisms has been studied also by Alexander and Lustigman (1966).  A small
amount of soil was suspended in water and the concentration of the nitro-
phenol was assayed by ultraviolet absorbance.  The 2-nitrophenol was the
most resistant to degradation and persisted unchanged for more than 64
days.  The 3-nitrophenol and 4-nitrophenol remained unchanged for 4 days
and 16 days, respectively, before being gradually biodegraded.  The three
biotransfonnation processes that have been observed for nitrophenols are
reduction of the nitro group (McCormick _et_ _al. 1976), hydroxylation of the
aromatic ring (Raymond and Alexander 1971), and displacement of the nitro
group by a hydroxy group (Raymond and Alexander 1971; Siddaramappa et al.
1973; Munnecke and Hsieh 1974).

89.5  Data Summary

    Table 89-1 summarizes the aquatic fate data for 4-nitrophenol.  Nitro-
phenols interfere with biological oxidative phosphorylation and thereby can
greatly inhibit the microbial growth of aquatic systems into which they are
introduced.  4-Nitrophenol will probably undergo slow photolytic destruc-
tion in ambient surface waters.  Volatilization is not thought to be an
important transport process but 4-nitrophenol should be strongly sorbed by
clay minerals.  There is a possibility that 4-nitrophenol could undergo
hydrolysis within the clay structure.  Although the biotransformation of
4-nitrophenol has been demonstrated, it appears to be very persistent in
aqueous mixed cultures.
                                      89-5

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89.6  Literature Cited

Alexander, M. and B.K. Lustigman.  1966.  Effect of chemical structure on
  microbial degradation of substituted benzenes.  J. Agr. Food. Chem.
  14(4):410-413.

Brebion, G.,  R. Cabridenc, and B. Huriet.  1967.  Studying the
  biodegradation possibilities of industrial effluents.  Application to the
  biodegradation of phenols.  Rev. Inst. Fr. Petrole Ann. Combust.
  Liquides.  22(6) :1029-1052.  Quoted by Howard _et _al. (1976).

Giinther, K. ,  W.G. Filby and K. Eiben.  1971.  Hydroxylation of substituted
  phenols:  an ESR-study in the Ti3+/H2C>2 system.  Tetrahedron Lett.
  (3):251-254.

Howard, P.H., J. Santodonato, J. Saxena, J.E. Mailing, and D. Greninger.
  1976.  Investigation of selected potential environmental contaminants:
  nitroaromatics.  U.S. Environmental Protection Agency  (Office of Toxic
  Substances), Washington, D.C. 600p. (EPA 560/2-76-010).

Johnson, K. and E.F. Degering.   1943.  Production of aldehydes and ketones
  from nitroparaffins.  J. Org. Chem.  8(1):10-11.

Leo, A., C. Hansch and D. Elkins.  1971.  Partition coefficients and their
  uses.  Chem. Rev. 71:  525-616.

Lu, P-Y. and  R. Metcalf.  1975.  Environmental fate and  biodegradability of
  benzene derivatives as studied in a model aquatic ecosystem.  Environ.
  Health Perspect.  19:269-273.

Makhinya, A.P.  1966.  Effect of o-, m-, and p-nitrophenols on the natural
  self-purification processes of reservoirs.  Vop. Kommunal. Gig. 6:76-79.
  (Abstract only).  CA 1968.  68:98508y.

McConnick, N.G., F.E. Feeherry, and H.S. Levinson.  1976.  Microbial
  transformation of 2,4,6-trinitrotoluene and other nitro aromatic
  compounds.   Appl. Environ. Microbiol. 31(6):949-958.

Mitchell, L.C.  1961.  The effect of ultraviolet light (25372) on 141
  pesticide chemicals by paper chromatography.  J. Assoc. Off. Agr. Chem.
  44:643-712.

Morrison, R.T. and R.N. Boyd.  1973.  Organic Chemistry  , 3rd edition.
  Allyn and Bacon, Inc., Boston.  1258 p.
                                     89-7

-------
Munnecke, D.E. and D.P.H. Hsieh.   1974.   Microbial decontamination of
  parathion and p-nitrophenol in aqueous media.   Appl.  Microbiol.
  28(2):212-217.

Nakagawa, M. and D.G. Crosby.  1974.  Photodecomposition of nitrofen. J.
  Agr. Food Chem. 22(5):849-853.

Neely, W.B., D.R. Branson, and G.E. Blau.  1974.   Partition coefficient to
  measure bioconcentration potential of  organic  chemicals in fish. Environ,
  Sci. Technol. 8:1113-1115.

Pearce, P.J. and R.J.J. Simkins.   1968.   Acid strengths of some
  substituted picric acids. Can.  J. Chem.  46(2):241-248.

Raymond, D.G.M. and M. Alexander.  1971.  Microbial metabolism and
  cometabolism of nitrophenols.  Pestic. Biochem.  Physiol. 1(2):123-130.

Saltzman, S. and S. Yariv.  1975.  Infrared study of the sorption of
  phenol and p-nitrophenol by montmorillonite.  Proc. Soil Sci. Soc. Amer.
  39(3):474-479.

Shackelford, W.M. and L.H. Keith.  1976.  Frequency of  organic compounds
  identified in water.  U.S. Environmental Protection Agency, (ERL),
  Athens, Ga. 617p. (EPA 600/4-76-062).

Siddaramappa, R., K.P. Rajaram, and N. Sethunathan.  1973.  Degradation of
  parathion by bacteria isolated from flooded soil.  Appl. Microbiol.
  26(6):846-847.

Suarez, C., R. Louys, K. Gunther and K.  Eiben.  1970.  Hydroxyl radical
  induced denitration of nitrophenols.  Tetrahedron Lett. (8):575-578.

Verschueren, K.  1977.  Handbook of environmental data on organic
  chemicals.  Elsevier/Van Nostrand, New York. 659 p.

Verzilina, M.K. and A.V. Belotsvetov.  1969.  Investigation of electronic
  absorption spectra of compounds with strongly polarized systems.  I.
  Esters and arylides of anisic acids.  J. Gen.  Chem. USSR. 39(3):626-629.

Weast, R.C. (ed).  1977.  Handbook of chemistry and physics. CRC Press,
  Inc. Cleveland, Ohio.  2398p.
                                     89-8

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                          90.  2.4-DINITROPHENOL


90.1  Statement of Probable Fate

    Based on the information gathered for 2-nitrophenol, 4-nitrophenol, and
this pollutant, 2,4-dinitrophenol will probably undergo slow photooxidation
in an aerated aquatic environment.  There is a possibility for photoreduc-
tion of the nitro group if the 2,4-dinitrophenol becomes absorbed by or-
ganic particulates.  2,4-Dinitrophenol should be strongly sorbed by clay
minerals, and may even undergo hydrolysis within the clay structure.
Volatilization and bioaccumulation appear to be unlikely processes, and
although biotransformation of 2,4-dinitrophenol has been demonstrated, the
nitrophenols are very persistent in aqueous mixed cultures and will inhibit
aerobic microbial growth in any natural aquatic systems in which they are
present.

90.2  Identification

    The chemical structure of 2,4-dinitrophenol is shown below.

            OH                               Alternate Names

                N02                          Aldifen
                                             2,4-DNP
    2,4-Dinitrophenol

    CAS NO. 51-28-5
    TSL NO. SL 28000

90.3  Physical Properties

    The general physical properties of 2,4-dinitrophenol are as follows.

    Molecular weight                         184.11
    (Weast 1977)

    Melting point                            114 C
    (Verschueren 1977)

    Boiling point                            No data found

    Vapor pressure                           No data found
                                      90-1

-------
    Solubility in water at 18C              5,600 mg/1
    (Verschueren 1977)

    Log octanol/water partition coefficient  1.53
    (Leo et_ al. 1971)

    pKa                                      4.09
    (Pearce and Simkins 1968)

90.4  Summary of Fate Data

    90.4.1  Photolysis

         2,4-Dinitrophenol is  a moderately acidic substance, pKa=4.09
(Pearce and Simkins 1968), and will exist substantially as an anion in en-
vironmental surface waters.  The ultraviolet absorption spectrum of 2,4-di-
nitrophenol in methanol exhibits a maximum at about 290 nm which extends
out beyond 400 nm (Verzilina and Belotsvetov 1969).

         Nakagawa and Crosby (1974) report that 4-nitrophenol was degraded
in aqueous solutions within a period of 1-2 months when it was exposed to
sunlight at a concentration of 200 mg/1.  The principal products were hy-
droquinone and 4-nitrocatechol.  A dark, acidic intractable polymer was
also produced.  Although no specific information was found in the reviewed
literature demonstrating that 2,4-dinitrophenol would also be photochemi-
cally hydroxylated in an analogous manner, the following mechanistic con-
siderations for this photochemical reaction indicate that, under similar
conditions, 2,4-dinitrophenol should be degraded to a mixture of compounds
which would include 4-nitrocatechol, 2-nitrohydroquinone, and 3,5-dinitro-
catechol.

         Nakagawa and Crosby (1974) proposed that the observed reaction
products arising from 4-nitrophenol were the result of photonucleophilic
displacement reactions involving water or hydroxide ion.  This type of re-
action mechanism could be rationalized for the formation of hydroquinone,
but it is probably not valid in the case of 4-nitrocatechol since it is
difficult to envision the development of a center of electron deficiency on
the unsubstituted carbon atom adjacent to the phenolic group, Suarez et al.
(1970) and Giinther e al. (1971) have demonstrated that hydroxyl radicals
preferentially attack 4-nitrophenol in water at the C-2 and C-4 positions,
and that the latter intermediate results in the displacement of the nitro
group by a hydroxy group.  Even though attack by hydroxyl radical can more
easily explain the observed products, it is uncertain whether the concen-
tration of hydroxyl radicals in the experiment of Nakagawa and Crosby
(1974) would have been sufficient to be responsible for the formation of
hydroquinone and 4-nitrocatechol.
                                     90-2

-------
         It has been reported that the ultraviolet irradiation of indivi-
dual nitrophenols at 254 nm was without effect (Mitchell 1961).  This ex-
periment, however, was conducted by irradiating the nitrosubstituted phenol
for 30 minutes after it had been applied in solution to a sheet of chroma-
tographic paper which was then dried.  The experiment was one of very short
duration and the conditions were anhydrous or hypohydrous.  For these
reasons it is probably not valid to draw any conclusions from this experi-
ment with respect to either the aquatic or terrestrial environment.

         A further photochemical reaction of 2,4-dinitrophenol that must be
considered is photoreduction of the nitro groups.  As an example, Nakagawa
and Crosby (1974) found that the nitro group of nitrofen, a nitroaromatic
pesticide, underwent photochemical reduction to amino and azo groups in an
aqueous solution that contained 10% methanol.  Aromatic nitro groups are
generally reduced photochemically in the presence of suitable hydrogen
donors (Morrison 1969) .  The log octanol/water partition coefficient of
2,4-dinitrophenol indicates a slight potential for absorption by suspended
aquatic organic matter.  In this way the organic material could serve as a
reducing agent in the photoreduction.  Massini and Voorn (1967) report that
the reduction of 2,4-dinitrophenol by ascorbic acid is sensitized by the
presence of chlorophyll and is enhanced further by ferrous ion.  The pro-
ducts of this reduction were not indicated but it is quite likely that a
complex mixture of nitrogen compounds was produced.

    90.4.2  Oxidation

         Suarez e_t a_l. (1970) and Gunther _e_t al. (1971) have demonstrated
that hydroxyl radicals attack both 2-nitrophenol and 4-nitrophenol at the
C-2 and C-4 positions resulting in the formation of a complex mixture that
follows from the generation of isomeric benzosemiquinones.  No information
was found, however, from which an environmentally relevant rate could be
estimated for this reaction.  It can be inferred that 2,4-dinitrophenol
could be oxidized in an analogous fashion.

    90.4.3  Hydrolysis

         The covalent bond of a substituent attached to an aromatic ring is
usually resistant to hydrolysis because of the high negative charge-density
of the aromatic nucleus.  Nonetheless, there is a possibility, albeit not
specifically substantiated in the reviewed literature, that nitrophenols
could undergo hydrolysis while sorbed by clay minerals.  Specifically,
4-nitrophenol is not only adsorbed by montmorillonite but also becomes
intercalated within the clay structure (Saltzman and Yariv 1975) and there
is no apparent reason to believe that 2,4-dinitrophenol would be less
likely to be similarly sorbed.  The water molecules within the clay struc-
ture can be highly acidic, and the phenolic group of the intercalated
2,4-dinitrophenol would become more acidic than when the molecule is in
                                     90-3

-------
aqueous solution due to coordination of the phenolic oxygen, atom with the
metal cations of the clay.  This acidic environment could conceivably lead
to hydrolysis of the 2,4-dinitrophenol anion in a manner similar to the
hydrolysis of aci-nitroparaffiris (Johnson and Degering 1943).
    90.4.4  Volatilization

         The vapor pressure of 4-nitrophenol at elevated temperatures
(146C) is only 2.2 torr.  As a more highly substituted nitrophenol, trie
vapor pressure of 2,4-dinitrophenol under ambient conditions would be ex-
pected to be less, and as a result, volatilization is a highly unlikely
transport process.  In addition, the high solubility of 2,4-dinitrophenol
in water (5,600 mg/1 at 18C) and its presence in solution primarily as an
anion strongly favor a partitioning tendency toward water rather than air.
It is reported that 4-nitrophenol does not even volatilize from boiling
water (Morrison and Boyd 1973) and 2,4-dinitrophenol would be expected to
behave similarly.

    90.4.5  Sorption

         The moderate value of the octanol/water partition coefficient
indicated  by log P = 1.53 (Leo et_ al. 1971) suggests only limited poten-
tial for sorption of 2,4-dinitrophenol by organic particulates.  The
stability of the clay complexes of 2-nitrophenol and 4-nitrophenol appears
to be considerable (Saltzman and Yariv 1975;  Chang and Anderson 1968),
however, and there is no reason to believe that 2,4-dinitrophenol will not
behave similarly.

    90.4.6  Bioaccumulation

         Based upon the observed relationship of the octanol/water parti-
tion coefficient and a compound's tendency to bioaccumulate in aquatic sys-
                                     90-4

-------
terns (Neely _e_t al.  1974), nitrophenols in general are not expected to
bioaccumulate in aquatic organisms.  Further support for this contention is
given by the study of Lu and Metcalf (1975) on the biomagnification of
nitroaromatic compounds.  Nitrobenzene itself was not found to be biomag-
nified and the fish in the model aquatic ecosystem excreted the
nitrobenzene primarily in the form of 4-nitrophenol.

    90.4.7  Biotransformation and Biodegradation

         All nitrophenols inhibit the microbial growth of natural aquatic
systems because they uncouple the metabolic process of oxidative phos-
phorylation (Howard t _al. 1976;  Makhinya 1966).  Most of the studies of
microbial degradation of.nitrophenols have concentrated on the isolation of
pure cultures which can utilize these pollutants as sources of energy, car-
bon, or nitrogen (Howard _e_t al. 1976).  It is questionable whether these
studies can be extrapolated to the environment of ambient surface waters
since the concentration of the test chemical employed for enrichment of an
organism and for obtaining a reasonable amount of cell growth is far above
the concentrations generally found in nature.

         There are only a. few published studies regarding the breakdown of
nitrophenolic compounds by natural communities of microorganisms (Howard e^t
al. 1976).  Brebion et _al. (1967) examined the ability of the microorgan-
isms taken from soil, water, and mud to degrade 4-nitrophenol.  The cells
were cultivated on a mineral nutrient solution in which nitrophenols were
added as the sole source of carbon.  The experimental findings revealed no
significant removal of the compounds under these conditions.  The fate of
2-, 3-, and 4-nitrophenol in the presence of natural communities of soil
organisms has been studied also by Alexander and Lustigman (1966).  A small
amount of soil was suspended in water and the degradation of the nitro-
phenol was assayed by ultraviolet absorbancy.  The 2-nitrophenol was the
most resistant to degradation and persisted unchanged for more than 64
days.  The 3-nitrophenol and the 4-nitrophenol remained unchanged for 4
days and 16 days, respectively, before being gradually biodegraded.  The
three biotransformation processes that have been observed for 2,4-dinitro-
phenol under optimal conditions in pure culture are reduction of the nitro
group (McCormick et_ al. 1976), hydroxylation of the aromatic ring (Raymond
and Alexander 1971), and displacement of the nitro group by a hydroxy group
(Raymond and Alexander 1971;  Siddaramappa _e_t _al. 1973;  Munnecke and Hsieh
1974).

90.5  Data Summary

    Table 90-1 summarizes the aquatic fate data for 2,4-dinitrophenol.
Nitrophenols interfere with biological oxidative phosphorylation and there-
by can greatly inhibit the microbial growth of aquatic systems into which
                                     90-5

-------
they are introduced.   2,4-Dinitrophenol will probably undergo slow pho-
tolytic destruction in ambient surface waters.   Volatilization is not
thought to be an important transport process but 2,4-dinitrophenol should
be strongly sorbed by clay minerals.  There is  a possibility that 2,4-di-
nitrophenol could undergo hydrolysis within the clay structure.  Although
the biotransformation of 2,4-dinitrophenol has  been demonstrated, this
compound is probably very persistent in aqueous mixed cultures.
                                       90-6

-------













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90.6  Literature Cited

Alexander, M. and B.K. Lustigman.   1966.   Effect  of  chemical  structure on
  microbial degradation of substituted benzenes.   J. Agr.  Food Chem.
  14(4):410-413.

Brebion, G., R. Cabridenc, and B.  Huriet.   1967.   Studying the
  biodegradation possibilities of  industrial effluents.   Application  to the
  biodegradation of phenols.   Rev. Inst.  Fr. Petrole Ann.   Combust.
  Liquides.  22(6) :1029-1052.   Quoted by  Howard t _al.  (1976).

Chang, C.W. and J.U. Anderson.  1968.  Flocculation  of  clays  and soils by
  organic compounds.  Proc. Soil.  Sci. Soc. Amer. 32(1):23-27.

Giinther, K., W.G. Filby and K. Eiben.  1971.  Hydroxylation of
  substituted phenols:  an ESR-study in the Ti^+/H202 system.
  Tetrahedron Lett.  (3)=251-254.

Howard, P.H., J. Santodonato,  J. Saxena,  J.E. Mailing,  and D. Greninger.
  1976.  Investigation of selected potential environmental contaminants:
  nitroaromatics.  U.S. Environmental Protection Agency (Office of Toxic
  Substances), Washington, D.C., 600p. (EPA-560/2-76-010).

Johnson, K. and E.F. Degering.  1943.  Production of aldehydes and ketones
  from nitroparaffins.  J. Org. Chem.  8(1):10-11.

Leo, A., C. Hansch and D. Elkins.   1971.   Partition  coefficients and  their
  uses.  Chem. Rev. 71:525-616.

Lu, P-Y, and R. Metcalf.  1975.  Environmental fate  and biodegradability of
  benzene derivatives as studied in a model aquatic  ecosystem.  Environ.
  Health Perspect.  19:269-273.

Makhinya, A.P.  1966.  Effect  of o-, m-,  and p-nitrophenols on the natural
  self-purification processes  of reservoirs.  Vop. Kotnmunal.  Gig.  6:76-79.
  (Abstract only).  CA 1968.   68:98508y.

Massini, P. and G. Voorn.  1967.  The effect of ferredoxin and ferrous ion
  on the chlorophyll sensitized photoreduction of dinitrophenol.
  Photochem. Photobiol.  6:851-856.

McCormick, N.G., F.E. Feeherry, and H.S.  Levinson.  1976.   Microbial
  transformation of 2,4,6-trinitrotoluene and other  nitroaromatic
  compounds. Appl. Environ. Microbiol.  31(6):949-958.
                                     90-8

-------
Mitchell, L.C.  1961.  The effect of. ultraviolet light (253?i)  on 141
  pesticide chemicals by paper chromatography.  J.  Assoc. Off.  Agr.  Chem.
  44:643-712.

Morrison, H.A.  1969.  The photochemistry of the nitro and nitroso groups.
  H. Feuer (ed).  The chemistry of the nitro and nitroso groups.   Part 1.
  Chap. 4.  pp. 165-212.  Interscience Publishers,  New York.

Morrison, R.T. and R.N. Boyd.  1973.  Organic Chemistry, 3rd  edition.
  Allyn and Bacon, Inc., Boston.  1258p.

Munnecke, D.E. and D.P.H. Hsieh.  1974.  Microbial  decontamination of
  parathion and p-nitrophenol in aqueous  media.  Appl. Microbiol.
  28(2):212-217.

Nakagawa, M. and D.G. Crosby.  1974.  Photodecomposition of nitrofen.  J.
  Agr. Food Chem.  22(5):849-853.

Neely, W.B., D.R. Branson, and G.E. Blau.  1974.  Partition coefficient to
  measure bioconcentration potential of organic chemicals in  fish.
  Environ. Sci. Technol.  8:1113-1115.

Pearce, P.J. and R.J.J. Simkins.  1968.  Acid strengths of some substituted
  picric acids.  Can. J. Chem.  46(2):241-248.

Raymond, D.G.M. and M. Alexander.  1971.   Microbial metabolism and
  cometabolism of nitrophenols.  Pestic.  Biochem.  Physiol.   1(2):123-130.

Saltzman, S. and S. Yariv.  1975.  Infrared study of the sorption of phenol
  and p-nitrophenol by montmorillonite.  Proc. Soil Sci. Soc. Amer.
  39(3):474-479.

Siddaramappa, R., K.P. Rajaram, and N. Sethunathan.  1973.  Degradation of
  parathion by bacteria isolated from flooded soil.  Appl. Microbiol.
  26(6):846-847.

Suarez, C., F. Louys, K. Gunther and K. Eiben.  1970.  Hydroxyl radical
  induced denitration of nitrophenols.  Tetrahedron Lett.  (8):575-578.

Verschueren, K.  1977.  Handbook of environmental data on organic
  chemicals.  Elsevier/Van Nostrand, New York.  659p.

Verzilina, M.K. and A.V. Belotsvetov.  1969.  Investigation of electronic
  absorption spectra of compounds with strongly polarized systems.  I.
  Esters and arylides of anisic acids.  J. Gen. Chem. USSR 39(3):626-629.

Weast, R.C.  (ed).  1977.  Handbook of chemistry and physics.   CRC Press,
  Inc.  Cleveland, Ohio.  2398p.
                                       90-9

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                   91.   2,4-DIMETHYLPHENOL (2,4-XYLENOL)


91.1  Statement of Probable Fate

    There is a lack of  data from which to assign a definitive environmental
fate for 2,4-dimethylphenol.   Based on the photolytic behavior of unsubsti-
tuted phenol and alkylbenzenes such as toluene,  it can be inferred that
2,4-dimethylphenol will undergo photooxidation in the aquatic environment.
It is not possible, however,  to determine the relative importance of photo-
oxidation in comparison to other possible fate processes.  2,4-Dimethyl-
phenol has been reported to be readily degraded by activated sludge cul-
tures but it appeared to be persistent in a simulated surface water en-
vironment .

91.2  Identification

    2,4-Dimethylphenol  has been detected in industrial effluents  and in
finished drinking water (Shackelford and Keith 1976).


                                             Alternate Names


                                             2,4-Xylenol
                                             l-Hydroxy-2,4-dimethylbenzene

     2,4-Dimethylphenol

    CAS NO.  105-67-9
    TSL NO.  ZE 56000

91.3  Physical Properties

    The general physical properties of 2,4-dimethylphenol are as  follows.

    Molecular weight                         122.16
    (Weast 1977)

    Melting point                            24.54C
    (Andon et_ al. 1960)

    Boiling point at 760 torr                210.93C
    (Andon et_ al. 1960)

    Vapor'pressure at 20C                   0.0621 torr*
    (Andon et al. 1960)
                                    91-1

-------
    Solubility in water at 160C             17,000 mg/1**
    (Erichsen and Dobbert  1955)

    Log octanol/water partition  coefficient  2.50
    (Calc.  by method of Tute 1971)

    pKa                                      10.60
    (Herington and Kynaston 1957)
* Vapor pressure of 2,4-dimethylphenol as a supercooled liquid.
**Solubilities of 3,5-dimethylphenol in water at 160C and 20C are 37,000
  mg/1 and 4,200 mg/1, respectively (Erichsen and Dobbert 1955).

91.4  Summary of Fate Data

    91.4.1  Photolysis

         2,4-Dimethylphenol is a very weak acid, pKa = 10.60 (Herington
and Kynaston 1957), and exists principally as its protonated, non-ionized
form when it is in true solution in environmental surface waters.  Coordi-
nation of the phenolic oxygen atom with dissolved or suspended di- and tri-
valent metal cations, however, can markedly increase the ionization of the
phenolic proton.  In the near ultraviolet spectral region the absorption
maximum of undissociated 2,4-dimethylphenol occurs at 277 nm and extends to
300 ran while the anion of 2,4-dimethylphenol has an absorption maximum at
296 nm which extends beyond 320 nm (Herington and Kynaston 1957). It should
be noted that complexes of phenolic compounds with metal cations, such as
iron(III), absorb light strongly at about 600 nm (Ackermann and Hesse
1970).  Any photolytic reactions of unsubstituted phenol that can occur in
surface waters would, therefore, probably occur at an enhanced  rate with
this pollutant.

         As an example, solid or liquid phenol has long been known to form
reddish high molecular weight material when exposed to sunlight and air
(Joschek and Miller  1966).  A possible explanation for these observations
could  be  the formation and photolysis of an oxygen-phenol charge-transfer
complex.  Joschek and Miller  (1966) report that  the steady irradiation of
aqueous solutions of unsubstituted phenol at 254 nm in the presence of oxy-
gen yields isolable  amounts of  4,4'-dihydroxybiphenyl, 2,4'-dihydroxybi-
phenyl, 2,2'-dihydroxybiphenyl, hydroquinone, and catechol as well as many
uncharacterized compounds.  The latter two compounds predominate as pro-
ducts  in  the more dilute  solutions.   The intermediate  that is postulated  to
explain this distribution of  products is the phenoxyl  radical.

          Inasmuch as 2,4-dimethylphenol  is a dimethyl  substituted mono-
cyclic aromatic compound, its photolytic behavior also should resemble that
of  toluene and m-xylene  (1,3-dimethylbenzene).   The primary  photochemical
                                     91-2

-------
process of toluene is generally regarded as a dissociation with formation
of a benzyl radical (Porter and Norman 1954).  Reaction of this benzyl
radical with molecular oxygen is reported to be extremely fast (k ~ 10
1. mole~l sec"^) resulting in the production of benzyl hydroperoxide
(Wei and Adelman 1969).  Although benzyl hydroperoxide has considerable
thermal stability, it can be photochemically transformed to benzyl alcohol
and benzaldehyde.  No information was found in the reviewed literature
specifically pertaining to the photolysis of 2,4-dimethylphenol, but it can
be inferred from the foregoing discussion that photolysis will be operative
as a degradative pathway in well aerated surface waters.

    91.4.2  Oxidation

         Hydroxylation of aqueous phenol at the C-2 position in the pres-
ence of air and iron(III) or copper(II) ions has been reported but at tem-
peratures and pressures far above what would be normally encountered in en-
vironmental surface waters (Makalets and Ivanova 1969).  In addition, un-
substituted phenol has been oxidized by passing molecular oxygen into an
aqueous solution at 25C and pH 9.5-13 (Kirso et al. 1967).  These observa-
tions, although not environmentally relevant in themselves, raise the
possibility that 2,4-dimethylphenol could be non-photolytically oxidized in
highly aerated waters which also contain iron and copper in solution or as
part of the suspended particulates.

    91.4.3  Hydrolysis

         There are no data to suggest that hydrolysis of 2,4-dimethylphenol
is an environmentally significant process.  The covalent bond of a substi-
tuent attached to an aromatic ring is usually resistant to hydrolysis be-
cause of the high negative charge-density of the aromatic nucleus (Morrison
and Boyd 1973) .

    91.4.4  Volatilization

         The vapor pressure of supercooled liquid 2,4-dimethylphenol at
20C is 0.0621 torr (Andon et_ _al. i960) and, based on the data of Erichsen
and Dobbert (1955), the solubility at 20C should be at least 1000 mg/1.
Low vapor pressure and a moderately high solubility usually imply that
there is little tendency for volatilization from water.  Furthermore, it
can be expected that aqueous 2,4-dimethylphenol also will be highly sol-
vated which will increase its persistence in water at low levels of con-
centration.

    91.4.5  Sorption

         Unsubstituted phenol apparently has very little affinity for
microcrystalline clays inasmuch as it can be almost completely desorbed
from a thin layer of montmorillonite that has been exposed for one week, to
                                     91-3

-------
the atmosphere at 40% relative humidity (Saltzman and Yariv 1975).  From
the data of Chang and Anderson (1968),  unsubstituted phenol also appears to
be ineffective as a flocculant of clays and soils.   This latter observation
implies that unsubstituted phenol does  not form stable organic-inorganic
aggregates in an aqueous medium.   2,4-Dimethylphenol should behave
similarly with regard to organic-inorganic interactions.  2,4-Dimethyl-
phenol, however, has a log octanol/water partition coefficient of 2.50
(Tute 1971) and may, therefore, have some defininte tendency to become
absorbed onto the organic detrius.

    91.4.6  Bioaccumulation

         Although the log octanol/water partition coefficient indicates
that 2,4-dimethylphenol may have a tendency to be absorbed by aquatic
biota, a review of the current literature revealed no information con-
cerning the bioaccumulation of phenol by aquatic microorganisms or by
aquatic invertebrates or vertebrates,

    91.4.7  Biotransformation and Biodegradation

         The microbial degradation of unsubstituted phenol has been ob-
served in many laboratory studies in which phenol represented the primary
carbon source provided for isolated and adapted microorganisms.  Alexander
and Lustigman (1966) observed that phenol was degraded rapidly by a mixed
population of soil microorganisms.  Their data suggested that the hydroxyl
group, compared to other benzene ring substituents, facilitated microbial
degradation.  Some species of soil bacteria that have been demonstrated to
be capable of utilizing toluene as a sole carbon source hydroxylate the
aromatic ring to a mixture of methyl substituted phenolic compounds which
are metabolized further to acetic acid and pyruvic acid (Glaus and Walker
1964;  Gibson _e_t _al. 1966).  In addition, enrichment cultures of micro-
organisms obtained from garden soil, compost, river mud, and the sediment
of a waste lagoon in a petroleum refinery were all shown to be capable of
degrading 2,4-dimethylphenol (Tabak et al. 1964).

         Biodegradation has been suggested as a mechanism for the degrada-
tion of phenolic wastes in natural waters (Streeter 1929;  Happold and Key
1932;  Mischonsniky 1934;  Krorabach and Barthel 1964;  Polisois et al.
1975;  Wuhrmann 1972), and recent studies have examined the importance of
microorganisms in this process.  Visser et_ a_l. (1977) conducted an in situ
investigation of the phenol-degrading activity of bacteria in river water.
Phenol (125 yg/1) was added to containers holding large quantities of river
water.  The containers were incubated in the river along with sterilized
controls.   The removal rate of phenol was 30 ug/1 per hour from the natural
samples compared to < 1 yg/1 per hour from the sterilized controls.
                                      91-4

-------
         2,4-Dimethylphenol has been reported to be almost completely de-
graded by the microorganisms of activated sludge cultures when the pollu-
tant was used as the sole source of carbon (Fitter and Kucharova-Rosolova
1974).  Kaplin e^ _al_. (1968), however, carried out a series of experiments
in which they sought to duplicate the conditions for biodegradability that
would occur  in a river that was receiving phenolic waste effluents from a
coking plant.  Unsubstituted phenol decomposed rapidly, cresols exhibited a
lag period of several days, but 2,4- and 2,3-dimethylphenol seemed to be
very persistent.

91.5  Data Summary

    Table 91-1 summarizes the aquatic fate data for phenol.  Photooxidation
appears to be a likely degradative pathway in the aquatic environment but
the data regarding biodegradation are somewhat conflicting and inconclu-
sive.  Metal-catalyzed oxidation may be important in some localized situa-
tions.  There may be some absorption by lypophilic materials but sorption
by clay minerals appears unlikely.  Volatilization and bioaccumulation are
probably not important processes in this pollutant's fate.
                                      91-5

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91.6  Literature Cited

Ackermann, G. and D. Hesse.  1970.   tJber eisen(III)-komplexe  mit  phenolen.
  III.  Die absorptionsspektren und deren auswertung.   Z.  Anorg.  Allg.
  Chem.  375(1)-.77-86.

Alexander, M. and B.K. Lustigman.  1966.  Effect of  chemical  structure  on
  microbial degradation of substituted benzenes.  J. Agr.  Food Chem.
  U(4):440-413.

Andon, R.J.L., D.P. Biddiscombe, J. D. Cox,  R.  Handley,  D. Harrop,  E.F.G.
  Herington, and J.F. Martin.  I960.  Thermodynamic  properties of organic
  oxygen compounds.  Part 1.  Preparation and physical properties of  pure
  phenol, cresols, and xylenols.  J. Chem. Soc. (Lond.)   5246-5254.

Chang, C.W. and J.U Anderson.  1S68.  Flocculation of  clays and soils by
  organic compounds.  Proc. Soil Sci.  Soc. Amer.  32(l):23-27.

Claus, D. and N. Walker.  1964.  The decomposition of  toluene by  soil
  bacteria.  J. Gen. Microbiol.  36:107-122.

Erichsen, L. and E. Dobbert.  1955.  Das gegenseitige  loslichkeitsver-
  halten von alkylphenolen und wasser.  Brennstoff-Chemie
  36(21/22):338-345.

Gibson, D.T., J.R. Koch, and R.E. Kallio.  1966.  Oxidative degradation of
  aromatic hydrocarbons by microorganisms.   Enzymatic formation  of
  catechol from benzene.  Biochemistry 7(7):2653-2662 .

Happold, F.C. and A. Key.  1932.  The bacterial purification  of gas works
  liquors.  The action of the liquors  on the bacterial flora  of sewage.
  J. Hyg.  32:573-577.

Herington, E.F.G. and W. Kynaston.   1957.  The  ultraviolet absorption
  spectra and dissociation constants of certain phenols  in aqueous
  solution.  Trans. Faraday Soc.  53:138-142.

JoscheK, H.I. and S.I. Miller.  1966.   Photooxidation  of phenol,  cresols,
  and dihydroxybenzenes.  J. Am. Chem. Soc.   88(14):3273-3281.

Kaplin, V.T., L.V. Semenchenko, and E.G. Ivanov.  1968.   Decomposition  of a
  phenol mixture in natural waters  (miniature-scale  operation).  Gidrokhim.
  Mater.  46:199-202.  (Abstract only).  CA  1968.  69:69568h.

Kirso, U., K. Kuiv, and M. Gubergrits.  1967.  Kinetics  of phenol and
  m-cresol oxidation by molecular oxygen in an  aqueous medium.  Zh.
  Prikl.  Khim.  40(7):1583-1589.  (Abstract only).   CA  1968.  68:12174b.
                                     91-7

-------
Krombach, H. and J. Barthel.   1964.   Investigation of a small watercourse
  accidentally polluted by phenol compounds.   Advan.  Water Pollut.  Res.
  1:191-224.

Makalets, B.I. and L.G. Ivanova.   1969.   Oxidation of phenol by atmospheric
  oxygen in aqueous solutions.  Neftekhimiya.   9(2):280-285 .  (Abstract
  only).  CA 1969.  71:29831y.

Mischonsniky,  S.  1934.  A study of  the  pollution of  fish-containing waters
  by waste phenolic waters.  14th Cong.  Chem.  Ind. (Paris, October  1934,
  Abstract only).  J. Am. Water Works Assoc.   1937.  29:304.,

iMorrison, R.T. and R.N. Boyd.  1973.   Organic  Chemistry,  3rd edition.
  Allyn and Bacon, Inc., Boston.   1258p.

Fitter, P. and P. Kucharova-Rosolova.  1974.   Relation between the
  structure and the biodegradability  of  organic compounds.  III.
  Biodegradability of aromatic hydroxy derivatives.  Sb.  Vys. Sk.
  Chem.-Technol.  Praze, Technol. Vody.   F19:43-57.  (Abstract only).  CA
  1976.  85:67656s.

Polisois, G.,  A. Tessier, P.G.C.  Campbell, and J.P. Villeneuve.  1975.
  Degradation of phenolic compounds  downstream from a petroleum refinery
  complex.  J. Fish. Res. Bd. Can.  32(11):2125-2131.

Porter, G. and I. Norman.  1954.   Trapped atoms and radicals in a glass
  cage.  Nature 174(4428):508-509.

Saltzman, S. and S. Yariv.  1975.  Infrared study of the sorption of phenol
  and p-nitrophenol by montmorillonite.   Proc. Soil Sci.  Soc. Amer.
  39(3):474-479.

Shackelford, W.M. and L.H. Keith.  1976.  Frequency of organic compounds
  identified in water.  U.S.  Environmental Protection Agency, (ERL),
  Athens, Ga.  617p.  (EPA 600/4-76-062).

Streeter, H.W.  1929.  Chlorophenol tastes and odors  in water supplies of
  Ohio River cities.  Am. J.  Pub. Health 19(8):929-934.

Tabak, H.H., C.W. Chambers, and P.W.  Kabler.   1964.  Microbial
  metabolism of aromatic compounds.   I.   Decomposition of phenolic
  compounds and aromatic hydrocarbons by phenol-adapted bacteria.  J.
  Bacteriol.  87(4):910-919.

Tute, M.S.  1971.  Principles and practice of Hansch analysis;  a guide  to
  structure-activity correlation for the medicinal chemist.  Adv. Drug
  Res. 6:1-77.
                                     91-8

-------
Visser, S.A., G. Lamontagne, V. Zoulalian, and A. Tessier.  1977.  Bacteria
  active in the degradation of phenols in polluted waters of the St.
  Lawrence River.  Arch. Environ. Contain. Toxicol.  6:455-469.

Weast, R.C. (ed.)  1977.  Handbook, of chemistry and physics.  CRC Press,
  Inc., Cleveland, Ohio.  2398p.

Wei, K.S. and A.H. Adelman.  1969.  The photooxidation of toluene.  The
  role of an excited charge transfer complex.  Tetrahedron Lett.
  (38):3297-3300.

Wuhrmann, K.  1972.  Stream purification.  In Mitchell, R. (ed.):  Water
  pollution microbiology.  Wiley Interscience, New York.  119p.
                                     91-9

-------
                          92.   p-CHLORO-m-CRESOL


92.1  Statement of Probable Fate

    The most probable aquatic fate of  p-chloro-m-cresol  is  intramolecular
photolysis.   Although no specific data pertaining  to  the photolysis  of
p-chloro-m-cresol were found in the reviewed literature,  this  supposition
is supported by the observed photolytic reductive  dechlorination  of
4-chlorophenol in 2-propanol and the well documented  photochemistry  of
aromatic methyl substituents.   No information was  found  in  the reviewed
literature in support of a role for oxidation and  hydrolysis as fate
pathways or for volatilization and sorption as transport  processes.   The
importance of bioaccumulation cannot be ascertained from the available
data, and it appears that although aerobic sewage  plant  treatment  can
effectively degrade p-chloro-m-cresol, it is uncertain whether biodegrada-
tion will be significant in ambient surface waters.

92.2  Identification

    p-Chloro-m-cresol has been detected in both primary  and secondary
effluents (Shackelford and Keith 1976).

                                             Alternate Names

                                             4-Chloro-m-cresol
                                             4-Chloro-3-methylphenol
                                             2~Chloro-5-hydroxytoluene
    p-Chloro-m-cresol

    CAS NO.  59-50-7
    TSL NO.  GO 71000

92.3  Physical Properties

    The general physical properties  of  p-chloro-m-cresol  are  as  follows.

    Molecular weight                         142.59
    (Weast 1977)

    Melting  point                            66C
    (Weast 1977)

    Boiling  point at 760 torr                235C
    (Weast 1977)

    Vapor pressure                           No  data  found
                                     92-1

-------
    Solubility in water at  20C              3850 mg/1
    (Windholz 1976)

    Log octanol/water partition coefficient   2.95
    (Calc.  by method of Tute 1971)

    pKa                                   No data  found

92.4  Summary of Fate Data

    92.4.1  Photolysis

         No specific information pertaining  to the  photolysis  of  p-chloro-
m-cresol was found in the reviewed  literature.   The ultraviolet  absorption
spectrum of the undissociated compound in methanol  has  a  maximum  at  281 nm
that extends out to 300 nm  while in basic solutions the maximum absorption
peak of the anion shifts to 298 nm  (Sadtler  Standard Spectra 1975).   These
electromagnetic absorption  characteristics are almost  identical to those
exhibited by the structurally similar compound,  4-chlorophenol (Drahonovsky
and Vacek 1971), and it can, therefore,  be expected by  inference  that the
photolytic degradation of p-chloro-m-cresol  will be similar in some  ways to
that of 4-chlorophenol.

         Omura and Matsuura (1971)  found that irradiation of 4-chlorophenol
in aqueous alkali produced  hydroquinone both at  254 nm  and above  290 nm.
When cyanide ion was present in the photolysis solutions, some of the sub-
stituent chlorine was replaced by cyano groups.   This  latter observation
supports a reaction mechanism involving interaction of  hydroxide  ion in
preference to water with the photolyzing carbon-chlorine  bond.   The  pho-
tolysis of 4-chlorophenol to hydroquinone also has  been reported  by
Grabowski (1961) at 313 nm indicating that it is probably the  anionic form
which undergoes carbon-chlorine fission.

         The fact that the  chlorine and methyl substituents occupy adjacent
positions on the aromatic ring of p-chloro-m-cresol opens the  possibility
for photolytically induced interaction between these two  groups.  Pinhey
and Rigby (1969) have reported that 4-chlorophenol  readily undergoes photo-
reduction in 2-propanol to yield phenol.  Analysis  of  the reaction mixture
indicated that hydrogen atoms were  being abstracted from  the solvent,
2-propanol, by what can be assumed  to be a phenoxyl radical and  a chlorine
atom.  Insofar as the hydrogen atoms of aromatic methyl groups are easily
abstracted, it is possible that intramolecular photolysis of p-chloro-m-
cresol could produce a mixture of compounds  from initial  intermediates in
which the methyl group had become chlorinated or had become oxidized to a
benzyl hydroperoxide.  Although no  specific  data pertaining to the aquatic
photolysis of p-chloro-m-cresol were found in the reviewed literature, this
supposition is supported by the observations of Pinhey  and Rigby  (1969) and
                                    92-2

-------
the well documented photochemistry of aromatic methyl substituents (Porter
and Norman 1954;  Morrison 1969;  Wei and Adelman 1969).

    92.4.2  Oxidation

         No information was found in the reviewed literature from which to
assess the possibility of direct oxidation in the aquatic environment.

    92.4.3  Hydrolysis

         There are no data to suggest that hydrolysis of p-chloro-m-cresol
is an environmentally significant process.  The covalent bond of a substi-
tuent attached to an aromatic ring is usually resistant to hydrolysis be-
cause of the high negative charge-density of the aromatic nucleus (Morrison
and Boyd 1973).

    92.4.4  Volatilization

         The fact that p-chloro-m-cresol has a boiling point of 235C
(Weast 1977) in addition to an aqueous solubility of 3850 mg/1 at 20C
(Windholz 1976) indicates that volatilization from water at ambient en-
vironmental temperatures will not exert an overbearing consequence on the
fate of this pollutant.  Furthermore, aqueous p-chloro-in-cresol should be a
weak acid that will be partially ionized and highly solvated, thus further
increasing its persistence with respect to volatilization at low levels of
concentration.

    92.4.5  Sorption

         Although the calculated value of the octanol/water partition
coefficient indicated by log P = 2.95 suggests a definite potential for
sorption of p-chloro-m-cresol by organic particulates, Zogorski and Faust
(1976) have reported that some substituted phenolate anions are not easily
sorbed from water by lipophilic materials.  No information was found
pertaining to the interaction of this pollutant with clay minerals.

    92.4.6  Bioaccumulation

         Calculation of the log octanol/water partition coefficient, using
the method of Tute (1971), yields a value for log P of 2.95.  This indi-
cates that, except for the limits imposed by ionization and toxicity, this
compound should normally exhibit a tendency to bioaccuraulate.  No informa-
tion was found in the reviewed literature regarding the bioaccumulation of
p-chloro-m-cresol.
                                     92-3

-------
    92.4.7  Biotransformation and Biodegradation

         The microbial breakdown of p-chloro-m-cresol has been examined
under the simulated conditions of a wastewater treatment plant by Voets et
al. (1976).  Although quite susceptible to aerobic degradation in an or-
ganic supplemented medium, it was relatively persistent in an inorganic
mineral medium, and it appeared to be completely resistant to anaerobic
digestion.

         Adapted mixed cultures, isolated by enrichment techniques from
garden soil, compost, river mud, and the sediment of a petroleum refinery
waste lagoon, were shown to be capable of partially degrading p-chloro-m-
cresol (Tabak ejt _a_l. 1964).  It is questionable, however, whether these
studies can be extrapolated to the environment of ambient surface waters
since the concentration of the substrate chemical employed for enrichment
of an organism and for obtaining a reasonable amount of cell growth is far
above the concentrations generally found in nature.

92.5  Data Summary

    Table 92-1 summarizes  the aquatic fate data of p-chloro-m-cresol.  The
most probable fate is intramolecular photolysis.  The operational likeli-
hood of other fate processes does not appear to be tenable in ambient sur-
face waters.  Although aerobic sewage plant treatment can effectively de-
grade p-chloro-m-cresol, it is uncertain whether biodegradation will be
significant in ambient surface waters.
                                       92-4

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92.6  Literature Cited

Drahonovsky,  J. and Z. Vacek.   1971.   Dissoziationskonstanten und
  austauscherchromatographie chlorierter  phenole.   Coll.  Czech.  Chem.
  Commun.  36(10):3431-3440.

Grabowski, Z.R.  1961.  Photochemical  reactions  of  some  aromatic halogen
  compounds.   Z. Physik. Chem.   27:239-243.

Morrison, H.A.  1969.  The photochemistry of  the nitro and  nitroso  groups.
  H. Feuer (ed.)  The chemistry of  the nitro  and nitroso  groups.  Part  I.
  Chap. 4.  pp.165-212.  Interscience  Publishers, New York.

Morrison, R.T. and R.N. Bpyd.   1973.   Organic chemistry,  3rd  Edition.
  Allyn and Bacon,  Boston, Mass.  1258p.

Omura, K. and T. Matsuura.  1971.   Photolysis of halogenophenols in aqueous
  alkali and  in aqueous cyanide. Tetrahedron  27:3101-3109.

Pinhey, J.T.  and R.D.G. Rigby.   1969.   Photoreduction of  chloro- and
  bromo-aromatic compounds.  Tetrahedron  Lett.  (16):1267-1270.

Porter, G. and I. Norman.  1954.  Trapped atoms  and radicals  in  a  glass
  cage.  Nature 174(4428) :508~509.

Sadtler Standard Spectra.  1975.  4-Chloro-m-cresol.  Sadtler Research
  Laboratories, Inc., a Subsidiary of  Block Engineering,  Inc.

Shackelford,  W.M. and L.H. Keith.   1976.   Frequency of  organic compounds
  identified  in water.  U.S. Environmental Protection Agency, (ERL),
  Athens, Ga.  617p.  (EPA 600/4-76-062) .

Tabak, H.H.,  C.W. Chambers, and P.W. Kabler.   1964.  Microbial metabolism
  of aromatic compounds.  I.  Decomposition of phenolic  compounds  and
  aromatic hydrocarbons by phenol-adapted bacteria.  J.  Bacteriol.
  87(4):910-919.

Tute, M.S.  1971.  Principles and practice of Hansch analysis:  a  guide to
  structure - activity correlation for the medicinal  chemist. Adv. Drug
  Res. 6:1-77.

Voets, J.P.,  P. Pipyn, P. Van Lancker, and W. Verstraete.  1976.  Degra-
  dation of microbicides under different  environmental  conditions.   J.
  Appl.  Bact.  40(l):67-72.
                                       92-6

-------
Weast, R.C. (ed).  1977.  Handbook of chemistry and physics.   CRC Press,
  Inc., Cleveland, Ohio.  2398p.

Wei, K.S. and A.H. Adelman.  1969.  The photooxidation of toluene.  The
  role of an excited charge transfer complex.  Tetrahedron Lett.
  (38):3297-3300.

Windholz, M. (ed).  1976.  The Merck Index.  Ninth Edition.  Merck and  Co.,
  Rahway, N.J.  1313p.

Zogorski, J.S. and S.D. Faust.  1976.  The effect of phosphate buffer on
  the adsorption of 2,4-dichlorophenol and 2,4-dinitrophenol.   J. Environ.
  Sci. Health 9:501-515.
                                     92-7

-------
                         93.  4,6-DINITRO-o-CRESOL


93.1  Statement of Probable Fate

    Based on the information gathered for nitrophenols and 4,6-dinitro-o-
cresol itself, this pollutant will probably undergo slow photooxidation in
an aerated aquatic environment.  There is a possibility for photoreduction
of the nitro group if the 4,6-dinitro-o-cresol becomes absorbed by organic
particulates.  4,6-Dinitro-o-cresol should be strongly sorbed by clay
minerals and may even undergo hydrolysis on the clay surface.  Volatiliza-
tion and bioaccumulation appear to be unlikely processes, and although
biotransformation of 4,6-dinitro-o-cresol has been demonstrated, it is un-
certain whether biodegradation will take place in ambient surface waters.

93.2  Identification

    4,6-Dinitro-o-cresol has been identified in industrial effluents
(Shackelford and Keith 1976).  The chemical structure is shown below.
                                             Alternate Names

                                             DNOC
                                             2,4-Dinitro-6-methylphenol
    4,6-Dinitro-o-cresol

    CAS NO. 534-52-1
    TSL NO. GO 96250

93.3  Physical Properties

    The physical properties of 4,6-dinitro-o-cresol are as follows.

    Molecular weight                         198.13
    (Verschueren 1977)

    Melting point                            85.8C
    (Verschueren 1977)

    Boiling point                            No data found

    Vapor pressure                           No data found*
                                      93-1

-------
    Solubility in water                      No data found

    Log octanol/water partition coefficient  2.85
    (Calc. by method of Tute 1971)

    PKa                                      4.35
    (Pearce and Simkins 1968)
*The vapor pressure of pure 3,5-dinitro-o-cresol has been determined to be
 5.2 x 10~5 torr at 20C (Balson 1947).

93.4  Summary of Fate Data

    93.4.1  Photolysis

         4,6-Dinitro-o-cresol is a moderately acidic substance, pKa=4.35
(Pearce and Simkins 1968), and will exist substantially as an. anion in en-
vironmental surface waters.  The ultraviolet absorption spectrum of 4,6-di-
nitro-o-cresol in dioxane exhibits a maximum at about 265 nm which extends
out beyond 400 nm (Bohmer _e_t al. 1972).  In aqueous sodium hydroxide there
is a strong maximum at 372 nm.

         Nakagawa and Crosby (1974) report that 4-nitrophenol was degraded
in aqueous solutions within a period of 1-2 months when it was exposed to
sunlight at a concentration of 200 mg/1.  The principal products were
hydroquinone and 4-nitrocatechol.  A dark, acidic intractable polymer was
also produced.  Although no specific information was found in the reviewed
literature demonstrating that 4,6-dinitro-o-cresol would also be photo-
chemically hydroxylated in an analogous manner, the following mechanistic
considerations for this photochemical reacton indicate that, under similar
conditions, 4,6-dinitro-o-cresol might be degraded to a mixture of com-
pounds which could possibly include 2-methyl-4-nitrocatechol and
2-methyl-6-nitrohydroquinone.

         Nakagawa and Crosby (1974) proposed that the observed reaction
products arising from 4-nitrophenol were the result of photonucleophilic
displacement reactions involving water or hydroxide ion.  This type of re-
action mechanism can be easily rationalized for the formation of hydro-
quinone, but it is probably not valid in the case of 4-nitrocatechol since
it is difficult to envision the development of a center of electron de-
ficiency on the unsubstituted carbon atom adjacent to the phenolic group.
Suarez _et _al. (1970) and Giinther et_ al. (1971) have demonstrated that
hydroxyl radicals preferentially attack 4-nitrophenol in water at the C-2
and C-4 positions, and that the latter intermediate results in the dis-
placement of the nitro group by a hydroxy group.  Even though attack by
hydroxyl radical can more easily explain the observed products, it is un-
certain whether the concentration of hydroxyl radicals in the experiment of
                                    93-2

-------
Nakagawa and Crosby (1974) would have been sufficient to be responsible for
the formation of hydroquinone and 4-nitrocatechol.

         It has been reported that the ultraviolet irradiation of 4,6-dini-
tro-o-cresol at 254 nm is without effect (Mitchell 1961).  This experiment,
however, was conducted by irradiating the compound for 30 minutes after its
solution had been applied to a sheet of chromatographic paper which was
then dried.  The experiment was one of very short duration and the condi-
tions were anhydrous or hypohydrous.  For these reasons it is probably not
valid to draw any conclusions from this experiment with respect to either
the aquatic or terrestrial environment.

         A further photochemical reaction of 4,6-dinitro-o-cresol that must
be considered is photored'uction of the nitro groups.  As an example,
Nakagawa and Crosby (1974) found that the nitro group of nitrofen, a nitro-
aromatic pesticide, underwent photochemical reduction to amino and azo
groups in an aqueous solution that contained 10% methanol.  Aromatic nitro
groups are, in general, reduced photochemically in the presence of suitable
hydrogen donors (Morrison 1969),  The log octanol/water partition coeffi-
cient of 4,6-dinitro-o-cresol indicates a definite potential for absorp-
tion by suspended organic matter.  In this way the organic material could
serve as a reducing agent in the photoreduction.  Massini and Voorn (1967)
report that the reduction of 2,4-dinitrophenol by ascorbic acid is sensi-
tized by the presence of chlorophyll and is enhanced further by ferrous
ion.  Inasmuch as the functional groups of 4,6-dinitro-o-cresol are in the
same relative positions as 2,4-dinitrophenol,  both compounds may undergo
this type of reduction.  Intramolecular reduction of 4,6-dinitro-o-cresol
involving the methyl group, however, is not as likely a process as it is
for the nitrotoluenes because the methyl group is not adjacent to either
nitro group (Morrison 1969).

    93.4.2  Oxidation

         Suarez _e_t_ al_. (1970) and Gunther _e_t_ al. (1971) have demonstrated
that hydroxyl radicals attack both 2-nitrophenol and 4-nitrophenol at the
C-2 and C-4 positions resulting in the formation of a complex mixture that
follows from the generation of isomeric benzosetniquinones.  No information
was found, however, from which an environmentally relevant rate could be
estimated for this reaction.  It can be inferred that 4,6-dinitro-o-cresol
might be oxidized in an analogous fashion.

    93.4.3  Hydrolysis

         The covalent bond of a substituent attached to an aromatic ring is
usually resistant to hydrolysis because of the high negative charge-density
of the aromatic nucleus.  Nonetheless, there is a possibility, albeit not
specifically substantiated in the reviewed literature, that nitrophenols
                                    93-3

-------
could undergo hydrolysis while sorbed by clay minerals.   Specifically, the
water molecules within the clay structure can be highly  acidic, and the
phenolic group of the sorbed 4,6-dinitro-o-cresol would  become more acidic
than when the molecule is in aqueous solution due to coordination of the
phenolic oxygen atom with the metal cations of the clay.  This acidic en-
vironment could conceivably lead to the hydrolysis of the 4,6-dinitro-o-
cresol anion in a manner similar to the hydrolysis of aci-nitroparaffins
(Johnson and Degering 1943).
    93.4.4  Volatilization

         Although phenol has a significant vapor pressure as a pure sub-
stance, nitrophenols can form intennolecular hydrogen bonds in the solid
state and can exhibit boiling points much above that of a phenol with the
same molecular weight and the absence of nitro groups.   Moreover, the vapor
pressure of an ionized substance in water is proportional to the percentage
of the compound that is not ionized.  Under most environmental conditions
this would imply that the vapor pressure in water of 4,6-dinitro-o-cresol
would be much less than the vapor pressure of the pure substance.  (The
vapor pressure of pure 3,5-dinitro-o-cresol has been determined to be 5.2 x
10~5 torr at 20C;  Balson 1947).  Solution interactions would reduce the
vapor pressure even more.  Thus, volatilization would probably not be an
important process for nitrophenols in general.  For exam-pie, 4-nitrophenol
does not volatilize from a solution of boiling water (Morrison and Boyd
1973).

    93.4.5  Sorption

         Although the value of the octanol/water partition coefficient
indicated by log P = 2.85 suggests a definite potential for sorption of
4,6-dinitro-o-cresol by organic particulates, Zogorski and Faust (1976)
have reported that substituted phenolate anions are not easily sorbed from
water by lipophilic materials.  Although no information was found per-
                                     93-4

-------
taining to the interaction of this pollutant with clay minerals, there is
no reason to believe that it will behave differently from the other nitro-
phenols (Chang and Anderson 1968).  Thus adsorption by clay can be expected
to be a significant transport process.

    93.4.6  Bioaccumulation

         Calculation of the log octanol/water partition coefficient, using
the method of Tute (1971), yields a value for log P of 2.85.  Within the
limits imposed by ionization and toxicity, this compound thus could exhibit
a tendency to bioaccumulate.  Although no specific information was found in
the reviewed literature regarding the bioaccumulation of 4,6-dinitro-o-
cresol, the compound cannot be expected to bioaccumulate because of its
marked toxicity.

    93.4.7  Biotransformation and Biodegradation

         Most of the studies of microbial degradation of nitrophenols have
concentrated on the isolation of pure cultures which can utilize these
pollutants as sources of energy, carbon, or nitrogen (Howard jejt _al. 1976).
It is questionable whether these studies can be extrapolated to the en-
vironment of ambient surface waters since the concentration of the test
chemical employed for enrichment of an organism and for obtaining a. reason-
able amount of cell growth is far above the concentrations generally found
in nature.

         4,6-Dinitro-o-cresol (DNOC) has been used as a herbicide for
several decades and its rate of decomposition in soil has been well
studied.  DNOC usually disappears from the soil within a few weeks to two
months (Gundersen and Jensen 1956;  Bruinsma I960;  Hurle and Pfefferkorn
1972).  Gundersen and Jensen (1956) isolated an Arthrobacter and a
Pseudomonas that grew on DNOC with the release of nitrite ion.  Tewfic and
Evans (1966) found that in pure culture Pseudomonas degraded DNOC to an
aminocresol whereas Arthrobacter initially hydroxylated the aromatic ring
before catabolism proceeded further.

93.5  Data Summary

    Table 93-1 summarizes the aquatic fate data for 4,6-dinitro-o-cresol.
This pollutant will probably undergo slow photolytic destruction in ambient
surface waters.  Volatilization is not thought to be an important transport
process but 4,6-dinitro-o-cresol might be strongly sorbed by clay minerals.
There is a possibility that 4,6-dinitro-o-cresol could undergo hydrolysis
while sorbed onto the clay structure.  Although the biotransformation of
4,6-dinitro-o-cresol has been demonstrated, it is uncertain whether biode-
gradation is an operational fate in ambient surface waters.
                                     93-5

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93.6  Literature Cited

Balson, E.W.  1947.  Studies in vapour pressure measurement.  Part III.  An
  effusion mamometer sensitive to 5 x 10~6 millimetres of mercury:
  vapour pressure of DDT and other slightly volatile substances.  Trans.
  Faraday Soc. 43:54-60.

Bb'hiner, V., J. Deveaux, and H. Kammerer.  1972.  Die additive
  zusammensetzung der UV-spektren phenolischer mehrkernverbindungen mit
  nitrogruppen aus den spektren entsprechend substituierter nitrophenole.
  Spectrochim. Acta 28A:1977-1985.

Bruinsma, J.  1960.  The action of 4,6-dinitro-o-cresol (DNOC) in soil.
  Plant and Soil 12:249-258.

Chang, C.W. and J.U. Anderson.  1968.  Flocculation of clays and soils by
  organic compounds.  Proc. Soil Sci. Soc. Amer.  32(l):23-27.

Gundersen, K. and H.L. Jensen.  1956.  Soil bacterium decomposing organic
  nitro compounds.  Acta Agr. Scand.  6:100-114,

Giinther, K., W.G. Filby and K. Eiben.  1971.  Hydroxylation of substituted
  phenols:  an ESR-study in the Ti3-4"/H202 system.  Tetrahedron Lett.
  (3):251-254.

Howard, P.H., J. Santodonato, J. Saxena, J.E. Mailing, and D. Greninger.
  1976.  Investigation of selected potential environmental contaminants:
  nitroaromatics.  U.S. Environmental Protection Agency (Office of Toxic
  Substances), Washington, B.C., 600 p.  (EPA 560/2-76-010).

Hurle, K. and V. Pfefferkorn.  1972.  Experiments on the rate of
  decomposition of DNOC in soils:  the effect of pre-treatments.  Proc. Br.
  Weed Control Conf.  11(2)=806-810 .

Johnson, K. and E.F. Degering.  1943.  Production of aldehydes and ketones
  from nitroparaffins.  J. Org. Chem.  8(1):10-11.

Massini, P. and G. Voorn.  1967.  The effect of ferredoxin and ferrous ion
  on  the chlorophyll sensitized photoreduction of dinitrophenol.
  Photochern. Photobiol.  6:851-856.

Mitchell, L.C.  1961.  The effect of ultraviolet light (2537R) on 141
  pesticide chemicals by paper chromatography.  J. Assoc. Off. Agr. Chem.
  44:643-712.

Morrison, H.A.  1969.  The photochemistry of the nitro and nitroso
  groups.  H. Feuer (ed.).  The chemistry of the nitro and nitroso groups.
  Part I.  Chap. 4.  pp.165-212.  Interscience Publishers, New York.
                                       93-7

-------
Morrison, R.T. and R.N.  Boyd.   1973.   Organic  Chemistry,  3rd edition.
  Allyn and Bacon, Inc.,  Boston 1258p.

Nakagawa, M. and D.G. Crosby.   1974.   Photodecomposition  of nitrofen.   J.
  Agr. Food Chem.  22(5):849-853.

Pearce, P.J. and R.J.J.  Simkins.  1968.   Acid  strengths of  some substituted
  picric acids. Can. J.  Chem.   46(2):241-248.

Shackelford, W.M. and L.H. Keith.   1976.   Frequency of organic compounds
  identified in water.  U.S.  Environmental Protection Agency, (ERL),
  Athens, Ga.  617p. (EPA 600/4-76-062).

Suarez, C., F. Louys, K.  "Giinther and  K.  Eiben.  1970.  Hydroxyl radical
  induced denitration of  nitrophenols.   Tetrahedron Lett.  (8):575-578.

Tewfic, M.S. and W.C. Evans.   1966.  The metabolism of 3,5-dinitro-o-cresol
  (DNOC) by soil microorganisms.  Biochem. J.  99:31P.

Tute, M.S.  1971.  Principles  and practice of  Hansch analysis:  a guide to
  structure-activity correlation for  the medicinal chemist.  Adv. Drug
  Res. 6:1-77.

Verschueren, K.  1977.  Handbook of environmental data on organic
  chemicals.  Elsevier/Van Nostrand,  New York.  659p.

Zogorski, J.J. and S.D.  Faust.  1976.  The effect of phosphate buffer  on
  the adsorption of 2,4-dichlorophenol and 2,4-dinitrophenol.  J. Environ.
  Sci. Health 9:501-15.
                                    93-8

-------
SECTION VIII:  PHTHALATE ESTERS
          Chapter 94

-------
                          94.  PHTHALATE ESTERS:
        DIMETHYL PHTHALATE,  DIETHYL PHTHALATE,  DI-n-BUTYL PHTHALATE.
DI-n-OCTYL PHTHALATE, BIS(2-ETHYLHEXYL) PHTHALATE, BUTYL BENZYL PHTHALATE
94.1  Statement of Probable Fate

    Bis(2-ethylhexyl) phthalate is the most well studied of the phthalate
esters.  For several of the phthalate esters, however, very little specific
data were found, and the aquatic fate of these compounds is to a large ex-
tent inferred from data for phthalate esters as a group.  Although their
solubilities vary from sparingly soluble to moderately soluble, they all
are probably readily adsorbed onto suspended particulates and biota and,
under certain conditions, are likely to form a water soluble complex with
humic substances.  Their transport will largely depend on the hydrogeologic
conditions of the aquatic system.  Volatilization is not considered to be a
competitive transport process, with the possible exception of those esters,
such as bis(2-ethylhexyl) and butyl benzyl phthalate, that have low solu-
bilities .

    A variety of organisms have demonstrated the ability to take up and ac-
cumulate phthalate esters; this is probably due to the esters lipophili-
city.  They have also been shown to become concentrated in animal and human
tissues and organs.  Mixed microbial systems can degrade phthalate esters
under aerobic conditions.  Degradation is generally slower under anaerobic
conditions and .ceases to be effective for bis(2-ethylhexyl) phthalate.   A
variety of multicellular organisms have demonstrated the ability to bio-
transform and eliminate phthalate esters.  Hydrolysis will occur in the
water column, but it may be too slow to be environmentally significant.
Bioaccu
-------






















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