Document ID: EPA-HQ-OPP-2004-0402-0035
Agency: epa
Document Type: Supporting & Related Material
Title: 
Posted Date: 2005-03-18T05:00Z

Page
1
of
49
ENVIRONMENTAL
FATE
SCIENCE
CHAPTER
FOR
CHLORINATED
DIBENZODIOXINS
AND
CHLORINATED
DIBENZOFURANS
(
CDD/
CDF)
AS
IMPURITIES
IN
TECHNICAL
GRADE
PESTICIDES
3/
5/
05
Page
2
of
49
EXECUTIVE
SUMMARY
Pentachlorophenol
(
PCP),
one
of
the
major
pesticides,
is
primarily
used
worldwide
as
a
wood
preservative
.
PCP
contains
chlorinated
dibenzodioxins
and
chlorinated
dibenzofurans
(
CDD/
CDFs)

as
contaminants
and
these
are
formed
during
the
manufacture
of
PCP.
The
CDD/
CDF
impurities
from
PCP
may
be
introduced
into
the
environment
when
pentachlorophenol
is
used
as
a
pesticide.

The
main
use
of
PCP
is
to
treat
utility
poles.
There
are
an
estimated
36
million
PCP­
treated
utility
poles
in
service
in
the
United
States.
Annually,
nearly
1
million
additional
utility
poles
are
replaced
(
3
percent
replacement
rate)
on
land
and
water
environment.
The
U.
S.
Environmental
Protection
Agency
has
estimated
that
the
utility
poles
in
service
contain
approximately
374
kg
of
dioxin
toxicity
equivalents
(
I­
TEQs).
The
CDD/
CDFs
in
these
poles
may
be
released
into
the
environment
via
volatilization
and
leaching.
In
addition,
CDD/
CDFs
may
enter
the
environment
during
the
pressuretreatment
of
the
utility
poles
when
the
utility
poles
are
removed
from
service
and
are
disposed
in
landfills.

CDD/
CDFs
In
Environmental
Compartments:

Once
released
into
the
environment,
CDD/
CDFs
are
primarily
associated
with
particulate
and
organic
matter
in
air,
soil,
sediment,
and
water
column
because
of
their
high
lipophilicity
and
low
water
solubility.
If
released
into
the
atmosphere,
CDD/
CDFs
will
partition
between
the
atmospheric
particles
and
gas
phase,
with
the
majority
of
CDD/
CDFs
being
adsorbed
onto
the
suspended
particles.

CDD/
CDFs
released
into
air
can
be
transported
long
distances
and
then
deposited
onto
vegetation,

soils,
and
water
via
dry/
wet
deposition.
CDD/
CDFs
released
into
soils
are
strongly
adsorbed
onto
soils
which
have
a
high
K
OC
as
these
dioxin
are
lipophilic
,
but
may
be
transported
into
water
bodies
due
to
storm
runoffs
or
soil
erosion.
CDD/
CDFs
sorbed
to
soil
exhibit
little
potential
for
significant
leaching
or
volatilization
once
sorbed
to
particulate
matter.
Upon
entering
the
water
column,

CDD/
CDFs
will
partition
between
water
and
suspended
particles
or
humic
substances
and
then
undergo
sedimentation
and
burial
with
some
uptake
by
aquatic
biota.
The
ultimate
environmental
sink
of
CDD/
CDFs
appears
to
be
aquatic
sediments.

A.
Abiotic
Degradation
of
CDD/
CDFs
in
Environmental
Compartments
a.
Hydrolysis
Page
3
of
49
Published
literature
does
not
report
any
hydrolytic
studies
on
any
of
the
congeners
of
CDDs
or
CDFs,
either
measured
or
estimated.

b.
Photolysis
in
Air/
Water/
Soils
The
available
evidence
indicates
that
CDDs
and
CDFs,
particularly
the
tetra­
and
higher
chlorinated
congeners,
are
extremely
stable
compounds
under
most
environmental
conditions.
The
only
environmentally
significant
transformation
processes
for
these
congeners
are
believed
to
be
photolysis
and
atmospheric
photooxidation
of
nonsorbed
species
in
the
gaseous
phase
or
at
the
soil
or
water­
air
interface.

The
mid­
summer
photolysis
half­
life
values
at
40
degrees
north
latitude
in
clear
near­
surface
water
for
2,3,7,8­
substituted
congeners,
estimates
based
on
one
laboratory
study,
ranged
from
6
hours
for
1,2,3,4,7,8­
HxCDF
to
47
days
for
1,2,3,4,6,7,8­
HpCDD.
However,
in
natural
water,
the
photolysis
half­
lives
are
expected
to
be
much
longer
because
the
majority
of
CDD/
CDFs
will
be
adsorbed
onto
particles
or
humic
substances.
Photolysis
of
CDD/
CDFs
in
soils
is
limited
only
to
the
near­
surface
area.
In
one
laboratory
study
(
conducted
with
two
soils
irradiated
for
15
days),

approximately
38
to
42
percent
of
OCDD
was
degraded
by
day
5
but
no
further
degradation
was
observed
in
the
following
10
days.
The
photolysis
half­
life
of
2,3,7,8­
TCDD
in
the
gaseous
phase
in
summer
sunlight
at
40
degrees
north
latitude
was
estimated
to
be
1
hour
as
an
upper
limit,
based
on
the
quantum
yield
for
photolysis
in
hexane.
However,
the
results
of
one
laboratory
study
indicated
that
CDD/
CDFs
adsorbed
onto
fly
ashes
are
very
stable
and
are
not
subject
to
significant
photodegradation
under
irradiation
with
a
medium
pressure
mercury
lamp
for
2
to
6
days.

Photo
oxidation
rates
of
vapor
phase
CDD/
CDFs
have
not
been
experimentally
measured.

However,
based
on
the
estimated
hydroxy
(
OH
)
radical
reaction
rate
constants
for
CDD/
CDFs,
one
researcher
concluded
that
the
OH
radical
reaction
is
likely
to
be
the
dominant
gas
phase
transformation
process
for
vapor
phase
CDD/
CDFs.
The
tropospheric
life
times
of
vapor
phase
CDD/
CDFs,
estimates
based
on
the
OH
radical
reaction
rate
constants,
ranged
from
2.0
days
for
2,3,7,8­
TCDD,
8.6
days
for
OCDD,
to
39
days
for
OCDF.
However,
because
the
majority
of
CDD/
CDFs
are
adsorbed
onto
the
air­
borne
particles,
the
lifetime
of
CDD/
CDFs
in
the
atmosphere
is
expected
to
be
much
longer.

B.
Biotic
Degradation
in
Environmental
Compartments
Page
4
of
49
CDD/
CDFs
congeners
are
highly
resistant
to
biodegradation.
One
study
indicated
that
of
the
approximately
100
strains
of
micro­
organisms
shown
previously
to
degrade
persistant
pesticides,
only
five
strains
showed
the
ability
to
degrade
2,3,7,8­
TCDD,
based
on
autoradiographs
of
thin­
layer
chromatograms.
In
another
study,
after
incubation
of
radiolabeled
2,3,7,8
­
TCDD
in
soils
for
one
year,
no
significant
degradation
was
observed.
In
an
investigation
into
the
persistence
of
CDD/
CDFs
in
contaminated
soils
in
Germany,
neither
an
appreciable
loss
of
CDD/
CDFs,
nor
significant
changes
in
the
congener
profiles,
either
by
volatilization
or
decomposition,
could
be
detected
over
a
period
of
9
years.
Several
studies
have,
however,
indicated
that
certain
ligninolytic
fungi
can
degrade
higherchlorinated
congeners
and
that
anaerobic
degradation
in
sediment
may
occur
at
a
slow
rate.
To
a
large
extent,
these
biodegradation
processes
involve
dechlorination
to
less­
chlorinated
(
and
possibly
more
toxic)
congeners.

Several
laboratory
studies
have
demonstrated
biochemical
formation
of
CDD/
CDFs
from
technical
grade
pentachlorophenol.
There
is
speculation
that
CDD/
CDFs
commonly
found
in
compost
might
result
in
part
from
this
mechanism,
based
on
the
data
where
compost
containers
were
found
to
contain
used
PCP­
treated
wood.
However,
the
extent
to
which
CDD/
CDFs
are
formed
in
the
environment
via
this
mechanism
can
not
be
estimated
at
this
time.

C.
Bioaccumulation
of
CDD/
CDFs
Fish
and
invertebrates
can
likely
bioaccumulate
2,3,7,8­
substituted
CDD/
CDFs
from
water
columns
and
sediments.
However,
because
most
CDD/
CDFs
in
a
water
column
and
sediment
are
associated
with
particulate
matter
and
dissolved
organic
matter,
bioaccumulation
most
likely
starts
with
uptake
of
CDD/
CDFs
by
benthic
organisms
directly
from
sediment
pore
waters
and
by
ingestion
of
contaminated
particles.
Organisms
preying
on
benthic
organisms
would
possibly
transfer
the
CDD/
CDFs
up
the
food
chain
but
no
sound
scientific
data
have
been
obtained.
Page
5
of
49
ENVIRONMENTAL
FATE
OF
CHLORINATED
DIBENZODIOXINS
AND
CHLORINATED
DIBENZOFURANS
(
CDD/
CDFS)
IMPURITIES
IN
PENTACHLOROPHENOL
AND
OTHER
CHLORINATED
PESITICIDES
Chlorinated
dibenzodioxins
and
chlorinated
dibenzofurans
are
not
classified
in
a
pesticide
category
and
are
not
registered
with
the
Agency
as
such.
The
Office
of
Pesticides
Programs
(
OPP)

does
not
have
a
database
on
environmental
fate
studies
on
CDDs/
CDFs.
Their
presence
in
the
environmental
compartments
is
through
many
sources.
Thus
environmental
fate
chapter
is
based
on
the
OPP's
search
of
published
literature
on
CDDs
and
CDFs.

Presence
of
CDDs
and
CDFs
in
the
Environmental
Compartments
A.
Release
of
CDD/
CDFs
from
Treated
Wood
Currently,
pentachlorophenol
is
used
mainly
in
the
United
States
as
a
wood
preservative,
for
treatment
of
utility
poles,
crossarms,
railroad
ties,
and
fenceposts.
The
major
current
use
of
PCP
in
the
U.
S.
is
to
treat
utility
poles.
Technical
grade
pentachlorophenol
contains
chlorinated
dibenzodioxins
and
chlorinated
dibenzofurans
(
CDD/
CDFs),
which
are
formed
during
the
manufacture
of
pentachlorophenol.
PCP
produced
prior
to
1987
contained
approximately
3
mg
I­
TEQ
(
International
Total
Equivalent
Quantity)
per
kg.
If
this
degree
of
contamination
still
applies
(
PCP
produced
today
contains
less
TEQ),
then
30
kg
TEQ
are
introduced
into
the
U.
S.
environment
annually
in
the
form
of
utility
poles
and
other
treated
wood.
Currently,
there
are
an
estimated
36
million
PCP­
treated
utility
poles
in
service
across
the
United
States.
EPA
estimated
that
these
utility
poles
contain
approximately
374
kg
of
I­
TEQ
(
Winters
et
al.,
1999).
CDD/
CDFs
may
be
released
into
the
environment
from
treated
wood
via
volatilization
and
leaching.
In
addition,
an
estimated
1.08
million
utility
poles
are
removed
from
service
per
year
in
the
United
States
(
3
percent
of
36
million
in
service).
Information
on
the
disposal
practices
for
used
utility
poles
in
the
United
States
has
not
been
documented.
However,
in
Canada,
utility
poles
removed
from
service
are
either
re­
processed
and
re­
used
as
utility
poles,
roof
shakes,
fences,
guide
rails,
sign
posts,
and
other
products,
or
are
disposed
of
in
landfills
(
Cooper
et
al.,
1996).

B.
Volatilization
and
Leaching:

No
guideline
studies
for
the
release
rates
of
dioxins
are
available
nor
were
studies
required
by
the
Agency
from
the
registrants.
Unequivocal
results
of
release
rates
have
not
been
determined.

However,
the
release
rates
of
CDD/
CDFs
from
utility
poles
(
specifically
volatilization
rate
and
leaching
rate
of
CDD/
CDFs)
are
the
most
critical
parameters
required
for
estimating
quantities
of
CDD/
CDFs
released
into
the
environment
from
use
of
pentachlorophenol
and
possibly
other
Page
6
of
49
chlorinated
phenols
on
utility
poles.
Several
approaches
have
been
used
to
estimate
the
volatilization
rate
of
CDD/
CDFs
from
poles.

Approach
#
1:
Eitzer
and
Hites
(
1987)
estimated
the
release
rate
via
volatilization
from
poles
to
be
0.1
percent
per
year,
based
on
the
assumption
that
CDD/
CDFs
in
the
poles
volatilize
at
the
same
rate
as
assumed
for
pentachlorophenol.

Approach
#
2:
Bremmer
et
al.
(
1994)
estimated
the
volatilization
rate
to
be
0.45
percent
per
year,

based
on
a
reported
half­
life
of
15
years
for
pentachlorophenol
in
poles
and
the
assumption
that
CDD/
CDFs
volatilize
ten
times
slower
than
pentachlorophenol.

Approach
#
3:
In
a
study
sponsored
by
the
Penta
Task
Force
(
Weinburg
Group,
Inc.,
1998),
a
soil
diffusion
model
was
used
to
estimate
volatilization
rates
of
specific
CDD/
CDF
congeners
from
poles.

The
estimated
volatilization
rates
ranged
from
0.001
percent
per
year
for
OCDF
to
0.004
percent
per
year
for
2,3,7,8­
HxCDD.

Winters
et
al.
(
1999)
measured
the
CDD/
CDF
content
and
profile
in
PCP­
treated
utility
poles
in
an
attempt
to
determine
the
movement
of
CDD/
CDFs
within
a
utility
pole.
Results
indicated
that
lower
chlorinated
CDD/
CDFs
migrated
towards
the
outside
of
the
poles,
but
a
conclusion
could
not
be
made
that
migration
enhances
the
release
of
CDD/
CDFs.
Various
methods
and
approaches
notwithstanding,
determination
of
reliable
release
rates
has
been
elusive.

A
Canadian
study
(
Gurprasad
et
al.,
1995)
on
leaching
of
CDD/
CDFs
from
PCP­
treated
poles
onto
the
soil
immediately
adjacent
to
the
utility
poles
has
shown
that
significant
levels
of
HxCDD,
HpCDD,
and
OCDD
were
found
in
soil
at
a
2
cm
distance
from
the
poles.
The
CDD/
CDF
levels
dropped
by
an
order
of
magnitude
at
a
distance
of
20
cm.
The
congener
profiles
of
CDD/
CDFs
in
soils
suggest
that
CDD/
CDFs
originated
from
the
utility
pole.
However,
a
leaching
rate
for
CDD/
CDFs
from
utility
poles
was
not
estimated.

No
environmental
related
study
to
date
has
estimated
the
exact
quantities
of
dioxins
released
to
the
environment­
soil/
air/
water
from
treated
woods.

C.
Transport
Mechanisms
in
Air
Although
not
directly
related
to
making
risk
assessment
based
on
fate
studies
required
by
OPP
guidelines
an
important
factor
to
consider
which
possibly
has
an
environmental
impact
is
the
longrange
atmospheric
transport
and
deposition
of
CDD/
CDF
released
to
the
atmosphere
by
combustion
and
other
sources
and
this
phenomenon
has
been
shown
to
be
a
major
contributor
to
the
background
Page
7
of
49
levels
of
CDD/
CDF
measured
in
environmental
media
(
Hites
and
Harless
1991;
Tysklind
et
al,
1993).

Long­
range
atmospheric
transport
and
deposition
of
CDD/
CDFs
are
mainly
controlled
by
three
physical
mechanisms:
(
1)
vapor/
particle
partitioning;
(
2)
dry
deposition;
and
(
3)
wet
deposition.

Table
1
presents
a
summary
of
some
of
the
deposition
rates
generated
by
investigators
in
the
United
States,
United
Kingdom,
Germany,
Sweden,
and
Belgium.
Based
on
the
limited
data
presented
in
Table
1,
CDD/
CDF
deposition
rates
are
apparently
higher
in
urban
areas
than
rural
areas.
(
Table
1
on
the
next
page).
Page
8
of
49
Table
1.
Summary
of
Selected
Deposition
Measurements
Reported
in
the
Literature
Author
Yeara
Sampling
Method
Analytes
Sampling
Locations
Range
of
Results
Horstmann
and
McLachlan
1996
Bergerhoff
CDD/
CDF
Germany
Rural
0.2­
2.3
ng
I­
TEQDF/
m2­
yr
Smith
et
al.
1995
Wet
deposition;
Ambient
air
samples
CDD/
CDF
New
York,
USA
Total
CDD/
CDF
flux
wet:
94
ng/
m2­
yr
dry:
100
ng/
m2­
yr
total:
194
ng/
m2­
yr
Wallenhorst
et
al.
1995
Bergerhoff
CDD/
CDF
Germany
Urban
Rural
11
ng
I­
TEQDF/
m2­
yr
2­
3
ng
I­
TEQDF/
m2­
yr
DeFré
et
al.
1994
Bergerhoff
CDD/
CDF
Flanders,
Belgium
Background
<
1
km
from
MSWI
Urban
0.7­
5.1
ng
I­
TEQDF/
m2­
yr
39­
374
ng
I­
TEQDF/
m2­
yr
13­
77
ng
I­
TEQDF/
m2­
yr
Hiester
et
al.
1993
Bergerhoff
CDD/
CDF/
PCB
Germany
Urban
Rural
3.6­
30.3
ng
I­
TEQDF/
m2­
yr
4.4
ng
I­
TEQDF/
m2­
yr
Liebl
et
al.
1993
Bergerhoff
CDD/
CDF
Germany
Urban
Rural/
Industrial
Rural
7.6
ng
I­
TEQDF/
m2­
yr
1.5
ng
I­
TEQDF/
m2­
yr
1.1
ng
I­
TEQDF/
m2­
yr
Andersson
et
al.
1992
Cotton
cloth;
snow
collector
CDD/
CDF
Umea,
Sweden
1
ng
I­
TEQDF/
m2­
yr
Fernandez
1992
Wet
and
dry
frisbee
collector
CDD/
CDF
United
Kingdom
Urban­
Semiurban
13­
17
ng
I­
TEQDF/
m2­
yr
Koester
and
Hites
1992a
Frisbees;
flat
glass
plates;

wet­
only
collector
CDD/
CDF
Indiana,
USA
Total
CDD/
CDF
flux
wet:
210­
220
ng/
m2­
yr
dry:
160­
320
ng/
m2­
yr
total:
370­
540
ng/
m2­
yr
Page
9
of
49
Mechanism
#
1:
Vapor/
Particle
(
V/
P)
Partitioning
Once
released
into
the
environment,
CDD/
CDFs
congeners
partition
between
vapor
and
airborne
particles
(
Hites
and
Harless,
1991;
Hippelein
et
al.,
1996).
Table
2
presents
a
summary
of
percentages
of
CDD/
CDFs
in
the
particulate
phase
measured
in
air
monitoring
studies
conducted
between
1989
and
1995.
The
key
parameters
controlling
the
vapor/
particle
partitioning
are
the
congener's
vapor
pressure,
the
atmospheric
temperature,
and
the
particulate
matter
concentration
in
the
atmosphere.
Congeners
with
higher
vapor
pressures
(
i.
e.,
the
less­
chlorinated
compounds)
are
found
to
a
greater
extent
in
the
vapor
phase,
as
indicated
in
Table
2.
For
a
given
congener,
the
fraction
in
the
vapor
phase
increases
with
increasing
ambient
temperature
and
decreases
with
increasing
particle
concentration.
One
portion
of
the
particle­
associated
compound
appears
to
be
freely
exchanged
between
the
particulate
and
vapor
phases
and
second
portion
may
be
irreversibly
sorbed
or
occluded
by
the
particles
and
not
in
equilibrium
with
the
gas
phase.

Mechanism
#
2:
Dry
Deposition
Dry
deposition
can
involve
two
phases,
dry
particulate
deposition
and
vapor
phase
deposition,

the
relative
importance
of
which
for
a
given
congener
is
dependent
primarily
on
the
vapor/
particle
(
V/
P)
partitioning.
First,
CDD/
CDFs
associated
with
particulate
matter
can
be
deposited
by
gravitational
settling
or
turbulent
diffusion.
Secondly,
CDD/
CDFs
can
be
deposited
by
vapor­
phase
diffusion
into
the
soil,
vegetation,
and
the
surface
layer
of
water
bodies.

As
noted
in
Table
2,
the
vast
majority
of
the
atmospheric
burden
of
hepta­
and
octachlorinated
CDD/
CDF
(
and,
to
a
lesser
extent,
the
burden
of
hexa­
and
penta­
chlorinated
congeners)

is
associated
with
particulate
matter.
As
such,
dry
particulate
deposition
is
a
major
mechanism
for
removal
of
these
congeners
from
the
atmosphere.
Few
studies
have
been
published
that
have
attempted
to
measure
only
dry
particulate
deposition
of
CDD/
CDFs.
Koester
and
Hites
(
1992a)
used
inverted
frisbees
and
flat
glass
plates
to
collect
dry
particulate
deposition.
Koester
and
Hites
(
1992a)

calculated
an
average
deposition
velocity
for
particulate­
associated
CDD/
CDFs
of
0.2
cm/
sec;

calculated
deposition
velocities
for
the
tetra­
through
octa­
chlorinated
congener
groups
ranged
from
0.086
to
0.6
cm/
sec.
Page
10
of
49
Table
2.
Percentages
of
CDD/
CDFs
in
Particulate
Phase
Measured
in
Air
Monitoring
Studies
Reference
Temp.

(

C)
Percent
of
Total
Congener
Group
Mass
in
Particulate
Phase
TCDD
PeCDD
HxCDD
HpCDD
OCDD
TCDF
PeCDF
HxCDF
HpCDF
OCDF
A
20
23
37
66
87
96
14
31
64
87
91
B
3
40
87
100
100
100
100
60
88
100
98
B
16
­
20
8
28
45
88
100
ND
28
30
93
100
B
>
28
5
13
45
60
100
ND
0
38
78
98
C
21
20
24
70
85
23
26
29
59
94
C
3
5
12
64
90
7
12
15
43
91
D
18
NR
NR
92
100
78
14
42
73
100
100
D
18
NR
NR
100
100
100
5
43
100
100
NR
E
(
urban)
NR
ND
0
65
82
100
20
71
100
100
100
E
(
rural)
NR
ND
ND
100
100
100
ND
ND
ND
ND
ND
F
18
10
28
45
77
93
9
22
48
77
89
G
9.5
31
59
82
>
96
>
97
18
55
79
>
93
>
94
NR
=
Not
reported.

ND
=
Not
detected.

Source:
Volume
IV.
References
are
as
follows:

Reference
A:
Eitzer
and
Hites
(
1989)

Reference
B:
Hites
and
Harless(
1991)

Reference
C:
Harless
and
Lewis
(
1992)

Reference
D:
Hunt
and
Maisel
(
1992)

Reference
E:
Bobet
et
al.
(
1990)

Reference
F:
Welsch­
Pausch
et
al.
(
1995)
(
data
provided
by
authors);
values
presented
for
HpCDD,
OCDD,
HpCDF,
and
OCDF
represent
lower
limits.
Page
11
of
49
Several
studies
have
shown
that
the
transfer
of
all
non­
hepta­
and
non­
octa­
chlorinated
CDD/
CDFs
to
leafy
vegetation
is
dominated
by
vapor
phase
deposition
which
involves
the
movement
of
vapor­
phase
dioxin
from
ambient
air
into
leafy
vegetation
(
Bacci,
et.
al.,
1990;
Gaggi
and
Bacci,

1985;
McLachlan,
et.
al.,
1995;
Rippen
and
Wesp,
1993;
Simonich
and
Hites,
1995).
Dry
particulate
and
wet
deposition
are
believed
to
be
the
dominant
mechanisms
by
which
vegetation
and
soil
are
exposed
to
hepta­
and
octa­
chlorinated
congeners.
Vapor
phase
deposition
directly
onto
soil
is
not
believed
to
be
a
dominant
process
in
most
settings
because
soil
is
usually
covered
by
vegetation
or
detritus
which
are
likely
to
serve
as
more
important
exchange
sites.

Mechanism
#
3
:
Wet
Deposition
Over
a
long
period
of
time,
wet
deposition
processes
are
believed
to
dominate
dry
deposition
in
terms
of
total
mass
deposition
of
CDD/
CDFs,
because
wet
deposition
is
the
primary
mechanism
responsible
for
removal
of
small
particulates
from
the
atmosphere.
For
removal
of
particulateassociated
chemicals,
wet
deposition
flux
is
the
product
of
the
particulate
scavenging
ratio
and
the
chemical
concentration
on/
in
various
particulate
size
fractions.
The
scavenging
ratio
is
calculated
as
the
product
of
the
scavenging
coefficient
and
precipitation
rate.

Listed
in
Table
3
are
the
average
precipitation
scavenging
ratios
for
congener
groups
reported
by
Hites
and
Harless
(
1991)
and
Koester
and
Hites
(
1992a)
for
Bloomington,
Indiana,
and
Indianapolis,
Indiana,
respectively.
Also
listed
in
Table
3
are
the
percentages
of
congener
groups
scavenged
as
particles
in
rain
rather
than
as
dissolved
solutes
in
rain.
Total
rain
scavenging
ratios
ranged
from
10,000
to
150,000;
hepta­
and
octa­
CDDs
(
i.
e.,
the
congeners
most
strongly
associated
with
particulates)
were
scavenged
most
efficiently.

A
phenomenon
that
is
not
common
to
other
pesticides
or
contaminant
pesticides
is
deposition.

Ths
process
appears
to
be
a
major
route
of
atmospheric
contamination.
However,
the
sources
of
dioxins
which
result
in
deposition
are
many
and
are
not
limited
to
pentachlorophenol
only.

Pentachlorphenol's
contribution
to
this
contamination
appears
minor.

D.
Transport
Pathways
of
CDDs/
CDFs
in
Soils
Open
literature
survey
on
dioxins
show
that
there
are
more
reports
on
the
quantitative
estimations
of
CDDs/
CDFs
and
their
interactions
with
soils.
In
the
discussion
that
follows
the
Agency
has
included
a
representative
literature
survey
on
dioxins
and
their
soil
interactions.
CDD/
CDFs
released
into
soils
are
strongly
adsorbed
onto
certain
types
of
soils
due
to
their
high
lipophility,
but
might
be
transported
into
water
bodies
due
to
storm
runoffs
or
soil
erosion.
CDD/
CDFs
sorbed
to
soil
exhibit
little
potential
for
significant
leaching
or
volatilization
once
sorbed
to
particulate
matter.

Upon
deposition
of
CDD/
CDFs
onto
soil
or
plant
surfaces,
there
can
be
an
initial
loss
due
to
Page
12
of
49
photodegradation
and/
or
volatilization.
The
extent
of
this
initial
loss
is
difficult
to
predict
and
is
controlled
by
climatic
factors,
soil
characteristics,
and
the
concentration
and
physical
form
of
the
deposited
CDD/
CDFs
(
i.
e.,
particulate­
bound,
dissolved
in
solvent,
etc.)
(
Freeman
and
Schroy,
1989;

Paustenbach
et
al.,
1992;
Nicholson
et
al.,
1993).
For
example,
observations
from
the
Seveso
incident
indicated
that
the
levels
of
2,3,7,8­
TCDD
aerially
deposited
on
the
soil
surface
decreased
over
a
period
of
time.

Although
few
studies
have
quantitatively
evaluated
the
transport
of
soil­
bound
CDD/
CDFs,

the
very
low
water
solubilities,
high
K
oc
s,
and
persistent
nature
of
these
chemicals
indicate
that
erosion
of
soil
to
water
bodies
may
be
the
dominant
surface
transport
mechanism
for
CDD/
CDFs
sorbed
to
soil
in
settings
where
erosion
is
possible
(
Paustenbach
et
al.,
1992;
Nicholson
et
al.,
1993).

CDD/
CDFs
below
the
soil
surface
(
i.
e.,
below
the
top
few
millimeters)
are
strongly
adsorbed
and
show
little
upward
or
downward
vertical
migration,
particularly
in
soils
with
a
high
organic
carbon
content
(
Yanders
et
al.,
1989).
Freeman
et
al.
(
1987)
found
no
statistically
meaningful
changes
in
the
concentration
profile
of
2,3,7,8­
TCDD
in
the
top
1
cm
of
Time
Beach
Soil
over
a
16­
month
period,
with
the
exception
of
the
top
3
mm
of
soil
exposed
to
water
and
sunlight
in
which
50
percent
reduction
in
2,3,7,8­
TCDD
concentration
was
observed.
In
addition,
highly
chlorinated
congeners
do
not
show
any
significant
degree
of
degradation
below
the
soil
surface,
substantially
in
the
first
6
months
(
diDomenico
et
al.,
1982),
but
that
rate
of
disappearance
then
slowed
by
over
two
orders
of
magnitude
(
diDomenico
et
al.,
1990).
Nash
and
Beall
(
1980)
reported
that
12
percent
of
the
2,3,7,8­

TCDD
applied
to
bluegrass
turf
as
a
component
(
7.5
ppm
concentration)
of
an
emulsifiable
Silvex
concentrate
volatilized
over
a
period
of
9
months.
Schwarz
and
McLachlan
(
1993)
observed
no
significant
changes
in
CDD/
CDF
concentrations
in
sewage
sludge
amended
soil
that
was
exposed
to
natural
sunlight
for
6
weeks
in
the
late
summer/
early
fall
in
Germany.
Similarly,
Cousins
et
al.
(
1996)

detected
no
volatilization
from
sludge­
amended
soils
through
which
air
was
pumped
for
30
days.
It
appears
that
the
volatilization
process
is
not
common
as
far
as
soils
are
concerned.

For
several
years
it
was
believed
that
near­
surface
(
i.
e.,
the
top
1cm)
CDD/
CDFs
could
volatilize
slowly
to
the
surface
(
Freeman
and
Schroy,
1985),
but
recent
research
has
indicated
that
CDD/
CDFs,
particularly
the
tetra
and
higher
chlorinated
congeners,
show
little
or
no
movement
upward
or
downward
in
the
subsurface
unless
surfactants
or
a
carrier
such
as
waste
oil
or
diesel
fuel
is
present
to
act
as
a
solvent
(
Kapila
et
al.,
1989;
Puri
et
al.,
1989;
Puri
et
al.,
1990;
Yanders
et
al.,

1989;
Schramm
et
al.,
1995).
For
example,
Palausky
et
al.
(
1986)
injected
2,3,7,8­
TCDD
dissolved
in
various
organic
solvents
into
soil
columns
to
determine
the
extent
of
vapor
phase
diffusion;
little
movement
due
to
volatilization
was
observed
unless
the
soil
was
incubated
at
40

C.
However,

laboratory
studies
have
shown
that
2,3,7,8­
TCDD
moves
readily
through
soil
with
waste
oil
Page
13
of
49
components
and
that
mobility
can
also
be
enhanced
by
the
presence
of
surfactants
such
as
sodium
lauryl
sulfate
(
Yanders
et
al.,
1989;
Puri
et
al.,
1989;
Schramm
et
al.,
1995).

Table
3.
Rain
Scavenging
Ratios
(
W)
and
Percent
Washout
Due
to
Particulates
(%
P)
for
CDDs
and
CDFs
in
Bloomington
and
Indianapolis
Ambient
Air
Congener
Group
Bloomington,
IN
Indianapolis,
IN
W
%
P
W
%
P
TCDD
PeCDD
HxCDD
HpCDD
OCDD
TCDF
PeCDF
HxCDF
HpCDF
OCDF
Total
CDD/
CDF
a
10,000
10,000
62,000
90,000
22,000
14,000
11,000
34,000
21,000
­­­
a
50
88
93
80
21
54
77
88
52
68
a
30,000
26,000
91,000
150,000
33,000
18,000
15,000
32,000
41,000
­­­
a
67
69
78
60
24
35
74
79
87
64
a
Rarely
detected;
no
calculations
performed.

Sources:
Hites
and
Harless
(
1991);
Koester
and
Hites
(
1992a).

Paustenbach
et
al.
(
1992)
reviewed
many
major
published
studies
on
dioxin
persistence
in
soil
and
concluded
that
2,3,7,8­
TCDD
probably
has
a
half­
life
of
25
to
100
years
in
subsurface
soil
and
9
to
15
years
at
the
soil
surface
(
i.
e.,
the
top
0.1
cm).
Several
major
studies
reviewed
by
Paustenbach
et
al.
(
1992)
and
additional
recent
studies
are
summarized
below.
Some
of
these
recent
studies
have
concluded
that
the
binding
of
dioxin­
like
compounds
to
soil
approaches
irreversibility
over
time
due
to
the
encapsulation
of
the
compounds
in
soil
organic
and
mineral
matter
(
Puri
et
al.,
1989;
Puri
et
al.,
1992;
Adriaens
and
Grbic­
Galic,
1992).

McLachlan
et
al.
(
1996)
presented
data
on
CDD/
CDF
persistence
in
a
sludge­
amended
soil
sampled
from
a
long­
term
field
experiment
started
in
1968.
Over
50
percent
of
the
CDD/
CDFs
present
in
the
soil
in
1972
were
still
present
in
1990.
The
concentrations
of
all
congeners
were
observed
to
decrease
gradually
and
in
the
same
manner
over
this
time,
indicating
that
either
physical
loss
of
material
from
the
experimental
plot
had
occurred
or
all
congeners
had
undergone
a
uniform
Page
14
of
49
reduction
in
extractability
over
time.
Half­
lives
for
the
disappearance
of
CDD/
CDFs
from
the
sludgeamended
soil
after
1972
were
on
the
order
of
20
years.
These
half­
lives
were
believed
by
McLachlan
et
al.
(
1996)
to
principally
reflect
physical
removal
rather
than
degradation.

Young
(
1983)
conducted
field
studies
on
the
persistence
and
movement
of
2,3,7,8­
TCDD
during
1973­
1979
on
a
military
test
area
that
had
been
aerially
sprayed
with
73,000
kg
of
2,4,5­
T
during
the
period
1962­
1970.
TCDD
levels
of
10
to
1,500
ng/
kg
could
be
found
in
the
top
15
cm
of
soil
14
years
after
the
last
application
of
herbicide
at
the
site.
Although
actual
data
were
not
available
on
the
amount
of
2,3,7,8­
TCDD
originally
applied
as
a
contaminant
of
the
2,4,5­
T,
best
estimates
indicated
that
less
than
one
percent
of
the
applied
2,3,7,8­
TCDD
remained
in
the
soil
after
14
years.
Photodegradation
at
the
time
of
and
immediately
after
aerial
application
was
believed
by
Young
(
1983)
to
be
responsible
for
most
of
the
disappearance.
However,
once
incorporated
into
the
soil,
the
data
indicated
a
half­
life
of
10
to
12
years.

Orazio
et
al.
(
1992)
studied
the
persistence
of
di­
to
octa­
chlorinated
CDDs
and
CDFs
in
sandy
loam
soil
held
in
laboratory
columns
under
water­
saturated
soil
conditions
for
a
period
of
15
months.
Measurable
upward
movement
was
reported
only
for
the
dichlorofurans
and
dichlorodioxins.

Downward
movement
was
only
noticeable
for
the
dichloro­
and
trichloro­
congeners.
The
mobility
of
the
CDDs
and
CDFs
was
not
significantly
affected
by
co­
contaminants
(
i.
e.,
pentachlorophenol
and
creosote
components)
present
at
concentrations
as
high
as
6,000
mg/
kg.
As
much
as
35
percent
loss
of
the
di­
and
trichloro­
congeners
due
to
degradation
was
observed;
no
significant
degradation
of
the
tetra­
through
octa­
chlorinated
congeners
was
reported
(
Orazio
et
al.,
1992).

Hagenmaier
et
al.
(
1992)
collected
soil
samples
around
two
industrial
plants
in
Germany
in
1981,
1987,
and
1989
at
the
same
site
and
from
the
same
depth,
using
the
same
sampling
method.

There
was
no
indication
(
within
the
limits
of
analytical
accuracy
(
±
20
percent))
of
appreciable
loss
of
CDDs
and
CDFs
by
vertical
migration,
volatilization,
or
degradation
over
the
8­
year
period.
Also,

there
were
no
significant
changes
in
the
congener
distribution
pattern
(
i.
e.,
tetra­
through
octa­)
over
this
time
period.

Yanders
et
al.
(
1989)
reported
that
12
years
after
oil
containing
2,3,7,8­
TCDD
was
sprayed
on
unpaved
roads
at
Times
Beach,
Missouri,
no
dioxin
was
discovered
deeper
than
20
cm.
However,

these
roads
were
paved
about
1
year
after
the
spraying
episode,
thus
preventing
volatilization
to
the
atmosphere.
Yanders
et
al.
(
1989)
excavated
this
soil
and
placed
the
soil
in
bins
located
outdoors,

subject
to
the
natural
conditions
of
sunlight
and
precipitation.
They
reported
no
appreciable
loss
nor
vertical
movement
of
2,3,7,8­
TCDD
from
the
soil,
even
in
the
uppermost
sections,
during
a
4­
year
study
period.
Puri
et
al.
(
1992)
reported
no
migration
or
loss
of
1,2,3,4­
TCDD,
1,2,3,7,8­
PeCDD,
Page
15
of
49
OCDD,
and
OCDF
from
samples
of
this
soil
which
were
examined
for
2
years
in
controlled
laboratory
column
experiments.

Hallett
and
Kornelson
(
1992)
reported
finding
2,3,7,8­
TCDD
at
levels
as
high
as
20
pg/
g
in
the
upper
2
inches
of
soil
obtained
from
areas
of
cleared
forest
in
New
Brunswick,
Canada,
where
the
pesticides
2,4­
D
and
2,4,5­
T
had
been
applied
in
one
or
more
applications
24
to
33
years
earlier.

Pereira
et
al.
(
1985)
reported
contamination
by
CDDs
of
the
sand
and
gravel
aquifer
underlying
unlined
surface
impoundments
at
a
wood­
treatment
facility
that
had
utilized
creosote
and
pentachlorophenol.
CDDs
migrated
both
vertically
and
horizontally
in
the
subsurface.
Puri
et
al.

(
1992),
using
soil
column
experiments
in
the
laboratory,
demonstrated
that
pentachlorophenol
and
naphthalene
and
methylnaphthalene
(
components
of
creosote)
readily
transported
CDD/
CDFs
through
soil.
Puri
et
al.
(
1989)
and
Kapila
et
al.
(
1989)
demonstrated
that
application
of
waste
oil
and
anionic
surfactant
solutions
to
field
and
laboratory
columns
of
Times
Beach
soil
can
move
2,3,7,8­
TCDD
through
soil.
Walters
and
Guiseppe­
Elie
(
1988)
showed
that
methanol/
water
solutions
(
1g/
L
or
higher)
substantially
increase
the
mobility
of
2,3,7,8­
TCDD
in
soils.

Migration
of
dioxins
in
cases
where
soil
erosion
is
possible,
some
quantitative
work
has
been
reported.
As
dioxins
are
highly
lipophilic,
soils
with
high
K
OC
can
retain
these
for
a
long
time
period
of
time
and
the
half
lives
of
dioxins
in
various
soils
are
long.
These
long
half
lives
are
not
necessarily
due
to
breakdown
of
dioxins
but
also
include
physical
removal
of
the
dioxins.
Because
of
high
presistence
in
soils
one
estimate
is
that
2,3,7,8­
TCDD
has
a
probable
half
life
of
25­
100
years
in
subsurface
soils
and
9­
15
years
on
the
soil
surface
(
top
0.10
cm).

Transport
Pathways
of
CDDs/
CDFs
in
Aqueous
Systems
CDD/
CDFs
have
been
shown
to
enter
aquatic
systems
directly
from
industrial
and
POTW
effluent
discharges,
from
deposition
of
CDD/
CDFs
in
the
atmosphere
directly
onto
water
bodies
(
of
importance
for
the
Great
Lakes),
and
in
erosion/
stormwater
runoff
from
areas
where
dioxincontaining
material
is
present
(
e.
g.,
a
contaminated
industrial
or
waste
disposal
site).
Thus,
for
any
given
water
body,
the
dominant
transport
mechanism
will
depend
on
site­
specific
conditions.
For
example,
Pearson
and
Swackhammer
(
1997)
report
that
atmospheric
deposition
is
the
dominant
source
of
CDD/
CDFs
to
Lake
Superior,
but
not
to
Lake
Michigan
or
Lake
Ontario.
However,
for
most
freshwater
bodies
today,
erosion/
stormwater
runoff
is
the
probable
dominant
mechanism
for
CDD/
CDF
input
and
the
CDD/
CDFs
present
in
that
runoff
can
be
attributed
to
atmospheric
deposition.
Several
studies
support
this
hypothesis.
Page
16
of
49
The
dominant
transport
mechanism
for
removal
of
CDD/
CDFs
from
the
water
column
is
believed
to
be
sedimentation
and,
ultimately,
burial
in
sediments.
Sediment
resuspension
and
remobilization
of
CDD/
CDFs
will
vary
on
a
site­
by­
site
basis
depending
on
the
nature
and
extent
of
physical
processes
(
e.
g.,
winds/
waves/
currents)
and
biological
processes
(
disturbance
by
benthic
organisms)
(
Fletcher
and
McKay,
1992).

Even
though
CDD/
CDFs
have
very
low
vapor
pressures,
they
can
volatilize
from
water.

However,
volatilization
is
not
expected
to
be
a
significant
loss
mechanism
for
the
tetra­
and
higher
chlorinated
CDD/
CDFs
from
the
water
column
under
most
non­
spill
scenarios.
Podoll
et
al.
(
1986)

calculated
volatilization
half­
lives
of
15
days
and
32
days
for
2,3,7,8­
TCDD
in
rivers
and
ponds/
lakes,

respectively.
Broman
et
al.
(
1992)
used
measured
concentrations
of
CDD/
CDFs
in
ambient
air
(
gaseous
phase)
and
in
Baltic
Sea
water
(
truly
dissolved
concentrations)
to
calculate
the
fugacity
gradient
over
the
air­
water
interface.
The
fugacity
ratios
obtained
indicated
a
net
transport
from
air
to
water
(
ratios
between
0.4
and
0.004).

Aquatic
organisms
can
bioaccumulate
significant
levels
of
CDD/
CDFs.
Although
the
mass
of
CDD/
CDFs
in
the
biota
in
a
given
water
body
will
account
for
only
a
small
fraction
of
the
total
mass
of
CDD/
CDFs
in
that
water
body
(
Mackay
et
al.,
1992a),
these
bioaccumulated
CDD/
CDFs
have
entered
the
aquatic
food
chain
and
can
lead
to
potentially
significant
human
and
wildlife
exposures
and
cause
sensitive
fish
species
to
be
at
increased
risk
(
U.
S.
EPA,
1993).

E.
Sorption
to
Particulates
and
Sedimentation
Most
CDD/
CDFs
entering
the
aquatic
environment
are
associated
with
particulate
matter
(
i.
e.,

dry
and
wet
deposition
of
atmospheric
particles,
eroded
soil/
stormwater
runoff
solids,
and
solids
in
municipal
and
industrial
discharges)
and
are
likely
to
remain
sorbed
to
the
particulate
matter
once
in
the
aquatic
environment.
Recent
studies
have
demonstrated
that
dissolved
CDD/
CDFs
entering
the
aquatic
environment
will,
like
other
lipophilic,
low
water
solubility
organic
compounds,
partition
to
suspended
solids
or
dissolved
organic
matter
such
as
humic
substances.

Muir
et
al.
(
1992)
and
Servos
et
al.
(
1992)
recently
reported
that
48
hours
after
the
addition
of
2,3,7,8­
TCDF,
1,3,6,8­
TCDD,
and
OCDD
in
a
sediment
slurry
to
natural
lake
water/
sediment
limnocorrals,
between
70
and
90
percent
had
partitioned
to
suspended
particulates.
The
proportion
freely
dissolved
in
water
ranged
from
<
2
percent
for
2,3,7,8­
TCDF
and
OCDD
to
10
to
15
percent
for
1,3,6,8­
TCDD.
The
remainder
was
associated
with
dissolved
organic
substances.
Page
17
of
49
Broman
et
al.
(
1992)
analyzed
water
collected
from
nine
sampling
points
in
the
Baltic
Sea
selected
to
be
representative
of
background
levels.
The
concentration
of
particle­
associated
(>
0.45mm)
total
CDD/
CDFs
varied
between
0.170
and
0.390
pg/
L
with
an
average
concentration
of
0.230
pg/
L
(
or
66
percent
of
total
CDD/
CDFs).
The
total
CDD/
CDF
concentration
of
the
"
apparently"
dissolved
fraction
varied
between
0.036
and
0.260
pg/
L
with
an
average
concentration
of
0.120
pg/
L
(
or
34
percent
of
the
total).
Subsequent
calculations
estimated
that,
on
average,
only
0.070
pg/
L
of
the
"
apparently"
dissolved
CDD/
CDFs
were
truly
dissolved.

Servos
et
al.
(
1992)
reported
that
1,3,6,8­
TCDD
and
OCDD
added
as
a
sediment
slurry
to
lake
Limoncorrals
rapidly
partitioned/
settled
to
surfacial
sediments
where
they
persisted
over
the
2
years
of
the
study.
The
half­
lives
of
1,3,6,8­
TCDD
and
OCDD
in
the
water
column
were
reported
as
2.6
and
4.0
days,
respectively.
Based
on
sediment
trap
and
mixed
surface
layer
studies
of
the
Baltic
Sea,
Broman
et
al.
(
1992)
report
that
the
mass
of
CDD/
CDFs
in
the
mixed
surface
layer
at
any
moment
represents
about
1
percent
of
the
total
flux
of
CDD/
CDFs
to
the
sediment
annually;
this
implies
little
recirculation
of
these
compounds
within
the
water
column
of
the
Baltic
Sea.
Broman
et
al.
(
1992)
also
reported
that
the
concentration
of
CDD/
CDFs
in
settling
solids
(
i.
e.,
sediment
trap
collected
material)
is
approximately
one
order
of
magnitude
greater
than
the
concentration
in
suspended
particulates.
They
attributed
this
elevated
concentration
to
the
capacity
of
settling
solids
to
scavenge
the
dissolved
fraction
as
the
solids
settle
through
the
water
column.

Most
of
the
literature
research
shows
that
the
although
atmospheric
deposition
is
a
major
source
of
dioxins
in
water
bodies,
soil
erosion
and
storm
water
runoffs
are
the
important
routes
for
the
presence
of
dioxins
in
water
systems.
In
these
systems
the
dioxins
have
a
tendency
to
partition
to
suspended
particles
and
ultimately
go
through
sedimentation.

F.
Bioaccumulation
Fish
and
invertebrates
can
strongly
bioaccumulate
2,3,7,8­
substituted
CDD/
CDFs,
although
the
benthic
and
pelagic
pathways
by
which
the
accumulation
occurs
are
not
well
understood.

Organisms
have
been
shown
to
accumulate
CDD/
CDFs
when
exposed
to
contaminated
sediments
and
also
to
bioconcentrate
CDD/
CDFs
which
have
dissolved
in
water.
However,
because
most
CDD/
CDFs
in
the
water
column
and
sediment
are
associated
with
particulate
matter
and
dissolved
organic
matter,
the
accumulation
observed
in
the
environment
may
be
primarily
food
chain­
based
starting
with
uptake
by
benthic
organisms
(
e.
g.,
mussels,
chironomids)
directly
from
sediment
pore
waters
and/
or
by
ingestion
or
filtering
of
contaminated
particles.
Those
organisms
consuming
benthic
organisms
(
e.
g.,
crayfish,
suckers)
would
then
pass
the
contaminants
up
the
food
chain
(
Muir
et
al.,

1992;
Fletcher
and
McKay,
1992;
U.
S.
EPA,
1993).
A
thorough
review
of
the
concepts
and
available
Page
18
of
49
information
on
the
bioaccumulation
of
2,3,7,8­
substituted
CDD/
CDFs
is
presented
in
U.
S.
EPA
(
1993)
and
U.
S.
EPA
(
1995)
documents.

The
U.
S.
EPA
(
1993)
document
presents
a
compilation
of
measured
steady­
state
BCFs
for
2,3,7,8­
TCDD.
The
log
BCFs
vary
by
more
than
an
order
of
magnitude
(
4.91
to
6.63
on
a
lipid
content
basis;
3.97
to
5.20
on
a
whole
body
basis).
This
variability
is
likely
due
to
incomplete
characterization
of
exposure
concentrations
or
experimental
shortcomings,
including
partitioning
onto
organic
matter
in
test
systems,
oversaturation
of
the
chemical
in
water,
and
time­
varying
concentrations
in
static
systems
(
U.
S.
EPA,
1993).
Table
4
presents
log
BCF
values
reported
by
various
researchers
for
CDD/
CDFs.

For
aquatic
organisms,
bioaccumulation
refers
to
the
net
accumulation
of
a
chemical
from
exposure
via
food
and
sediments
as
well
as
water.
A
bioaccumulation
factor
(
BAF)
is
the
ratio
(
in
L/
kg)
of
a
chemical's
concentration
in
the
tissue
of
an
aquatic
organism
to
its
concentration
in
the
ambient
water,
in
situations
where
both
the
organism
and
its
food
are
exposed
and
the
ratio
does
not
change
substantially
over
time
(
U.
S.
EPA,
1993;
1995).
The
difference
between
BAFs
and
BCFs
is
primarily
in
the
routes
of
exposure
involved
and
the
levels
of
accumulation
attained
because
of
these
exposure
routes.
BCFs
are
measured
in
laboratory
experiments
designed
to
measure
the
chemical
uptake
by
the
organism
only
from
water.
BAFs
are
usually
determined
from
measurements
of
chemical
concentration
in
water
and
organism
tissue
samples
that
are
obtained
in
the
field
from
aquatic
systems
at
presumed
steady­
state
exposure
conditions.
Thus,
BAFs
include
both
direct
uptake
from
the
water
as
well
as
uptake
from
intake
of
food
and
sediments.

Because
reliable
measurements
of
trace
levels
of
CDD/
CDFs
in
ambient
water
are
generally
not
available,
it
is
not
practical
to
develop
measured
BAFs
for
these
compounds.
However,

detectable
concentrations
of
CDD/
CDFs
are
generally
measurable
in
sediments.
The
relationship
between
chemical
concentrations
in
organisms
and
sediments
is
defined
as
the
biota­
sediment
accumulation
factor
(
BSAF).
BSAFs
can
be
used
to
measure
and
predict
bioaccumulation
directly
from
measured
concentrations
of
chemicals
in
surface
sediments
and
biota.
They
may
also
be
used
to
estimate
BAFs.
Because
BSAFs
are
based
on
field
measurements
and
incorporate
uptake
from
water
and
food,
and
the
effects
of
metabolism
and
growth,
BAFs
estimated
from
BSAFs
will
incorporate
the
net
affect
of
all
these
factors
(
U.
S.
EPA,
1993;
1995).
The
ratios
of
BSAFs
of
CDD/
CDFs
to
a
BSAF
for
2,3,7,8­
TCDD
yield
bioaccumulation
equivalency
factors
(
BEFs)
which
can
be
used
to
estimate
the
combined
toxic
potential
of
CDD/
CDFs
as
a
toxic
equivalence
concentration.
Table
5
presents
BSAFs
and
BEFs
derived
for
CDD/
CDFs
from
Lake
Ontario
lake
trout
and
sediment.
Page
19
of
49
Table
4.
Log
BCF
Values
for
CDD/
CDFs
in
Fish
Congener
Measured
Log
BCFs
Various
Species
(
Reference
A)
Measured
Log
BCFs
Guppy
(
Reference
B)
Calculated
Log
BCFs
Guppy
(
Reference
C)

2,3,7,8­
TCDD
3.73­
5.90
5.24
5.48
1,2,3,7,8­
PeCDD
5.27
5.34
1,2,3,4,7,8­
HxCDD
3.23­
4.00
5.01
5.07
1,2,3,6,7,8­
HxCDD
4.94
5.08
1,2,3,7,8,9­
HxCDD
4.93
5.18
1,2,3,4,6,7,8­
HpCDD
2.71­
3.32
4.68
4.79
OCDD
1.90­
3.97
4.13
4.39
2,3,7,8­
TCDF
3.39­
4.82
4.93
1,2,3,7,8­
PeCDF
4.84
2,3,4,7,8­
PeCDF
3.70
5.14
4.79
1,2,3,4,7,8­
HxCDF
4.57
1,2,3,6,7,8­
HxCDF
4.95
4.58
1,2,3,7,8,9­
HxCDF
4.71
2,3,4,6,7,8­
HxCDF
4.59
1,2,3,4,6,7,8­
HpCDF
4.46
4.26
1,2,3,4,7,8,9­
HpCDF
4.32
OCDF
2.77
3.90
3.88
Reference
A:
Mackay
et
al.
(
1992a);
wet
weight
BCFs.
Reference
B:
Govers
and
Krop
(
1996);
lipid­
adjusted
BCFs.
Reference
C:
Govers
and
Krop
(
1996);
values
calculated
with
the
Solubility
Parameters
for
Fate
Analysis
model.
Page
20
of
49
Table
5.
CDD/
CDF
BSAFs
and
BEFs
for
Lake
Ontario
Lake
Trout
Congener
Estimated
Log
Kow
a
BSAF
BEF
2,3,7,8­
TCDD
7.02
0.059
1.0
1,2,3,7,8­
PeCDD
7.50
0.054
0.92
1,2,3,4,7,8­
HxCDD
7.80
0.018
0.31
1,2,3,6,7,8­
HxCDD
7.80
0.0073
0.12
1,2,3,7,8,9­
HxCDD
7.80
0.0081
0.14
1,2,3,4,6,7,8­
HpCDD
8.20
0.0031
0.051
OCDD
8.60
0.00074
0.012
2,3,7,8­
TCDF
6.5b
0.047
0.80
1,2,3,7,8­
PeCDF
7.0b
0.013
0.22
2,3,4,7,8­
PeCDF
7.0b
0.095
1.6
1,2,3,4,7,8­
HxCDF
7.5b
0.0045
0.076
1,2,3,6,7,8­
HxCDF
7.5b
0.011
0.19
2,3,4,6,7,8­
HxCDF
7.5b
0.040
0.67
1,2,3,7,8,9­
HxCDF
7.5b
0.037
0.63
1,2,3,4,6,7,8­
HpCDF
8.0b
0.00065
0.011
1,2,3,4,7,8,9­
HpCDF
8.0b
0.023
0.39
OCDF
8.80
0.001
0.016
Source:
U.
S.
EPA
(
1995).

a
Burkhard
and
Kuehl
(
1986).
b
Estimated
based
on
degree
of
chlorination
and
Burkhard
and
Kuehl
(
1986).
Page
21
of
49
In
most
of
the
literature
surveyed,
rates
of
accumulation
as
well
as
depuration
rates
either
were
not
determined
or
were
not
discussed.
Although
these
dioxins
are
highly
bioaccumulative,
the
data
on
their
foodchain
transfer
is
scant.

G.
Degradation
of
CDDs
and
CDFs
in
the
Environmental
Compartments
The
available
evidence
indicates
that
CDDs
and
CDFs,
particularly
the
tetra­
and
higher
chlorinated
congeners,
are
extremely
stable
compounds
under
most
environmental
conditions.
The
only
environmentally
significant
transformation
processes
for
these
congeners
are
believed
to
be
photolysis
and
atmospheric
photo
oxidation
of
nonsorbed
species
in
the
gaseous
phase
or
at
the
soil
or
water­
air
interface.
Several
studies
have,
however,
indicated
that
certain
ligninolytic
fungi
can
degrade
higher­
chlorinated
congeners
and
that
anaerobic
degradation
in
sediment
may
occur
at
a
slow
rate.
To
a
large
extent,
these
biodegradation
processes
involve
dechlorination
to
less­
chlorinated
(
and
possibly
more
toxic)
congeners.

Abiotic
Degradation:

a.
Hydrolysis
There
is
no
available
evidence
indicating
that
hydrolysis
would
be
an
operative
environmental
process
for
degradation
of
CDDs
or
CDFs.
The
attachment
of
chlorines
directly
to
the
aromatic
ring
in
CDDs
and
CDFs
confers
hydrolytic
stability.
Specifically,
S
N
1
and
S
N
2
reactions
do
not
take
place
readily
at
sp2
hybridized
carbons
(
Leifer
et
al.,
1983;
Miller
and
Zepp,
1987).

b.
Photolysis
Photolysis
appears
to
be
one
of
the
most
environmentally
significant
degradation
mechanisms
for
CDD/
CDFs
in
water
and
soil,
and
possibly
in
the
atmosphere.
CDD/
CDFs
absorb
electromagnetic
radiation
at
wavelengths
greater
than
290
nm
(
i.
e.,
the
lower
bound
of
the
Sun's
radiation
reaching
the
Earth's
surface)
and,
therefore,
can
be
expected
to
be
subject
to
photolysis
by
sunlight.
Numerous
studies
have
demonstrated
that
CDD/
CDFs
undergo
relatively
rapid
photolysis
in
a
variety
of
organic
solvents
and
at
much
slower
rates
in
water,
typically
following
first
order
kinetics.
Because
of
the
difficulties
inherent
in
controlling
experimental
variables
for
nonvolatile
and
highly
lipophilic
compounds
like
CDD/
CDFs,
few
photolysis
studies
have
been
performed
with
natural
waters,
with
CDD/
CDFs
sorbed
to
soil
or
particulate
matrices,
and
with
gas
phase
CDD/
CDFs.
The
photochemistry
of
CDD/
CDFs
has
been
reviewed
by
Miller
and
Zepp
(
1987),
Choudry
and
Webster
(
1987),
Atkinson
(
1991;
1996),
and
others.
Page
22
of
49
The
major
products
from
photolysis
are
complex
and,
in
most
studies,
a
reasonable
mass
balance
has
not
been
obtained.
Although
the
photolytic
pathway(
s)
for
CDD/
CDFs
has
not
been
fully
identified,
a
major
photolysis
pathway
appears
to
be
photodechlorination
resulting
in
formation
of
lower
chlorinated
CDD/
CDFs.
Several
researchers
have
reported
that
carbon­
oxygen
cleavage
and
other
mechanisms
may
be
similarly
or
more
important
pathways
for
CDD/
CDFs
containing
four
or
fewer
chlorines
(
Choudhry
and
Webster,
1989;
Dulin
et
al.,
1986;
Miller
et
al.,
1989;
Kieatiwong
et
al.,
1990).
Studies
performed
to
date
suggest
that
the
photolytic
degradation
products
from
irradiation
of
CDD/
CDFs
in
organic
solvents
may
be
significantly
different
than
those
observed
when
surface­
sorbed
and
gas­
phase
CDD/
CDFs
are
irradiated.
A
preferential
loss
of
chlorines
from
the
lateral
positions
(
i.
e.,
chlorines
at
the
2,
3,
7,
and
8
positions)
rather
than
from
the
peri
positions
(
i.
e.,

chlorines
at
the
1,
4,
6,
and
9
positions)
has
been
reported
for
solution
studies
(
Crosby
et
al.,
1973;

Dobbs
and
Grant,
1979;
Tysklind
and
Rappe,
1991);
the
opposite
trend
is
observed
for
some
congener
groups
when
irradiated
as
dry
films,
sorbed
to
soil,
and
as
gas­
phase
CDD/
CDFs
(
Choudry
and
Webster,
1989;
Kieatiwong
et
al.,
1990;
Sivils
et
al.,
1994;
Sivils
et
al.,
1995).

Most
studies
performed
to
date
have
used
artificial
light,
pure
congeners,
and
reaction
media
consisting
of
homogenous
solutions
in
aqueous­
organic
solvent
mixtures,
silica
gel,
or
clean
solid
surfaces.
Thus,
although
photolysis
of
CDD/
CDFs
at
environmentally
significant
rates
has
been
observed
in
laboratory
studies,
the
results
of
these
studies
may
not
be
representative
of
photolysis
rates
that
occur
under
actual
environmental
conditions.
The
following
subsections
summarize
the
key
findings
of
recent
environmentally
significant
studies
for
the
water,
soil,
and
air
media.

c.
Photolysis
in
Solution.

Numerous
studies
have
demonstrated
that
CDD/
CDFs
will
undergo
photodechlorination
following
first
order
kinetics
in
solution
with
preferential
loss
of
chlorine
from
the
lateral
positions.

Photolysis
is
slow
in
pure
water
but
increases
dramatically
when
solvents
serving
as
hydrogen
donors
are
present
such
as
hexane,
benzene,
methanol,
acetonitrile,
hexadecane,
ethyl
oleate,
dioxane,
and
isooctane
(
Buser,
1976;
Buser,
1988;
Choudry
and
Webster,
1987;
Choudry
et
al.,
1990;
Crosby
et
al.,
1971;
Crosby,
1978;
Crosby,
1981;
Dobbs
and
Grant,
1979;
Dougherty
et
al.,
1991;
Dulin
et
al.,

1986;
Friesen
et
al.,
1990;
Hutzinger,
1973;
Koester
and
Hites,
1992b;
Koshioka
et
al.,
1990;

Wagenaar
et
al.,
1995;
and
others).

The
photolytic
behavior
of
CDD/
CDFs
in
organic
solvents
or
in
pure
water,
however,
is
not
expected
to
accurately
reflect
the
photolytic
behavior
of
these
compounds
in
natural
waters.
Natural
waters
have
differing
quantities
and
types
of
suspended
particulates
and
dissolved
organic
material
that
could
either
retard
or
enhance
the
photolysis
of
CDD/
CDFs.
However,
only
a
few
studies
have
been
performed
that
have
examined
the
photolysis
of
CDD/
CDFs
using
natural
waters
and
sunlight.
Page
23
of
49
Several
other
published
studies
have
used
a
mixture
of
water
and
acetonitrile
to
enhance
the
solubility
of
CDD/
CDFs.
The
following
paragraphs
summarize
the
results
of
these
studies.
Table
6
presents
selected
results
from
these
studies.

Dulin
et
al.
(
1986)
studied
the
photolysis
of
2,3,7,8­
TCDD
in
a
water:
acetonitrile
solution
(
1:
1,
v/
v)
under
sunlight
and
artificial
light
(
i.
e.,
a
mercury
lamp).
The
quantum
yield
for
photodegradation
of
2,3,7,8­
TCDD
in
water
was
three
times
greater
under
artificial
light
at
313
nm
than
under
sunlight
which
suggests
that
the
medium­
pressure
mercury
lamp
used
was
a
more
efficient
light
source
than
the
sun
(
Koester
and
Hites,
1992b).
Podoll
et
al.
(
1986)
used
the
Dulin
et
al.
(
1986)

reaction
rate
data
from
the
artificial
light
experiment
to
calculate
seasonal
half­
life
values
for
2,3,7,8­

TCDD
at
40
degrees
north
latitude
in
clear
near­
surface
water.
The
calculated
seasonal
values
for
half­
lives
ranged
from
0.9
days
in
summer
to
4.9
days
in
winter.
Using
the
Dulin
et
al.
(
1986)
reaction
rate
data
from
the
sunlight
experiment
yields
calculated
seasonal
values
for
half­
lives
ranging
from
2.7
days
in
summer
to
16
days
in
winter.

Choudhry
and
Webster
(
1989)
studied
the
photolytic
behavior
of
a
series
of
CDDs
in
a
water:
acetonitrile
solution
(
2:
3,
v/
v)
using
a
low­
pressure
mercury
lamp.
Choudry
et
al.
(
1990)

investigated
the
photolytic
behavior
of
two
CDFs
(
1,2,4,7,8­
PeCDF
and
1,2,3,4,7,8­
HxCDF)
using
a
similar
experimental
setup
but
with
a
1:
1
water:
acetonitrile
solution.
Assuming
that
the
quantum
yields
observed
in
these
studies
are
the
same
as
would
be
observed
in
natural
waters,
Choudry
and
Webster
(
1989)
and
Choudry
et
al.
(
1990)
estimated
the
mid­
summer
half­
life
values
at
40
degrees
north
latitude
in
clear
near­
surface
water
to
be
as
follows:
1,2,3,7­
TCDD
(
1.8
days);
1,3,6,8­
TCDD
(
0.3
days);
1,2,3,4,7­
PeCDD
(
15
days);
1,2,3,4,7,8­
HxCDD
(
6.3
days);
1,2,3,4,6,7,8­
HpCDD
(
47
days);
OCDD
(
18
days);
1,2,4,7,8­
PeCDF
(
0.2
hours);
and
1,2,3,4,7,8­
HxCDF
(
6
hours).
However,

Choudry
and
Webster
(
1989)
also
experimentally
determined
the
sunlight
photolysis
half­
life
of
1,3,6,8­
TCDD
in
pond
water
to
be
3.5
days
(
i.
e.,
more
than
10
times
greater
than
the
half­
life
predicted
by
laboratory
experiments).
The
authors
attributed
this
significant
difference
in
photolysis
rates
to
the
light
screening/
quenching
effects
of
dissolved
organic
matter.

Friesen
et
al.
(
1990)
examined
the
photolytic
behavior
of
1,2,3,4,7­
PeCDD
and
1,2,3,4,6,7,8­

HpCDD
in
water:
acetonitrile
(
2:
3,
v/
v)
and
in
pond
water
under
sunlight
at
50
degrees
north
latitude.

The
observed
half­
lives
of
these
two
compounds
in
the
water:
acetonitrile
solution
were
12
and
37
days,
respectively,
which
are
very
similar
to
the
results
observed
by
Choudry
and
Webster
(
1989)
for
these
two
congeners
using
artificial
light.
However,
in
contrast
to
the
results
observed
by
Choudhry
and
Webster
(
1989)
for
1,3,6,8­
TCDD,
Friesen
et
al.
(
1990)
found
that
the
half­
lives
of
1,2,3,4,7­

PeCDD
and
1,2,3,4,6,7,8­
HpCDD
were
much
shorter
in
pond
water
(
0.94
and
2.5
days,
respectively)

than
in
the
water:
acetonitrile
solution.
Similarly,
Friesen
et
al.
(
1993)
studied
the
photodegradation
Page
24
of
49
of
2,3,7,8­
TCDF
and
2,3,4,7,8­
PeCDF
by
sunlight
using
water:
acetonitrile
(
2:
3,
v/
v)
and
lake
water.

The
observed
half­
lives
of
the
2,3,7,8­
TCDF
and
2,3,4,7,8­
PeCDF
in
the
water:
acetonitrile
solution
were
6.5
and
46
days,
respectively,
and
1.2
and
0.19
days
in
lake
water,
respectively.
Friesen
et
al.

(
1990)
and
Friesen
et
al
(
1993)
attributed
the
significant
differences
between
the
natural
water
and
water:
acetonitrile
solution
results
to
indirect
or
sensitized
photolysis
due
to
the
presence
of
naturally
occurring
components
in
the
lake
and
pond
water.

Dung
and
O'Keefe
(
1992),
in
their
investigation
of
aqueous
photolysis
of
2,3,7,8­
TCDF
and
1,2,7,8­

TCDF,
reported
findings
similar
to
those
of
Friesen
et
al.
(
1993).
Dung
and
O'Keefe
prepared
aqueous
solutions
by
pumping
water
through
"
generator
columns"
containing
particles
coated
with
a
thin
film
of
the
respective
TCDF
congener.
Solutions
in
high
purity
HPLC
water,
distilled
water,

Hudson
River
water,
and
Saratoga
Lake
water
were
exposed
to
sunlight
during
September
at
approximately
42.5
degrees
north
latitude.
The
photolysis
rates
of
the
two
TCDF
congeners
observed
in
the
river
and
lake
water
(
half­
lives
of
about
4
to
6
hours)
were
double
the
rates
observed
in
pure
water
(
half­
lives
of
about
8
to
11
hours).
Dung
and
O'Keefe
(
1992)
attribute
the
difference
in
rates
to
the
presence
of
natural
organics
in
the
river
and
lake
water
that
may
be
acting
as
sensitizers.

Kim
and
O'Keefe
(
1998)
confirmed
the
results
observed
in
Dung
and
O'Keefe
(
1992)
by
photolyzing
1,2,7,8­
TCDD;
2,3,7,8­
TCDF;
OCDD;
and
OCDF
in
natural
water
from
seven
different
locations.
The
values
reported
in
Table
7
are
the
average
values
for
the
seven
natural
waters.
The
half­
lives
of
TCDD
and
TCDF
were
at
least
two
times
shorter
in
all
the
natural
waters
than
in
pure
water.
However,
the
natural
water
half­
lives
of
OCDD
and
OCDF
showed
greater
variability
with
some
half­
lives
greater
than
pure
water
and
some
less
than
pure
water.

Because
there
was
no
apparent
relationship
between
the
dissolved
organic
content
of
the
water
and
the
observed
half­
lives,
it
was
hypothesized
that
certain
waters
may
contain
organic
molecules
which
either
act
as
photosensitizers
or
as
photodesensitizers.
Page
25
of
49
Table
6.
Photolysis
Rates
of
CDDs/
CDFs
in
Water
and
Water:
Acetonitrile
Mixtures
CONGENER
LIGHT
SOURCE
REACTION
MEDIUM
PHOTOLYSIS
RATE
CONSTANT
(
1/
day)
HALF­
LIFE
(
days)

DURING
SUMMER
REFERENCE
CDDs
1,2,7,8­
TCCDD
sunlight
water
from
7
ponds/
lakes
4.06
0.17
Kim
and
O'Keefe
(
1998)

1,3,6,8­
TCDD
Hg
lamp
pond
water
0.198
3.5
Choudry
and
Webster
(
1989)

2,3,7,8­
TCDD
sunlight
water:
acetonitrile
(
1:
1
v/
v)
0.255
2.7
Podoll
et
al.
(
1986)

2,3,7,8­
TCDD
Hg
lamp
water:
acetonitrile
(
1:
1
v/
v)
0.78
0.9
Podoll
et
al.
(
1986)

1,2,3,4,7,8­
HxCDD
Hg
lamp
water:
acetonitrile
(
2:
3
v/
v)
0.111
6.3
Choudry
and
Webster
(
1989)

1,2,3,4,6,7,8­
HpCDD
Hg
lamp
water:
acetonitrile
(
2:
3
v/
v)
0.0148
47
Choudry
and
Webster
(
1989)

OCDD
Hg
lamp
water:
acetonitrile
(
2:
3
v/
v)
0.0397
18
Choudry
and
Webster
(
1989)

OCDD
sunlight
water
from
7
ponds/
lakes
1.04
0.67
Kim
and
O'Keefe
(
1998)

CDFs
2,3,7,8­
TCDF
sunlight
water
from
7
ponds/
lakes
3.87
0.18
Kim
and
O'Keefe
(
1998)

1,2,7,8­
TCDF
sunlight
HPLC
water
1.96
0.35
Dung
and
O'Keefe
(
1992)

1,2,7,8­
TCDF
sunlight
distilled
water
2.18
0.32
Dung
and
O'Keefe
(
1992)

1,2,7,8­
TCDF
sunlight
Saratoga
Lake
3.53
0.20
Dung
and
O'Keefe
(
1992)

1,2,7,8­
TCDF
sunlight
Hudson
River
3.96
0.18
Dung
and
O'Keefe
(
1992)

1,2,7,8­
TCDF
Hg
lamp
HPLC
water
24.5
0.03
Dung
and
O'Keefe
(
1992)

2,3,7,8­
TCDF
sunlight
water:
acetonitrile
(
1:
2.5
v/
v)
0.106
6.5
Friesen
et
al.
(
1993)

2,3,7,8­
TCDF
sunlight
lake
water
0.58
1.2
Friesen
et
al.
(
1993)

2,3,7,8­
TCDF
sunlight
distilled
water
1.49
0.47
Dung
and
O'Keefe
(
1992)

2,3,7,8­
TCDF
sunlight
HPLC
water
1.56
0.44
Dung
and
O'Keefe
(
1992)

2,3,7,8­
TCDF
sunlight
Saratoga
Lake
2.64
0.26
Dung
and
O'Keefe
(
1992)

2,3,7,8­
TCDF
sunlight
Hudson
River
2.83
0.25
Dung
and
O'Keefe
(
1992)

2,3,7,8­
TCDF
Hg
lamp
HPLC
water
16.8
0.04
Dung
and
O'Keefe
(
1992)

2,3,4,7,8­
PeCDF
sunlight
water:
acetonitrile
(
1:
2.5
v/
v)
0.015
46.2
Friesen
et
al.
(
1993)

2,3,4,7,8­
PeCDF
sunlight
lake
water
3.59
0.19
Friesen
et
al.
(
1993)

OCDF
sunlight
water
from
7
ponds/
lakes
1.19
0.58
Kim
and
O'Keefe
(
1998)
Page
26
of
49
d.
Photolysis
on
Soil.

Photolysis
of
CDD/
CDFs
on
soil
has
been
reported
but
the
factors
affecting
the
rate
and
extent
of
photolysis
have
not
been
well
characterized.
Based
on
the
data
generated
to
date,

photolysis
is
an
operative
degradation
process
only
in
the
near­
surface
soil
where
UV
light
penetrates
(
i.
e.,
the
top
few
millimeters
or
less
of
soil)
and
dechlorination
of
peri­
substituted
chlorines
appears
to
occur
preferentially.
Even
within
this
near
surface
area,
the
rate
of
photolysis
is
substantially
slower
than
the
rate
of
photolysis
that
would
be
observed
in
a
solution
of
same
depth
presumably
as
a
result
of
the
light­
attenuating
effects
of
soils.
Below
this
near­
surface
level,
photolysis
is
not
a
significant
process,
and
CDD/
CDFs
present
in
soil
at
any
greater
depth
are
likely
to
persist
(
Miller
et
al.,
1989;
Puri
et
al.,
1989;
Yanders
et
al.,
1989;
Kieatiwong
et
al.,
1990).
The
substantial
research
performed
on
the
environmental
persistence
of
2,3,7,8­
TCDD
in
the
area
around
the
ICMESA
chemical
plant
in
Seveso,
Italy,
demonstrates
that
photolysis
in
soils
is
a
near­
surface
process.
The
Seveso
area
was
contaminated
when
a
trichlorophenol
reaction
vessel
overheated
in
1976,
blowing
out
the
safety
devices
and
spraying
2,3,7,8­
TCDD­
contaminated
material
into
the
environment.
The
levels
of
dioxin
in
the
soil
decreased
substantially
during
the
first
6
months
following
the
accident
before
reaching
a
relatively
steady
state
of
1/
5
to
1/
11
of
the
initial
levels
(
DiDomenico
et
al.,
1982).

Kieatiwong
et
al.
(
1990)
investigated
the
photolysis
of
2,3,7,8­
TCDD
added
to
two
agricultural
soils
(
350
ug/
kg)
of
approximately
0.3
mm
depth
and
irradiated
for
15
days
with
a
mercury
lamp.
A
loss
of
approximately
15
percent
of
2,3,7,8­
TCDD
was
observed
on
the
soil
of
higher
organic
carbon
and
clay
content.
A
loss
of
approximately
45
percent
of
2,3,7,8­
TCDD
was
observed
on
the
soil
of
lower
organic
carbon
and
clay
content.
There
was
no
significant
loss
on
either
soil
after
the
first
5
days.

Miller
et
al.
(
1989)
studied
the
CDD
degradation
products
resulting
from
irradiation
of
13Clabeled
OCDD
on
two
soil
types
using
sunlamps.
Approximately
38
to
42
percent
of
the
OCDD
were
degraded
by
day
five
of
the
experiment;
no
significant
further
loss
of
OCDD
was
observed
over
the
following
10
days.
Although
determined
not
to
be
the
dominant
photolysis
pathway,

photodechlorination
was
observed
in
both
soils;
approximately
10
to
30
percent
of
the
lower
chlorinated
congeners
were
produced
from
the
immediate
higher
chlorinated
congeners.
The
HpCDD
and
HxCDD
congeners
observed
as
degradation
products
were
present
in
approximately
similar
proportions
to
the
number
of
congeners
in
each
congener
group.
However,
Miller
et
al.
(
1989)
found
that,
unlike
the
results
reported
for
photolysis
of
OCDD
in
solution
by
Choudry
and
Webster
(
1989)

and
Dobbs
and
Grant
(
1979),
2,3,7,8­
TCDD
and
1,2,3,7,8­
PeCDD
were
observed
in
greater
yields
than
would
be
expected
on
the
basis
of
the
number
of
potential
TCDD
and
PeCDD
congeners.
One­
Page
27
of
49
fifth
to
one­
third
of
the
total
yield
of
PeCDDs
was
1,2,3,7,8­
PeCDD,
and
one­
half
of
the
total
yield
of
TCDDs
was
2,3,7,8­
TCDD.

Kieatiwong
et
al.
(
1990)
performed
similar
experiments
to
those
of
Miller
et
al.
(
1989)
using
natural
sunlight
rather
than
sunlamps
for
irradiation
of
13C­
labeled
OCDD
on
soils.

Photodechlorination
was
estimated
to
account
for
approximately
10
percent
of
the
loss
of
OCDD.

One­
third
to
one­
half
of
the
total
yield
of
PeCDDs
was
1,2,3,7,8­
PeCDD,
and
one­
half
of
the
total
yield
of
TCDDs
was
2,3,7,8­
TCDD.
The
findings
of
Miller
et
al.
(
1989)
and
Kieatiwong
et
al.
(
1990)

indicate
that
the
2,3,7,8­
substituted
TCDD
and
PeCDD
congeners
were
either
preferentially
formed
or
were
photochemically
less
reactive
than
the
other
congeners
that
were
formed.

Tysklind
et
al.
(
1992)
also
studied
the
sunlight
photolysis
of
OCDD
on
soil
and
reported
results
in
good
agreement
with
those
of
Miller
et
al.
(
1989)
and
Kieatiwong
et
al.
(
1990).

Photodechlorination
was
observed
with
production
of
HpCDDs,
HxCDDs,
PeCDDs,
and
TCDDs
over
the
16­
day
irradiation
period.
Photodechlorination
at
the
peri­
substituted
positions
was
the
preferred
photodechlorination
mechanism;
the
proportions
of
2,3,7,8­
substituted
congeners
present
in
the
soils
after
16
days
for
each
congener
group
were
as
follows:
HxCDD
­
65
percent;
PeCDD
­

40
percent;
and
TCDD
­
75
percent.

The
sunlight
photolysis
of
OCDF
on
soil
was
also
studied
by
Tysklind
et
al.
(
1992).

Photodechlorination
was
observed.
However,
unlike
the
case
with
OCDD,
photodechlorination
of
the
lateral­
substituted
positions
was
found
to
be
the
dominant
photodechlorination
mechanism
resulting
in
a
relative
decreasing
proportion
of
2,3,7,8­
substituted
congeners
during
the
irradiation
period.
2,3,7,8­
TCDF
was
not
observed
in
any
of
the
irradiated
samples.

Schwarz
and
McLachlan
(
1993)
studied
the
photolysis
of
CDD/
CDFs
in
an
experiment
designed
to
simulate
the
application
of
sewage
sludge
to
an
agricultural
field.
No
significant
changes
in
CDD/
CDF
concentrations
were
observed
during
the
43­
day
exposure
period
to
late
summer/
early
fall
natural
sunlight.
Although
the
OCDD
concentration
in
the
sludge
had
been
increased
by
a
factor
of
1,000
through
spiking,
no
increase
in
HpCDD
concentrations
were
observed.
The
absence
of
any
changes
indicates
that
neither
photodegradation
nor
volatilization
are
important
mechanisms
in
the
fate
of
CDD/
CDF
in
sewage
sludge
following
agricultural
applications.

The
addition
of
solvents
to
soil
can
increase
the
rate
and
extent
of
photolysis.
Botre
et
al.

(
1978)
reported
the
destruction
of
8

g/
mL
of
2,3,7,8­
TCDD
in
0.02
M
hexadecylpyridinium
chloride
(
an
aqueous
surfactant)
applied
to
soil
in
4
hours.
Kieatiwong
et
al.
(
1990)
reported
that
addition
of
hexadecane
to
soil
containing
2,3,7,8­
TCDD
resulted
both
in
an
increased
rate
of
photolysis
and
Page
28
of
49
continued
photochemical
loss
of
2,3,7,8­
TCDD
beyond
the
point
at
which
soil
containing
no
hexadecane
showed
photochemical
loss.

Although
only
minimally
related
to
soil
environmental
conditions,
Buser
(
1988)
studied
the
photolytic
decomposition
rates
of
2,3,7,8­
TCDF,
1,2,3,4­
TCDF,
and
1,2,7,8­
TCDF
dried
as
thin
films
on
quartz
vials.
When
exposed
to
sunlight,
the
substances
slowly
degraded
with
reported
halflives
of
5
days,
4
days,
and
1.5
days,
respectively.
Similarly,
Koester
and
Hites
(
1992b)
studied
the
photodegradation
of
a
series
of
tetra­
through
octa­
chlorinated
CDDs
and
CDFs
on
silica
gel.
In
general,
the
CDFs
degraded
much
more
rapidly
than
the
CDDs,
and
half­
lives
increased
with
increasing
level
of
chlorination
(
1,2,7,8­
TCDF
excluded).
The
half­
lives
for
CDDs
ranged
from
3.7
days
for
1,2,3,4­
TCDD
to
11.2
days
for
OCDD.
The
half­
lives
for
CDFs
ranged
from
0.1
day
for
1,2,3,8,9­
PeCDF
to
0.4
days
for
OCDF.

e.
Photolysis
on
Vegetation.

Photolysis
of
CDD/
CDFs
sorbed
on
the
surface
of
vegetation
has
not
been
well
characterized
and
the
findings
to
date
are
somewhat
contradictory.
McCrady
and
Maggard
(
1993)
reported
that
2,3,7,8­
TCDD
sorbed
on
the
surface
of
reed
canary
grass
(
Phalaris
arundinacea
L.)
undergoes
photolytic
degradation
with
a
half­
life
of
44
hours
in
natural
sunlight.
In
contrast,
Welsch­
Pausch
et
al.
(
1995)
found
little
difference
in
the
CDD/
CDF
congener
patterns
between
grass
(
Lolium
multiflorum)
grown
on
an
outdoor
plot
and
grass
grown
in
a
greenhouse
(
i.
e.,
UV­
light
transmission
blocked).
In
an
attempt
to
clarify
this
contradiction,
Welsch­
Pausch
and
McLachlan
(
1995)
studied
the
photodegradation
of
CDD/
CDFs
on
pasture
grass
(
Arrhenatherion
elatioris)
during
two
growing
cycles
(
summer
and
autumn)
using
two
greenhouses.
One
greenhouse
was
constructed
of
glass
which
blocks
UV
transmission
and
the
other
was
constructed
of
plexiglass
(
4
mm)
with
an
UV­
light
transmission
of
greater
than
50
percent
in
the
280­
320
mm
range.
In
both
the
summer
and
autumn
exposure
periods,
the
concentrations
of
CDD/
CDFs
(
on
a
congener
group
basis)
were
similar
in
the
grass
exposed
to
UV­
light
and
the
grass
that
was
not
exposed.
Welsch­
Pausch
and
McLachlan
(
1995)
concluded
that
if
photodegradation
is
occurring,
it
is
a
relatively
insignificant
factor
in
the
accumulation
of
CDD/
CDF
in
pasture
grass.

f.
Photolysis
in
Air.

Photolysis
of
CDD/
CDFs
in
the
atmosphere
has
not
been
well­
characterized.
Based
on
the
data
generated
to
date,
however,
photolysis
appears
to
be
a
significant
mechanism
for
degradation
(
i.
e.,
principally
dechlorination
of
the
peri­
substituted
chlorines)
of
those
CDD/
CDFs
present
in
the
atmosphere
in
the
gas
phase.
For
airborne
CDD/
CDFs
sorbed
to
particulates,
photolysis
appears
to
Page
29
of
49
proceed
very
slowly
and
may
be
influenced
by
the
organic
content
of
the
particle.
Because
of
the
low
volatility
of
CDD/
CDFs,
few
studies
have
been
attempted
to
measure
actual
rates
of
photodegradation
of
gaseous­
phase
CDD/
CDF,
and
only
recently
have
studies
been
undertaken
to
examine
the
relative
importance
of
photolysis
for
particulate­
bound
CDD/
CDFs.

g.
Gas­
Phase
Photolysis
Podoll
et
al.
(
1986)
estimated
the
photolysis
rate
of
2,3,7,8­
TCDD
vapors
in
the
atmosphere
based
on
the
quantum
yield
for
photolysis
in
hexane.
The
half­
life
in
summer
sunlight
at
40
degrees
north
latitude
was
estimated
to
be
1
hour
as
an
upper
limit.

Mill
et
al.
(
1987)
reported
that
photolysis
of
vapor
phase
2,3,7,8­
TCDD
at
145

C
in
a
flow
system
using
a
pulsed
308
nm
laser
light
gave
a
photolysis
quantum
yield
ranging
from
0.013
to
0.03,

equivalent
to
an
atmospheric
half­
life
of
a
few
hours.
However,
the
true
gas
phase
quantum
yield
at
25

C
is
uncertain,
and,
therefore,
the
atmospheric
lifetime
in
sunlight
is
uncertain.

Orth
et
al.
(
1989)
conducted
photolysis
experiments
with
vapor­
phase
2,3,7,8­
TCDD
under
illumination
with
a
Hg­
Xe
light
source
and
filters
to
achieve
radiation
in
the
UV
region
from
250
nm
to
340
nm.
The
temperature
in
the
photolysis
cell
was
approximately
150

C.
Carrier
gases
included
air
and
helium.
No
significant
difference
in
the
rate
constants
was
observed
in
helium
and
air,
5.4
x
10­
3
sec­
1
and
5.9
x
10­
3
sec­
1,
respectively,
which
correspond
to
half­
lives
of
a
few
minutes.
The
average
quantum
yield
in
air
over
the
absorption
band
was
found
to
be
0.033
+
0.046.
As
with
the
results
of
Mill
et
al.
(
1987),
the
true
gas
phase
quantum
yield
at
25

C
is
uncertain,
and,
therefore,
the
atmospheric
lifetime
in
sunlight
is
uncertain.

Sivils
et
al.
(
1994;
1995)
studied
the
gas
phase
photolysis
of
several
CDDs
(
2,3,7­
TrCDD;

2,3,7,8­
TCDD;
1,2,3,4­
TCDD;
1,2,3,7,8­
PeCDD,
and
1,2,4,7,8­
PeCDD)
by
irradiating
the
effluent
from
a
gas
chromatograph
with
broadband
radiation
in
the
UV/
visible
region
for
periods
of
time
up
to
20
minutes.
The
irradiated
sample
was
then
introduced
into
a
second
gas
chromatograph
to
measure
the
extent
of
dechlorination.
The
results
showed
that
degradation
followed
first
order
kinetics
and
that
there
was
an
inverse
relationship
between
the
degree
of
chlorination
and
the
rate
of
disappearance.
Although
the
lack
of
photoproducts
prevented
an
independent
confirmation
of
the
preferential
loss
mechanism,
the
results
indicated
that
laterally­
substituted
congeners
(
i.
e.,
chlorines
at
the
2,
3,
7,
and
8
positions)
degrade
at
a
slower
rate
than
the
peri­
substituted
congeners
(
i.
e.,

chlorines
at
the
1,
4,
6,
and
9
positions).
Although
the
rate
constants
were
not
presented
in
Sivils
et
al.
(
1994),
the
degradation
rate
for
2,3,7,8­
TCDD
(
30
percent
loss
in
20
minutes)
was
reported
to
Page
30
of
49
be
slower
than
the
rates
for
all
other
tested
CDDs.
Also,
1,2,4,7,8­
PeCDD
(
with
2
peri­
chlorines)

degraded
significantly
faster
than
1,2,3,7,8­
PeCDD
(
with
only
1
peri­
chlorine).

h.
Particulate
Phase
Photolysis
The
photolysis
of
2,3,7,8­
TCDD
sorbed
onto
small
diameter
fly
ash
particulates
suspended
in
air
was
measured
by
Mill
et
al.
(
1987).
The
results
indicated
that
fly
ash
appears
to
confer
photostability
on
2,3,7,8­
TCDD.
Little
(
8
percent)
to
no
loss
was
observed
on
the
two
fly
ash
samples
after
40
hours
of
illumination.

Koester
and
Hites
(
1992b)
studied
the
photodegradation
of
CDD/
CDFs
naturally
adsorbed
to
five
fly
ashes
collected
from
electrostatic
precipitators
(
one
from
a
hospital
incinerator,
two
from
municipal
incinerators,
and
two
from
coal­
fired
power
plants)
using
a
rotary
reactor
and
a
mediumpressure
mercury
lamp.
Although
they
found
that
CDD/
CDFs
underwent
photolysis
in
solution
and
when
spiked
onto
silica
gel,
no
significant
degradation
was
observed
in
11
photodegradation
experiments
performed
on
the
ashes
for
periods
ranging
from
2
to
6
days.
Three
additional
experiments
were
performed
to
determine
what
factors
may
have
been
inhibiting
photolysis.
From
the
results
of
these
additional
experiments,
Koester
and
Hites
(
1992b)
concluded
that:
(
1)
the
absence
of
photodegradation
was
not
due
to
the
absence
of
a
hydrogen­
donor
organic
substance;
(
2)
other
molecules
or
the
ash,
as
determined
by
a
photolysis
experiment
with
an
ash
extract,
inhibit
photodegradation
either
by
absorbing
light
and
dissipating
energy
or
by
quenching
the
excited
states
of
the
CDD/
CDFs;
and
(
3)
the
surface
of
the
ash
itself
may
hinder
photolysis
by
shielding
the
CDD/
CDFs
from
light.

Pennise
and
Kamens
(
1996)
injected
the
emissions
from
the
combustion
of
wood
chips
treated
with
PCP,
plastic
PVC
pipe
shavings,
and
solid
2,4,6­
trichlorophenol
into
25­
m3
outdoor
Teflon
film
chambers.
The
behavior
of
the
eight
congener
groups
was
monitored
for
approximately
one
day.

Experiments
were
performed
during
January,
June,
and
October
of
1994
in
North
Carolina.

Experiments
with
combustion
temperatures
ranging
from
350
to
380

C
were
categorized
as
lowtemperature
experiments,
and
those
ranging
from
760
to
800

C
were
categorized
as
high
temperature
experiments.
Little
or
no
reactivity
was
observed
for
CDD/
CDFs
sorbed
to
particles
resulting
from
the
high
temperature
experiments.
Greater
photochemical
reactivity
was
observed
for
CDD/
CDFs
sorbed
to
particles
resulting
from
the
low
combustion
experiments,
where
photolysis
appeared
to
compete
with
a
CDD/
CDF
production
mechanism
believed
to
be
associated
with
non­
combusted
PCP.
Photolysis
rates
appeared
to
decrease
with
increasing
levels
of
chlorination.
On
low
temperature
combustion
particles,
estimated
TCDD
half­
lives
(
excluding
the
impact
of
the
observed
formation
from
PCP)
increased
from
0.4
hours
under
summer
conditions
to
17
hours
under
winter
Page
31
of
49
conditions.
On
high­
temperature
particles,
estimated
TCDD
half­
lives
ranged
from
6.8
to
62
hours.

Estimated
OCDD
half­
lives
ranged
from
5
to
38
hours
in
low
combustion
temperature
experiments
to
36
to
257
hours
in
high
temperature
combustion
experiments.

i.
Photo
oxidation
Until
recently,
the
reaction
rates
of
hydroxyl
(
OH)
radicals,
ozone
(
O
3),
and
nitrate
(
NO
3)
radicals
with
CDDs
and
CDFs
had
not
been
measured
because,
in
large
part,
the
low
vapor
pressures
of
these
compounds
make
direct
measurements
very
difficult
with
currently
available
techniques.
In
the
absence
of
experimental
data,
Podoll
et
al.
(
1986)
and
Atkinson
(
1987)
estimated
the
half­
life
of
2,3,7,8­
TCDD
vapor
via
OH
oxidation
in
the
atmosphere
to
be
8.3
days
and
3
days,
respectively.

In
a
subsequent
study,
Atkinson
(
1991)
used
published
reaction
rate
data
for
other
organic
compounds
to
estimate
the
OH
radical
reaction
rate
constants
for
vapor­
phase
dibenzofuran
and
dibenzo­
p­
dioxin,
and
from
these
estimates,
Atkinson
(
1991)
estimated
the
OH
radical
reaction
rate
constants
for
the
CDDs
and
CDFs.
Based
on
these
empirical
estimates,
Atkinson
(
1991)
concluded
that
the
OH
radical
reaction
is
likely
to
be
the
dominant
gas­
phase
transformation
process
for
vapor
phase
CDDs
and
CDFs.
The
tropospheric
lifetimes
calculated
by
Atkinson
(
1991)
from
the
rate
constant
estimates
increased
with
increasing
levels
of
chlorination
from
2.0
days
for
2,3,7,8­
TCDD
and
4.4
days
for
2,3,7,8­
TCDF
to
9.6
days
for
OCDD
and
39
days
for
OCDF.

Kwok
et
al.
(
1994)
expanded
the
work
of
Atkinson
(
1991)
by
experimentally
determining
the
room
temperature
gas­
phase
reaction
rate
constants
of
dibenzofuran
and
dibenzo­
p­
dioxin
with
hydroxyl
radical
to
be
3.9E­
12
cm3
molecule­
1
s­
1
and
1.48E­
11
cm3
molecule­
1
s­
1,
respectively.

These
measured
rate
constants
for
dibenzo­
p­
dioxin
and
dibenzofuran
are
lower
than
those
estimated
by
Atkinson
(
1991)
by
factors
of
2.5
and
8.
Assuming
a
12­
hour
average
daytime
OH
radical
concentration
of
8
x
105
molecule/
cm3,
Kwok
et
al.
(
1994)
estimated
the
atmospheric
lifetimes
for
gas
phase
reactions
of
dibenzofuran
and
dibenzo­
p­
dioxin
with
OH
radicals
to
be
3.7
days
and
1.0
day,
respectively.
Also,
based
on
experimental
data,
Kwok
et
al.
(
1994)
calculated
lifetimes
for
the
gas
phase
reaction
of
dibenzofurans
with
NO
3
and
O
3
to
be
greater
than
7
years
and
greater
than
205
days,
respectively;
the
calculated
lifetimes
for
the
gas
phase
reaction
of
dibenzo­
p­
dioxin
with
NO
3
and
O
3
were
4.9
days
and
greater
than
330
days,
respectively.
The
latter
results
indicate
that
reaction
with
the
OH
radical
is
the
dominant
photo
oxidation
mechanism.

Kwok
et
al.
(
1995)
extended
the
work
of
Kwok
et
al.
(
1994)
by
measuring
the
OH
radical
reaction
rate
constants
for
1­
chlorodibenzo­
p­
dioxin,
diphenyl
ether,
and
1,2­
dimethoxybenzene.

These
new
reaction
rate
data
when
taken
together
with
the
measurements
of
Kwok
et
al.
(
1994)
and
Page
32
of
49
the
estimation
method
described
in
Atkinson
(
1991)
were
used
to
generate
more
reliable
estimates
of
the
reaction
rate
constants
for
the
2,3,7,8­
substituted
CDDs
and
CDFs
(
Atkinson,
1996).
Table
7
presents
these
recalculated
rate
constants
and
tropospheric
lifetimes
and
half­
lives.
As
can
be
seen
from
Table
7,
the
persistence
of
CDD/
CDFs
increases
with
increasing
degree
of
chlorination.

Recently,
Brubaker
and
Hites
(
1997)
measured
the
OH
radical
reaction
rate
constants
for
dibenzo­
p­
dioxin,
dibenzofuran,
2,7­
dichlorodibenzo­
p­
dioxin
(
2,7­
D),
2,8­
dichlorodibenzofuran
(
2,8­
F),
and
1,2,3,4­
dibenzo­
p­
dioxin
(
1,2,3,4­
D)
over
temperatures
ranging
from
72
to
159

C.

From
these
results,
OH
reaction
rate
constants
were
estimated
for
each
compound
at
25

C.
When
these
estimated
values
were
compared
to
the
previously
measured
rate
constants
reported
by
Kwok
et
al.
(
1994;
1995)
and
the
values
predicted
by
Atkinson
(
1995;
1996),
Brubaker
and
Hites
(
1997)

concluded
that
Atkinson's
structure
activity
method
is
reliable.

Biotic
Degradation:

Biotransformation
and
Biodegradation
Most
investigations
examining
the
biodegradability
of
CDDs
and
CDFs
have,
until
recently,

focused
on
the
microbial
degradation
of
2,3,7,8­
TCDD.
Arthur
and
Frea
(
1989)
provided
a
comprehensive
review
of
studies
conducted
during
the
1970s
and
1980s
and
concluded
that
2,3,7,8­

TCDD
is
recalcitrant
to
microbial
degradation.
Several
major
studies
conducted
during
that
time
period
as
well
as
more
recent
studies
are
discussed
below.
Page
33
of
49
Table
7.
Estimated
Tropospheric
Half­
Lives
of
CDDs/
CDFs
with
Respect
to
Gas­
Phase
Reaction
with
the
OH
Radical
Congener
Group
2,3,7,8­
Substituted
Congener
Estimated
OH
Reaction
Rate
Constant
(
cm3/
molecule­
sec)
Estimated
Tropospheric
Lifetimea,
b
(
days)
Estimated
Tropospheric
Half­
Lifea,
c
(
days)

TCDD
2,3,7,8­
TCDD
7.08E­
13
17
12
PeCDD
1,2,3,7,8­
PeCDD
4.59E­
13
26
18
HxCDD
1,2,3,4,7,8­
HxCDD
1,2,3,6,7,8­
HxCDD
1,2,3,7,8,9­
HxCDD
1.97E­
13
2.95E­
13
2.95E­
13
61
40
40
42
28
28
HpCDD
1,2,3,4,6,7,8­
HpCDD
1.30E­
13
92
64
OCDD
OCDD
5.09E­
14
234
162
TCDF
2,3,7,8­
TCDF
4.26E­
13
28
19
PeCDF
1,2,3,7,8­
PeCDF
2,3,4,7,8­
PeCDF
2.65E­
13
2.49E­
13
45
48
31
33
HxCDF
1,2,3,4,7,8­
HxCDF
1,2,3,6,7,8­
HxCDF
1,2,3,7,8,9­
HxCDF
2,3,4,6,7,8­
HxCDF
1.06E­
13
1.51E­
13
1.62E­
13
1.40E­
13
113
79
74
85
78
55
51
59
HpCDF
1,2,3,4,6,7,8­
HpCDF
1,2,3,4,7,8,9­
HpCDF
6.04E­
14
6.78E­
14
198
176
137
122
OCDF
OCDF
2.58E­
14
462
321
a
Calculated
using
a
24­
hour,
seasonal,
and
global
tropospheric
average
OH
radical
concentration
of
9.7
X
105molecule/
cm3
(
Prinn
et
al.,
1995).

b
Lifetime
=
[(
reaction
rate
constant)(
OH
concentration)]­
1.

c
Half­
life
=
0.693/[(
reaction
rate
constant)(
OH
concentration)].

Source:
Based
on
Atkinson
(
1996).
Page
34
of
49
Matsumura
and
Benezet
(
1973)
tested
approximately
100
strains
of
micro­
organisms
that
were
shown
previously
to
degrade
persistent
pesticides;
only
five
strains
showed
any
ability
to
metabolize
2,3,7,8­
TCDD,
based
on
autoradiographs
of
thin­
layer
chromatograms.
Hutter
and
Philippi
(
1982)
concluded
that
although
it
is
possible
that
the
less
chlorinated
dioxins
are
more
susceptible
to
biodegradation,
microbial
action
on
2,3,7,8­
TCDD
is
very
slow
under
optimum
conditions;
the
long­
term
incubations
of
radiolabeled
2,3,7,8­
TCDD
yielded
no
radioactivity
in
carbon
dioxide
traps
after
1
year,
and
analyses
of
the
cultures
showed
that,
at
most,
1
to
2
percent
of
the
initial
2,3,7,8­
TCDD
were
recovered
as
a
potential
metabolite
(
assumed
to
be
a
hydroxylated
derivative
of
2,3,7,8­
TCDD).
Camoni
et
al.
(
1982)
added
organic
compost
to
contaminated
soil
from
the
Seveso
area
to
enrich
the
soil
and
enhance
the
2,3,7,8­
TCDD
biodegradation
rate;
however,
the
soil
amendment
had
no
clear
effect
on
degradation.
Quensen
and
Matsumura
(
1983)
reported
that
low
concentrations
(
5
ppb)
of
radiolabeled
2,3,7,8­
TCDD
were
metabolized
by
pure
cultures
of
Nocardiopsis
spp.
and
Bacillus
megaterium
that
had
been
isolated
from
farm
soil.
The
extent
of
metabolism
after
1­
week
incubation
was
strongly
dependent
on
the
carrier
solvent
used
to
dissolve
and
introduce
the
2,3,7,8­
TCDD
to
the
culture
medium.
The
solvent
ethyl
acetate
gave
the
best
results;
52
percent
of
14C
were
recovered
as
2,3,7,8­
TCDD
out
of
a
total
of
77
percent
14C
recovered.

However,
incubation
of
2,3,7,8­
TCDD
in
farm
soil,
garden
soil,
and
forest
soil
resulted
in
little,
if
any,

metabolism
of
2,3,7,8­
TCDD.

Bumpus
et
al.
(
1985)
tested
the
white
rot
fungus,
Phanerochaete
chrysosporium,
which
secretes
a
unique
H
2
0
2­
dependent
extracellular
lignin­
degrading
enzyme
system
capable
of
generating
carbon­
centered
free
radicals
(
Tien
and
Kirk,
1983;
Tien
and
Kirk,
1984).
Lignin
is
resistant
to
attack
by
all
microorganisms
except
some
species
of
fungi
and
a
relatively
small
number
of
bacteria
species.

Radiolabeled
2,3,7,8­
TCDD
was
oxidized
to
labeled
C0
2
by
nitrogen­
deficient,
ligninolytic
cultures
of
P.
chrysosporium;
because
the
label
was
restricted
to
the
ring,
it
was
concluded
that
the
strain
was
able
to
degrade
halogenated
aromatic
rings.
In
10
mL
cultures
containing
1,250
pmol
of
substrate,

27.9
pmol
of
2,3,7,8­
TCDD
were
converted
to
labeled­
CO
2
during
the
30­
day
incubation
period;

thus,
only
about
2
percent
of
the
starting
material
were
converted.

Hofmann
et
al.
(
1992)
demonstrated
that
the
fungi,
Fusarium
redolens,
could
degrade
3­

chlorodibenzofuran
and,
to
a
lesser
degree,
mono­
and
di­
CDDs.
Hoffman
et
al.
(
1992)
also
identified
14
other
strains
of
fungi
that
demonstrated
the
capability
to
degrade
dibenzofuran
(
nonchlorinated).
The
strains
are
members
of
the
following
genera:
Mucor,
Chaetomium,
Phoma,

Fusarium,
Paecilomyces,
Papulaspora,
Inonotus,
Lentinus,
Phanerochaete,
Polyporus,
Pycnoporus,

Schizophyllum,
and
Trametes.
Page
35
of
49
Takada
et
al.
(
1994;
1996)
reported
significant
degradation
of
2,3,7,8­
substituted
CDDs
and
CDFs
by
low­
nitrogen
medium
cultures
of
the
white
rot
fungus,
Phanerochaete
sordida
YK­
624
strain.
Tetra­
through
octa­
CDDs
and
CDFs
were
incubated
for
14
days
in
glucose­
amended
cultures
at
30

C.
For
both
CDDs
and
CDFs,
the
1,2,3,6,7,8­
congeners
showed
the
highest
degradation
values,
75
percent
and
70
percent,
respectively.
The
lowest
degradation
values
were
for
2,3,7,8­
TCDD
(
40
percent),
1,2,3,7,8­
TCDF
(
45
percent),
and
1,2,3,7,8­
PeCDF.
Similar
results
were
obtained
under
the
same
conditions
for
P.
chrysoporium
IFO
31249
strain.

Several
recent
reports
indicate
that
CDDs
and
CDFs,
like
PCBs,
may
undergo
microbial
dechlorination
in
anaerobic
sediments.
Adriaens
and
Grbic­
Galic
(
1992;
1993)
and
Adriaens
et
al.

(
1995)
have
reported
the
results
of
a
series
of
microcosm
studies
utilizing
Hudson
River
sediment
(
contaminated
with
Aroclor
1242)
and
aquifer
material
(
contaminated
with
CDDs)
from
Pensacola,

Florida.
Both
types
of
substrates
were
spiked
with
several
CDDs
(
1,2,3,4,6,7,8­
HpCDD;
1,2,3,4,7,8­

HxCDD;
and
1,2,4,6,8,9­/
1,2,4,6,7,9­
HxCDD)
and
CDFs
(
1,2,3,4,6,7,8­
HpCDF
and
1,2,4,6,8­

PeCDF)
and
monitored
over
a
period
of
16
months
at
an
incubation
temperature
of
30

C.
The
Hudson
River
sediment
was
spiked
with
144

g/
kg
of
each
congener
and
the
Pensacola
aquifer
material
was
spiked
with
63

g/
kg
of
each
congener.
Recoveries
of
the
CDD/
CDF
congeners
from
the
control
samples
decreased
with
increasing
incubation
time
indicating
that
these
congeners
are
strongly
sorbed
to
the
substrates.
For
example,
after
50
days
of
incubation,
the
fraction
of
CDD/
CDF
that
could
be
recovered
by
manual
extraction
had
already
decreased
to
20­
40
percent.

All
of
the
congeners,
with
the
exception
of
HpCDF,
showed
a
slow
decrease
in
concentration
over
time
attributed
to
biologically
mediated
reductive
dechlorination
with
net
disappearance
rates
ranging
from
0.0031
wk­
1
to
0.0175
wk­
1
(
i.
e.,
half­
lives
of
approximately
1
to
4
years).
However,

Adriaens
et
al.
(
1995)
conclude
that
the
actual
half­
lives
may
be
orders
of
magnitude
higher.
If
it
is
assumed
that
transformation/
degradation
occurs
only
for
CDD/
CDF
in
the
aqueous
phase,
then
the
CDD/
CDF
that
sorb
to
the
sediments
may
never
be
biologically
available
because
of
the
apparent
very
slow
rates
of
desorption.
The
experiment
with
1,2,3,4,6,7,8­
HpCDD
yielded
formation
of
two
HxCDD
(
1,2,3,4,7,8­
and
1,2,3,6,7,8­).
Thus,
removal
of
the
peri­
substituted
(
1,4,6,9)
chlorines
was
favored
with
enrichment
of
2,3,7,8­
substituted
congeners.
No
lesser
chlorinated
congeners
were
identified
from
incubations
with
the
other
tested
congeners.
1,2,4,6,8­
PeCDF
was
also
examined
in
dichlorophenol­
enriched
cultures.
After
6
months
incubation,
several
TCDFs
were
identified
which
also
indicated
that
peri­
dechlorination
was
the
preferred
route
of
reduction.

Barkovskii
et
al.
(
1994)
expanded
the
testing
of
Adriaens
and
Grbic­
Galic
(
1992;
1993)
by
spiking
the
sediments
with
higher
doses
to
determine
if
faster
rates
could
be
achieved.
Passaic
River
sediments
(
contaminated
with
CDD/
CDFs)
were
spiked
with
4,500

g/
kg
of
OCDD
and
incubated
Page
36
of
49
under
anaerobic
conditions
for
6
months.
Although
no
significant
degradation
of
OCDD
was
observed,
significant
reductions
in
the
concentrations
of
the
hepta­,
hexa­,
penta­,
and
tetra­
CDDs
were
observed.

Barkovskii
and
Adriaens
(
1995;
1996)
reported
that
2,3,7,8­
TCDD
(
extracted
from
Passaic
River
sediments)
was
susceptible
to
reductive
dechlorination
when
incubated
at
30

C
under
methanogenic
conditions
in
a
mixture
of
aliphatic
and
organic
acids
inoculated
with
microorganisms
obtained
from
Passaic
River
sediments.
The
initial
concentration
of
2,3,7,8­
TCDD
(
20
±
4

g/
L)

decreased
by
30
percent
to
14
±
2

g/
L
over
a
period
of
7
months
with
the
consecutive
appearance
and
disappearance
of
tri­,
di­,
and
mono­
CDDs.
Experiments
were
also
conducted
by
spiking
the
sediment
with
HxCDDs,
HpCDDs,
and
OCDD.
Up
to
10
percent
of
the
spiked
OCDD
was
converted
to
hepta­,
hexa­,
penta­,
tetra­,
tri­,
di­,
and
monochlorinated
isomers,
but
the
reaction
stoichiometry
was
not
determined.
Two
distinct
pathways
of
dechlorination
were
observed:
the
peri­
dechlorination
pathway
of
2,3,7,8­
substituted
hepta­
to
penta­
CDDs,
resulting
in
the
production
of
2,3,7,8­
TCDD,

and
the
peri­
lateral
dechlorination
pathway
of
non­
2,3,7,8­
substituted
congeners.
Direct
evidence
of
further
lateral
dechlorination
of
2,3,7,8­
TCDD
was
obtained
from
the
historically
contaminated
incubations.
Pasteurized
cells
exhibited
no
peri­
dechlorination
pathway,
and
triCDDs
were
the
leastchlorinated
congeners
produced
in
these
treatments.
These
results
demonstrate
that:
(
1)
both
freshly
spiked
and
aged
CDDs
are
available
to
microbial
reductive
dechlorination;
(
2)
the
peri
and
triCDD
dechlorinations
are
attributed
to
activities
of
nonmethanogenic,
non­
spore­
forming
microbial
subpopulations;
and
(
3)
the
2,3,7,8­
residue
patterns
in
historically
contaminated
sediments
are
likely
affected
by
microbial
activity.

Biochemical
Formation
of
CDD/
CDFs
from
Chlorophenols
Many
researchers
demonstrated
in
laboratory
conditions
that
biochemical
formation
of
CDD/
CDFs
from
chlorophenol
precursors
is
possible.
However,
the
extent
to
which
CDD/
CDFs
are
formed
in
the
environment
via
this
mechanism
can
not
be
estimated
at
this
time.
This
finding
implies
that
potentially
the
PCP
that
is
volatilized
or
leached
from
the
utility
poles
might
convert
to
CDD/
CDFs
in
the
environment,
thereby
providing
additional
CDD/
CDF
inputs
into
the
environment.

The
studies
of
these
researchers
are
described
briefly
below.

Several
researchers
(
Svenson
et
al.,
1989;
Oberg
et
al.,
1990;
Wagner
et
al.,
1990;
Oberg
and
Rappe,
1992;
Moritomo
and
Kenji,
1995)
have
demonstrated
the
biochemical
formation
of
CDD/
CDFs
from
chlorophenols
in
laboratory
studies
conducted
with
solutions
of
trichlorophenol
and
pentachlorophenol
in
the
presence
of
peroxidase
enzymes
and
hydrogen
peroxide,
and
with
sewage
sludge
spiked
with
pentachlorophenol.
Peroxidase
are
common
enzymes
in
nature.
The
actual
Page
37
of
49
conversion
efficiency
of
chlorinated
phenols
to
CDD/
CDFs
observed
in
these
studies
was
low,

however.
In
the
solution
studies,
Oberg
and
Rappe
(
1992)
reported
a
conversion
efficiency
of
PCP
to
OCDD
of
about
0.01
percent;
Morimoto
and
Kenji
(
1995)
reported
a
conversion
effficiency
of
PCP
to
OCDD
of
about
0.8
percent;
and
Wagner
et
al.
(
1990)
reported
a
conversion
efficiency
of
trichlorophenol
to
HpCDD
of
about
0.001
percent.
Oberg
et
al.
(
1990)
reported
a
conversion
efficiency
of
trichlorophenol
to
CDD/
CDFs
of
about
0.001
percent.
In
the
sewage
sludge
study
(
Oberg
et
al.,
1992),
a
conversion
efficiency
of
PCP
to
total
CDDs
of
0.0002
to
0.0004
percent
was
reported.

Some
researchers
have
speculated
that
CDD/
CDFs
commonly
found
in
compost
might
result
in
part
from
the
biochemical
formation
of
CDD/
CDFs
from
PCP
in
treated
wood
(
Goldfarb
et
al.,

1992;
Malloy
et
al.,
1993).
Oberg
et
al.
(
1993)
measured
the
extent
of
CDD/
CDF
formation
in
three
conventional
garden
composts:
two
were
spiked
with
PCP,
and
one
was
spiked
with
hexachlorobenzene.
The
two
PCP­
spiked
composts
were
monitored
for
periods
of
55
days
and
286
days,
respectively.
A
significant
increase
in
the
concentrations
of
the
higher
chlorinated
congeners,

particularly
the
HpCDDs,
OCDD,
and,
to
a
lesser
extent,
OCDF,
were
observed.
Similar
results
were
reported
for
the
hexachlorobenzene­
spiked
compost,
which
was
monitored
for
a
period
of
49
days.

However,
in
a
study
conducted
with
sewage
sludges
from
two
German
communities,
Weber
(
1995)

did
not
observed
any
significant
changes
in
CDD/
CDF
concentrations
after
the
chlorophenol­
spiked
sewage
sludges
were
subjected
to
anaerobic
digestion
in
laboratory
reactors
for
60
days.
The
sewage
sludges
were
spiked
with
2,3,5­
trichlorophenol
(
10
to
25
mg/
kg),
a
mixture
of
2,3,5­
trichlorophenol
and
diochlorophenols
(
2.5
to
35
mg/
kg),
and
a
mixture
of
di­,
tri­,
and
tetra­
chlorobenzenes
(
4
to
40
mg/
kg).
The
initial
CDD/
CDFs
concentrations
in
the
two
sewage
sludges
tested
were
9­
and
20­
ng
TEQ/
kg.

Several
researchers
at
the
U.
S.
Department
of
Agriculture
(
USDA)
reported
that
dairy
cows
fed
with
PCP­
treated
wood
excreted
OCDD
in
amounts
almost
four
fold
over
what
they
had
ingested
(
Fries
et
al.,
1997).
Feil
and
Tiernan
(
1997)
reported
that
rats
fed
with
technical
PCP
had
liver
concentrations
of
HxCDD,
HpCDD,
HpCDF,
OCDD,
and
OCDF
two
or
three
orders
of
magnitude
higher
than
rates
fed
with
purified
PCP.
These
results
suggest
the
in
vivo
formation
of
CDD/
CDFs
from
pre­
dioxins
(
i.
e.,
chlorinated
phenoxy
phenols
present
as
contaminants
in
PCP).
A
follow­
on
study
(
Huwe
et
al.,
1998)
investigated
the
metabolic
conversion
of
a
pre­
dioxin
(
monochloro­
2­

phenoxyphenol)
to
OCDD
in
a
feeding
study
with
rats.
The
results
of
this
study
demonstrated
the
formation
of
OCDD
from
the
pre­
dioxin,
although
the
conversion
was
estimated
to
be
less
than
2
percent.

Wittsiepe
et
al.
(
1998)
demonstrated
that
CDD/
CDF
can
be
formed
through
the
reaction
of
chlorophenols
with
myeloperioxidase
(
a
component
of
neutrophilic
granulocytes,
a
subgroup
of
human
leucocytes).
The
CDD/
CDFs
formed
showed
different
homolog
patterns
and
formation
rates
Page
38
of
49
depending
on
the
degree
of
chlorination
of
the
chlorophenol
substrate.
The
formation
rates
ranged
from
1
to
16
umole
of
CDD/
CDFs
per
mole
of
chlorophenol
substrate.

Environmental
Fate
Assessment:

Presence
of
CDDs
and
CDFs
in
the
environmental
compartments
is
due
to
volatilization
into
air,
leaching
from
wood
preservative
treated
poles
into
water
and
soil,
dry
and
wet
deposition
onto
air,
water
and
soils,
and
sorption
into
soils.
A
quantitative
estimate
of
the
dioxins
in
the
environmental
compartments
is
an
unresolved
issue.
The
available
scientific
data
indicate
that
CDDs
and
CDFs,

particularly
the
tetras­
and
higher
chlorinated
congeners,
are
extremely
stable
under
most
environmental
conditions.
However,
some
of
these
congeners,
under
certain
conditions,
are
photolytically
unstable
and
in
some
cases
undergo
photoooxidation.
Most
of
the
congeners
are
also
resistant
to
biodegradation
under
aerobic
or
anaerobic
soil
conditions
and
most
are
persistent
in
soils.

The
process
of
bioaccumulation
has
been
observed
in
the
benthic
organisms
but
biotransformation
process
up
the
food
chain
has
not
been
observed.
A
sound
environmental
fate
assessment
of
CDDs
and
CDFs'
release
into
the
environmental
compartments
of
water
and
soil
due
to
the
sole
use
of
PCP
(
in
which
the
dioxins
are
present
as
micro
contaminants)
as
a
wood
preservative
can
not
be
made
due
to
the
unknown
source
and
exact
quantitation
of
CDDs
and
CDFs
in
the
environment.
Page
39
of
49
BIBLIOGRAPHY
Adriaens,
P.;
Grbic­
Galic,
D.
(
1992)
Effect
of
cocontaminants
and
concentration
on
the
anaerobic
biotransformation
of
PCDD/
F
in
methanogenic
river
sediments.
Organohalogen
Compounds
8:
209­
212.

Adriaens,
P.;
Grbic­
Galic,
D.
(
1993)
Reductive
dechlorination
of
PCDD/
F
by
anaerobic
cultures
and
sediments.
Organohalogen
Compounds
12:
107­
110.

Adriaens,
P.;
Fu,
Q.;
Grbic­
Galic,
D.
(
1995)
Bioavailability
and
transformation
of
highly
chlorinated
dibenzo­
p­
dioxins
and
dibenzofurans
in
anaerobic
soils
and
sediments.
Environ.
Sci.
Technol.
29(
9):
2252­
2260.

Andersson,
P.;
Marklund,
S.;
Rappe,
C.
(
1992)
Levels
and
profiles
of
PCDDs
and
PCDFs
in
environmental
samples
as
determined
in
snow
deposited
in
northern
Sweden.
Organohalogan
Compounds
8:
307­
310.

Arthur,
M.
F.;
Frea,
J.
I.
(
1989)
2,3,7,8­
Tetrachlorodibenzo­
p­
dioxins,
Aspects
of
its
important
properties
and
its
potential
biodegradation
in
soils.
J.
Environ.
Qual.
18:
1­
11.

Atkinson,
R.
(
1987)
Estimation
of
OH
radical
reaction
rate
constants
and
atmospheric
lifetimes
for
polychlorobiphenyls,
dibenzo­
p­
dioxins,
and
dibenzofurans.
Environ.
Sci.
Technol.
21:
305­
307.

Atkinson,
R.
(
1991)
Atmospheric
lifetimes
of
dibenzo­
p­
dioxins
and
dibenzofurans.
The
Science
of
the
Total
Environment.
104:
17­
33.

Atkinson,
R.
(
1995)
Personal
communication
with
Greg
Schweer
(
Versar,
Inc.)
Estimated
congenerspecific
gas­
phase
reaction
rate
constants
of
OH
radical
with
PCBs,
PCDDs,
PCDFs.

Atkinson,
R.
(
1996)
Atmospheric
chemistry
of
PCBs,
PCDDs
and
PCDFs.
Issues
in
Environmental
Science
and
Technology.
6:
53­
72.

Bacci,
E.;
Cerejeira,
M.
J.;
Gaggi,
C.;
Chemello,
G.;
Calamari,
D.;
Vighi,
M.
(
1990)
Bioconcentration
of
organic
chemical
vapours
in
plant
leaves:
the
azalea
model.
Chemosphere
21(
4­
5):
525­
535.

Barkovskii,
A.;
Adriaens,
P.
(
1996)
Microbial
dechlorination
of
historically
present
and
freshly
spiked
chlorinated
dioxins
and
diversity
of
dioxin­
dechlorinating
populations.
Applied
and
Environmental
Microbiology
62(
12):
4556­
4562.

Barkovskii,
A.;
Fu,
Q.;
Adriaens,
P.
(
1994)
Biological
and
abiotic
dechlorination
of
highly
chlorinated
PCDD/
PCDF:
Issues
of
bioavailability
and
pathways.
Organohalogen
Compounds
21:
469­
473.
Page
40
of
49
Barkovskii,
A.
L.;
Adriaens,
P.
(
1995)
Reductive
dechlorination
of
tetrachloro­
dibenzo­
p­
dioxin
partitioned
from
Passaic
River
sediments
in
an
autochthonous
microbial
community.
Organohalogen
Compounds
24:
17­
21.

Bobet,
E.;
Berard,
M.
F.;
Dann,
T.
(
1990)
The
measurement
of
PCDD
and
PCDF
in
ambient
air
in
southwestern
Ontario.
Chemosphere
20(
10­
12):
1439­
1445.

Bremmer,
H.
J.;
Troost,
L.
M.;
Kuipers,
G.;
de
Koning,
J.;
Sein,
A.
A.
(
1994)
Emissions
of
dioxin
in
The
Netherlands.
Report
No.
770501018.
Bilthoven
Netherlands:
RIVM,
1994.

Broman,
D.;
Naf,
C.;
Zebühr,
Y.
(
1992)
Occurrence
and
dynamics
of
polychlorinated
dibenzo­
pdioxins
and
dibenzofurans
and
other
combustion
related
organic
pollutants
in
the
aquatic
environment
of
the
Baltic.
Chemosphere
25(
1­
2):
125­
128.

Brubaker,
W.
W.;
Hites,
R.
A.
(
1997)
Gas­
phase
hydroxyl
radical
reactions
and
related
atmospheric
removal
of
polychlorinated
dibenzo­
p­
dioxins
and
dibenzofurans.
Organohalogen
Compounds
33:
235­
239.

Brzuzy,
L.
P.;
Hites,
R.
A.
(
1995)
Estimating
the
atmospheric
deposition
of
polychlorinated
dibenzop
dioxins
and
dibenzofurans
from
soil.
Environ.
Sci.
Technol.
29:
2090­
2098.

Bumpus,
J.
A.;
Tien,
M.;
Wright,
D.;
Aust,
S.
D.
(
1985)
Oxidation
of
persistent
environmental
pollutants
by
a
white
rot
fungus.
Science
228:
1434­
1436.

Burkhard,
L.
P.;
Kueh.,
D.
W.
(
1986)
N­
octanol/
water
partition
coefficients
by
reverse
phase
liquid
chromatography/
mass
spectrometry
for
eight
tetrachlorinated
planar
molecules.
Chemosphere.
15(
2):
163­
167.

Buser,
H.
R.
(
1976)
Preparation
of
qualitative
standard
mixtures
of
polychlorinated
dibenzo­
p­
dioxins
and
dibenzofurans
by
ultraviolet
and
gamma
irradiation
of
the
octachlorocompounds.
J.
Chromatogr.
192:
303­
307.

Buser,
H.
R.
(
1988)
Rapid
photolytic
decomposition
of
brominated
and
brominated/
chlorinated
dibenzodioxins
and
dibenzofurans.
Chemosphere
17:
889­
903.

Camoni,
I.;
Dimuccio,
A.;
Pontecorvo,
D.;
Taggi,
F.;
Vergori,
I.
(
1982)
Laboratory
investigation
for
the
microbial
degradation
of
2,3,7,8­
tetrachlorodibenzo­
p­
dioxin
in
soil
by
addition
of
organic
compost.
Pergammon
Ser.
Environ.
Sci.
5:
95­
103.

Choudry,
G.
G.;
Webster,
G.
R.
B.
(
1987)
Environmental
photochemistry
of
polychlorinated
dibenzofurans
(
PCDFs)
and
dibenzo­
p­
dioxins
(
PCDDs):
A
review.
Toxicol.
Environ.
Chem.
14:
43­
61.

Choudry,
G.
G.;
Webster,
G.
R.
B.
(
1989)
Environmental
photochemistry
of
PCDDs.
2.
Quantum
yields
of
direct
phototransformation
of
1,2,3,7­
tetra­,
1,3,6,8­
tetra­,
1,2,3,4,6,7,8­
hepta­,
and
Page
41
of
49
1,2,3,4,6,7,8,9­
octachlorodibenzo­
p­
dioxin
in
aqueous
acetonitrile
and
their
sunlight
halflives
J.
Agric.
Food
Chem.
37:
254­
261.

Choudry,
G.
G.;
Foga,
M.;
Webster,
G.
R.
B.;
Muir,
D.
C.
G.;
Friesen,
K.
(
1990)
Quantum
yields
of
the
direct
phototransformation
of
1,2,4,7,8­
penta­
and
1,2,3,4,7,8­
hexa
chlorodibenzofuran
in
aqueous
acetonitrile
and
their
sunlight
half­
lives.
Toxicological
and
Environmental
Chemistry
26:
181­
195.

Cooper,
P.;
Ung,
T.;
Aucoin,
J­
P.;
Timusk,
C.
(
1996)
The
potential
for
re­
use
of
preservativetreated
utility
poles
removed
from
service.
Waste
Management
&
Research.
14:
263­
279.

Cousins,
I.
T.;
Hartlieb,
N.;
Teichmann,
C.;
Jones,
K.
C.
(
1996)
Volatilization
of
polychlorinated
biphenyls
from
sludge­
amended
soils.
Organohalogen
Compounds
28:
58­
63.

Crosby,
D.
G.;
Wong,
A.
S.;
Plimmer,
J.
R.;
Woolson,
E.
A.
(
1971)
Photodecomposition
of
chlorinated
dibenzo­
p­
dioxins.
Science
173:
748.

Crosby,
D.
G.;
Moilanen,
K.
W.;
Wong,
A.
S.
(
1973)
Environmental
generation
and
degradation
of
dibenzodioxins
and
dibenzofurans.
Environ.
Health
Perspect.,
Exp.
Issue.
5:
259­
266.

Crosby,
D.
G.
(
1978)
Conquering
the
monster
­
the
photochemical
destruction
of
chlorodioxins.
In:
Disposal
and
decontamination
of
pesticides.
M.
V.
Kennedy,
Ed.
ACS
Symposium
Series
73:
1­
12.

Crosby,
D.
G.
(
1981)
Methods
of
photochemical
degradation
of
halogenated
dioxins
in
view
of
environmental
reclamation.
Paper
presented
on
"
Human
health
aspects
of
accidental
exposure
to
dioxins.
Strategy
for
environmental
reclamation
and
community
protection"
Bethesda,
MD,
October
5­
7,
1982.

De
Fre,
R.;
Van
Cleuvenbergen,
R.;
Schoeters,
J.
(
1994)
Measurement
of
deposition
of
dioxins
in
Flanders,
Belgium.
Organohalogen
Compounds
20:
9­
14.

diDomenico,
A.;
Viviano,
G.;
Zapponi,
G.
(
1982)
Environmental
persistence
of
2,3,7,8­
TCDD
at
Seveso.
In:
Chlorinated
dioxins
and
related
compounds,
impact
on
the
environment.
O.
Hutzinger
et
al.,
Eds.
Elmsford,
NY:
Pergamon
Press,
pp.
105­
113.

diDomenico,
A.;
Cerlesi,
S.;
Ratti,
S.
(
1990)
A
two­
exponential
model
to
describe
the
vanishing
trend
of
2,3,7,8­
tetrachlorodibenzodioxin
(
TCDD)
in
the
soil
at
Seveso,
Northern
Italy.
Chemosphere
20(
10­
12):
1559­
1566.

Dobbs,
A.
J.;
Grant,
C.
(
1979)
Photolysis
of
highly
chlorinated
dibenzo­
p­
dioxins
by
sunlight.
Nature
278(
8):
163­
165.

Dougherty,
E.
J.;
McPeters,
A.
L.;
Overcash,
M.
R.
(
1991)
Kinetics
of
photodegradation
of
2,3,7,8­
tetrachlorodibenzo­
p­
dioxins:
theoretical
maximum
rate
of
soil
decontamination.
Chemosphere
23(
5):
589­
600.
Page
42
of
49
Dulin,
D.;
Drossman,
H.;
Mill,
T.
(
1986)
Products
and
quantum
yields
for
photolysis
of
chloroaromatics
in
water.
Environ.
Sci.
Technol.
20:
72­
77.

Dung,
M.;
O'Keefe,
P.
W.
(
1992)
Comparative
rates
of
photolysis
of
polychlorinated
dibenzofurans
in
organic
solvents
and
in
aqueous
solutions.
Organohalogen
Compounds
8:
233­
236.

Eitzer,
B.
D.;
Hites,
R.
A.
(
1987)
Reply
to
the
comment
on
"
airborne
dioxins
and
furans:
source
and
fate".
Environ
Sci.
Technol.
21:
924.

Eitzer,
B.
D.;
Hites,
R.
A.
(
1989)
Polychlorinated
dibenzo­
p­
dioxins
and
dibenzofurans
in
the
ambient
atmosphere
of
Bloomington,
Indiana.
Environ.
Sci.
Technol.
23:
1389­
1395.

Feil,
V.
J.;
Tiernan,
T.
(
1997)
Pentachlorophenol
as
a
source
of
dioxins
and
furans.
Organohalogen
Compounds
33:
353­
354.

Fernandez
(
1992)
The
analysis
of
toxic
organic
micro
pollutants
in
ambient
air
and
deposition.
Chemosphere
26(
7­
10):
1311­
1316.

Fletcher,
C.
L.;
McKay,
W.
A.
(
1992)
Polychlorinated
dibenzo­
p­
dioxins
(
PCDDs)
and
dibenzofurans
(
PCDFs)
in
the
aquatic
environment
­
a
literature
review.
AEA
Environment
and
Energy.
Report
No.
AEA­
EE­
0241.

Freeman,
R.
A.;
Schroy,
J.
M.
(
1985)
Environmental
mobility
of
TCDD.
Chemosphere
14:
873­
876.

Freeman,
R.
A.;
Hileman,
F.
D.;
Noble,
R.
W.;
Schroy,
J.
M.
(
1987)
Experiments
on
the
mobility
of
2,3,7,8­
tetrachlorodibenzo­
p­
dioxin
at
Times
Beach,
Missouri.
In:
Exner,
J.
H.
ed.,
Solving
Hazardous
Waste
Problems,
ACS
Symposium
Series
Number
338.

Freeman,
R.
A.;
Schroy,
J.
M.
(
1989)
Comparison
of
the
rate
of
TCDD
transport
at
Times
Beach
and
at
Elgin
AFB.
Chemosphere
18(
1­
6):
1305­
1312.

Fries,
G.
F.;
Dawson,
T.
E.;
Paustenbach,
D.
J.;
Mathur,
D.
B.;
Luksemburg,
W.
J.
(
1997)
Biosynthesis
of
hepta­
and
octa­
chlorodioxins
in
cattle
and
evidence
for
lack
of
involvement
by
rumen
microorganisms.
Organohalogen
Compounds
33:
296­
301.

Friesen,
J.
K.;
Muir,
D.
C.
G.;
Webster,
G.
R.
B.
(
1990)
Evidence
of
sensitized
photolysis
of
polychlorinated
dibenzo­
p­
dioxins
in
natural
waters
under
sunlight
conditions.
Environ.
Sci.
Technol.
24(
11):
1739­
1744.

Friesen,
K.
J.;
Loewen,
M.
D.;
Foga,
M.
M.
(
1993)
Environmental
aquatic
photodegradation
of
chlorinated
dibenzofurans
and
their
photoproducts.
Organohalogen
Compounds
12:
135­
137.

Gaggi,
C.;
Bacci,
E.
(
1985)
Accumulation
of
chlorinated
hydrocarbon
vapours
in
pine
needles.
Chemosphere
14(
5):
451­
456.
Page
43
of
49
Goldfarb,
T.
D.;
Malloy,
T.
A.;
Surico,
M.
(
1992)
PCDDs,
PCDFs,
PCBs,
chlorophenols
(
CPs)
and
chlorobenzenes
(
CBzs)
in
samples
from
various
types
of
composting
facilities
in
the
United
States.
Organohalogen
Compounds
8:
253­
256.

Govers,
H.
A.
J.;
Krop,
H.
B.
(
1996)
Partition
constants
of
chlorinated
dibenzofurans
and
dibenzo­
pdioxins
Organohalogen
Compounds
28:
5­
10.

Gurprasad,
N.;
Constable,
M.;
Haidar,
N.;
Cabalo,
E.
(
1995)
Polychlorinated
dibenzo­
p­
dioxins
(
PCDDs)
leaching
from
pentachlorophenol­
treated
utility
poles.
Organohalogen
Compounds.
24:
501­
504.

Hagenmaier,
H.;
She,
J.;
Linidig,
C.
(
1992)
Persistence
of
polychlorinated
dibenzo­
p­
dioxins
and
polychlorinated
dibenzofurans
in
contaminated
soil
at
Maulach
and
Rastatt
in
Southwest
Germany.
Chemosphere
25(
7­
10):
1449­
1456.

Hallet,
D.
J.;
Kornelson,
P.
J.
(
1992)
Persistence
of
2,4­
dichlorophenol,
2,4,5­
trichlorophenol,
2,3,7,8­
tetrachlorodibenzo­
p­
dioxin,
and
2,3,7,8­
tetrachlorodibenzofuran
in
soils
of
a
forest
ecosystem
treated
with
2,4­
D/
2,4,5­
T
herbicide
in
Eastern
Canada.
Presented
at:
Dioxin
'
92,
12th
International
Symposium
on
Chlorinated
Dioxins
and
Related
Compounds;
Tampere,
Finland;
August
1992.

Harless,
R.
L.;
Lewis,
R.
G.
(
1992)
Evaluation
of
a
sampling
and
analysis
method
for
determination
of
polyhalogenated
dibenzo­
p­
dioxins
and
dibenzofurans
in
ambient
air.
Chemosphere
25(
7­
10):
1317­
1322.

Hiester,
E.;
Böhm,
R.;
Eynck,
P.;
Gerlock,
A.;
Mülder,
W.;
Restow,
H.
(
1993)
Longterm
monitoring
of
PCDD,
PCDF
and
PCB
in
bulk
deposition
samples.
Organohalogen
Compounds
12:
147­
150.

Hippelein,
M.;
Kaupp,
H.;
Dörr,
G.;
McLachlan,
M.;
Hutzinger,
O.
(
1996)
Baseline
contamination
assessment
for
a
new
resource
recovery
facility
in
Germany.
Part
II:
Atmospheric
concentrations
of
PCDD/
F.
Chemosphere
32(
8):
1605­
1616.

Hites,
R.
A.;
Harless,
R.
L.
(
1991)
Atmospheric
transport
and
deposition
of
polychlorinated
dibenzop
dioxins
and
dibenzofurans.
Research
Triangle
Park,
NC:
U.
S.
Environmental
Protection
Agency,
Office
of
Research
and
Development.
EPA/
600/
3­
91/
002.

Hofmann,
K.
H.;
Polnisch,
E.;
Kreisel,
H.;
Mach,
H.;
Schubert,
M.
(
1992)
Degradation
of
dibenzofuran
and
dibenzodioxins
by
fungi.
Presented
at:
Dioxin
'
92,
12th
International
Symposium
on
Chlorinated
Dioxins
and
Related
Compounds;
Tampere,
Finland;
August
1992.

Horstmann,
M.;
McLachlan,
M.
S.
(
1996)
Evidence
of
a
novel
mechanism
of
semivolatile
organic
compound
deposition
in
coniferous
forests.
Environ.
Sci.
Tech.
30(
5):
1794­
1796.
Page
44
of
49
Hunt,
G.
T.;
Maisel,
B.
E.
(
1992)
Atmospheric
concentrations
of
PCDD
and
PCDFs
in
southern
California.
J.
Air
Waste
Manag.
Assoc.
42:
672­
680.

Hutter,
R.;
Philippi,
M.
(
1982)
Studies
in
microbial
metabolism
of
TCDD
under
laboratory
conditions.
In:
Chlorinated
dioxins
and
related
compounds,
impact
on
the
environment.
O.
Hutzinger
et
al.
eds.
Elmsford,
NY:
Pergammon
Press.
pp.
87­
93.

Huwe,
J.
K.;
Feil,
V.
J.;
Tiernan,
T.
O.
(
1998)
In
vivo
formation
of
octachlorodibenzo­
p­
dioxin
from
a
predioxin.
Organohalogen
Compounds
36:
93­
95.

Kapila,
S.;
Yanders,
A.
F.;
Orazio,
C.
E.;
Meadows,
J.
E.;
Cerlesi,
S.;
Clevenger,
T.
E.
(
1989)
Field
and
laboratory
studies
on
the
movement
and
fate
of
tetrachlorodibenzo­
p­
dioxin
in
soil.
Chemosphere
18:
1297­
1304.

Kieatiwong,
S.;
Nguyen,
L.
V.;
Hebert,
V.
R.;
Hackett,
M.;
Miller,
G.
C.;
Miille,
M.
J.;
Mitzel,
R.
(
1990)
Photolysis
of
chlorinated
dioxins
in
organic
solvents
and
on
soils.
Environ.
Sci.
Technol.
24(
10):
1575­
1580.

Kim,
M.;
O'Keefe,
P.
(
1998)
The
role
of
natural
organic
compounds
in
photosensitized
degradation
of
polychlorinated
dibenzo­
p­
dioxins
and
dibenzofurans.
Organohalogen
Compounds.
36:
377­
380.

Koester,
C.
J.;
Hites,
R.
A.
(
1992a)
Wet
and
dry
deposition
of
chlorinated
dioxins
and
furans.
Environ.
Sci.
Technol.
26(
7):
1375­
1382.

Koester,
C.
J.;
Hites,
R.
A.
(
1992b)
Photodegradation
of
polychlorinated
dioxins
and
dibenzofurans
adsorbed
to
fly
ash.
Environ.
Sci.
Technol.
26(
3):
502­
507.

Koshioka,
M.;
Ishizoka,
M.;
Yamada,
T.;
Kanazawa,
J.;
Murai,
T.
(
1990)
Quantum
yields
of
the
photodegradation
of
1,2,3,4­,
1,3,6,8­,
and
2,3,7,8­
tetrachlorodibenzo­
p­
dioxins
and
their
half­
life
periods.
J.
Pesticide
Sci.
15:
39­
45.

Kwok,
E.
S.
C.;
Arey,
J.;
Atkinson,
R.
(
1994)
Gas­
phase
atmospheric
chemistry
of
dibenzo­
p­
dioxin
and
dibenzofuran.
Environ.
Sci.
Technol.
28(
3):
528­
533.

Kwok,
E.
S.
C.;
Atkinson,
R.;
Arey,
J.
(
1995)
Rate
constants
for
the
gas­
phase
reactions
of
the
OH
radical
with
dichlorobiphenyls,
1­
chlorodibenzo­
p­
dioxin,
1,2­
dimethoxybenzene,
and
diphenyl
ether:
estimation
of
OH
radical
reaction
rate
constants
for
PCBs,
PCDDs,
and
PCDFs.
Environ.
Sci.
Technol.
29(
6):
1591­
1598.

Leifer,
A.;
Brink,
R.
H.;
Thom,
G.
C.;
Partymiller,
K.
G.
(
1983)
Environmental
transport
and
transformation
of
polychlorinated
biphenyls.
Washington,
D.
C.:
U.
S.
Environmental
Protection
Agency,
Office
of
Toxic
Substances.
EPA­
560/
5­
83­
025.
Page
45
of
49
Lieble,
K.;
Büchen,
M.;
Ott,
W.;
Fricke,
W.
(
1993)
Polychlorinated
dibenzo(
p)
dioxins
and
dibenzofurans
in
ambient
air;
concentrations
and
deposition
measurements
in
Hessen,
Germany.
Organohalogen
Compounds
12:
85­
88.

Mackay,
D.;
Shiu,
W.
Y.;
Ma,
K.
C.
(
1992a)
Illustrated
handbook
of
physical­
chemical
properties
and
environmental
fate
for
organic
chemicals:
polynuclear
aromatic
hydrocarbons,
polychlorinated
dioxins,
and
dibenzofurans.
Chelsea,
MI:
Lewis
Publishers.

Malloy,
T.
A.;
Goldfarb,
T.
D.;
Surico,
M.
T.
J.
(
1993)
PCDDs,
PCDFs,
PCBs,
chlorophenols
(
CPs)
and
chlorobenzenes
(
CBzs)
in
samples
from
various
types
of
composting
facilities
in
the
United
States.
Chemosphere
27(
1­
3):
325­
334.

Matsumura,
F.;
Benezet,
J.
H.
(
1973)
Studies
on
the
bioaccumulation
and
microbial
degradation
of
2,3,7,8­
tetrachlorodibenzo­
p­
dioxin.
Environ.
Health
Perspect.
5:
253­
258.

McCrady,
J.
K.;
Maggard,
S.
P.
(
1993)
Uptake
and
photodegradation
of
2,3,7,8­
tetrachlorodibenzo­
pdioxin
sorbed
to
grass
foliage.
Environ.
Sci.
Technol.
27:
343­
350.

McLachlan,
M.
S.;
Welsch­
Pausch,
K.;
Tolls,
J.
(
1995)
Field
validation
of
a
model
of
the
uptake
of
gaseous
SOC
in
Lolium
multiflorum
(
rye
grass).
Environ.
Sci.
Technol.
29(
8):
1988­
2004.

McLachlan,
M.
S.;
Sewart,
A.
P.;
Bacon,
I.
R.;
Jones,
K.
C.
(
1996)
Persistence
of
PCDD/
Fs
in
a
sludge­
amended
soil.
Environ.
Sci.
Technol.
30(
8):
2567­
2571.

Mill,
T.;
Rossi,
M.;
McMillen,
D.;
Coville,
M.;
Leung,
D.;
Spang,
J.
(
1987)
Photolysis
of
tetrachlorodioxin
and
PCBs
under
atmospheric
conditions.
Internal
report
prepared
by
SRI
International
for
USEPA,
Office
of
Health
and
Environmental
Assessment,
Washington,
D.
C.

Miller,
G.
C.;
Zepp,
R.
G.
(
1987).
2,3,7,8­
Tetrachlorodibenzo­
p­
dioxin:
environmental
chemistry.
In:
Exner,
J.
H.
ed.,
Solving
hazardous
waste
problems
­­
learning
from
dioxins.
Washington,
D.
C.:
American
Chemical
Society.

Miller,
G.
C.;
Hebert,
V.
R.;
Miille,
M.
J.;
Mitzel,
R.;
Zepp,
R.
G.
(
1989)
Photolysis
of
octachlorodibenzo­
p­
dioxin
on
soils:
production
of
2,3,7,8­
TCDD.
Chemosphere
18:
1265­
1274.

Morimoto,
K.;
Kenji,
T.
(
1995)
Effect
of
humic
substances
on
the
enzymatic
formation
of
OCDD
from
PCP.
Organohalogen
Compounds
23:
387­
392.

Mousa,
M.;
Ganey,
P.
E.;
Quensen,
J.
F.;
Madhukar,
B.
V.;
Chou,
K.;
Giesy,
J.
P.;
Fischer,
L.
J.;
Boyd,
S.
A.
(
1997)
Altered
biological
activity
of
commercial
PCB
mixtures
due
to
microbial
reductive
dechlorination.
Organohalogen
Compounds.
34:
1­
5.

Muir,
D.
C.
G.;
Lawrence,
S.;
Holoka,
M;
Fairchild,
W.
L.;
Segstro,
M.
D.;
Webster,
G.
R.
B.;
Servos,
M.
R.
(
1992)
Partitioning
of
polychlorinated
dioxins
and
furans
between
water,
sediments
and
biota
in
lake
mesocosms.
Chemosphere
25(
1­
2):
199­
124.
Page
46
of
49
Nash,
R.
G.;
Beall,
M.
L.
(
1980)
Distribution
of
Silvex,
2,4­
D,
and
TCDD
applied
to
turf
in
chambers
and
field
plots.
J.
Agric.
Food
Chem.
28(
3):
614­
623.

Nicholson,
K.
W.;
Rose,
C.
L.;
Lee,
D.
S.;
Pomeroy,
I.
R.
(
1993)
Behaviour
of
polychlorinated
dibenzo­
p­
dioxins
(
PCDDs)
and
dibenzofurans
(
PCDFs)
in
the
terrestrial
environment:
a
review.
AEA
Environment
and
Energy.
Report
No.
AEA­
EE­
0519.

Oberg,
L.
G.;
Glas,
B.;
Swanson,
S.
E.;
Rappe,
C.;
Paul,
K.
G.
(
1990)
Peroxidase­
catalyzed
oxidation
of
chlorophenols
to
polychlorinated
dibenzo­
p­
dioxins
and
dibenzofurans.
Arch.
Environ.
Contam.
Toxicol.
19:
930­
938.

Oberg,
L.
G.;
Rappe,
C.
(
1992)
Biochemical
formation
of
PCDD/
Fs
from
chlorophenols.
Chemosphere
25(
1­
2):
49­
52.

Oberg,
L.
G.;
Andersson,
R.;
Rappe,
C.
(
1992)
De
novo
formation
of
hepta­
and
octachlorodibenzop
dioxins
from
pentachlorophenol
in
municipal
sewage
sludge.
Organohalogen
Compounds
9:
351­
354.

Oberg,
L.
G.;
Wagman,
N.;
Andersson,
R.;
Rappe,
C.
(
1993)
De
novo
formation
of
PCDD/
Fs
in
compost
and
sewage
sludge
­
a
status
report.
Organohalogen
Compounds
11:
297­
302.

Orazio,
C.
E.;
Kapila,
S.;
Puri,
R.
K.;
Yanders,
A.
F.
(
1992)
Persistence
of
chlorinated
dioxins
and
furans
in
the
soil
environment.
Chemosphere
25(
7­
10):
1469­
1474.

Orth,
R.
G.;
Ritchie,
C.;
Hileman,
F.
(
1989)
Measurement
of
the
photoinduced
loss
of
vapor
phase
TCDD.
Chemosphere.
18:
1275­
1282.

Palausky,
J.;
Kapila,
S.;
Manahan,
S.
E.;
Yanders,
A.
F.;
Malhotra,
R.
K.;
Clevenger,
T.
E.
(
1986)
Studies
on
vapor
phase
transport
and
role
of
dispersing
medium
on
mobility
of
2,3,7,8­
TCDD
in
soil.
Chemosphere
15:
1387­
1396.

Paustenbach,
D.
J.;
Wenning,
R.
J.;
Lau,
V.;
Harrington,
N.
W.;
Rennix,
D.
K.;
Parsons,
A.
H.
(
1992)
Recent
developments
on
the
hazards
posed
by
2,3,7,8­
tetrachlorobenzo­
p­
dioxin
in
soil:
implications
for
setting
risk­
based
cleanup
levels
at
residential
and
industrial
sites.
J.
Toxicol.
and
Environ.
Health
36:
103­
149.

Pearson,
R.
F.;
Swackhammer,
D.
L.
(
1997)
Assessing
the
importance
of
atmospheric
deposition
of
PCDD/
Fs
to
the
Great
Lakes:
compositional
comparison
of
PCDD/
F
sedimentary
accumulations.
Organohalogen
Compounds.
33:
76­
81.

Pereira,
W.
E.;
Rostad,
C.
E.;
Sisak,
M.
E.
(
1985)
Geochemical
investigations
of
polychlorinated
dibenzo­
p­
dioxins
in
the
subsurface
environment
at
an
abandoned
wood­
treatment
facility.
Environ.
Toxicol.
and
Chem.
4:
629­
639.

Podoll,
R.
T.;
Jaber,
H.
M.;
Mill,
T.
(
1986)
Tetrachlorodibenzodioxin:
Rates
of
volatilization
and
photolysis
in
the
environment.
Environ.
Sci.
Technol.
20:
490­
492.
Page
47
of
49
Puri,
R,
K.;
Clevenger,
T.
E.;
Kapila,
S.;
Yanders,
A.
F.;
Malhotra,
R.
K.
(
1989)
Studies
of
parameters
affecting
translocation
of
tetrachlorodibenzo­
p­
dioxin
in
soil.
Chemosphere.
18:
1291­
1296.

Puri,
R.
K.;
Kapila,
S.;
Lo,
Y.
H.;
Orazio,
C.;
Clevenger,
T.
E.;
Yanders,
A.
F.
(
1990)
Effect
of
cocontaminants
on
the
disposition
of
polychlorinated
dibenzofurans
in
saturated
soils.
Chemosphere
20(
10­
12):
1589­
1596.

Puri,
R.
K.;
Quiping,
Y.;
Orazio,
C.
E.;
Yanders,
A.
F.;
Kapila,
S.;
Cerlesi,
S.;
Facchetti,
S.
(
1992)
Transport
and
persistence
of
chlorinated
organics
in
varied
soil
environments.
Presented
at:
Dioxin'
92,
12th
International
Symposium
on
Chlorinated
Dioxins
and
Related
Compounds;
Tampere,
Finland;
August
1992.

Quensen,
J.
F.;
Matsumura,
F.
(
1983)
Oxidative
degradation
of
2,3,7,8­
tetrachlorodibenzo­
p­
dioxin
by
microorganisms.
Environmental
Toxicology
and
Chemistry.
2:
261­
268.

Rippen,
G.;
Wesp,
H.
(
1993)
Kale
update
of
PCDD/
PCDF,
PCB
and
PAH
under
field
conditions:
importance
of
gaseous
dry
deposition.
Organohalogen
Compounds
12:
111­
114.

Schramm,
K.
W.;
Wu,
W.
Z.;
Henkelmann,
B.;
Merk,
M.;
Xu,
Y.;
Zhang,
Y.
Y.;
Kettrup,
A.
(
1995)
Influence
of
linear
alkybenzene
sulfonate
(
LAS)
as
organic
cosolvent
on
leaching
behavior
of
PCDD/
Fs
from
fly
ash
and
soil.
Organohalogen
Compounds
24:
513­
516.

Schwarz,
K.;
McLachlan,
M.
S.
(
1993)
The
fate
of
PCDD/
F
in
sewage
sludge
applied
to
an
agricultural
field.
Organohalogen
Compounds
12:
155­
158.

Servos,
M.
R.;
Muir,
D.
C.
G.;
Webster,
G.
R.
B.
(
1992)
Environmental
fate
of
polychlorinated
dibenzo­
p­
dioxins
in
lake
enclosures.
Can.
J.
Fish.
Aquat.
Sci.
49:
722­
734.

Simonich,
S.
L.;
Hites,
R.
A.
(
1995)
Global
distribution
of
persistent
organochlorine
compounds.
Science
269:
1851­
1854.

Sivils,
L.
D.;
Kapila,
S.;
Yan,
Q.;
Elseewi,
A.
A.
(
1994)
Studies
on
gas­
phase
phototransformation
of
polychlorinated
dibenzo­
p­
dioxins.
Organohalogen
Compounds
19:
349­
353.

Sivils,
L.
D.;
Kapila,
S.;
Yan,
Q.;
Zhang,
X.;
Elseewi,
A.
A.
(
1995)
Studies
on
vapor
phase
phototransformation
of
polychlorinated
dibenzo­
p­
dioxins
(
PCDDs):
effect
of
environmental
parameters.
Organohalogen
Compounds
24:
167­
172.

Smith,
D.
G.;
Chaudhuri,
I.
S.;
Heinold,
D.;
Ruffle,
B.
(
1995)
An
alternative
approach
for
estimating
plant
uptake
of
dioxin
vapors.
Organohalogen
Compounds
24:
187­
193.

Svenson,
A.;
Kjeller,
L.
O.;
Rappe,
C.
(
1989)
Enzyme­
mediated
formation
of
2,3,7,8­
tetrasubstituted
chlorinated
dibenzodioxins
and
dibenzofurans.
Environ.
Sci.
Technol.
23(
7):
900­
902.
Page
48
of
49
Takada,
S.;
Nakamura,
M.;
Matsueda,
T.;
Kurokawa,
Y.;
Fukamati,
K.,
Kondo,
R.;
Sakai,
K.
(
1994)
Degradation
of
PCDDs/
PCDFs
by
ligninolytic
fungus
Phanerochaete
sordida
YK­
624.
Organohalogen
Compounds.
20:
195­
198.

Takada,
S.;
Nakamura,
M.;
Matsueda,
T.;
Kondo,
R.;
Sakai,
K.
(
1996)
Degradation
of
polychlorinated
dibenzo­
p­
dioxins
and
polychlorinated
dibenzofurans
by
the
white
rot
fungus
Phanerochaete
sordida
YK­
624.
Applied
and
Environmental
Microbiology
62(
12):
4323­
4328.

Tien,
M.;
Kirk,
T.
K.
(
1983)
Lignin­
degrading
enzyme
from
the
Hymenomycete
Phanerochaete
chrysosporium
Burds.
Science
221:
661­
663.

Tien,
M.;
Kirk,
T.
K.
(
1984)
Lignin­
degrading
enzyme
from
Phanerochaete
chrysosporium:
purification,
characterization,
and
catalytic
properties
of
a
unique
H
2
O
2­
requiring
oxygenase.
Proc.
Natl.
Acad.
Sci.
USA
81:
2280­
2284.

Tysklind,
M.;
Rappe,
C.
(
1991)
Photolytic
transformation
of
polychlorinated
dioxins
and
dibenzofurans
in
fly
ash.
Chemosphere
23(
8­
10):
1365­
1375.

U.
S.
Environmental
Protection
Agency.
(
1993)
Interim
report
on
data
and
methods
for
assessment
of
2,3,7,8­
tetrachlorodibenzo­
p­
dioxin
risks
to
aquatic
life
and
associated
wildlife.
Duluth,
MN:
U.
S.
Environmental
Protection
Agency,
Office
of
Research
and
Development.
EPA/
600/
R­
93­
055.

U.
S.
Environmental
Protection
Agency
(
1995)
Great
Lakes
Water
Quality
Initiative
technical
support
document
for
the
procedure
to
determine
bioaccumulation
factors.
Washington,
D.
C.:
U.
S.
Environmental
Protection
Agency,
Office
of
Water.
EPA­
820­
B­
95­
005.

Wagenaar,
W.
J.
Boelhouwers,
E.
J.;
de
Kok,
H.
A.
M.;
Groen,
C.
D.;
Govers,
H.
A.
J.;
Olie,
K.;
de
Gerlache,
J.;
Rooij,
C.
(
1995)
A
comparative
study
of
the
photolytic
degradation
of
octachlorodibenzofuran
(
OCDF)
and
octachlorodibenzo­
p­
dioxin
(
OCDD).
Chemosphere
31(
4):
2983­
2992.

Wagner,
H.
C.;
Schramm,
K.
W.;
Hutzinger,
O.
(
1990)
Biogenic
polychlorinated
dioxin
and
furan
from
trichlorophenol.
Organohalogen
Compounds
3:
453­
456.

Wallenhorst,
T.;
Krauss,
P.;
Hagenmaier,
H.
(
1995)
PCDD/
F
in
ambient
air
and
deposition
in
Baden­
Wurttenburg,
Germany.
Organohalogen
Compounds
24:
157­
161.

Walters,
R.
W.;
Guiseppi­
Elie,
A.
(
1988)
Sorption
of
2,3,7,8­
tetrachlorobenzo­
p­
dioxin
to
soils
from
water/
methanol
mixtures.
Environ.
Sci.
Technol.
22(
7):
819­
825.

Weber,
H.;
Disse,
G.;
Hamann,
R.;
Haupt,
H.
J.
(
1995)
Influence
of
the
aerobic
and
anaerobic
digestion
on
the
levels
of
chlorinated
dibenzofurans
and
dibenzodioxins
in
sewage
sludge.
Organohalogen
Compounds
24:
81­
85.
Page
49
of
49
Weinberg
Group,
Inc.
(
1998)
Volatilization
of
microcontaminants
from
pentachlorophenol­
treated
utility
poles.
Prepared
for
Penta
Task
Force
by
the
Weinberg
Group,
Inc.
May
28,
1998.

Welsch­
Pausch,
K.;
McLachlan,
M.
S.
(
1995)
Photodegradation
of
PCDD/
Fs
on
pasture
grass.
Organohalogen
Compounds
24:
509­
512.

Welsch­
Pausch,
K.;
McLachlan,
M.
S.;
Umlauf,
G.
(
1995)
Determination
of
the
principal
pathways
of
polychlorinated
dibenzo­
p­
dioxins
and
dibenzofurans
to
Lolium
multiflorum
(
Welsh
Ray
Grass).
Environ.
Sci.
Technol.
29:
1090­
1098.

Winters,
D.
L.;
Roberts,
G.
B.;
Boggess,
K.
E.;
Davis,
M.;
Alburty,
D.
S.;
Lorber,
M.
N.
(
1999)
A
field
study
to
evaluate
the
potential
for
the
release
of
dioxins
from
pentachlorophenol­
treated
utility
poles.
Organohalogen
Compounds.
41:
35­
39.

Yanders,
A.
F.;
Orazio,
C.
E.;
Puri,
R.
K.;
Kapila,
S.
(
1989)
On
translocation
of
2,3,7,8­
tetrachlorodibenzo­
p­
dioxin:
time
dependent
analysis
at
the
Times
Beach
experimental
site.
Chemosphere
19(
1­
6):
429­
432.

Young,
A.
L.
(
1983)
Long­
term
studies
on
the
persistence
and
movement
of
TCDD
in
a
natural
ecosystem.
In:
Human
and
environmental
risks
of
chlorinated
dibenzodioxins
and
related
compounds.
Tucker,
R.
E.;
Young,
A.
L.;
Gray,
A.
P.,
Eds.
Plenum
Press.