Document ID: EPA-HQ-OW-2002-0033-0338
Agency: epa
Document Type: Supporting & Related Material
Title: 
Posted Date: 2003-04-14T04:00Z

United
Stater
Office
of
Water
Environmental
Protection
Regulatmr
and
Standards
Agency
Warhmgton.
DC
20460
May
1,
1986
e,
Ep',
Water
EPA
440/
5­
86­
001
QUALlTY
CRITERIA
for
W
A
m
1986
­
TO
INTERESTED
PARTIES
Section
304(
a)(
l)
of
the
Clean
Water
Act
(
33
U.
S.
C.

1314(
a)
(
1)
requires
the
Environmental
Protection
Agency
(
EPA)
to
pub1
ish
and
periodically
update
ambient
water
qua1
ity
criteria.

These
criteria
are
to
accurately
reflect
the
latest
scientific
knowledge
(
a)
on
the
kind
and
extent
of
all
identifiable
effects
on
health
and
welfare
including,
but
not
limited
to,
plankton,

fish
shellfish,
wildlife,
plant
life,
shorelines,
beaches,

aesthetics,
and
recreation
which
may
be
expected
from
the
presence
of
pollutants
in
any
body
of
water
including
ground
water;
(
b)
on
the
concentration
and
dispersal
of
pollutants,
or
their
byproducts,
through
biological,
physical,
and
chemical
processes;
and
(
c)
on
the
effects
of
pollutants
on
biological
community
diversity,
productivity,
and
stability,
including
information
on
the
factors
affecting
rates
of
eutrophication
and
organic
and
inorganic
sedimentation
for
varying
types
of
receiving
waters.
These
criteria
are
not
rules
and
they
do
not
have
regulatory
impact.
Rather,
these
criteria
present
scientific
data
and
guidance
of
the
environmental
effects
of
pollutants
which
can
be
useful
to
derive
regulatory
requirements
based
on
considerations
of
water
quality
impacts.
When
additional
data
has
become
available,
these
summaries
have
been
updated
to
reflect
the
latest
Agency
recommendations
on
acceptable
limits
for
aquatic
life
and
human
health
protection.

Periodically
EPA
and
its
predecessor
agencies
has
issued
ambient
water
quality
criteria,
beginning
in
1968
with
the
"
Green
Book"
followed
by
the
1973
publication
of
the
"
Blue
Book"
(
Water
Quality
Criteria
1972).
In
1976,
the
"
Red
Book"
(
Quality
For
aalc
by
the
Svpmlendsnt
of
Documents,
US
Mwmmsnt
RlnUng
ORW
Waahpton.
DC
M02
Criteria
for
Water)
was
published.
On
November
28,
1980
(
45
FR
79318),
and
February
15,
1984
(
49
FR
5831),
EPA
announced
through
Federal
­
Register
notices,
the
publication
of
65
individual
ambient
water
quality
criteria
documents
 or
pollutants
listed
as
toxic
under
section
307(
a)(
l)
of
the
Clean
Water
Act.
on
July
29,
1985
(
50
FR
30784),
EPA
published
additional
water
quality
criteria
documents.

The
development
and
publication
of
ambient
water
quality
criteria
has
been
pursued
over
the
past
10
years
and
is
an
ongoing
process.
EPA
expects
to
publish
about
10
final
criteria
documents
each
year.
Some
of
these
will
update
and
revise
existing
criteria
recommendations
and
others
will
be
issued
for
the
first
time.

In
a
continuing
effort
to
provide
those
who
use
EPA'S
water
quality
and
human
health
criteria
with
up­
to­
date
criteria
values
and
associated
information,
this
document
­­
Q
u
a
l
m
Criteria
__
 or
Water
­
1986
was
assembled.
This
document
includes
summaries
of
all
the
contaminants
for
which
EPA
has
developed
criteria
recom­

mendations
(
Appendix
A­
C)
.
The
appropriate
appendix
is
identified
at
the
end
of
each
summary.
A
more
detailed
description
of
these
procedures
can
be
found
in
the
appropriate
Appendix.
Copies
of
this
document
can
be
obtained
by
contacting
the
U.
S.
Government
Printing
Office
at:

U.
S.
Government
Printing
Office
Superintendent
of
Documents
N.
Capitol
and
H
Street
N.
W.
Washington,
D.
C.
20401
A
fee
is
charged
f
o
r
this
document.

Copies
of
the
complete
background
ambient
water
quality
criteria
documents
containing
all
the
data
used
to
develop
the
criteria
recommendations
summarized
herein
and
the
"
Red
Bookt8,

including
complete
bibliographies
are
available
only
from:
0
National
Technical
Information
Service
5285
Port
Royal
Road
Springfield,
VA
22161
Telephone:
(
703)
487­
4650
The
NTIS
order
numbers
for
the
criteria
documents
can
be
found
in
the
Index.
A
fee
is
charged
for
copies
of
these
documents.

As
new
criteria
are
developed
and
existing
criteria
revised,

updated
criteria
summaries
will
be
made
available
once
a
year
to
those
who
purchase
this
document
through
the
U.
S,
Government
Printing
office.
You
will
automatically
be
placed
on
the
mailing
list
to
receive
annual
updates.
The
cost
for
receiving
annual
updates
is
included
in
the
purchase
price
of
the
document.

­­
Quality
Criteria
­
f
o
r
Water,
1986
is
designed
to
be
easily
updated
to
reflect
EPA`
s
continuing
work
to
present
the
latest
scientific
information
and
practices,
Our
planned
schedule
 or
future
criteria
development
in
the
next
few
years
is
attached
for
your
information.
a
The
Agency
is
current1
y
developing
Acceptable
Daily
Intake
(
ADI)
or
Verified
Reference
Dose
(
RfD)
values
on
a
number
of
chemicals
for
Agency­
wide
use.
Based
upon
this
new
analysis
the
values
have
changed
significantly
for
5
chemicals
from
those
used
in
the
original
human
health
criteria
calculation
done
in
1980.

The
chemicals
affected
are
as
follows:

0
chemical
.
1980
WQC
1.
cyanide
200
ug/
L
2.
Ethylbenzene
1.4
mg/
L
4.
Phenol
3.5
mg/
L
3
.
Nitrobenzene
19.8
mg/
L
5.
Toluene
14.3
mg/
L
Draft
RfD
.02
mg/
kg/
day
.
Q1
mg/
kg/
day
.
Q005
mg/
kg/
day
0.1
mg/
kg/
day
0.3
mg/
kg/
day
FOR
FORTHER
INFORMATION
CONTACT:

Dr.
Frank
Gostomski
at
the
above
address
or
by
phoning
(
202)
245­

3030.

It
is
EPA's
goal
to
continue
to
develop
and
make
available
ambient
water
quality
criteria
reflecting
the
latest
scientific
practices
and
information.
In
this
way
we
can
continue
to
improve
and
protect
the
quality
of
the
Nation's
waters.

James
M.
Conlon
i/
and
Standards
DRAFT
CRITERIA
DOCUMENTS
TO
BE
PROPOSED
LATE
FY
86/
EARLY
­­
FY
87
­­
0
Diethyhexylphthalate
1
,
2
,
4
,
Trichlorobenzene
Silver
Phenanthrene
2
,
4
,
5
,
Trichlorophenol
Organotins
Tributyltin
Selenium
(
no
saltwater
criteria)
Hexachlorobenzene
Antimony
111
Acrolein
(
no
saltwater
criteria)

­­
LATE
FY
87/
EARLY
Thallium
(
no
saltwater
criteria)
Tetrachloroethylene
(
no
saltwater
criteria)
Phenol
Toluene
Chloroform
(
no
saltwater
criteria)
'
imaline
Acrvlontrile
w
Hexachlorocyclopentadiene
(
no
saltwater
criteria)
Dimethylphenol
Hexachlorobutadiene
(
no
saltwater
criteria)

­
Both
lists
will
incorporate
aquatic
and
human
health
values.

­
All
above
are
toxic
pollutants
except
f
o
r
organotins
and
analine
which
are
non­
conventionals.
INDEX
INTRODUCTION
SUMMARY
CHART
.,'
Acenaphthene
Acrolein
Acrylonitrile
Aesthetics
Alkalinity
Aldrin/
Dieldrin
Ammonia
Antimony
Arsenic
Asbestos
Bacteria
Barium
Benzene
Benzidine
Beryllium
Boron
Cadmium
Carbon
Tetrachloride
Chlordane
Chlorinated
Benzenes
Chlorinated
Ethanes
Chlorinated
Naphthalenes
Chlorine
Chlorinated
Phenols
Chloroalkyl
Ethers
Chlorof
o
m
Chlorophenoxy
Herbicides
Chromium
2­
Chlorophenol
Color
Copper
Cyanide
DDT
and
Metabolites
Demeton
Dichlorobenzenes
Dichlorobenzidine
Dichloroethylenes
2,4,
­
Dichlorophenol
Dichloropropanes/
Dichloropropenes
2,4,
­
Dimethylphenol
Dinitrotoluene
Diphenylhydrazine
Endosulfan
Endrin
Ethylbenzene
Fluoranthene
Gasses,
Total
Dissolved
Guthion
PB­
263943
PB
81­
117293
PB
81­
117343
PB
81­
117350
PB­
2
6394
3
PB
85­
227031
PB
81­
117376
PB
81­
117384
PB
81­
117392
PB
81­
117400
PB
81­
117426
PB
85­
227429
PB
81­
117434
PB
81­
117418
PB
81­
117442
PB­
2
63
94
3
PB
85­
227478
PB
81­
117459
PB­
2
63
94
3
PB
85­
227023
PB
85­
227460
PB
81­
117491
PB­
263943
PB
81­
117509
PB
81­
117517
PB
81­
117525
PB
81­
117533
PB
81­
117541
PB
81­
117558
PB
81­
117566
PB
81­
117731
PB
81­
117582
PB
81­
117590
PB
81­
117608
PB­
2
63
943
PJ3­
2
639
4
3
PB
81­
117574
NTIS
No.
­­

PB
81­
117269
PB
81­
117277
PB
81­
117285
FB­
2
63
94
3
PB­
2
63
94
3
PB
81­
117301
PB
85­
227114
PB
81­
117319
PB
85­
227445
PB
81­
117335
PB
86­
158­
045
h
PB­
263943
Haloethers
Halomethanes
Hardness
Heptachlor
Hexachlorobutadiene
Hexachlorocyclohexane
Hexachlorocyclopentadiene
Iron
Isophorone
Lead
Malathion
Manganese
Mercury
Methoxychlor
Mirex
Naphthalene
Nickel
Nitrates,
Nitrites
Nitrobenzene
Nitrophenols
Nitrosamines
Oil
and
Grease
Oxygen,
Dissolved
Parathion
Pentachlorophenol
Ph
Phenol
Phosphorus
Phthalate
Esters
Polychlorinated
Biphenyls
Polynuclear
Aromatic
Hydrocarbons
Selenium
Silver
Solids
(
Dissolved)
h
Salinity
Solids
(
Suspended)
&
Turbidity
Sulfides,
Hydrogen
Sulfide
Taintina
Substances
PB
81­
117616
PB
81­
117624
PB­
263943
PB
81­
117632
PE
81­
117640
PB
81­
117657
PB
81­
117665
PB­
263943
PB
81­
117673
PB
85­
227437
PB­
263943
PB­
263943
PB
85­
227452
PB­
2
63
943
PB­
263943
PB
81­
117707
PB
81­
117715
PB­
263943
PB
81­
117723
PB
81­
117749
PB
81­
117756
PB­
263943
PB
86­
208253
PB­
263943
PB
81­
117764
PB­
263943
PB
81­
117772
PB­
263943
PB
81­
117780
PB
81­
117798
PB
81­
117806
PB
81­
117814
PB
81­
117822
PB­
263943
PB­
263943
PB­
263943
PB­
263943
Temperagure
PB­
263943
Tetrachloroethylene
PB
81­
117830
Thallium
PB
81­
117848
Toluene
PB
81­
117855
Toxaphene
PB
81­
117863
Trichloroethylene
PB
81­
117871
Vinyl
Chloride
PB
81­
117889
Zinc
PB
81­
117897
2,3,7,8­
Tetrachlorodibenzo­
p­
dioxin
EPA
#
440/
5­
84­
007
APPENDIX
A
APPENDIX
B
APPENDIX
C
BIBLIOGRAPHY
Methodology
for
Developing
Criteria
Methodology
for
Developing
Criteria
Methodology
for
Developing
Criteria
I
SUMMARY
CHART
1
0
CRITERIA:
ACENAPHTHENE
Aquatic
Life
The
available
data
for
acenaphthene
indicate
that
acute
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
1,700
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
acenaphthene
to
sensitive
freshwater
aquatic
animals
but
toxicity
to
freshwater
algae
occur
at
concentrations
as
low
as
520
ug/
L.

The
available
data
for
acenaphthene
indicate
that
acute
and
chronic
toxicity
to
saltwater
aquatic
life
occurs
at
concentrations
as
low
as
970
and
710
ug/
L,
respectively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
Toxicity
to
algae
occurs
at
concentrations
as
low
as
500
ug/
L.

Human
Health
Sufficient
data
are
not
available
for
acenaphthene
to
derive
a
level
which
would
protect
against
the
potential
toxicity
of
this
compound.
Using
available
organoleptic
data,
to
control
undesirable
taste
and
odor
quality
of
ambient
water
the
estimated
level
is
0.02
mg/
L.
It
should
be
recognized
that
organoleptic
data,
have
limitations
as
a
basis
for
establishing
water
quality
criteria,
and
have
no
demonstrated
relationship
to
potential
adverse
human
health
effects.

0
(
45
F
.
R
.
79318,
November
28,
1980)
,
SEE
APPENDIX
B
FOR
METHODOLOGY
ACROLEIN
Aquatic
Life
The
available
data
for
acrolein
indicate
that
acute
and
chronic
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
68
and
21
ug/
L,
respectively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.

The
available
data
for
acrolein
indicate
that
acute
toxicity
to
saltwater
aquatic
life
occurs
at
concentrations
as
low
as
55
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
acrolein
to
sensitive
saltwater
aquatic
life.

0
Human
Health
For
the
protection
of
human
health
from
the
toxic
properties
of
acrolein
ingested
through
contamins'ed
aquatic
organisms,
the
ambient
water
criterion
is
determined
o
be
320
ug/
L.

For
the
protection
of
human
health
from
the
toxic
properties
of
acrolein
ingested
through
contaminated
aquatic
organisms
alone,
the
ambient
water
criterion
is
determined
to
be
780
ug/
L*

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
0
,
CRITERIA:
ACRYMNITRILE
Aquatic
Life
The
available,
data
for
acrylonitrile
indicate
that
acute
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
7,550
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
definitive
data
are
available
concerning
the
chronic
toxicity
of
acrylonitrile
to
sensitive
freshwater
aquatic
life
but
mortality
occurs
at
concentrations
as
low
as
2,600
ug/
L
with
a
fish
species
exposed
for
30
days.

Only
one
saltwater
species
has
been
tested
with
acrylonitrile
and
no
statement
can
be
made
concerning
acute
or
chronic
Human
Health
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
resulting
from
exposure
to
acrylonitrile
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentrations
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
and
lo­'.

The
corresponding
recommended
criteria
are
0.58
ug/
L,
0.058
ug/
L,
and
0.006
ug/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
0
,
of
water,
the
levels
are
6.5
ug/
L,
0.65
ug/
L,
and
0.065
ug/
L,

respectively.

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
AESTHETIC
QUALITIES
CRITERIA:

a
A
l
l
waters
free
from
substances
attributable
to
wastewater
or
other
discharges
that:

settle
to
form
objectionable
deposits;

float
as
debris,
scum,
oil,
or
other
matter
to
form
nuisances;

produce
objectionable
color,
odor,
taste,

or
turbidity;

injure
or
are
toxic
or
produce
adverse
physiological
responses
in
humans,

animals
or
plants:
and,

produce
undesirable
or
nuisance
aquatic
life.

RATIONALE:

Aesthetic
qualities
of
water
address
the
general
principles
laid
down
in
common
law.
They
embody
the
beauty
and
quality
of
water
and
their
concepts
may
vary
within
the
minds
of
individuals
encountering
the
waterway.
A
rationale
for
these
qualities
cannot
be
developed
with
quantifying
definitions;
however,

decisions
concerning
such
quality
factors
can
portray
the
best
in
the
public
interest.

_.,
Aesthetic
q
u
a
l
i
t
i
e
s
provide
t
h
e
general
rules
t
o
protect
water
against
environmental
insults:
they
provide
minimal
freedom
requirements
from
p
o
l
l
u
t
i
o
n
;
they
are
e
s
s
e
n
t
i
a
l
properties
to
protect
the
Nation's
waterways.

(
QUALITY
CRITERIA
FOR
WATER,
JULY
1976)
PB­
263943
SEE
APPENDIX
C
FOR
METHODOLOGY
ALKALINITY
CRITERION:

20
mg/
L
or
more
as
CaC03
freshwater
aquatic
life
except
where
natural
concentrations
are
less.

INTRODUCTION:

Alkalinity
is
the
sum
total
of
components
in
the
water
that
tend
to
elevate
the
pH
of
the
water
above
a
value
of
about
4.5.

It
is
measured
by
titration
with
standardized
acid
to
a
pH
value
of
about
4.5
and
it
is
expressed
commonly
as
milligrams
per
liter
of
calcium
carbonate.
Alkalinity,
therefore,
is
a
measure
of
the
buffering
capacity
of
the
water,
and
since
pH
has
a
direct
effect
on
organisms
as
well
as
an
indirect
effect
on
the
toxicity
of
certain
other
pollutants
in
the
water,
the
buffering
capacity
is
important
to
water
quality.
Examples
of
commonly
occurring
materials
in
natural
waters
that
increase
the
alkalinity
are
carbonates,
bicarbonates,
phosphates
and
hydroxides.

RATIONALE
:

The
alkalinity
of
water
used
for
municipal
water
supplies
is
important
because
it
affects
the
amounts
of
chemicals
that
need
to
be
added
to
accomplish
calculation,
softening
and
control
of
corrosion
in
distribution
systems.
The
alkalinity
of
water
assists
in
the
neutralization
of
excess
acid
produced
during
the
addition
of
such
materials
as
aluminum
sulfate
during
chemical
coagulation.
Waters
having
sufficient
alkalinity
do
not
have
to
be
supplemented
with
artificially
added
materials
to
increase
the
i
alkalinity.
Alkalinity
resulting
from
naturally
occurring
a
materials
such
a
s
carbonate
and
bicarbonate
is
not
considered
a
h
e
a
l
t
h
hazard
i
n
drinking
water
supplies,
per
se,
and
n
a
t
u
r
a
l
l
y
occurring
maximum
l
e
v
e
l
s
up
t
o
approximately
400
mg/
L
as
calcium
carbonate
are
n
o
t
considered
a
problem
t
o
human
h
e
a
l
t
h
(
NAS,

1974).

A
l
k
a
l
i
n
i
t
y
is
important
f
o
r
f
i
s
h
and
o
t
h
e
r
a
q
u
a
t
i
c
l
i
f
e
i
n
freshwater
systems
because
it
b
u
f
f
e
r
s
pH
changes
t
h
a
t
occur
­
n
a
t
u
r
a
l
l
y
a
s
a
r
e
s
u
l
t
o
f
p
h
o
t
o
s
y
n
t
h
e
t
i
c
a
c
t
i
v
i
t
y
of
t
h
e
chlorophyll­
bearing
vegetation.
Components
of
a
l
k
a
l
i
n
i
t
y
such
as
carbonate
and
biocarbonate
w
i
l
l
complex
some
t
o
x
i
c
heavy
metals
and
reduce
their
t
o
x
i
c
i
t
y
markedly.
For
these
reasons,
t
h
e
National
Technical
Advisory
Committee
(
NATC,
1968)
recommended
a
minimum
a
l
k
a
l
i
n
i
t
y
of
20
mg/
L
'
and
t
h
e
subsequent
NAS
Report
(
1974)
recommended
t
h
a
t
natural
a
l
k
a
l
i
n
i
t
y
not
be
reduced
by
more
than
25
percent
but
did
not
place
an
absolute
minimal
value
f
o
r
it.
T
h
e
u
s
e
of
t
h
e
2
5
p
r
e
s
e
n
t
r
e
d
u
c
t
i
o
n
a
v
o
i
d
s
t
h
e
problem
of
establishing
standards
on
waters
where
n
a
t
u
r
a
l
a
l
k
a
l
i
n
i
t
y
is
a
t
o
r
below
2
0
mg/
L.
For
such
waters,
a
l
k
a
l
i
n
i
t
y
should
n
o
t
be
f
u
r
t
h
e
r
reduced.

The
NAS
Report
recommends
t
h
a
t
adequate
amounts
of
a
l
k
a
l
i
n
i
t
y
be
maintained
t
o
buffer
the
pH
within
t
o
l
e
r
a
b
l
e
l
i
m
i
t
s
f
o
r
marine
waters.
It
has
been
noted
a
s
a
c
o
r
r
e
l
a
t
i
o
n
t
h
a
t
productive
waterfowl
h
a
b
i
t
a
t
s
are
above
2
5
mg/
L
w
i
t
h
h
i
g
h
e
r
a
l
k
a
l
i
n
i
t
i
e
s
r
e
s
u
l
t
i
n
g
i
n
better
waterfowl
h
a
b
i
t
a
t
s
(
NATC,
1968).
Excessive
alkalinity
can
cause
problems
for
swimmers
by
altering
the
pH
of
the
lacrimal
fluid
around
the
eye,
causing
irritation.
0
For
industrial
water
supplies,
high
alkalinity
can
be
damaging
to
industries
involved
in
food
production,
especially
those
in
which
acidity
accounts
for
flavor
and
stability,
such
as
the
carbonated
beverages.
In
other
instances,
alkalinity
is
desirable
because
water
with
a
high
alkalinity
is
much
less
corrosive.

A
brief
summary
of
maximum
alkalinities
accepted
as
a
source
of
raw
water
by
industry
is
included
in
Table
1.
The
concentrations
listed
in
the
table
are
for
water
prior
to
treatment
and
thus
are
only
desirable
ranges
and
not
critical
ranges
for
industrial
use.
0
The
effect
of
alkalinity
in
water
used
for
irrigation
may
be
important
in
some
instances
because
it
may
indirectly
increase
the
relative
proportion
of
sodium
in
s
o
i
l
water.
As
an
example,

when
bicarbonate
concentrations
are
high,
calcium
and
magnesium
ions
that
are
in
solution
precipitate
as
carbonates
in
the
soil
water
gs
the
water
becomes
more
concentrated
through
evaporation
and
transpiration.
A
s
the
calcium
and
magnesium
ions
decrease
in
concentration,
the
percentage
of
sodium
increases
and
results
in
soil
and
plant
damage.
Alkalinity
may
also
lead
to
chlorosis
in
plants
because
it
causes
the
iron
to
precipitate
as
a
hydroxide
(
NAS,
1974).
Hydroxyl
ions
react
with
available
iron
in
the
soil
TABLE
I*

Maximum
Alkalinity
In
Waters
Used
As
A
Source
Of
Supply
Prior
To
Treatment
Industry
Alkalinity
mg/
L
as
CaC03
­

Steam
generation
boiler
makeup.......
.....
350
Steam
generation
cooling
..................
500
Textile
mill
products
.....................
50­
200
Paper
and
allied
products
.................
75­
150
Chemical
and
Allied
Products..............
500
Petroleum
refining
........................
'
500
Primary
metals
industries.................
200
Food
canning
industries.........
..........
300
1
Bottled
and
canned
soft
drinks............

*
NAS,
1974
water
and
make
the
iron
unavailable
to
.
plants.
Such
deficiencies
induce
chlorosis
and
further
plant
damage.
Usually
alkalinity
must
exceed
6
mg/
L
before
such
effects
are
noticed,
however.

(
QUALITY
CRITERIA
FOR
WATER,
JULY
1976)
PB­
263943
SEE
APPENDIX
C
FOR
METHODOLOGY
*
ALDRIN­
DIELDRIN
Aquatic
Life
Dieldrin
For
dieldrin
the
criterion
to
protect
freshwater
aquatic
life
as
derived
using
the
Guidelines
is
0.0019
ug/
L
as
a
24­
hour
average,
and
the
concentration
should
not
exceed
1.0
ug/
L
at
any
time.

For
dieldrin
the
criterion
to
protect
saltwater
aquatic
life
as
derived
using
the
Guidelines
is
0.0019
ug/
L
as
a
24­
hour
average,
and
the
concentration
should
not
exceed
0.71
ug/
L
at
any
time
­
Aldrin
For
freshwater
aquatic
life
the
concentration
of
aldrin
should
not
exceed
4.0
ug/
L
at
any
time.
No
data
are
available
concerning
the
chronic
toxicity
of
aldrin
to
sensitive
freshwater
aquatic
life.

For
saltwater
aquatic
life
the
concentration
of
aldrin
should
not
exceed
1.3
ug/
L
at
any
time.
No
data
are
available
concerning
the
chronic
toxicity
of
aldrin
to
sensitive
saltwater
aquatic
life.

Human
Health
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
aldrin
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
I
*
Indicates
suspended,
canceled
or
restricted
by
U.
S.
EPA
Office
of
Pesticides
and
Toxic
Substances
.
a
ambient
water
concentration
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,

the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
and
The
corresponding
recommended
criteria
are
0.74
ng/
L,
0.074
ng/
L,

and
0.0074
ng/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
0.79
ng/
L,
0.079
ng/
L,
and
0.0079
ng/
L,
respectively.

FOP
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
dieldrin
through
ingestion
of
contaminated
water
and
Contaminated
aquatic
organisms,
the
ambient
water
concentration
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
,
lifetime
are
estimated
at
and
The
corresponding
recommended
criteria
are
0.71
ng/
L,
0.071
ng/
L,
and
0.0071
ng/
L,
respectively.
If
these
above
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
0.76
ng/
L,
0.076
ng/
L,
and
0.0076
ng/
L,

respectively.
0
(
45
F
.
R
.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
AMMONIA
SUMMARY
:

All
concentrations
used
herein
are
expressed
as
un­
ionized
ammonia
(
NH3),
because
NH3,
not
the
ammonium
ion
(
NH4+)
has
been
demonstrated
to
be
the
principal
toxic
form
of
ammonia.
The
data
used
in
deriving
criteria
are
predominantly
from
flow
through
tests
in
which
ammonia
concentrations
were
measured.

Ammonia
was
reported
to
be
acutely
toxic
to
freshwater
organisms
at
concentrations
(
uncorrected
 or
pH)
ranging
from
0.53
to
22.8
mg/
L
NH3
for
19
invertebrate
species
representing
14
families
and
16
genera
and
from
0.083
to
4.60
mg/
L
NH3
 or
29
fish
species
from
9
families
and
18
genera.
Among
fish
species,
reported
96­

hour
LC50
ranged
from
0.083
to
1.09
mg/
L
for
salmonids
and
from
0.14
to
4.60
mg/
L
NH3
for
nonsalmonids.
Reported
data
from
chronic
tests
on
ammonia
with
t
w
o
freshwater
invertebrate
species,
both
daphnids,
showed
effects
at
concentrations
(
uncorrected
for
pH)
ranging
from
0.304
to
1.2
mg/
L
NH3,
and
with
nine
freshwater
fish
species,
from
five
families
and
seven
genera,
ranging
from
0.0017
to
0.612
mg/
L
NH3.
I
Concentrations
of
ammonia
acutely
toxic
to
fishes
may
cause
loss
of
equilibrium,
hyperexcitability,
increased
breathing,

cardiac
output
and
oxygen
uptake,
and,
in
extreme
cases,

convulsions,
coma,
and
death.
At
lower
concentrations
ammonia
has
many
effects
on
fishes,
including
a
reduction
in
hatching
success,
reduction
in
growth
rate
and
morphological
development,

and
pathologic
changes
in
tissues
of
gills,
livers,
and
kidneys.
Several
factors
have
been
shown
to
modify
acute
NH3
toxicity
in
fresh
water.
Some
factors
alter
the
concentration
of
un­

ionized
ammonia
in
the
water
by
affecting
the
aqueous
ammonia
equilibrium,
and
some
factors
affect
the
toxicity
of
un­
ionized
ammonia
itself,
either
ameliorating
or
exacerbating
the
effects
of
ammonia.
Factors
that
have
been
shown
to
affect
ammonia
toxicity
include
dissolved
oxygen
concentration,
temperature,

pH,
previous
acclimation
to
ammonia,
fluctuating
or
intermittent
exposures,
carbon
dioxide
concentration,
salinity,
and
the
presence
of
other
toxicants.

The
most
well­
studied
of
these
is
pH:
the
acute
toxicity
of
NH3
has
been
shown
to
increase
as
pH
decreases.
Sufficient
data
exist
from
toxicity
tests
conducted
at
different
pH
values
to
formulate
a
mathematical
expression
to
describe
pH­
dependent
acute
NH3
toxicity.
The
very
limited
amount
o
f
data
regarding
effects
of
pH
on
chronic
NH3
toxicity
also
indicates
increasing
NH3
toxicity
with
decreasing
pH,
but
the
data
are
insufficient
to
derive
a
broadly
applicable
toxicity/
pH
relationship.
Data
on
temperature
effects
on
acute
NH3
toxicity
are
limited
and
somewhat
variable,
but
indications
are
that
NH3
toxicity
to
fish
is
greater
as
temperature
decreases.
There
is
no
information
available
regarding
temperature
effects
on
chronic
NH3
toxicity.

Examination
of
pH
and
temperature­
corrected
acute
NH3
toxicity
values
among
species
and
genera
of
freshwater
organisms
showed
that
invertebrates
are
generally
more
tolerant
than
fishes,
a
notable
exception
being
the
fingernail
clam.
There
is
no
clear
trend
among
groups
of
fish;
the
several
most
sensitive
a
tested
s
p
e
c
i
e
s
and
genera
include
r
e
p
r
e
s
e
n
t
a
t
i
v
e
s
from
diverse
f
a
m
i
l
i
e
s
(
Salmonidae,
Cyprinidae,
Percidae,
and
Centrarchidae).

A
v
a
i
l
a
b
l
e
c
h
r
o
n
i
c
t
o
x
i
c
i
t
y
data
f
o
r
f
r
e
s
h
w
a
t
e
r
organisms
a
l
s
o
i
n
d
i
c
a
t
e
i
n
v
e
r
t
e
b
r
a
t
e
s
(
c
l
a
d
o
c
e
r
a
n
s
,
one
i
n
s
e
c
t
s
p
e
c
i
e
s
)
t
o
be
more
t
o
l
e
r
a
n
t
t
h
a
n
f
i
s
h
e
s
,
a
g
a
i
n
w
i
t
h
t
h
e
e
x
c
e
p
t
i
o
n
o
f
t
h
e
f
i
n
g
e
r
n
a
i
l
clam.
When
c
o
r
r
e
c
t
e
d
f
o
r
t
h
e
presumed
effects
of
t
e
m
p
e
r
a
t
u
r
e
andpH,
there
is
a
l
s
o
no
clear
t
r
e
n
d
among
groups
of
f
i
s
h
f
o
r
c
h
r
o
n
i
c
t
o
x
i
c
i
t
y
v
a
l
u
e
s
,
t
h
e
most
s
e
n
s
i
t
i
v
e
s
p
e
c
i
e
s
i
n
c
l
u
d
i
n
g
r
e
p
r
e
s
e
n
t
a
t
i
v
e
s
from
f
i
v
e
f
a
m
i
l
i
e
s
(
Salmonidae,

Cyprinidae,
I
c
t
a
l
u
r
i
d
a
e
,
C
e
n
t
r
a
r
c
h
i
d
a
e
,
and
Catostomidae)
and
h
a
v
i
n
g
c
h
r
o
n
i
c
v
a
l
u
e
s
r
a
n
g
i
n
g
by
n
o
t
much
more
t
h
a
n
a
f
a
c
t
o
r
o
r
two.
The
range
of
a
c
u
t
e­
c
h
r
o
n
i
c
r
a
t
i
o
s
f
o
r
1
0
s
p
e
c
i
e
s
from
6
0
f
a
m
i
l
i
e
s
was
3
t
o
43,
and
acute­
khronic
r
a
t
i
o
s
were
h
i
g
h
e
r
f
o
r
t
h
e
s
p
e
c
i
e
s
h
a
v
i
n
g
c
h
r
o
n
i
c
t
o
l
e
r
a
n
c
e
b
e
l
o
w
t
h
e
median.

A
v
a
i
l
a
b
l
e
d
a
t
a
i
n
d
i
c
a
t
e
t
h
a
t
d
i
f
f
e
r
e
n
c
e
s
i
n
s
e
n
s
i
t
i
v
i
t
i
e
s
between
warm
and
coldwater
f
a
m
i
l
i
e
s
of
a
q
u
a
t
i
c
organisms
are
inadequate
t
o
w
a
r
r
a
n
t
d
i
s
c
r
i
m
i
n
a
t
i
o
n
i
n
t
h
e
n
a
t
i
o
n
a
l
ammonia
c
r
i
t
e
r
i
o
n
between
b
o
d
i
e
s
of
water
w
i
t
h
I'warm1'
and
'
lcoldwaterll
f
i
s
h
e
s
;

r
a
t
h
e
r
,
effects
o
f
organism
s
e
n
s
i
t
i
v
i
t
i
e
s
on
t
h
e
c
r
i
t
e
r
i
o
n
are
most
a
p
p
r
o
p
r
i
a
t
e
l
y
handled
by
s
i
t
e­
s
p
e
c
i
f
i
c
criteria
d
e
r
i
v
a
t
i
o
n
procedures.

Data
f
o
r
c
o
n
c
e
n
t
r
a
t
i
o
n
s
o
f
NH3
t
o
x
i
c
t
o
f
r
e
s
h
w
a
t
e
r
phytoplankton
and
v
a
s
c
u
l
a
r
p
l
a
n
t
s
,
a
l
t
h
o
u
g
h
l
i
m
i
t
e
d
,
i
n
d
i
c
a
t
e
t
h
a
t
freshwater
p
l
a
n
t
s
p
e
c
i
e
s
are
a
p
p
r
e
c
i
a
b
l
y
more
t
o
l
e
r
a
n
t
t
o
NH3
t
h
a
n
are
i
n
v
e
r
t
e
b
r
a
t
e
s
o
r
f
i
s
h
e
s
.
T
h
e
ammonia
c
r
i
t
e
r
i
o
n
appropriate
f
o
r
the
protection
of
a
q
u
a
t
i
c
animals
w
i
l
l
therefore
i
n
a
l
l
l
i
k
e
l
i
h
o
o
d
be
s
u
f
f
i
c
i
e
n
t
l
y
p
r
o
t
e
c
t
i
v
e
of
p
l
a
n
t
l
i
f
e
.
0
,,
Available
acute
and
chronic
data
for
ammonia
with
saltwater
organisms
are
very
limited,
and
insufficient
to
derive
a
saltwater
criterion.
A
few
saltwater
invertebrate
species
have
been
tested,
and
the
prawn
Macrobrachiurn
rosenberqil
was
the
most
sensitive.
The
few
saltwater
fishes
tested
suggest
greater
sensitivity
than
freshwater
fishes.
Acute
toxicity
of
NH3
appears
to
be
greater
at
low
pH
values,
similar
to
findings
in
freshwater.
Data
for
saltwater
plant
species
are
limited
to
diatoms,
which
appear
to
be
more
sensitive
than
the
saltwater
invertebrates
for
which
data
are
available.

More
quantitative
information
needs
to
be
published
on
the
toxicity
of
ammonia
to
aquatic
life.
Several
key
research
needs
must
be
addressed
to
provide
a
more
complete
assessment
of
ammonia
toxicity.
These
are:
(
1)
acute
tests
with
additional
saltwater
fish
species
and
saltwater
invertebrate
species:
(
2)

life­
cycle
and
early
life­
stage
tests
with
representative
freshwater
and
saltwater
organisms
from
different
families,
with
particular
attention
to
trends
of
acute­
chronic
ratios;
(
3)

fluctuating
and
intermittent
exposure
tests
with
a
variety
of
species
and
exposure
patterns:
(
4
)
more
complete
tests
of
the
individual
and
combined
effects
of
pH
and
temperature,
especially
for
chronic
toxicity:
(
5)
more
histopathological
and
histochemical
research
with
fishes,
which
would
provide
a
rapid
means
of
identifying
and
quantifying
sublethal
ammonia
effects;

and
(
6)
studies
on
effects
of
dissolved
and
suspended
solids
on
acute
and
chronic
toxicity.
NATIONAL
CRITERIA:

The
p
r
o
c
e
d
u
r
e
s
d
e
s
c
r
i
b
e
d
i
n
t
h
e
G
u
i
d
e
l
i
n
e
s
f
o
r
D
e
r
i
v
i
n
g
Numerical
National
Water
Q
u
a
l
i
t
y
C
r
i
t
e
r
i
a
f
o
r
t
h
e
P
r
o
t
e
c
t
i
o
n
of
Aquatic
Organisms
and
T
h
e
i
r
U
s
e
s
i
n
d
i
c
a
t
e
t
h
a
t
,
except
p
o
s
s
i
b
l
y
where
a
l
o
c
a
l
l
y
important
species
is
v
e
r
y
s
e
n
s
i
t
i
v
e
,
freshwater
a
q
u
a
t
i
c
o
r
g
a
n
i
s
m
s
and
t
h
e
i
r
u
s
e
s
s
h
o
u
l
d
n
o
t
be
a
f
f
e
c
t
e
d
unacceptably
i
f
:

(
1)
t
h
e
1­
hour*
average
concentration
of
un­
ionized
ammonia
(
i
n
mg/
L
NH3)
does
n
o
t
exceed,
more
often
t
h
a
n
once
every
3
years
on
t
h
e
average,
t
h
e
numerical
value
given
by
0.52/
FT/
FPH/
2,

where:

FT
=
10­
0.03(
20­
TCAP);
TCAP
­
<
T
­
<
30
10­
0.03(
20­
T)
;
0
­
<
T
­
<
TCAP
FPH
=
1
:
a
<
p
~
<
g
1+
10­
7.4­
PH
1.25
;
6.5
TCAP
=
2
0
C:
S
a
l
m
o
n
i
d
s
o
r
o
t
h
e
r
coldwater
s
p
e
c
i
e
s
p
r
e
s
e
n
t
=
25
C;
Salmonids
and
o
t
h
e
r
coldwater
species
absent
<
p
H
<
8
s
e
n
s
i
t
i
v
e
s
e
n
s
i
t
i
v
e
(*
An
averaging
p
e
r
i
o
d
of
1
h
o
u
r
may
n
o
t
b
e
appropriate
i
f
e
x
c
u
r
s
i
o
n
s
of
c
o
n
c
e
n
t
r
a
t
i
o
n
s
t
o
greater
t
h
a
n
1.5
t
i
m
e
s
t
h
e
average
occur
during
t
h
e
hour;
i
n
such
cases,
a
s
h
o
r
t
e
r
a
v
e
r
a
g
i
n
g
p
e
r
i
o
d
may
be
needed.)

(
2
)
t
h
e
4­
day
average
c
o
n
c
e
n
t
r
a
t
i
o
n
of
un­
i
o
n
i
z
e
d
ammonia
(
i
n
mg/
L
NH3)
does
n
o
t
exceed,
more
o
n
t
h
e
average,
t
h
e
average*

0.80/
FT/
FPH/
RATIO,
where
FT
and
FPH
,
o
f
t
e
n
t
h
a
n
once
every
3
years
n
u
m
e
r
i
c
a
l
v
a
l
u
e
g
i
v
e
n
b
y
are
a
s
above
and:
RATIO
=
16
;
7.7
­
<
pH
­
<
9
=
24
10­
7.7­
ph
1+
10­
7.4ph
;
6.5<
­
ph
­
<
7.7
TCAP
=
15
C;
Salmonids
or
other
sensitive
coldwater
species
present
coldwater
species
absent
=
20
C:
Salmonids
and
other
sensitive
(*
Because
these
formulas
are
nonlinear
in
pH
and
temperature,
the
criterion
should
be
the
average
of
separate
evaluations
of
the
formulas
reflective
of
the
fluctuations
of
flow,
pH,
and
temperature
within
the
averaging
period;
it
is
not
appropriate
in
general
to
simply
apply
the
formula
to
average
pH,
temperature,

and
flow.)

The
extremes
for
temperature
(
0,
30)
and
pH
(
6
.
5
,
9)
given
in
the
above
formulas
are
absolute.
It
is
not
permissible
with
current
data
to
conduct
any
extrapolations
beyond
these
limits.

In
particular,
there
is
reason
to
believe
that
appropriate
criteria
at
pH
>
9
will
be
lower
than
the
plateau
between
pH
8
and
9
given
above.

Criteria
concentrations
for
the
pH
range
6.5
to
9.0
and
the
temperature
range
0
C
to
30
C
are
provided
in
the
following
tables.
Total
ammonia
concentrations
equivalent
to
each
un­

ionized
ammonia
concentration
are
also
provided
in
these
tables.

There
are
limited
data
on
the
effect
of
temperature
on
chronic
toxicity.
EPA
will
be
conducting
additional
research
on
the
effects
of
temperature
on
ammonia
toxicity
in
order
to
fill
perceived
data
gaps.
Because
of
this
uncertainty,
additional
site­
specific
information
should
be
developed
before
these
criteria
are
used
in
wasteload
allocation
modeling.
For
example,

the
chronic
criteria
tabulated
for
sites
lacking
salmonids
are
less
certain
at
temperatures
much
below
20
C
than
those
tabulated
at
temperatures
near
20
C.
Where
the
treatment
levels
needed
to
meet
these
criteria
below
20
C
may
be
substantial,
use
of
site­

specific
criteria
is
strongly
suggested.
Development
of
such
criteria
should
be
based
upon
site­
specific
toxicity
tests.

Data
available
for
saltwater
species
are
insufficient
to
derive
a
criterion
for
saltwater.

The
recommended
exceedence
frequency
of
3
years
is
the
Agency's
best
scientific
judgment
of
the
average
amount
of
time
it
will
take
an
unstressed
system
to
recover
from
a
pollution
event
in
which
exposure
to
ammonia
exceeds
the
criterion.
A
stressed
system,
for
example,
one
in
which
several
outfalls
occur
in
a
limited
area,
would
be
expected
to
require
more
time
for
e
recovery.
The
resilience
of
ecosystems
and
their
ability
to
recover
differ
greatly,
however,
and
site­
specific
criteria
may
be
established
if
adequate
justification
is
provided.

The
use
of
criteria
in
designing
waste
treatment
facilities
requires
the
selection
of
an
appropriate
wasteload
allocation
model.
Dynamic
models
are
preferred
for
the
application
of
these
criteria.
Limited
data
or
other
factors
may
make
their
use
impractical,
in
which
case
one
should
rely
on
a
steady­
state
model.
The
Agency
recommends
the
interim
use
of
145
or
lQlO
for
Criterion
Maximum
Concentration
design
flow
and
745
or
7410
for
the
Criterion
Continuous
Concentration
design
flow
in
steady­

state
models
for
unstressed
and
stressed
systems
respectively.
1
(
I
)
Onchow
avuagm
ancmntratlonr
for
mmonla.*
,
r\
PH
o
c
5
c
10
c
[
IS
c
M
C
25
c
M
C
A.
Salmmldr
or
omr
Smnrltlvm
Coldwatr
Speclos
Present
Un­
1onIz.
d
A
m
n
i
a
(
mqllltor
NH,)

6.54
0.0091
0.0129
0.0182
0.026
O.
OS6
0
.
O
Y
0.036
6.75
0.0149
0.02
I
0.030
0.042
0.059
0.059
O.
OS9
7
.
oo
0.025
0.033
0
.
Mb
0.066
0.09)
0.093
0.093
0.054
0.048
0.068
0.095
0.135
0.135
0.135
7
3
0
0.045
0
.
w
0.091
0.128
0.181
0.181
0.181
7
3
5
n93
0.056
0.080
0.113
0.159
0.22
0.22
0.22
8.00
0.065
0
092
0.130
0.184
0.26
0.26
0
3
6
0.065
0.092
0.130
0.184
0.26
0.26
0.26
0.065
0.092
0.130
0.184
0
­
26
0
3
6
0.26
8
3
5
8.54
8.15
0.065
0.092
0.130
0.184
0.26
0.26
0.26
9
.
oo
0.065
0.092
0.130
0.184
0
2
6
0
3
6
0
2
6
Total
Amonla
(
mg/
Iltu
NH3)

35
32
28
25
17.4
12.2
8.0
4
.
I
2.6
I
.47
0
.
a
33
_­
w)
26
22
1
6
3
11.4
7
­
5
4
3
2.4
1
.
w
0.83
31
28
25
20
15.5
10.9
7.1
4.1
2.3
1.37
0
A3
30
n
­.
24
19.7
14.9
1
0
3
6
9
4
;
o
2.3
I
.
s
0
.86
29
27
­.
23
19
.2
14.6
10.3
6
­
0
.
_.
3.9
2.3
1.42
0.91
20
14.31
18.6
13
3
16.4
11.6
13.4
9.5
10.1
7.3
4.8
3
.5
2
.8
2.1
1.71
I
.28
I
­
07
0.83
0.72
0
3
8
7.2md
5.2
8.
Sallnonlds
and
O
t
h
U
S.
nrltlv*
Coldratw
Speclor
Absmt
UWl0nlz.
d
AnnuxlIa
(
m
g
/
I
l
t
r
NH,)

0.0091
0.0129
0.0182
0.026
0
.036
0.051
o.
os1
0.0149
0.02
1
0.030
0.042
O.
OS9
0.084
0.084
0.023
0.033
0.046
0.066
0.093
0.131
0.131
0.034
0.048
0.068
0.095
0.135
0.190
0.190
0.045
0.064
0.091
0.128
0.181
0.26
0.26
0.056
0.000
0.1
13
0.159
0.22
0.32
0.32
0.065
0.092
0.130
0.184
0.26
0
5
7
0
5
7
0.065
0.092
0.1SO
0.184
0.26
0.37
0.37
6
­
54
6.75
7
.
oo
7
2
5
7.50
­
7.75
8.00
8.25
830
8.75
9
.00
_­.
0
;
065
0
;
092
Oil30
Oil84
0.26
0
;
37
0;
37
0.065
0.092
0.130
0.184
0.26
0.37
0.37
0.06s
0
492
0.130
0.184
0
3
6
0.37
0.37
Total
A
m
n
i
a
(
mg/
llter
NWg)

35
32
28
2
3
17.4
12.2
8
.
o
4.
s
2.6
1.47
0.86
33
30
26
22
16.3
11.4
7.5
4.2
2
­
4
I
.40
0.83
31
28
25
20
15.5
10.9
7.1
4.1
2.3
1.31
0.83
30
27
24
19.7
14.9
10.5
6.9
4.0
2.3
1
.
JB
0.86
29
27
23
19.2
14.6
105
6.8
3.9
2.3
1.42
0.91
29
26
23
19.0
!
4
.
s
10.2
6.8
4
.
O
2.4
1.52
1.01
20
18.6
16.4
13.5­

4.
2.9
1
.
el
1.18
0.82
(
2)
4­
day
average
cmcentratlons
for
annonIa.*

PH
o
c
5
c
10
c
15
C
20
c
is
c
M
C
A.
S
a
l
m
l
d
s
or
Other
Sensltlve
ColdMater
SDecles
Present
Un­
Ionized
Amonla
(
r
n
g
/
l
l
t
r
NH3)

x
6
3
0
0.0007
6.75
0.0012
7
.00
0.0021
7.25
0.0037
7
.
so
0.0066
7.75
0.0109
8
.
oo
0.0126
8.25
0.0126
830
0.0126
8.75
0.0126
9
­
00
0.0126
0.0009
0.0017
0.0029
0.0052
0.0093
0.01
53
0.0177
0.0177
0.0177
0.0171
0.0177
0.0013
0.0023
0.0042
0.0074
0.0132
0.022
0.025
0.025
0.025
0.025
0.025
0.0019
0.0033
0.0059
0.0105
0.0186
0.03
1
0.035
0.035
0.035
0.035
0.035
0.0019
0.0033
0.0059
0.0105
0.01%
0.031
0.035
0.035
0.035
0.035
0
.
Of5
0.0019
0.0019
0.0033
0.0033
O.
CO59
0.0059
0.0105
0.0105
0.0186
0.0186
0.031
0.031
0.035
0.035
0.035
0.035
0.035
0.035
0.03s
0.035
0.035
0.035
6
3
0
2
3
2.4
2.2
2
3
1.49
1
.04
0.73
6.73
2.5
2.4
2
.2
2.2
1.49
I
.04
0.73
7.00
2.5
2.4
2
3
2
3
1­
49
1
.
w
0.74
7.25
2.5
2.4
2.2
2.2
1
.
M
1
.04
0.74
7.50
2.5
2.4
2
2
2
3
1
.
M
1.05'
0.74
7.75
2.3
2.2
2.1
2.0
1.40
0.99
0.71
0.47
8
.
a0
1.53
1.44
1.37
1.33
0.93
8.25
0.87
0.82
0.78
0.76
0.54
8.50
0.49
0.47
0.45
0.44
0
2
2
0
23
0.17
8
.75
0.28
0.27
0.26
0.27
0.19
OS15
0.11
9
.
W
0.16
0.16
0.16
0.16
0.13
0.10
0.08
00:
s
0.28
8.
Salmonlds'and
Other
Sensltlve
Coldwater
Species
Absent?

U
n
­
l
o
n
i
r
e
d
'
A
~
l
a
(
m
g
/
l
l
t
r
NH3)

6.50
0.0007
0.0009
0.0013
0.0019
0.0026
0.0026
0.0026
6.75
0.0012
0.0017
0.0023
0.0033
O.
oQ47
0.0047
0­
0047
7
.00
0.0021
0.0029
0.0042
0.0059
0.0083
0.0083
0.0083
7.25
0.0037
0.0052
0.0074
0.0105
0.0148
0.0148
0.0148
7
3
0
0.0066
0.0093
0.0132
0.0186
0.026
0.026
0.026
7.75
0.0109
0.01%
0.022
0.031
0.043
0.043
0.043
8.00
0.0126
0.0177
0.025
0.035
0.050
0.050
0.050
8.25
0.0126
0.0177
0.025
0.035
0.050
0.050
0.050
8.50
0.0126
0.0177
0.025
0.035
0
.
ow
0.0%
0.0%
8.75
0.0126
0.0177
0.025
0.035
0.050
0.0%
0.
oso
9.
W
0.0126
0.0177
0.025
0.035
0.0%
0.0%
0.0%

6.50
6.7s
7
.00
7
2
s
7
3
0
7.75
8
.
oo
8.25
8.50
8.75
9
.
oo
2.5
2.5
2.4
2.5
2
5
2.3
1.53
0
3
7
0.49
0.28
0.16
2
.4
2
­
4
2
­
4
2
­
4
2.4
2
­
2
1
­
44
0.82
0.47
0
3
7
0.16
2
.2
2
3
2
3
2.2
2.2
2.1
1.37.
0.78
0
­
45
0.26
0.16
2
3
2
3
2
3
2­
2
2
3
2.0
1.33
0.76
0.44
0.27
0.16
2.1
2.1
2.1
2.
I
2.1
1.98
1.31
0.76
0.45
037
0.17
Bs
d
1.46
l
,
Z
D
1.03
0
8c
1.47
1.
z:
1.04
D
.
@
g
1.47
,
tl
1
.04
0,
BC
1.48
l>
2P
1.05
0
0(
P
1.491.22.
I.
WO,@?
I.
39
/,/
U
1.00
0.82.
0.93
0.7@
0
6
1
Oa<
s'
0
3
4
0.40
0
*
33
0
3
5
0.21
0.16
0.14
0.1
I
+
To
convert
ihw
values
to
mg/
llter
N.
multlply
by
0.822.

t
Sltcspocltlc
erltala
dwelopmeni
I
s
shongly
suggested
a
t
tqm*
Yros
above
20
C
because
of
ihe
l
l
m
l
t
d
data
aveilable
to
gwuate
the
c
r
l
t
r
l
a
rac­
tlan,
and
a
t
tanporatures
b.
lW
20
C
boCauSe
Of
tha
llnltsd
data
and
brause
WO
aangn
I
n
the
crltwla
m
q
have
algnltlcant
Impact
on
the
level
of
trmtunt
r
q
u
l
r
d
I
n
meeting
t
h
e
recamended
crltorie.
The
Agency
acknowledges
that
the
Criterion
Continuous
Concentration
stream
flow
averaging
period
used
for
steady­
state
wasteload
allocation
modeling
may
be
as
long
as
30
days
in
situations
involving
POTWs
designed
to
remove
ammonia
where
limited
variability
of
effluent
pollutant
concentration
and
resultant
concentrations
in
receiving
waters
can
be
demonstrated.

In
cases
where
l
o
w
variability
can
be
demonstrated,
longer
averaging
periods
for
the
ammonia
Criterion
Continuous
Concentration
(
e.
g.,
30­
day
averaging
periods)
would
be
acceptable
because
the
magnitude
and
duration
of
exceedences
above
the
Criterion
Continuous
Concentration
would
be
sufficiently
limited.
These
matters
are
discussed
in
more
detail
in
the
Technical
Support
Document
for
Water
Quality­
Based
Toxics
Control
(
U.
S.
EPA,
1985a).

(
50
F.
R.
30784,
July
29,
1985)
SEE
APPENDIX
A
FOR
METHODOLOGY
ANTIMONY
CRITERIA:

Aquatic
Life
The
available
data
for
antimony
indicate
that
acute
and
chronic
toxicity
to
freshwater
aquatic
life
occur
at
concentrations
as
low
as
9,000
and
1,600
ug/
L,
respectively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
Toxicity
to
algae
occurs
at
concentrations
as
low
as
610
ug/
L.

No
saltwater
organisms
have
been
adequately
tested
with
antimony,
and
no
statement
can
be
made
concerning
acute
or
chronic
toxicity.

Human
Health
For
the
protection
of
human
health
from
the
toxic
properties
of
antimony
ingested
through
water
and
contaminated
aquatic
organisms,
the
ambient
water
criterion
is
determined
to
be
146
UWL.

For
the
protection
of
human
health
from
the
toxic
properties
of
antimony
ingested
through
contaminated
aquatic
organisms
alone,
the
ambient
water
criterion
is
determined
to
be
45
mg/
L.

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
J
ARSENIC
AQUATIC
LIFE
SUMMARY:
0
The
chemistry
of
arsenic
in
water
is
complex
and
the
form
present
in
solution
is
dependent
on
such
environmental
conditions
as
Eh,
pH,
organic
content,
suspended
solids,
and
sediment.
The
relative
toxicities
of
the
various
forms
of
apsenic
apparently
vary
from
species
to
species.
For
inorganic
arsenic(
II1)
acute
values
for
16
freshwater
animal
species
ranged
from
812
ug/
L
for
a
cladoceran
to
97,000
ug/
L
for
a
midge,
but
the
three
acute­

chronic
ratios
only
ranged
from
4.660
to
4.862.
The
five
acute
values
for
inorganic
arsenic(
V)
covered
about
the
same
range,
but
the
single
acute­
chronic
ratio
was
28.71.
The
six
acute
values
for
MSMA
ranged
from
3,243
to
1',
403,000
ug/
L.
The
freshwater
residue
data
indicated
that
arsenic
is
not
bioconcentrated
to
a
high
degree
but
that
lower
forms
of
aquatic
life
may
accumulate
higher
arsenic
residues
than
fish.
The
low
bioconcentration
factor
and
short
half­
life
of
arsenic
in
fish
tissue
suggest
that
residues
should
not
be
a
problem
to
predators
of
aquatic
life.
a
The
available
data
indicate
that
freshwater
plants
differ
a
great
deal
as
to
their
sensitivity
to
arsenic(
II1)
and
arsenic(
V).
In
comparable
tests,
the
alga,
­­­­­___­­­
Selenastrum
capricornutum,
was
45
times
more
sensitive
to
arsenic(
V)
than
to
arsenic(
III),
although
other
data
present
conflicting
information
on
the
sensitivity
of
this
alga
to
arsenic(
V).
Many
plant
values
for
inorganic
arsenic(
II1)
were
in
the
same
range
as
the
available
chronic
values
for
freshwater
animals;
several
0
­
plant
values
for
arsenic(
V)
were
lower
than
the
one
available
chronic
value.

The
other
toxicological
data
revealed
a
wide
range
of
toxicity
based
on
tests
with
a
variety
of
freshwater
species
and
endpoints.
Tests
with
early
life
stages
appeared
to
be
the
most
sensitive
indicator
of
arsenic
toxicity.
Values
obtained
from
this
type
of
test
with
inorganic
arsenic(
II1)
were
lower
than
chronic
values
contained
in
Table
2.
For
example,
an
effect
concentration
of
40
ug/
L
was
obtained
in
a
test
on
inorganic
arsenic(
II1)
with
embryos
and
larvae
of
a
toad.

Twelve
species
of
saltwater
animals
have
acute
values
for
inorganic
arsenic(
II1)
from
232
to
16,030
ug/
L
and
the
single
acute­
chronic
ratio
is
1.945.
The
only
values
available
for
inorganic
arsenic(
V)
are
for
two
invertebrate
and
are
between
2,000
and
3,000
ug/
L.
Arsenic(
II1)
and
arsenic(
V)
are
equally
toxic
to
various
species
of
saltwater
algae,
but
the
sensitivities
of
the
species
range
from
19
ug/
L
to
more
than
1,000
ug/
L.
In
a
test
with
an
oyster,
a
BCF
of
350
was
obtained
for
inorganic
arsenic(
II1).

NATIONAL
CRITERIA:

The
procedures
described
in
the
Guidelines
for
Deriving
Numerical
National
Water
Quality
Criteria
for
the
Protection
of
Aquatic
Organisms
and
Their
Uses
indicate
that,
except
possibly
where
a
locally
important
species
is
very
sensitive,
freshwater
aquatic
organisms
and
their
uses
should
not
be
affected
unacceptably
if
the
4­
day
average
concentration
of
arsenic(
II1)

does
not
exceed
190
ug/
L
more
than
once
every
3
years
on
the
average
and
if
the
1­
hour
average
concentration
does
not
exceed
360
ug/
L
more
than
once
every
3
years
on
the
average.

The
procedures
described
in
the
Guidelines
indicate
that,

except
possibly
where
a
locally
important
species
is
very
sensitive,
saltwater
aquatic
organisms
and
their
uses
should
not
be
affected
unacceptably
if
the
4­
day
average
concentration
of
arsenic(
II1)
does
not
exceed
36
ug/
L
more
than
once
every
3
years
on
the
average
and
if
the
1­
hour
average
concentration
does
not
exceed
69
ug/
L
more
than
once
every
3
years
on
the
average.
This
criterion
might
be
too
high
wherever
Skeletonema
cosrarum
or
Thalassiosira
aestivalis
­­­
are
ecologically
important.

Not
enough
data
are
available
to
allow
derivation
of
numerical
national
water
quality
criteria
for
freshwater
aquatic
life
for
inorganic
arsenic(
V)
or
any
organic
arsenic
compound.

Inorganic
arsenic(
V)
is
acutely
toxic
to
freshwater
aquatic
animals
at
concentrations
as
low
as
850
ug/
L
and
an
acute­
chronic
ratio
of
2
8
was
obtained
with
the
fathead
minnow.
Arsenic(
V)

affected
freshwater
aquatic
plants
at
concentrations
as
low
as
48
ug/
L.
Monosodium
methanearsenace
(
MSMA)
is
acutely
toxic
to
aquatic
animals
at
concentrations
as
low
as
1,900
ug/
L,
but
no
data
are
available
concerning
chronic
toxicity
to
animals
or
toxicity
to
plants.

Very
few
data
are
available
concerning
the
toxicity
of
any
form
of
arsenic
other
than
inorganic
arsenic(
II1)
to
saltwater
aquatic
life.
The
available
data
do
show
that
inorganic
arsenic(
v)
is
acutely
toxic
to
saltwater
animals
at
concentrations
as
low
as
2,319
ug/
L
and
affected
some
saltwater
p
l
a
n
t
s
a
t
13
t
o
56
ug/
L.

c
h
r
o
n
i
c
t
o
x
i
c
i
t
y
of
any
arsenic(
II1)
t
o
saltwater
N
o
data
a
r
e
a
v
a
i
l
a
b
l
e
concerning
t
h
e
form
of
a
r
s
e
n
i
c
o
t
h
e
r
than
inorganic
aquatic
l
i
f
e
.

EPA
b
e
l
i
e
v
e
s
t
h
a
t
a
measurement
such
as
tlacid­
solublett
would
provide
a
more
s
c
i
e
n
t
i
f
i
c
a
l
l
y
c
o
r
r
e
c
t
b
a
s
i
s
upon
which
t
o
e
s
t
a
b
l
i
s
h
c
r
i
t
e
r
i
a
f
o
r
metals.
T
h
e
c
r
i
t
e
r
i
a
were
developed
on
t
h
i
s
b
a
s
i
s
.
However,
a
t
t
h
i
s
t
i
m
e
,
no
EPA
approved
methods
f
o
r
such
a
measurement
a
r
e
a
v
a
i
l
a
b
l
e
t
o
implement
t
h
e
c
r
i
t
e
r
i
a
through
t
h
e
r
e
g
u
l
a
t
o
r
y
programs
of
t
h
e
Agency
and
t
h
e
S
t
a
t
e
s
.

The
Agency
is
considering
development
and
approval
of
methods
f
o
r
a
measurement
such
as
acid­
soluble.
U
n
t
i
l
a
v
a
i
l
a
b
l
e
,
however,

EPA
recommends
applying
the
criteria
using
t
h
e
t
o
t
a
l
recoverable
method.
This
has
two
impacts:
(
1)
c
e
r
t
a
i
n
species
of
some
metals
cannot
be
analyzed
d
i
r
e
c
t
l
y
because
t
h
e
t
o
t
a
l
recoverable
method
does
not
distinguish
between
individual
oxidation
states,
and
(
2)

these
c
r
i
t
e
r
i
a
may
be
overly
protective
when
based
on
the
t
o
t
a
l
recoverable
method.

The
recommended
exceedence
frequency
of
3
y
e
a
r
s
is
t
h
e
Agency's
b
e
s
t
s
c
i
e
n
t
i
f
i
c
judgment
of
t
h
e
average
amount
of
t
i
m
e
it
w
i
l
l
t
a
k
e
an
unstressed
system
t
o
r
e
c
o
v
e
r
from
a
p
o
l
l
u
t
i
o
n
event
i
n
which
exposure
t
o
arsenic(
II1)
exceeds
t
h
e
c
r
i
t
e
r
i
o
n
.
a
stressed
system,
f
o
r
example,
one
i
n
which
several
o
u
t
f
a
l
l
s
occur
i
n
a
l
i
m
i
t
e
d
a
r
e
a
,
would
be
expected
t
o
r
e
q
u
i
r
e
more
t
i
m
e
f
o
r
recovery.
The
r
e
s
i
l
i
e
n
c
e
of
ecosystems
and
t
h
e
i
r
a
b
i
l
i
t
y
t
o
recover
differ
greatly,
however,
and
s
i
t
e­
s
p
e
c
i
f
i
c
criteria
may
be
established
i
f
adequate
j
u
s
t
i
f
i
c
a
t
i
o
n
is
provided.

0
The
use
of
criteria
i
n
designing
waste
treatment
facilities
,

r
e
q
u
i
r
e
s
t
h
e
s
e
l
e
c
t
i
o
n
of
an
a
p
p
r
o
p
r
i
a
t
e
wasteload
a
1
l
o
c
a
t
i
o
n
model.
Dynamic
models
are
preferred
for
the
application
of
these
criteria.
Limited
data
or
other
factors
may
make
their
use
impractical,
in
which
case
one
should
rely
on
a
steady­
state
model,
The
Agency
recommends
the
interim
use
of
145
or
lQlO
for
Criterion
Maximum
Concentration
design
flow
and
745
or
7Q10
for
the
Criterion
Continuous
Concentration
design
flow
in
steady­

state
models
for
unstressed
and
stressed
systems
respective1
y.

These
matters
are
discussed
in
more
detail
in
the
Technical
Support
Document
fox
Water
Quality­
Based
Toxics
Control
(
U.
S.

EPA,
1985).

HUMAN
HEALTH
CRITERIA:

For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
due
to
expos,
ure
of
arsenic
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentration
should
be
zero
based
on
the
non­

threshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
and
The
corresponding
criteria
are
2
2
ng/
L,
2.2
ng/
L,
and
.22
ng/
L,
respectively.
If
the
above
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
175
ng/
L,
17.5
ng/
L,
and
1.75
ng/
L,
respectively.
Other
concentrations
representing
different
risk
levels
may
be
e
calculated
by
use
of
the
Guidelines.
The
risk
estimate
range
is
presented
for
information
purpoes
and
does
not
represent
an
,,

Agency
judgment
on
an
81acceptable8t
risk
level.
(
45
SEE
F.
R.
79318
APPENDIX
A
N
o
v
.
28,1980)
(
50
FOR
METHODOLOGY
F.
R.
30784,
J
u
l
y
a
ASBESTOS
CRITERIA:

Aquatic
Life
No
freshwater
organisms
have
been
tested
with
any
asbestifom
mineral
and
no
statement
can
be
made
concerning
acute
or
chronic
toxicity
No
saltwater
organisms
have
been
tested
with
any
asbestiform
mineral
and
no
statement
can
be
made
concerning
acute
or
chronic
toxicity.

Human
Health
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
asbestos
through
ingestion
of
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentration
should
be
zero.
The
estimated
levels
which
would
result
in
increased
lifetime
cancer
risks
of
and
are
300,000
fibers/
L,
30,000
fibers/
L,
and
3,000
fibers/
L,
respectively.
Estimates
for
consumption
of
aquatic
organisms
only,
excluding
the
consumption
of
water
cannot
be
made.

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
MmnoDomeY
BACTERIA
Freshwater
Bathinq
Based
on
a
statistically
sufficient
number
of
samples
(
generally
not
less
than
5
samples
equally
spaced
over
a
30­
day
period),
the
geometric
mean
of
the
indicated
bacterial
densities
should
not
exceed
one
or
the
other
of
the
following:(')

E.
coli
126
per
100
ml;
or
enterococci
33
per
100
ml;

no
sample
should
exceed
a
one
sided
confidence
limit
(
C.
L.)

calculated
using
the
following
as
guidance:

designated
bathing
beach
75%
C.
L.

moderate
use
for
bathing
82%
C.
L.

light
use
for
bathing
90%
C.
L.

infrequent
use
for
bathing
95%
C.
L.

based
on
a
site­
specific
log
standard
deviation,
or
if
site
data
are
insufficient
to
establish
a
log
standard
deviation,
then
using
0.4
as
the
log
standard
deviation
for
both
indicators.

Marine
Water
Bathinq
Based
on
a
statistically
sufficient
number
of
samples
(
generally
not
less
than
5
samples
equally
spaced
over
a
30­
day
period),
the
geometric
mean
of
the
enterococci
densities
should
not
exceed
35
per
100
ml;
no
sample
should
exceed
a
one
sided
confidence
limit
using
the
following
as
guidance:

designated
bathing
beach
75%
C.
L.

moderate
use
for
bathing
82%
C.
L.

light
use
for
bathing
90%
C.
L.

infrequent
use
for
bathing
95%
C.
L.
based
on
a
site­
specific
log
standard
deviation,
or
if
site
data
are
insufficient
to
establish
a
log
standard
deviation,
then
using
0.7
as
the
log
standard
deviation.

Note
(
1)
­
Only
one
indicator
should
be
used.
The
Regulatory
agency
should
select
the
appropriate
indicator
for
its
conditions.

Shellfish
Harvesting
Waters
The
median
fecal
coliform
bacterial
concentration
should
not
exceed
14
MPN
per
100
ml
with
not
more
than
10
percent
of
samples
exceeding
43
MPNperlOOml
forthetakingof
shellfish.

RATIONALE
Bathing
Waters
A
recreational
water
quality
criterion
can
be
defined
as
a
"
quantifiable
relationship
between
the
density
of
an
indicator
in
the
water
and
the
potential
human
health
risks
involved
in
the
water's
recreational
use."
From
such
a
definition,
a
criterion
can
be
adopted
which
establishes
upper
limits
for
densities
of
indicator
bacteria
in
waters
that
are
associated
with
acceptable
health
risks
for
swimmers.

The
Environmental
Protection
Agency,
in
1972,
initiated
a
series
of
studies
at
marine
and
fresh
water
bathing
beaches
which
were
designed
to
determine
if
swimming
in
sewage­
contaminated
marine
and
fresh
water
carries
a
health
risk
for
bathers;
and,
if
so,
to
what
type
of
illness.
Additionally,
the
Agency
wanted
to
determine
which
bacterial
indicator
is
best
correlated
to
swimming­
associated
health
effects
and
if
the
relationship
is
strong
enough
to
provide
a
criterion.
(
l)
The
quantitative
relationships
between
the
rates
of
swimming­

associated
health
effects
and
bacterial
indicator
densities
were
determined
using
regression
analysis.
Linear
relationships
were
estimated
from
data
grouped
on
the
basis
of
summers
or
trials
with
similar
indicator
densities.
The
data
for
each
summer
were
analyzed
by
pairing
the
geometric
mean
indicator
density
for
a
summer
bathing
season
at
each
beach
with
the
corresponding
swimming­
associated
gastrointestinal
illness
rate
for
the
same
summer.
The
swimming­
associated
illness
rate
was
determined
by
subtracting
the
gastrointestinal
illness
rate
in
nonswimmers
from
that
for
swimmers.
Ihese
two
variables
from
multiple
beach
sites
were
used
to
calculate
a
regression
coefficient,
y­
intercept
and
95%
confidence
intervals
for
the
paired
data.
In
the
marine
studies
the
total
number
of
points
for
use
in
regression
analysis
was
increased
by
collecting
trial
days
with
similar
indicator
densities
from
each
study
location
and
placing
them
into
groups.

The
swimming­
associated
illness
rate
was
determined
as
before,
by
suk&
racting
the
nonswimmer
illness
rate
of
all
the
individuals
included
in
the
grouped
trial
days
from
the
swimmer
illness
rate
during
these
safe
grouped
trial
days.
The
grouping
by
trial
days
with
similar
indicator
densities
approach
was
not
possible
with
the
freshwater
data
because
the
variation
of
bacterial
indicator
densities
in
freshwater
samples
was
not
larqe
enouqh
to
allow
such
an
adjustment
to
be
made.
For
the
saltwater
studies
the
results
of
the
regression
analyses
of
illness
rates
against
indicator
density
data
was
very
similar
using
the
"
by
summer"
or
Itby
grouped
trial
days"
approaches.
0
1
The
methods
used
to
enumerate
t
h
e
b
a
c
t
e
r
i
a
l
i
n
d
i
c
a
t
o
r
d
e
n
s
i
t
i
e
s
which
showed
t
h
e
best
r
e
l
a
t
i
o
n
s
h
i
p
to
swimming­

associated
g
a
s
t
r
o
e
n
t
e
r
i
t
i
s
rates
were
s
p
e
c
i
f
i
c
a
l
l
y
developed
 or
t
h
e
recreational
water
q
u
a
l
i
t
y
studies.'

These
membrane
f
i
l
t
e
r
methods
have
s
u
c
c
e
s
s
f
u
l
l
y
undergone
p
r
e
c
i
s
i
o
n
and
b
i
a
s
t
e
s
t
i
n
g
by
t
h
e
EPA
Environmental
Monitoring
and
Support
Laboratory.
(
2)

Several
monitoring
s
i
t
u
a
t
i
o
n
s
to
assess
b
a
c
t
e
r
i
a
l
q
u
a
l
i
t
y
are
encountered
by
r
e
g
u
l
a
t
o
r
y
agencies.
The
s
i
t
u
a
t
i
o
n
needing
t
h
e
most
rigorous
monitoring
is
the
designated
swimming
beach.
Such
a
r
e
a
s
are
f
r
e
q
u
e
n
t
l
y
l
i
f
e
g
u
a
r
d
p
r
o
t
e
c
t
e
d
,
p
r
o
v
i
d
e
parking
and
other
public
access
and
a
r
e
heavily
used
by
the
public.
Public
b
e
a
c
h
e
s
o
f
t
h
i
s
t
y
p
e
were
used
by
EPA
i
n
d
e
v
e
l
o
p
i
n
g
t
h
e
r
e
l
a
t
i
o
n
s
h
i
p
described
i
n
t
h
i
s
document.

Other
r
e
c
r
e
a
t
i
o
n
a
l
a
c
t
i
v
i
t
i
e
s
may
i
n
v
o
l
v
e
bodies
of
water
which
are
regulated
by
individual
State
water
q
u
a
l
i
t
y
standards.

These
recreational
resources
may
be
n
a
t
u
r
a
l
wading
ponds
used
by
c
h
i
l
d
r
e
n
or
waters
where
i
n
c
i
d
e
n
t
i
a
l
f
u
l
l
body
c
o
n
t
a
c
t
occurs
because
of
water
skiing
o
r
other
similar
activities.

EPA's
evaluation
of
the
bacteriological
data
indicated
t
h
a
t
using
the
f
e
c
a
l
coliform
indicator
group
a
t
the
maximum
geometric
mean
of
200perlOOm1,
recommended
i
n
Q
u
a
l
i
w
C
r
i
t
e
r
i
a
___­
f
o
r
w
a
t
e
r
would
cause
an
estimated
8
i
l
l
n
e
s
s
p
e
r
1
,
0
0
0
s
w
i
m
m
e
r
s
a
t
fresh
water
beaches
and
1
9
i
l
l
n
e
s
s
p
e
r
1
,
0
0
0
s
w
i
m
m
e
r
s
a
t
marine
beaches.
These
relationships
are
only
approximate
and
are
based
on
a
p
p
l
y
i
n
g
r
a
t
i
o
s
of
t
h
e
g
e
o
m
e
t
r
i
c
means
o
f
t
h
e
v
a
r
i
o
u
s
i
n
d
i
c
a
t
o
r
s
from
t
h
e
EPA
s
t
u
d
i
e
s
to
t
h
e
200
p
e
r
100
m
l
fecal
coliform
criterion.
However,
these
are
EPA's
best
estimates
of
the
accepted
illness
rates
for
areas
which
apply
the
EPA
fecal
coliform
criterion.

The
E.
coli
and
enterococci
criteria
presented
in
Table
1
were
developed
using
these
currently
accepted
illness
rates.
The
equations
developed
by
D~
four(~)
and
Cabelli(*)
were
used
to
calculate
the
geometric
mean
indicator
densities
corresponding
to
the
accepted
gastrointestinal
illness
rates.
These
densities
are
for
steady
state
dry
weather
conditions.
The
beach
is
in
noncom­

pliance
with
the
criteria
if
the
geometric
mean
of
several
bacterial
density
samples
exceeds
the
value
listed
in
Table
1.

Noncompliance
is
also
signaled
by
an
unacceptably
high
value
for
any
single
bacterial
sample.
The
maximum
acceptable
bacterial
density
for
a
single
sample
is
set
higher
than
that
for
the
geometric
mean,
in
order
to
avoid
necessary
beach
closings
based
on
single
samples.
In
deciding
whether
a
beach
should
be
left
open,
it
is
the
long
term
geometric
mean
bacterial
density
that
is
of
interest.
Because
of
day­
to­
day
fluctuations
around
this
mean,
a
decision
based
on
a
single
sample
(
or
even
several
samples)
may
be
erroneous,
i.
e.,
the
sample
may
exceed
the
recommended
mean
criteria
even
though
the
long­
term
geometric
mean
is
protective,
or
may
fall
below
the
maximum
even
if
this
mean
is
in
the
nonprotective
range.

To
set
the
single
sample
maximum,
it
is
necessary
to
specify
the
desired
chance
that
the
beach
will
be
left
open
when
the
protection
is
adequate.
This
chance,
or
confidence
level,
was
based
on
Agency
judgment.
For
the
simple
decision
rule
considered
here,
a
smaller
confidence
level
corresponds
to
a
more
/
stringent
(
i.
e.
lower)
single
sample
maximum.
Conversely,
a
greater
confidence
level
corresponds
to
less
stringent
(
i.
e.

higher)
maximum
values.
This
technique
reduces
the
chances
of
single
samples
inappropriately
indicating
violations
of
the
recommended
criteria.

By
using
a
control
chart
analogy
(
5)
and
the
actual
log
standard
deviations
from
the
EPA
studies,
single
sample
maximum
densities
for
various
confidence
levels
were
calculated.
EPA
then
assigned
qualitative
use
intensities
to
those
confidence
levels.
A
low
confidence
level
(
75%)
was
assigned
to
designated
beach
areas
because
a
high
degree
of
caution
should
be
used
to
evaluate
water
quality
for
heavily
used
areas.
Less
intensively
used
areas
would
allow
less
restrictive
single
sample
limits.

Thus,
95%
confidence
might
be
appropriate
for
swimmable
water
in
remote
areas.
Table
1
summarizes
the
results
of
these
calculations.
These
single
sample
maximum
levels
should
be
recalculated
for
individual
areas
if
significant
differences
in
log
standard
deviations
occur.

The
levels
displayed
in
Table
1
depend
not
only
on
the
assumed
standard
deviation
of
log
densities,
but
also
on
the
chosen
level
of
acceptable
risk.
While
this
level
was
based
on
the
historically
accepted
risk,
it
is
still
arbitrary
insofar
as
the
historical
risk
was
itself
arbitrary.

Currently
EPA
is
not
recommending
a
change
in
the
stringency
of
its
bacterial
criteria
for
recreational
waters.
Such
a
change
does
not
appear
warranted
until
more
information
based
on
greater
experience
with
the
new
indicators
can
be
accrued.
EPA
and
the
S
t
a
t
e
Agencies
can
then
evaluate
t
h
e
impacts
of
change
i
n
terms
of
beach
closures
and
other
restricted
uses.

S
h
e
l
l
f
i
s
h
Harvesting
Waters
The
microbiological
c
r
i
t
e
r
i
o
n
f
o
r
shellfish
water
q
u
a
l
i
t
y
has
been
accepted
by
international
agreement
t
o
be
70
t
o
t
a
l
coliforms
p
e
r
100
m
l
,
using
a
median
MPN,
with
no
more
than
1
0
p
e
r
c
e
n
t
of
t
h
e
values
exceeding
230
t
o
t
a
l
coliforms.
No
evidence
of
disease
outbreak
from
consumption
of
raw
s
h
e
l
l
f
i
s
h
which
were
grown
i
n
w
a
t
e
r
s
meeting
t
h
i
s
b
a
c
t
e
r
i
o
l
o
g
i
c
a
l
c
r
i
t
e
r
i
o
n
has
been
demonstrated.
This
standard
has
proven
t
o
be
a
p
r
a
c
t
i
c
a
l
l
i
m
i
t
when
supported
by
s
a
n
i
t
a
r
y
surveys
of
t
h
e
growing
waters,

acceptable
q
u
a
l
i
t
y
i
n
she1
l
f
i
s
h
meats,
and
good
epidemiological
evidence.
However,
evidence
from
f
i
e
l
d
investigations
suggests
t
h
a
t
n
o
t
a
l
l
t
o
t
a
l
coliform
occurrences
can
be
a
s
s
o
c
i
a
t
e
d
w
i
t
h
f
e
c
a
l
p
o
l
l
u
t
i
o
n
.
Thus,
a
t
t
e
n
t
i
o
n
h
a
s
been
d
i
r
e
c
t
e
d
toward
t
h
e
adoption
of
t
h
e
f
e
c
a
l
coliform
t
e
s
t
t
o
measure
more
a
c
c
u
r
a
t
e
l
y
t
h
e
occurrence
and
magnitude
of
fecal
p
o
l
l
u
t
i
o
n
i
n
s
h
e
l
l
f
i
s
h
­

growing
waters.

A
series
of
studies
was
i
n
i
t
i
a
t
e
d
by
the
National
S
h
e
l
l
f
i
s
h
S
a
n
i
t
a
t
i
o
n
Program
and
d
a
t
a
r
e
l
a
t
i
n
g
t
h
e
occurrence
of
t
o
t
a
l
coliforms
t
o
numbers
of
fecal
coliforms
were
compiled.
The
data
s
h
o
w
t
h
a
t
a
70
coliform
MPNperlOOmlatthe
5
O
t
h
p
e
r
c
e
n
t
i
l
e
w
a
s
e
q
u
i
v
a
l
e
n
t
t
o
a
fecal
coliform
MPN
of
14
p
e
r
1
J
O
m
l
.
The
data,

t
h
e
r
e
f
o
r
e
,
i
n
d
i
c
a
t
e
t
h
a
t
a
median
v
a
l
u
e
f
o
r
a
f
e
c
a
l
coliform
standard
is
15
and
t
h
e
9
O
t
h
p
e
r
c
e
n
t
i
l
e
s
h
o
u
l
d
n
o
t
exceed
43
f
o
r
a
0
5­
tubeI
3­
dilution
method.

EPA
i
s
c
u
r
r
e
n
t
l
y
(
1986)
co­
sponsoring,
with
t
h
e
National
Oceanic
and
Atmospheric
Administration,
research
into
the
application
of
the
enterococci
and
E.
coli
indicators
for
assessing
the
quality
of
shellfish
harvesting
waters.
The
Food
and
Drug
Administration
is
also
reviewing
the
results
of
these
studies.
A
change
to
the
new
indicators
may
be
forthcoming
if
the
studies
show
a
correlation
between
gastrointestinal
disease
and
the
consumption
of
raw
shellfish
from
waters
with
defined
densities
of
the
new
indicators.
However,
these
studies
have
not
sufficiently
progressed
to
justify
any
change
at
this
time.

Thus,
evaluation
of
the
microbiological
suitability
of
waters
for
recreational
taking
of
shellfish
should
be
based
upon
the
fecal
colifonn
bacterial
levels.
c6)
CIIITERIA
FOR
INDICAllR
FOR
B
r
n
I
O
r
n
C
A
L
IEWITIES
s
h
l
e
sanple
Maximun
Allowable
Oensity
(
41,
(
5)

Acceptable
Srhh
Steady
State
Designated
Msderate
Full
Lightly
Used
Infrequmtly
Used
e
n
t
e
r
i
t
i
s
Fate
per
Indicator
(
upper
75%
C.
L.)
Recreation
Contact
Rxreation
1000
swimners
Density
(
upper
82%
CL.)
Rereation
(
upper
95%
CL.)
Assaiated
Qstro­
Geawtric
Mean
madl
Area
m
y
contact
Full
m
y
m11
m
y
ONltact
(
u
p
r
90%
C.
L.
Freshwater
enterococci
8
33(
1)

­­
E.
coli
8
126P)

Marine
Bter
61
235
89
108
151
298
406
576
enterococci
19
35(
3)
104
124
276
N
O
t
e
s
t
(
1
)
Calculated
to
nearest
m
l
e
nunber
using
quation:

Calculated
to
nearest
whole
nunber
using
equation:

Calculated
to
nearest
w
b
l
e
nunber
using
equation:
(
mean
entemxcci
density)
=
i"
tilogl0
(
mean
­
­
E.
coli
density)
=
antFloglO
(
mean
enterococci
density)
=
antiloglo
illness
rate/
lOOO
people
+
6.28
9.40
(
2)
illness
rate/
l000
pqi
le
+
11.74
9.40
(
3)
illness
rate/
lOOO
people
­
0.20
12.17
500
(
4)
sirgle
sanple
l
h
i
t
=
m
t
i
l
q
l
0
(
laglo
indicator
geamtric
+
Factor
determined
frm
x
(
lq110
stand,
mean
density/
100
m
l
)
areas
rnder
the
Wnnal
deviation)
prcbability
cuwe
for
tb
ass­
level
of
pmbability
The
appmpdata
factors
for
tb
indicated
one
sided
confidence
levels
are:
75%
C.
L.
­
­
675
82%
C.
L.
­
­
935
90%
C.
L.
­
1.28
95%
C.
L.
­
1.65
(
5)
Based
on
the
observed
log
standard
deviations
dwhq
tk
EPA
studies:
and
enterccccci;
and
0.7
for
mq*
e
mter:
enterococci.
standard
dwiation
for
its
coditions
&
ich
muld
then
vary
t&
sig(
ie
mple
m
t.
0.4
f
o
r
freshwater
E.
&
Each
'
uridi
tion
s
b
u
p
fstablish­
its
om
1.

2
.

3
.

4.

5.

6.
References
Ambient
Water
Quality
Criteria
for
Bacteria
­
1986,
EPA
440/
5­
84­
002,
U.
S.
Environmental
Protection
Agency,
Office
of
Water
Regulations
and
Standards,
Washington,
DC.
(
NTIS
access
#:
PB
86­
158­
045)

Test
Methods
for
Escherichia
coli
and
Enterococii
in
Water
By
The
Membrane
Filter
procedure,
EPA
600/
4­
85­
076,
U.
S.

Environmental
Protection
Agency,
Cincinnati,
OH.
(
NTIS
access
#:
PB
86­
158­
052)

Dufour,
A.
P.
1983.
Health
Effects
Criteria
for
Fresh
Recreat
iona
1
Waters.
EPA­
6
0
O/
1­
8
4
­
0
04
,
U.
S.
Env
ironmenta
1
Protection
Agency,
Cincinnati,
OH.

Cabelli,
V.
J.
1981.
Health
Effects
Criteria
for
Marine
Recreational
Waters.
EPA­
600/
1­
80­
031,
U.
S.
Environxental
Protection
Agency,
Cincinnati,
OH.

ASTM.
1951.
Manual
on
Quality
Control
of
Materials.

Special
Technical
Publication
15­
C,
American
Society
for
Testing
and
Materials,
Philadelphia,
PA.

U.
S.
Environmental
Protection
Agency.
1976.
Quality
Criteria
for
Water.
U.
S.
Environmental
Protection
Agency,

Washington,
DC.
BARIUM
1
mg/
L
f
o
r
domestic
water
supply
(
h
e
a
l
t
h
)
.

INTRODUCTION:

Barium
is
a
yellowish­
white
metal
of
t
h
e
a
l
k
a
l
i
n
e
e
a
r
t
h
group.
I
t
o
c
c
u
r
s
i
n
n
a
t
u
r
e
c
h
i
e
f
l
y
a
s
b
a
r
i
t
e
,
BaS04
and
witherite,
BaC03,
both
of
which
are
highly
insoluble
s
a
l
t
s
.
The
metal
is
stable
i
n
dry
a
i
r
,
but
r
e
a
d
i
l
y
oxidized
by
humid
a
i
r
o
r
w
a
t
e
r
.

Many
of
t
h
e
s
a
l
t
s
of
barium
are
s
o
l
u
b
l
e
i
n
both
w
a
t
e
r
and
a
c
i
d
,
and
s
o
l
u
b
l
e
barium
s
a
l
t
s
a
r
e
reported
t
o
be
poisonous
(
Lange,
1965:
NAS,
1
9
7
4
)
.
However,
barium
i
o
n
s
g
e
n
e
r
a
l
l
y
a
r
e
thought
t
o
be
r
a
p
i
d
l
y
p
r
e
c
i
p
i
t
a
t
e
d
o
r
removed
from
s
o
l
u
t
i
o
n
by
absorption
and
sedimentation
(
M
c
K
e
e
and
Wolf,
1963
NAS,
1974).
I
While
barium
is
a
m
a
l
l
e
a
b
l
e
,
d
u
c
t
i
l
e
metal,
i
t
s
major
commercial
value
is
i
n
its
compounds.
Barium
compounds
a
r
e
used
i
n
a
v
a
r
i
e
t
y
of
i
n
d
u
s
t
r
i
a
l
a
p
p
l
i
c
a
t
i
o
n
s
i
n
c
l
u
d
i
n
g
t
h
e
metallurgic,
paint,
g
l
a
s
s
and
e
l
e
c
t
r
o
n
i
c
s
i
n
d
u
s
t
r
i
e
s
,
a
s
w
e
l
l
a
s
f
o
r
medicinal
purposes.

RATIONALE:

Concentrations
of
barium
drinking
water
supplies
generally
range
from
less
than
0.6
ug/
L
t
o
approximately
1
0
ug/
L
with
upper
l
i
m
i
t
s
i
n
a
fewmidwesternandwesternStates
ranging
f
r
o
m
1
0
0
t
o
3,000
ug/
L
(
PHS,
1962/
1963;
K
a
t
z
,
1
9
7
0
;
L
i
t
t
l
e
,
1971).
Barium
e
n
t
e
r
s
t
h
e
body
p
r
i
m
a
r
i
l
y
t
h
r
o
u
g
h
a
i
r
and
w
a
t
e
r
,
s
i
n
c
e
appreciable
amounts
are
not
contained
i
n
foods
(
NAS,
1974).
0
The
fatal
dose
of
barium
for
man
is
reported
to
be
550
to
600
mg.
Ingestion
of
soluble
barium
compounds
may
also
result
in
effects
on
the
gastrointestinal
tract,
causing
vomiting
and
diarrhea,
and
on
the
central
nervous
system,
causing
violent
tonic
and
clonic
spasms
followed
in
some
cases
by
paralysis
(
Browning,
1961;
Patty,
1962,
cited
in
Preliminary
Air
Pollution
Survey
of
Barium
and
Its
Compounds,
1969).
Barium
salts
are
considered
to
be
muscle
stimulants,
especially
for
the
heart
muscle
(
Sollman,
1957).
By
constricting
blood
vessels,
barium
may
cause
an
increase
in
blood
pressure.
On
the
other
hand,
it
is
not
likely
that
barium
accumulates
in
the
bone,
muscle,
kidney
or
other
tissues
because
it
is
readily
excreted
(
Browning,
1961;

McKee
and
Wolf,
1963).

Stokinger
and
Woodward
(
1958)
developed
a
safe
concentration
for
barium
in
drinking
water
based
on
the
limiting
values
for
industrial
atmospheres,
an
estimate
of
the
amount
absorbed
into
the
blood
stream,
and
daily
consumption
of
2
liters
of
water.

From
other
factors
they
arrived
at
a
limiting
concentration
of
2
mg/
L
for
a
healthy
adult
human
population,
to
which
a
safety
factor
was
applied
to
allow
for
any
possible
accumulation
in
the
body.
Since
barium
is
not
removed
by
conventional
water
treatment
processes
and
because
of
the
toxic
effect
on
the
heart
andbloodvessels,
a
limit
oflmg/
L
is
recommended
for
barium
in
domestic
water
supplies.

Experimental
data
indicate
that
the
soluble
barium
concentration
in
fresh
and
marine
water
generally
would
have
to
exceed
5
0
mg/
L
before
toxicity
to
aquatic
life
would
be
expected.

In
most
natural
waters,
there
is
sufficient
sulfate
or
carbonate
to
precipitate
the
barium
present
in
the
water
as
a
virtually
insoluble,
non­
toxic
compound.
Recognizing
that
the
physical
and
chemical
properties
of
barium
generally
wil
P
preclude
the
existence
of
the
toxic
soluble
form
under
usual
marine
and
fresh
water
conditions,
a
restrictive
criterion
f
o
r
aquatic
life
appears
unwarranted.
0
(
QUALITY
CRITERIA
FOR
WATER,
JULY
1976)
PB­
263943
SEE
APPENDIX
C
FOR
METHODOLOGY
CRITERIA:

Aquatic
Life
The
available
data
for
benzene
indicate
that
acute
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
5,300
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
benzene
to
sensitive
freshwater
aquatic
life.

The
available
data
for
benzene
indicate
that
acute
toxicity
to
saltwater
aquatic
life
occurs
at
concentrations
as
low
as
5,100
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
definitive
data
are
available
concerning
the
chronic
toxicity
of
benzene
to
sensitive
saltwater
aquatic
life,
but
adverse
effects
occur
at
concentrations
as
lowas700ug/
Lwitha
fish
species
exposed
for
168
days.

Human
Health
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
benzene
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentrations
should
be
zero,
based
on
the
non
threshold
assumption
Tor
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
and
The
corresponding
recommended
criteria
are
6.6
ug/
L,
0.66
ug/
L,
and
0.066
ug/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
i
water,
the
levels
are
400
ug/
L,
40.0
ug/
L,
and
4.0
ug/
L,

respectively.

(
45
F
.
R
.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
BENZIDINE
Aquatic
Life
The
available
data
for
beniidine
indicate
that
acute
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
2
,
5
0
0
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested
NO
data
are
available
concerning
the
chronic
toxicity
of
benzidine
to
sensitive
freshwater
aquatic
life.

No
saltwater
organisms
have
been
tested
with
benzidine
and
no
statement
can
be
made
concerning
acute
and
chronic
toxicity.

Human
Health
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
benzidine
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentrations
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.

However,
zero
level
may
not
be
attainable
at
the
present
time.

Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
loq5,

The
corresponding
recommended
criteria
are
1.2
ng/
L,
0.12
ng/
L,
and
0.01
ng/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,

excluding
consumption
of
water,
the
levels
are
5
.
3
ng/
L,
0.53
and
ng/
L,
and
0.05
ng/
L,
respectively.

145
F
.
R
.
79318,
November
2
8
.
1980)
0
_#

SEE
APPENDIX
B
'
FOR
METHODOLOGY
BERYLLIUM
CRITERIA:

Aquatic
Life
The
available
data
for
beryllium
indicate
that
acute
and
chronic
toxicity
to
freshwater
aquatic
life
occur
at
concentrations
as
low
as
130
and
5.3
ug/
L,
respectively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
Hardness
has
a
substantial
effect
on
acute
toxicity.

The
limited
saltwater
data
base
available
for
beryllium
does
not
permit
any
statement
concerning
acute
or
chronic
toxicity.

Human
For
the
maximum
protection
of
carcinogenic
effects
of
exposure
Health
human
health
from
the
potential
to
beryllium
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentration
should
be
zero,
based
on
the
non
threshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
10'­
6t
and
10­
70
The
corresponding
recommended
criteria
are
37
ng/
L,
3.7
ng/
L,
and
0.37
ng/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
641
ng/
L,
64.1
ng/
L,
and
6.41
ng/
L,

respectively.

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
CRITERION:

750
mg/
L
for
long­
term
irrigation
on
sensitive
crops.

INTRODUCTION:

Boron
is
not
found
in
its
elemental
form
in
nature:
it
is
usually
found
as
a
sodium
or
calcium
borate
salt.
Boron
salts
are
used
in
fire
retardants,
the
production
of
glass,
leather
tanning
and
finishing
industries,
cosmetics,
photographic
materials,
metallurgy
and
for
high
energy
rocket
fuels.

Elemental
boron
also
can
be
used
in
nuclear
reactors
for
neutron
absorption.
Borates
are
used
as
"
burnable"
poisons.

RATIONALE
:

Boron
is
an
essential
element­
for
growth
of
plants
but
there
is
no
evidence
that
it
is
required
by
animals,
The
maximum
concentration
found
in
1,546
samples
of
river
and
lake
waters
from
various
parts
of
the
United
States
was
5.0
mg/
L;
the
mean
Value
was
0.1
mg/
L
(
Kopp
and
Kroner,
1967).
Ground
waters
could
contain
substantially
higher
concentrations
at
certain
places.

The
concentration
in
seawater
is
reported
as
4.5
mg/
L
in
the
form
of
borate
(
NAS,
1974).
Naturally
occurring
concentrations
of
boron
should
have
no
effects
on
aquatic
life.

The
minimum
lethal
dose
for
minnows
exposed
to
boric
acid
at
20
OC
for
6
hours
was
reported
to
be
18,000
to
19,000
mg/
L
in
distilled
water
and
19,000
to
19,500
mg/
L
in
hard
water
(
Le
clerc
and
Devlaminck,
1955:
Le
Clerc,
1960).
I
n
t
h
e
d
a
i
r
y
cow,
1
6
t
o
2
0
g/
day
of
b
o
r
i
c
acid
f
o
r
4
0
days
produced
no
ill
e
f
f
e
c
t
s
(
McKee
and
Wolf,
1963).

S
e
n
s
i
t
i
v
e
crops
have
shown
t
o
x
i
c
effects
a
t
1000
ug/
L
o
r
less
of
boron
(
Richards,
1954).
Bradford
(
1
9
6
6
)
,
i
n
a
review
of
boron
d
e
f
i
c
i
e
n
c
i
e
s
and
t
o
x
i
c
i
t
i
e
s
,
s
t
a
t
e
d
t
h
a
t
when
t
h
e
boron
c
o
n
c
e
n
t
r
a
t
i
o
n
i
n
i
r
r
i
g
a
t
i
o
n
waters
w
a
s
g
r
e
a
t
e
r
than
0.75
ug/
L,

some
s
e
n
s
i
t
i
v
e
p
l
a
n
t
s
such
a
s
c
i
t
r
u
s
began
t
o
show
i
n
j
u
r
y
.

Biggar
and
Fireman
(
1960)
showed
t
h
a
t
w
i
t
h
n
e
u
t
r
a
l
and
a
l
k
a
l
i
n
e
s
o
i
l
s
of
h
i
g
h
absorption
c
a
p
a
c
i
t
i
e
s
,
water
containing
2
ug/
L
boron
might
be
used
f
o
r
some
t
i
m
e
without
i
n
j
u
r
y
t
o
s
e
n
s
i
t
i
v
e
p
l
a
n
t
s
.
T
h
e
c
r
i
t
e
r
i
o
n
of
750
ug/
L
i
s
t
h
o
u
g
h
t
t
o
p
r
o
t
e
c
t
s
e
n
s
i
t
i
v
e
crops
during
long­
term
i
r
r
i
g
a
t
i
o
n
.

(
QUALITY
CRITERIA
FOR
WATER,
JULY
1976)
PB­
263943
SEE
APPENDIX
C
FOR
METHODOLOGY
CADMIUM
AQUATIC
LIFE
SUMMARY:
­
Freshwater
acute
values
for
cadmium
are
available
for
species
in
4
4
genera
and
range
from
1.0
ug/
L
for
rainbow
trout
to
28,000
ug/
L
for
a
mayfly.
The
antagonistic
effect
of
hardness
on
acute
toxicity
has
been
demonstrated
with
five
species.
Chronic
tests
have
been
conducted
on
cadmium
with
12
freshwater
fish
species
and
4
invertebrate
species
with
chronic
values
ranging
from
0.15
ug/
L
for
_­
Daphnia
­­
mas5
­
to
156
ug/
L
for
the
Atlantic
salmon.

Acute­
chronic
ratios
are
available
for
eight
species
and
range
from
0.9021
for
the
chinook
salmon
to
433.8
for
the
flagfish.

Freshwater
aquatic
plants
are
affected
by
cadmium
at
concentrations
ranging
from
2
to
7,400
ug/
L.
These
values
are
in
the
same
range
as
the
acute
toxicity
values
for
fish
and
invertebrate
species,
and
are
considerably
above
the
chronic
values.
Bioconcentration
factors
(
BCFs)
for
cadmium
in
fresh
water
range
from
164
to
4,190
for
invertebrates
and
from
3
to
2,213
for
fishes.

Saltwater
acute
values
for
cadmium
and
five
species
of
fishes
range
from
577
ug/
L
for
larval
Atlantic
silverside
to
114,000
ug/
L
for
juvenile
mummichog.
Acute
values
for
30
species
of
invertebrates
range
from
15.5
ug/
L
for
a
mysid
to
135,000
ug/
L
for
an
oligochaete
worm.
The
acute
toxicity
of
cadmium
generally
increases
as
salinity
decreases.
The
effect
of
temperature
seems
to
be
species­
specif
ic.
Two
life­
cycle
tests
with
­
Mysidopsis
bahia
under
different
test
conditions
resulted
in
similar
chronic
values
of
8.2
and
7.1
ug/
L,
but
the
acute­
chronic
ratios
were
1.9
and
15,
respectively.
The
acute
values
appear
to
0
reflect
effects
of
salinity
and
temperature,
whereas
the
few
available
chronic
values
apparently
do
not.
A
life­
cycle
test
with
Mysidopsis
bigelowi
also
resulted
in
a
chronic
value
of
7.1
ug/
L
and
an
acute­
chronic
ratio
of
15.

and
macroalgae
revealed
effects
at
22.8
to
860
ug/
L.
Studies
with
microalgae
BCFs
determined
with
a
variety
of
saltwater
invertebrates
ranged
from
5
to
3,160.
BCFs
for
bivalve
molluscs
were
above
1,000
in
long
exposures,
with
no
indication
that
steady­
state
had
been
reached.
Cadmium
mortality
is
cumulative
for
exposure
periods
beyond
4
days.
Chronic
cadmium
exposure
resulted
in
significant
effects
on
the
growth
of
bay
scallops
at
78
ug/
L
and
on
reproduction
of
a
copepod
at
44
ug/
L.

NATIONAL
CRITERIA:
1
The
procedures
described
in
the
Guidelines
for
Deriving
Numerical
National
Water
Quality
Criteria
for
the
Protection
of
Aquatic
Organisms
and
Their
Uses
indicate
that,
except
possibly
where
a
locally
important
species
is
very
sensitive,
freshwater
aquatic
organisms
and
their
uses
should
not
be
affected
unacceptably
if
the
4­
day
average
concentration
(
in
ug/
L)
of
cadmium
does
not
exceed
the
numerical
value
given
by
.(
0.7852
[
ln(
hardness)
1­
3.490)
more
than
once
every
3
years
on
the
average
and
if
the
one­
hour
average
concentration
(
in
ug/
L)
does
not
exceed
the
numerical
value
given
by
.(
1.128[
ln(
hardness)]­

3.828)
more
than
once
every
3
years
on
the
average.
For
example,
at
hardnesses
of
50,
100,
and
200
mg/
L
as
CaC03
the
4­

day
average
concentrations
of
cadmium
are
0.66,
1.1,
and
2.0
ug/
L,
respectively,
and
the
1­
hour
average
concentrations
are
1.8,
3.9
and
8.6
ug/
L.
I
f
brook
t
r
o
u
t
,
brown
t
r
o
u
t
,
and
s
t
r
i
p
e
d
b
a
s
s
a
r
e
a
s
s
e
n
s
i
t
i
v
e
a
s
some
d
a
t
a
i
n
d
i
c
a
t
e
,
t
h
e
y
might
n
o
t
be
protected
by
t
h
i
s
c
r
i
t
e
r
i
o
n
.

T
h
e
procedures
described
i
n
t
h
e
G
u
i
d
e
l
i
n
e
s
i
n
d
i
c
a
t
e
t
h
a
t
,

e
x
c
e
p
t
p
o
s
s
i
b
l
y
where
a
l
o
c
a
l
l
y
important
s
p
e
c
i
e
s
i
s
v
e
r
y
s
e
n
s
i
t
i
v
e
,
saltwater
aquatic
organisms
and
t
h
e
i
r
uses
should
not
be
a
f
f
e
c
t
e
d
unacceptably
i
f
t
h
e
4­
day
a
v
e
r
a
g
e
c
o
n
c
e
n
t
r
a
t
i
o
n
o
f
cadmium
does
not
exceed
9.3
ug/
L
more
than
once
every
3
years
on
t
h
e
average
and
i
f
t
h
e
1­
hour
a
v
e
r
a
g
e
c
o
n
c
e
n
t
r
a
t
i
o
n
does
n
o
t
exceed
43
ug/
L
more
than
once
every
3
years
on
t
h
e
average.
The
l
i
t
t
l
e
information
t
h
a
t
is
a
v
a
i
l
a
b
l
e
concerning
t
h
e
s
e
n
s
i
t
i
v
i
t
y
of
t
h
e
American
l
o
b
s
t
e
r
t
o
cadmium
i
n
d
i
c
a
t
e
s
t
h
a
t
t
h
i
s
important
s
p
e
c
i
e
s
might
n
o
t
be
p
r
o
t
e
c
t
e
d
by
t
h
i
s
c
r
i
t
e
r
i
o
n
.
I
n
a
d
d
i
t
i
o
n
,

d
a
t
a
s
u
g
g
e
s
t
t
h
a
t
t
h
e
a
c
u
t
e
t
o
x
i
c
i
t
y
of
cadmium
is
s
a
l
i
n
i
t
y
dependent:
therefore,
the
1­
hour
average
concentration
might
be
u
n
d
e
r
p
r
o
t
e
c
t
i
v
e
a
t
low
s
a
l
i
n
i
t
i
e
s
and
o
v
e
r
p
r
o
t
e
c
t
i
v
e
a
t
h
i
g
h
s
a
l
i
n
i
t
i
e
s
.

EPA
b
e
l
i
e
v
e
s
t
h
a
t
a
measurement
such
as
t'acid­
solublell
would
p
r
o
v
i
d
e
a
more
s
c
i
e
n
t
i
f
i
c
a
l
l
y
c
o
r
r
e
c
t
b
a
s
i
s
upon
which
t
o
e
s
t
a
b
l
i
s
h
c
r
i
t
e
r
i
a
f
o
r
metals.
The
c
r
i
t
e
r
i
a
w
e
r
e
developed
on
t
h
i
s
basis.
However,
a
t
t
h
i
s
t
i
m
e
,
no
EPA­
approved
methods
f
o
r
such
a
measurement
are
a
v
a
i
l
a
b
l
e
t
o
implement
t
h
e
c
r
i
t
e
r
i
a
through
t
h
e
r
e
g
u
l
a
t
o
r
y
programs
of
t
h
e
Agency
and
t
h
e
S
t
a
t
e
s
.

The
Agency
is
considering
development
and
approval
of
methods
f
o
r
a
measurement
such
as
acid­
soluble.
U
n
t
i
l
a
v
a
i
l
a
b
l
e
,
however,

EPA
recommends
applying
t
h
e
criteria
using
t
h
e
t
o
t
a
l
recoverable
method.
T
h
i
s
has
two
impacts:
(
1)
c
e
r
t
a
i
n
species
of
some
metals
0
cannot
be
analyzed
directly
because
the
total
recoverable
method
does
not
distinguish
between
individual
oxidation
states,
and
(
2)

these
criteria
may
be
overly
protective
when
based
on
the
total
recoverable
method.

The
recommended
exceedence
frequency
of
3
years
is
the
Agency's
best
scientific
judgment
of
the
average
amount
of
time
it
will
take
an
unstressed
system
to
recover
from
a
pollution
event
in
which
exposure
to
cadmium
exceeds
the
criterion.
A
stressed
system,
for
example,
one
in
which
several
outfalls
occur
in
a
limited
area,
would
be
expected
to
require
more
time
for
recovery.
The
resilience
of
ecosystems
and
their
ability
to
recover
differ
greatly,
however,
and
site­
specific
criteria
may
be
established
if
adequate
justification
is
provided.
I
The
use
of
criteria
in
designing
waste
treatment
facilities
requires
the
selection
of
an
appropriate
wasteload
allocation
model.
Dynamic
models
are
preferred
for
the
application
of
these
criteria.
Limited
data
or
other
factors
may
make
their
use
impractical,
in
which
case
one
should
rely
on
a
steady­
state
model.
The
Agency
recommends
the
interim
use
of
145
or
lQlO
for
Criterion
Maximum
Concentration
design
flow
and
745
or
7410
for
the
Criterion
Continuous
Concentration
design
flow
in
steady­

state
models
for
unstressed
and
stressed
systems,
respectively.

These
matters
are
discussed
in
more
detail
in
the
Technical
Support
Document
for
Water
Quality­
Based
Toxics
Control
(
U.
S.

EPA,
1985).

HUMAN
HEALTH
CRITERIA:

The
ambient
water
quality
criterion
for
cadmium
is
recommended
to
be
identical
to
the
existing
drinking
water
s
t
a
n
d
a
r
d
which
i
s
1
0
ug/
L.
A
n
a
l
y
s
i
s
of
t
h
e
t
o
x
i
c
effects
d
a
t
a
r
e
s
u
l
t
e
d
i
n
a
c
a
l
c
u
l
a
t
e
d
l
e
v
e
l
which
is
p
r
o
t
e
c
t
i
v
e
of
human
0
h
e
a
l
t
h
a
g
a
i
n
s
t
t
h
e
i
n
g
e
s
t
i
o
n
o
f
c
o
n
t
a
m
i
n
a
t
e
d
w
a
t
e
r
and
c
o
n
t
a
m
i
n
a
t
e
d
a
q
u
a
t
i
c
o
r
g
a
n
i
s
m
s
.
The
c
a
l
c
u
l
a
t
e
d
v
a
l
u
e
is
comparable
t
o
t
h
e
present
standard.
For
t
h
i
s
reason
a
s
e
l
e
c
t
i
v
e
c
r
i
t
e
r
i
o
n
based
on
exposure
s
o
l
e
l
y
from
consumption
of
6.5
grams
of
aquatic
organisms
was
not
derived.

(
45
F.
R.
79318
Nov.
28,1980)
(
50
F
.
R
.
30784,
J
u
l
y
29,
1985)
SEE
APPENDIX
A
FOR
METHODOLOGY
CARBON
TETRACHLORIDE
CRITERIA:

Aquatic
Life
The
available
data
for
carbon
tetrachloride
indicate
that
acute
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
35,200
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
carbon
tetrachloride
to
sensitive
freshwater
aquatic
life.

The
available
data
for
carbon
tetrachloride
indicate
that
acute
toxicity
to
saltwater
aquatic
life
occurs
at
concentrations
as
low
as
50,000
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
carbontetrachloride
to
sensitive
saltwater
aquatic
life.

Human
Health
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
carbon
tetrachloride
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentrations
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
and
lo­'.
The
corresponding
recommended
criteria
are
4.0
ug/
L,
0.40ug/
L,
and
0.04
ug/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
I
water,
the
levels
are
69.4
ug/
L,
6.94
ug/
L,
and
0.69
ug/
L
respectively.
0
(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
CRITERIA:

0
CHLORDANE
Aquatic
Life
For
chlordane
the
criterion
to
protect
freshwat
tic
life
as
derived
using
the
Guidelines
is
0.0043
ug/
L
as
a
24­
hOUr
average,
and
the
concentration
should
not
exceed
2.4
ug/
L
at
any
time.

For
chlordane
the
criterion
to
protect
saltwater
aquatic
life
as
derived
using
the
Guidelines
is
0.0040
ug/
L
as
a
24­
hour
average,
and
the
concentration
should
not
exceed
0.09
ug/
L
at
any
time.

Human
Health
r
aqu
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
chlordane
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentration
should
be
zero
based
on
the
nonthreshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
loq5,
lom6,
and
lo­'.
The
corresponding
recommended
criteria
are
4.6
ng/
L,
0.46
ng/
L,
and
0.046
ng/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
4.8
ng/
L,
0.48
ng/
L,
and
0.048
ng/
L,

respectively.

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
CHLORINATED
BENZENES
CRITERIA:
Aquatic
­
Life
The
available
data
for
chlorinated
benzenes
indicate
that
acute
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
250
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
the
more
toxic
of
the
chlorinated
benzenes
to
sensitive
freshwater
aquatic
life,
but
toxicity
occurs
at
concentrations
as
low
as
50
ug/
L
for
a
fish
species
exposed
for
7.5
days.

The
available
data
for
chlorinated
benzenes
indicate
that
acute
and
chronic
toxicity
to
saltwater
aquatic
life
occur
at
concentrations
as
low
as
160
and
129
ug/
L,
respectively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.

Human
Health
Xonochlorobenzene
For
comparison
purposes,
two
approaches
were
used
to
derive
criterion
levels
for
monochlorobenzene.
Based
on
available
toxicity
data,
for
the
protection
of
public
health
the
derived
level
is
4
8
8
ug/
L.
Using
available
organoleptic
data,
to
control
undesirable
taste
and
odor
quality
of
ambient
water
the
estimated
level
is
20
ug/
L.
It
should
be
recognized
that
organoleptic
data
have
limitations
as
a
basis
for
establishing
water
quality
criteria,
and
have
no
demonstrated
relationship
to
potential
adverse
human
health
effects.
a
(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
0
CRITERIA:
CHLQRINATED
ETHANES
Aquatic
­
Life
The
available
freshwater
data
for
chlorinated
ethanes
indicate
that
toxicity
increases
greatly
with
increasing
chlorination,
and
that
acute
toxicity
occurs
at
concentrations
as
low
as
118,000
ug/
L
for
l,
Z­
dichloroethane,
18,000
ug/
L
for
two
trichloroethanes,
9,320
ug/
L
for
two
tetrachloroethanes,
7,240
ug/
L
for
pentachloroethane,
and
980
ug/
L
for
hexachloroethane.

Chronic
toxicity
occurs
at
concentrations
as
low
as
20,000
ug/
L
for
1,2­
dichloroethane,
9,400
ug/
L
for
l,
l,
Z­
trichloroethane,

2,400
ug/
L
for
1,1,2,2­
tetrachloroethane,
1,100
ug/
L
for
pentachloroethane,
and
540
ug/
L
for
hexachloroethane.
Acute
and
chronic
toxicity
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
0
The
available
saltwater
data
for
chlorinated
ethanes
indicate
that
toxicity
increases
greatly
with
increasing
chlorination
and
that
acute
toxicity
to
fish
and
invertebrate
species
occurs
at
concentrations
as
low
as
113,000
ug/
L
for
lI2­
dichloroethane,

31,200
ug/
L
for
l,
l,
l­
trichloroethane,
9,020
ug/
L
for
1,1,2,2­
tetrachloroethane,
390
ug/
L
for
pentachloroethane,

and
940
ug/
L
for
hexachloroethane.
Chronic
toxicity
occurs
at
concentrations
as
low
as
281
ug/
L
for
pentachloroethane.
Acute
and
chronic
toxicity
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
Human
Health
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
1,2­
dichloroethane
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentration
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
and
The
corresponding
recommended
criteria
are
9.4
ug/
L,
0.94
ug/
L,
and
0.094
ug/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,'
the
levels
are
2,430
ug/
L,
243
ug/
L,
and
24.3
ug/
L,

respectively.

For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
l,
l,
2­
trichloroethane
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentration
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.
However,

zero
level
may
not
be
attainable
at
the
present
time.
Therefore,

the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
and
lo­'.

The
corresponding
recommended
criteria
are
6.0
ug/
L,
0.6
ug/
L,

and
0.06
ug/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
418
ug/
L,
41.8
ug/
L,
and
4.18
ug/
L,

respectively.
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
1,1,2,2­
tetrachloroethane
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentration
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.
However,

zero
level
may
not
be
attainable
at
the
present
time.
Therefore,

the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
and
The
corresponding
recommended
criteria
are
1.7
ug/
L,
0.17
ug/
L,

and
0.017
ug/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
107
ug/
L,
10.7
ug/
L,
and
1.07
ug/
L,

respectively.
I
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
hexachloroethane
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentration
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
and
loe7.
The
corresponding
recommended
criteria
are
19
ug/
L,
1.9
ug/
L,
and
0.19
ug/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
87.4
ug/
E,
8.74
ug/
L,
and
0.87
ug/
L,

respectively.

0
J
For
the
protection
of
human
health
from
the
toxic
properties
of
l,
l,
l­
trichloroethane
ingested
through
water
and
contaminated
aquatic
organisms,
the
ambient
water
criterion
is
determined
to
be
18.4
mg/
L.

For
the
protection
of
human
health
from
the
toxic
properties
of
l,
l,
l­
trichloroethane
ingested
through
contaminated
aquatic
organisms
alone,
the
ambient
water
criterion
is
determined
to
be
1.03
ug/
l.

Because
of
insufficient
available
data
 or
monochloroethane,

1,
l­
dichloroethane,
1,1,
l12­
tetrachloroethane,
and
pentachloroethane,
satisfactory
criteria
cannot
be
derived
at
this
time,
using
the
present
guidelines.

(
4
5
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
CRITERIA:
CHLORINATED
NAPHTHALENES
Aquatic
Life
The
available
data
for
chlorinated
naphthalenes
indicate
that
acute
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
l
o
w
as
1,600
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
chlorinated
naphthalenes
to
sensitive
freshwater
aquatic
life.

The
available
data
for
chlorinated
naphthalenes,
indicate
that
acute
toxicity
to
saltwater
aquatic
life
occurs
at
concentrations
as
low
as
7.5
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
chlorinated
naphthalenes
to
sensitive
saltwater
aquatic
life.
0
Human
Health
Using
the
present
guidelines,
a
satisfactory
criterion
cannot
be
derived
at
this
time
because
of
insufficient
available
data
for
chlorinated
naphthalenes.

(
4
5
F.
R.
79318,
November
2
8
,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
0
2
CHLORINE
SUMMARY
:

T
h
i
r
t
y­
t
h
r
e
e
freshwater
s
p
e
c
i
e
s
i
n
28
genera
have
been
exposed
t
o
TRC
and
t
h
e
a
c
u
t
e
v
a
l
u
e
s
range
from
28
ug/
L
f
o
r
­
D
a
e
n
i
a
­­
magna
t
o
710
Ug/
L
f
o
r
t
h
e
t
h
r
e
e
s
p
i
n
e
s
t
i
c
k
l
e
b
a
c
k
.
F
i
s
h
and
i
n
v
e
r
t
e
b
r
a
t
e
species
had
s
i
m
i
l
a
r
ranges
of
s
e
n
s
i
t
i
v
i
t
y
.

Freshwater
c
h
r
o
n
i
c
t
e
s
t
s
h
a
v
e
been
c
o
n
d
u
c
t
e
d
w
i
t
h
two
i
n
v
e
r
t
e
b
r
a
t
e
and
one
f
i
s
h
s
p
e
c
i
e
s
and
t
h
e
c
h
r
o
n
i
c
v
a
l
u
e
s
f
o
r
these
three
s
p
e
c
i
e
s
ranged
from
l
e
s
s
t
h
a
n
3.4
t
o
26
ug/
L,
w
i
t
h
acute­
chronic
r
a
t
i
o
s
from
3.7
t
o
g
r
e
a
t
e
r
than
78.

The
acute
s
e
n
s
i
t
i
v
i
t
i
e
s
of
24
species
of
saltwater
animals
i
n
2
1
genera
have
been
determined
f
o
r
CPO,
and
t
h
e
LC50
range
from
26
ug/
L
f
o
r
t
h
e
e
a
s
t
e
r
n
o
y
s
t
e
r
t
o
1,418
ug/
L
f
o
r
a
mixture
of
two
shore
crab
species.
T
h
i
s
range
is
very
s
i
m
i
l
a
r
t
O
t
h
a
t
observed
with
f
r
e
s
h
w
a
t
e
r
s
p
e
c
i
e
s
,
and
f
i
s
h
and
i
n
v
e
r
t
e
b
r
a
t
e
s
p
e
c
i
e
s
had
s
i
m
i
l
a
r
s
e
n
s
i
t
i
v
i
t
i
e
s
.
Only
one
chronic
test
has
been
conducted
w
i
t
h
a
s
a
l
t
w
a
t
e
r
s
p
e
c
i
e
s
,
Menidia
E
e
n
i
n
s
u
l
a
e
­­
I­__
I
and
i
n
t
h
i
s
t
e
s
t
t
h
e
acute
chronic
r
a
t
i
o
w
a
s
1.162.

The
a
v
a
i
l
a
b
l
e
data
i
n
d
i
c
a
t
e
t
h
a
t
a
q
u
a
t
i
c
p
l
a
n
t
s
are
more
r
e
s
i
s
t
a
n
t
t
o
c
h
l
o
r
i
n
e
than
f
i
s
h
and
i
n
v
e
r
t
e
b
r
a
t
e
species.

NATIONAL
CRITERIA:

The
procedures
d
e
s
c
r
i
b
e
d
i
n
t
h
e
G
u
i
d
e
l
i
n
e
s
f
o
r
D
e
r
i
v
i
n
g
Numerical
National
Water
Q
u
a
l
i
t
y
C
r
i
t
e
r
i
a
f
o
r
t
h
e
Protection
of
Aquatic
Organisms
and
Their
U
s
e
s
i
n
d
i
c
a
t
e
t
h
a
t
,
except
possibly
where
a
l
o
c
a
l
l
y
important
species
is
very
s
e
n
s
i
t
i
v
e
,
freshwater
a
q
u
a
t
i
c
organisms
and
t
h
e
i
r
u
s
e
s
s
h
o
u
l
d
n
o
t
b
e
a
f
f
e
c
t
e
d
unacceptably
i
f
t
h
e
4­
day
average
concentration
of
t
o
t
a
l
residual
chlorine
does
not
exceed
11
ug/
L
more
than
once
every
3
years
on
J­­

t
h
e
average
and
i
f
t
h
e
1­
hour
average
c
o
n
c
e
n
t
r
a
t
i
o
n
does
n
o
t
exceed
19
ug/
L
more
than
once
every
3
years
on
t
h
e
average.

T
h
e
procedures
described
i
n
t
h
e
Guidelines
i
n
d
i
c
a
t
e
t
h
a
t
,

except
p
o
s
s
i
b
l
y
where
a
l
o
c
a
l
l
y
important
s
p
e
c
i
e
s
i
s
very
s
e
n
s
i
t
i
v
e
,
saltwater
aquatic
organisms
and
t
h
e
i
r
uses
should
not
be
a
f
f
e
c
t
e
d
unacceptably
i
f
t
h
e
4­
day
average
c
o
n
c
e
n
t
r
a
t
i
o
n
of
chlorine­
produced
o
x
i
d
a
n
t
s
does
n
o
t
exceed
7.5
ug/
L
more
than
once
e
v
e
r
y
3
y
e
a
r
s
on
t
h
e
average
and
i
f
t
h
e
one­
hour
average
c
o
n
c
e
n
t
r
a
t
i
o
n
does
n
o
t
exceed
13
ug/
L
more
t
h
a
n
once
every
3
years
on
t
h
e
average.

The
recommended
exceedence
frequency
o
f
3
y
e
a
r
s
is
t
h
e
Agency's
best
s
c
i
e
n
t
i
f
i
c
judgment
of
t
h
e
average
amount
of
t
i
m
e
,

it
w
i
l
l
take
an
u
n
s
t
r
e
s
s
e
d
system
t
o
r
e
c
o
v
e
r
f
r
o
m
a
p
o
l
l
u
t
i
o
n
e
v
e
n
t
i
n
which
exposure
t
o
c
h
l
o
r
i
n
e
exceeds
t
h
e
c
r
i
t
e
r
i
o
n
.
A
stressed
system,
f
o
r
example,
one
i
n
which
several
o
u
t
f
a
l
l
s
occur
i
n
a
l
i
m
i
t
e
d
a
r
e
a
,
would
be
expected
t
o
r
e
q
u
i
r
e
more
t
i
m
e
f
o
r
recovery.
The
r
e
s
i
l
i
e
n
c
e
of
ecosystems
and
t
h
e
i
r
a
b
i
l
i
t
y
t
o
recover
differ
g
r
e
a
t
l
y
,
however,
and
s
i
t
e­
s
p
e
c
i
f
i
c
c
r
i
t
e
r
i
a
may
be
established
i
f
adequate
j
u
s
t
i
f
i
c
a
t
i
o
n
is
provided.

The
use
of
c
r
i
t
e
r
i
a
i
n
designing
waste
treatment
f
a
c
i
l
i
t
i
e
s
r
e
q
u
i
r
e
s
t
h
e
s
e
l
e
c
t
i
o
n
o
f
an
a
p
p
r
o
p
r
i
a
t
e
wasteload
a
1
l
o
c
a
t
i
o
n
model.
Dynamic
models
a
r
e
preferred
f
o
r
t
h
e
application
of
these
c
r
i
t
e
r
i
a
.
L
i
m
i
t
e
d
d
a
t
a
o
r
o
t
h
e
r
f
a
c
t
o
r
s
may
make
t
h
e
i
r
u
s
e
a
i
m
p
r
a
c
t
i
c
a
l
,
i
n
which
c
a
s
e
one
should
r
e
l
y
on
a
s
t
e
a
d
y­
s
t
a
t
e
model.
The
Agency
recommends
t
h
e
i
n
t
e
r
i
m
use
of
145
o
r
lQlO
f
o
r
Criterion
Maximum
Concentration
design
flow
and
745
o
r
7Q10
f
o
r
the
Criterion
Continuous
Concentration
design
flow
in
steady­

state
models
for
unstressed
and
stressed
systems,
respectively.

These
matters
are
discussed
in
more
detail
in
the
Technical
Support
Document
f
o
r
Water
Quality­
Based
Toxics
Control
(
U
.
S
.

EPA,
1985).

(
50
F.
R.
30784,
July
29,
1985)
SEE
APPENDIX
A
FOR
METHODOLOGY
CHLORINATED
PHENOLS
Aquatic
­
Life
The
available
freshwater
data
for
chlorinated
phenols
indicate
that
toxicity
generally
increases
with
increasing
chlorination,
and
that
acute
toxicity
occurs
at
concentrations
as
low
as
30
ug/
L
for
4­
chloro­
3­
methylphenol
to
greater
than
5
0
0
,
0
0
0
ug/
L
for
other
compounds.
Chronic
toxicity
occurs
at
concentrations
as
low
as
970
ug/
L
for
2,4,6­
trichlorophenol.

Acute
and
chronic
toxicity
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.

The
available
saltwater
data
for
chlorinated
phenols
indicate
that
toxicity
generally
increases
with
increasing
chlorination
and
that
acute
toxicity
occurs
at
concentrations
as
low
as
4
4
0
ug/
L
for
2,3,5,6­
tetrachlorophenol
and
29,700
ug/
L
for
4
­

chlorophenol.
Acute
toxicity
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
chlorinated
phenols
to
sensitive
saltwater
aquatic
life.

Human
Health
Sufficient
data
are
not
available
f
o
r
3­
chlorophenol
to
derive
a
level
which
would
protect
against
the
potential
toxicity
of
this
compound.
Using
available
organoleptic
data,
to
control
undesirable
taste
and
odor
qualities
of
ambient
water,

the
estimated
level
is
0,1
ug/
L.
It
should
be
recognized
that
organoleptic
data
have
limitations
as
a
basis
for
establishing
a
water
quality
criterion,
and
have
no
demonstrated
a
/
relationship
to
potential
adverse
human
health
effects.

Sufficient
data
are
not
available
for
4­
chlorophenol
to
derive
a
level
which
would
protect
against
the
potential
toxicity
of
this
compound.
Using
available
organoleptic
data,
to
control
undesirable
taste
and
odor
qualities
of
ambient
water
the
estimated
level
is
0.1
ug/
L.
It
should
be
recognized
that
organoleptic
data
have
limitations
as
a
basis
for
establishing
a
water
quality
criterion,
and
have
no
demonstrated
relationship
to
potential
adverse
human
health
effects.

Sufficient
data
are
not
available
for
2,3­
dichlorophenol
to
derive
a
level
which
would
protect
against
the
potential
toxicity
of
this
compound.
Using
available
organoleptic
data,
to
control
undesirable
taste
and
odor
qualities
of
ambient
water
the
estimated
level
is
0.04
ug/
L.
It
should
be
recognized
that
organoleptic
data
have
limitations
as
a
basis
for
establishing
a
water
quality
criterion,
and
have
no
demonstrated
relationship
to
potential
adverse
human
health
effects.

Sufficient
data
are
not
available
for
2,5­
dichlorophenol
to
derive
a
level
which
would
protect
against
the
potential
toxicity
of
this
compound.
Using
available
organoleptic
data,
to
control
undesirable
taste
and
odor
qualities
of
ambient
water
the
estimated
level
is
0.5
ug/
L.
It
should
be
recognized
that
organoleptic
data
have
limitations
as
a
basis
for
establishing
a
water
quality
criterion,
and
have
no
demonstrated
relationship
to
potential
adverse
human
health
effects.

Sufficient
data
are
not
available
for
2,6­
dichlorophenol
to
derive
a
level
which
would
protect
against
the
potential
toxicity
of
this
compound.
Using
available
organoleptic
data,
to
control
undesirable
taste
and
odor
qualities
of
ambient
water
the
estimated
level
is
0.2
ug/
L.
It
should
be
recognized
that
organoleptic
data
have
limitations
as
a
basis
for
establishing
a
water
quality
criterion,
and
have
no
demonstrated
relationship
to
potential
adverse
human
health
effects.

Sufficient
data
are
not
available
for
3,4­
dichlorophenol
to
derive
a
level
which
would
protect
against
the
potential
toxicity
of
this
compound.
Using
available
organoleptic
data,
to
control
undesirable
taste
and
odor
qualities
of
ambient
water
the
estimated
level
is
0.3
ug/
L.
It
should
be
recognized
that
organoleptic
data
have
limitations
as
a
basis
for
establishing
a
water
quality
criterion,
and
have
no
demonstrated
relationship
to
potential
adverse
human
health
ef'fects.

For
comparison
purposes,
two
approaches
were
used
to
derive
criterion
levels
for
2,4,5­
trichlorophenol.
Based
on
available
toxicity
data,
to
protect
public
health
the
derived
level
is
2.6
mg/
L.
Using
available
organoleptic
data,
to
control
undesirable
taste
and
odor
quality
of
ambient
water
the
estimated
level
is
1.0
ug/
L.
It
should
be
recognized
that
organoleptic
data
have
limitations
as
a
basis
for
establishing
a
water
quality
criterion,
and
have
no
demonstrated
relationship
to
potential
adverse
human
health
effects.

For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
2,4,6­
trichlorophenoI
through
the
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentration
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
10­
5,
10­
6,
and
10­
7.
The
corresponding
recommended
criteria
are
12
ug/
L,
1.2
ug/
L,
and
0.12
ug/
L,

respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
36
ug/
L,
3.6
ug/
L,
and
0.36
ug/
L,
respectively.
Using
available
organoleptic
data,
to
control
undesirable
taste
and
odor
qualities
of
ambient
water
the
estimated
level
is
2
ug/
L.

It
should
be
recognized
that
organoleptic
data
have
limitations
as
a
basis
for
establishing.
a
water
quality
criterion,
and
have
no
demonstrated
relationship
to
potential
adverse
human
health
effects.

Sufficient
data
are
not
available
for
2,3,4,6­

tetrachlorophenol
to
derive
a
level
which
would
protect
against
the
potential
toxicity
of
this
compound.
Using
available
organoleptic
data,
to
control
undesirable
taste
and
odor
qualities
of
ambient
water
the
estimated
level
is
1.0
ug/
L.
It
should
be
recognized
that
organoleptic
data
have
limitations
as
a
basis
for
establishing
a
water
quality
criterion,
and
have
demonstrated
relationship
to
potential
adverse
human
health
effects.

Sufficient
data
are
not
available
for
2­
methyl­
4­
chlorophenol
to
derive
a
criterion
level
which
would
protect
against
any
potential
toxicity
of
this
compound.
Using
available
organoleptic
data,
to
control
undesirable
taste
and
odor
qualities
of
ambient
water
the
estimated
level
is
1,800
ug/
L.
It
should
be
recognized
t
h
a
t
organoleptic
data
have
l
i
m
i
t
a
t
i
o
n
s
a
s
a
b
a
s
i
s
f
o
r
e
s
t
a
b
l
i
s
h
i
n
g
a
water
q
u
a
l
i
t
y
c
r
i
t
e
r
i
o
n
and
have
no
demonstrated
r
e
l
a
t
i
o
n
s
h
i
p
t
o
p
o
t
e
n
t
i
a
l
a
d
v
e
r
s
e
human
h
e
a
l
t
h
effects
.

S
u
f
f
i
c
i
e
n
t
data
a
r
e
not
a
v
a
i
l
a
b
l
e
f
o
r
3­
methyl­
4­
chlorophenol
t
o
derive
a
c
r
i
t
e
r
i
o
n
l
e
v
e
l
which
would
p
r
o
t
e
c
t
a
g
a
i
n
s
t
any
p
o
t
e
n
t
i
a
l
t
o
x
i
c
i
t
y
o
f
t
h
i
s
compound.
Using
a
v
a
i
l
a
b
l
e
o
r
g
a
n
o
l
e
p
t
i
c
d
a
t
a
,
t
o
c
o
n
t
r
o
l
u
n
d
e
s
i
r
a
b
l
e
t
a
s
t
e
and
odor
q
u
a
l
i
t
i
e
s
of
ambient
water
t
h
e
estimated
l
e
v
e
l
is
3,000
ug/
L.
It
should
be
recognized
t
h
a
t
organoleptic
data
have
l
i
m
i
t
a
t
i
o
n
s
a
s
a
b
a
s
i
s
f
o
r
e
s
t
a
b
l
i
s
h
i
n
g
a
water
q
u
a
l
i
t
y
c
r
i
t
e
r
i
o
n
,
and
have
no
demonstrated
r
e
l
a
t
i
o
n
s
h
i
p
t
o
p
o
t
e
n
t
i
a
l
a
d
v
e
r
s
e
human
h
e
a
l
t
h
e
f
f
e
c
t
s
,

S
u
f
f
i
c
i
e
n
t
data
a
r
e
not
a
v
a
i
l
a
b
l
e
f
o
r
3­
methyl­
6­
chlorophenol
t
o
d
e
r
i
v
e
a
c
r
i
t
e
r
i
o
n
l
e
v
e
l
which
would
p
r
o
t
e
c
t
a
g
a
i
n
s
t
any
p
o
t
e
n
t
i
a
l
t
o
x
i
c
i
t
y
o
f
t
h
i
s
compound.
Using
a
v
a
i
l
a
b
l
e
o
r
g
a
n
o
l
e
p
t
i
c
d
a
t
a
,
t
o
c
o
n
t
r
o
l
u
n
d
e
s
i
r
a
b
l
e
t
a
s
t
e
and
odor
q
u
a
l
i
t
i
e
s
of
ambient
water
t
h
e
estimated
l
e
v
e
l
is
2
0
ug/
L.
It
should
be
recognized
t
h
a
t
organoleptic
data
have
l
i
m
i
t
a
t
i
o
n
s
a
s
a
b
a
s
i
s
f
o
r
e
s
t
a
b
l
i
s
h
i
n
g
a
water
q
u
a
l
i
t
y
c
r
i
t
e
r
i
o
n
,
and
have
no
demonstrated
r
e
l
a
t
i
o
n
s
h
i
p
t
o
p
o
t
e
n
t
i
a
l
a
d
v
e
r
s
e
human
h
e
a
l
t
h
e
f
f
e
c
t
s
.

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
CRITERIA
:

a
CHLOROALKYL
ETHERS
Aquatic
­
Life
The
available
data
for
chloroalkyl
ethers
indicate
that
acute
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
238,000
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
NO
definitive
data
are
chloroalkyl
ethers
to
sensitive
freshwater
aquatic
life.
available
concerning
the
chronic
toxicity
of
No
saltwater
organism
has
been
tested
with
any
chloroalkyl
ether
and
therefore,
no
statement
can
be
made
concerning
acute
or
chronic
toxicity.

Human
Health
For
the
protection
of
human
health
from
the
toxic
properties
of
bis(
2­
chloroisopropy1)
ether
ingested
through
water
and
contaminated
aquatic
organisms,
the
ambient
water
criterion
is
determined
to
be
34.7
ug/
L.

For
the
protection
of
human
health
from
the
toxic
properties
of
bis(
2­
chloroisopropy1)
ether
ingested
through
contaminated
aquatic
organisms
alone,
the
ambient
water
criterion
is
determined
to
be
4.36
mg/
L.

For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
bis(
chloromethy1)
ether
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentrations
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
10­
5,
and
The
corresponding
recommended
criteria
are
37.6
x
ug/
L,
3.76
x
ug/
L,
and
0.376
x
ug/
L,

respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
18.4
x
ug/
L,
1.84
x
ug/
L,
and
0.184
x
10­
3
ug/
L,
respectively.

For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
bis(
2­
chloroethyl)
ether
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentrations
should
be
zero
based
on
the
nonthreshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
and
10­
7.

The
corresponding
recommended
criteria
are
0.30
ug/
L,
0.030
ug/
L,
and
0.003
ug/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
13.6
ug/
L,
1.36
ug/
L,
and
0.136
ug/
L,
respectively.

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
CRITERIA:
CHLOROFORM
Aquatic
Life
The
available
data
for
chloroform
indicate
that
acute
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
28,900
ug/
L,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
the
three
tested
species
Twenty­
seven­
day
LC50
values
indicate
that
chronic
toxicity
occurs
at
concentrations
as
low
as
1,240
ug/
L,
and
could
occur
at
lower
concentrations
among
species
or
other
life
stages
that
are
more
sensitive
than
the
earliest
life
cycle
stages
of
the
rainbow
trout.
The
data
base
for
saltwater
species
is
limited
to
one
test
and
therefore,
no
statement
can
be
made
concerning
acute
or
1
chronic
toxicity.

Human
Health
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
chloroform
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentrations
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.

However,
zero
level
may
not
be
attainable
at
the
present
time.

Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
loe5,

The
corresponding
recommended
criteria
are
.
1.90
ug/
L,
0.19
ug/
L,
and
0.019
ug/
L,
respectively.
If
these
and
estimates
are
made
for
consumption
of
aquatic
organisms
only,

excluding
consumption
of
water,
the
levels
are
157
ug/
L,
15.7
ug/
L,
and
1.57
ug/
L,
respectively.

(
45
F
.
R
.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
CHMROPHENOXY
HERBICIDES
2,4­
D;
2,4,5­
TP
CRITERIA:

2,4­
D
100
ug/
L
for
domestic
water
supply
(
health)

2,4,5­
TP
1
0
ug/
L
f
o
r
domestic
water
supply
(
health)

RATIONALE:

T
w
o
w
i
d
e
l
y
u
s
e
d
h
e
r
b
i
c
i
d
e
s
a
r
e
2
,
4
­
D
(
2
,
4
­

dichlorophenoxyacetic
a
c
i
d
)
and
2
,
4
,
5
­
T
P
(
s
i
l
v
e
x
)
[
2
­
(
2
'
4
,
5­

trichlorophenoxy)
propionic
acid.
Each
of
t
h
e
s
e
compounds
is
formulated
i
n
a
v
a
r
i
e
t
y
of
s
a
l
t
s
and
esters
t
h
a
t
may
have
a
marked
d
i
f
f
e
r
e
n
c
e
i
n
h
e
r
b
i
c
i
d
a
l
p
r
o
p
e
r
t
i
e
s
,
b
u
t
a
l
l
a
r
e
hydrolyzed
rapidly
to
t
h
e
corresponding
acid
i
n
the
body.

The
subacute
oral
t
o
x
i
c
i
t
y
of
chlorophenoxy
herbicides
has
been
investigated
i
n
a
number
of
species
of
experimental
animals
(
Palmer
and
Radeleff,
1964;
Lehman,
1965).
The
dog
w
a
s
found
t
o
be
s
e
n
s
i
t
i
v
e
and
often
displayed
m
i
l
d
injury
i
n
response
to
doses
of
10
mg/
kg/
day
f
o
r
90
days,
and
s
e
r
i
o
u
s
effects
from
a
dose
of
2
0
mg/
kg/
day
f
o
r
90
days.
Lehman
(
1965)
reported
t
h
a
t
t
h
e
no­

effect
l
e
v
e
l
of
2,4­
D
is
0.5
mg/
kg/
day
i
n
t
h
e
r
a
t
,
and
8.0
mg/
kg/
day
i
n
the
dog.

Data
are
a
v
a
i
l
a
b
l
e
on
t
h
e
t
o
x
i
c
i
t
y
of
2,4­
D
to
man.
A
d
a
i
l
y
dosage
of
500
mg
(
about
7
mg/
kg)
produced
no
apparent
ill
effects
i
n
a
v
o
l
u
n
t
e
e
r
o
v
e
r
a
21­
day
period
(
Kraus,
1
9
4
6
)
.
When
2,4­
D
was
i
n
v
e
s
t
i
g
a
t
e
d
a
s
a
p
o
s
s
i
b
l
e
t
r
e
a
t
m
e
n
t
 or
disseminated
coccidioidomycosis,
t
h
e
p
a
t
i
e
n
t
had
no
s
i
d
e
effects
from
18
intravenous
doses
during
3
3
days;
each
of
t
h
e
l
a
s
t
1
2
doses
i
n
a
t
h
e
series
w
a
s
800
m
g
(
about
15
mg/
kg)
o
r
more,
t
h
e
l
a
s
t
being
2000
mg
(
about
37
mg/
kg)
(
Seabury,
1
9
6
3
)
.
A
1
9
t
h
and
f
i
n
a
l
dose
of
3600
m
g
(
67
mg/
kg)
produced
mild
symptoms.

The
long­
term
no­
e
f
f
e
c
t
s
l
e
v
e
l
s
(
mg/
kg/
day)
are
l
i
s
t
e
d
f
o
r
t
h
e
r
a
t
and
t
h
e
dog.
Those
v
a
l
u
e
s
are
a
d
j
u
s
t
e
d
by
a
f
a
c
t
o
r
of
1/
500
f
o
r
2
,
4
­
D
and
2,4,5­
TP.
The
s
a
f
e
l
e
v
e
l
s
a
r
e
t
h
e
n
readjusted
t
o
r
e
f
l
e
c
t
t
o
t
a
l
allowable
intake
p
e
r
person.
Since
l
i
t
t
l
e
2
,
4
­
D
o
r
2,4,5­
TP
is
expected
t
o
occur
i
n
foods,
2
0
percent
of
t
h
e
safe
exposure
level
can
reasonably
be
a
l
l
o
c
a
t
e
d
t
o
water
without
jeopardizing
t
h
e
h
e
a
l
t
h
of
the
consumer.

(
QUALITY
CRITERIA
FOR
WATER,
JULY
1976)
PB­
263943
SEE
APPENDIX
C
FOR
METHODOLOGY
CHROMIUM
rvI)

AQUATIC
LIFE
SUMMARY:
0
Acute
toxicity
values
for
chromium(
V1)
are
available
f
o
r
freshwater
animal
species
in
27
genera
and
range
from
23.07
ug/
L
for
a
cladoceran
to
1,870,000
ug/
L
for
a
stonefly.
These
species
include
a
wide
variety
of
animals
that
perform
a
wide
spectrum
of
ecological
functions.
All
five
tested
species
of
daphnids
are
especially
sensitive.
The
few
data
that
are
available
indicate
that
the
acute
toxicity
of
chromium(
V1)

decreases
as
hardness
and
pH
increase.

The
chronic
value
for
both
rainbow
trout
and
brook
trout
is
264.6
uq/
L,
which
is
much
lower
than
the
chronic
value
of
1,987
ug/
L
for
the
fathead
minnow.
The
acute­
chronic
ratios
for
these
three
fishes
range
from
18.55
to
260.8.
In
all
three
chronic
tests
a
temporary
reduction
in
growth
occurred
at
low
concentrations.
Six
chronic
tests
with
five
species
of
daphnids
gave
chronic
values
that
range
from
<
2.5
to
40
ug/
L
and
the
acute­
chronic
ratios
range
from
1.130
to
>
9.680.
Except
for
the
fathead
minnow,
all
the
chronic
tests
were
conducted
in
soft
water.
Green
algae
are
quite
sensitive
to
chromium(
V1).
The
bioconcentration
factor
obtained
with
rainbow
trout
is
less
than
3
.
Growth
of
chinook
salmon
was
reduced
at
a
measured
concentration
of
16
ug/
L.
0
The
acute
toxicity
of
chromium
(
VI)
to
23
saltwater
vertebrate
and
invertebrate
species
ranges
from
2,000
ug/
L
 or
a
polychaete
worm
and
a
mysid
to
105,000
ug/
L
for
the
mud
Snail.
The
chronic
values
 or
a
polychaete
range
from
<
13
to
36.74
ug/
L,
whereas
that
for
a
mysid
is
132
ug/
L.
The
acute­
chronic
ratios
range
from
15.38
to
>
238.5.
Toxicity
to
macroalgae
was
reported
at
1,000
and
5,000
ug/
L.
Bioconcentration
factors
for
chromium(
V1)

range
from
125
to
236
for
bivalve
molluscs
and
polychaetes.

CHROMIUM
1111)

Acute
values
for
chromium(
II1)
are
available
for
20
freshwater
animal
species
in
la
genera
ranging
from
2,221
ug/
L
for
a
mayfly
to
71,060
ug/
L
for
caddisfly.
Hardness
has
a
significant
influence
on
toxicity,
with
chromium(
II1)
being
more
toxic
in
soft
water.

A
life­
cycle
test
with
__
Daphnia
____
maqna
__
__
in
soft
water
gave
a
chronic
value
of
66
ug/
L.
In
a
comparable
test
in
hard
water
the
lowest
test
concentration
of
44
ug/
L
inhibited
reproduction
of
­
Dap&
nia
­­­
m
g
E
,
but
this
effect
may
have
resulted
from
ingested
precipitated
chromium.
In
a
life­
cycle
test
with
the
fathead
minnow
in
hard
water
the
chronic
value
was
1,025
ug/
L.
Toxicity
data
are
available
for
only
two
freshwater
plant
species.
A
concentration
of
9,900
ug/
L
inhibited
growth
of
roots
of
Eurasian
watermilfoil.
A
freshwater
green
alga
was
affected
by
a
concentration
of
397
ug/
L
in
soft
water.
No
bioconcentration
factor
has
been
measured
for
chromium(
II1)
with
freshwater
organisms.

Only
two
acute
values
are
available
for
chromium
(
111)
in
saltwater
10,300
ug/
L
for
the
eastern
oyster
and
31,500
ug/
L
for
the
mummichog.
In
a
chronic
test
effects
were
not
observed
on
a
polychaete
worm
at
50,400
ug/
L
at
pH
=
7.9,
but
acute
lethality
occurred
when
pH
=
4.5.
Bioconcentration
factors
for
saltwater
organisms
and
chromium(
II1)
range
from
86
to
153,
similar
to
the
bioconcentration
factors
for
chromium(
V1)
and
saltwater
species.

NATIONAL
CRITERIA:

CHROMIUM
(
VI)

The
procedures
described
in
the
Guidelines
for
Deriving
Numerical
National
Water
Quality
Criteria
for
the
Protection
of
Aquatic
Organisms
and
Their
Uses
indicate
that,
except
possibly
where
a
locally
important
species
is
very
sensitive,
freshwater
aquatic
organisms
and
their
uses
should
not
be
affected
unacceptably
if
the
4­
day
average
concentration
of
chromium(
V1)

does
not
exceed
11
ug/
L
more
than
once
every
3
years
on
the
average
and
if
the
1­
hour
avera
e
concentration
does
not
exceed
7
16
ug/
L
more
than
once
every
3
years
on
the
average.

The
procedures
described
in
the
Guidelines
indicate
that,

except
possibly
where
a
locally
important
species
is
very
sensitive,
saltwater
aquatic
organisms,
and
their
uses
should
not
be
affected
unacceptably
if
the
4­
day
average
concentration
of
chromium(
V1)
does
not
exceed
50
ug/
L
more
than
once
every
3
years
on
the
average
and
if
the
1­
hour
average
concentration
does
not
exceed
1,100
ug/
L
more
than
once
every
3
years
on
the
average.

Data
suggest
that
the
acute
toxicity
of
chromium
(
VI)
is
salinity
dependent;
therefore,
the
1­
hour
average
concentration
might
be
underprotective
at
low
salinities.

CHROMIUM(
1111
The
procedures
described
in
the
Guidelines
indicate
that,

.
I
except
possibly
where
a
locally
important
species
is
very
sensitive,
freshwater
aquatic
organisms
and
their
uses
should
not
be
affected
unacceptably
if
the
4­
day
average
concentration
(
in
ug/
L)
of
chromium(
II1)
does
not
exceed
the
numerical
value
given
by
e(
0.8190[
ln(
hardness)]+
l.
561)
more
than
once
every
3
years
on
the
average
and
if
the
1­
hour
average
concentration
(
in
ug/
L)

does
not
exceed
the
numerical
value
given
by
(
0.8190[
ln(
hardness)
]+
3.688)
more
than
once
every
3
years
on
the
e
average.
For
example,
at
hardnesses
of
50,
100,
and
200
mg/
L
as
CaC03
the
&
day
average
concentrations
o
f
chromium(
II1)
are
120,

210,
and
370
ug/
L,
respectively,
and
the
1­
hour
average
concentrations
are
980,
1,700,
and
3,100
ug/
L.

No
saltwater
criterion
can
be
derived
for
chromium(
III),
but
10,300
ug/
L
is
the
EC50
for
eastern
oyster
embryos,
whereas
50,400
ug/
L
did
test.

EPA
believes
provide
a
more
not
affect
a
polychaete
worm
in
a
life­
cycle
a
that
a
measurement
such
as
18acid­
soluble1'
would
scientifically
correct
basis
upon
which
to
establish
criteria
for
minerals.
The
criteria
were
developed
on
this
basis.
However,
at
this
time,
no
EPA­
approved
methods
for
such
a
measurement
are
available
to
implement
the
criteria
through
the
regulatory
programs
of
the
Agency
and
the
States.

The
Agency
is
considering
development
and
approval
of
methods
for
a
measurement
such
as
acid­
soluble.
Until
available,

however,
EPA
recommends
applying
the
criteria
using
the
total
recoverable
method.
This
has
two
impacts:
(
1)
certain
species
of
some
metals
recoverable
cannot
be
analyzed
directly
method
does
not
distinguish
because
the
total
a
between
individual
oxidation
states,
and
(
2)
these
criteria
may
0
when
based
on
the
total
recoverable
method.

The
recommended
exceedence
frequency
be
overly
protective
of
3
years
is
the
'
Agency's
best
scientific
judgment
of
the
average
amount
of
time
it
will
take
an
unstressed
system
to
recover
from
a
pollution
event
in
which
exposure
to
chromium
exceeds
the
criterion.
A
stressed
system,
for
example,
one
in
which
several
outfalls
occur
in
a
limited
area,
would
be
expected
or
require
more
time
for
recovery.
The
resilience
of
ecosystems
and
their
ability
to
recover
differ
greatly,
however,
and
site­
specific
criteria
may
be
established
if
adequate
justification
is
provided­

The
use
of
criteria
in
designing
waste
treatment
facilities
requires
the
selection
of
an
appropriate
wasteload
a1
location
model.
Dynamic
models
are
preferred
for
the
application
of
these
criteria.
Limited
data
or
other
factors
may
make
their
use
impractical,
in
which
case
one
should
rely
on
a
steady­
state
model.
The
Agency
recommends
the
interim
use
of
1Q5
or
lQlO
for
Criterion
Maximum
Concentration
design
flow
and
745
or
7410
or
the
Criterion
Continuous
Concentration
design
flow
in
steady­

state
models
for
unstressed
and
stressed
systems,
respectively.

These
matters
are
discussed
in
more
detail
in
the
Technical
Support
Document
for
water
Quality­
Based
Toxics
Control
(
U.
S.

EPA,
1985).

HUMAN
HEALTH
CRITERIA:
0
For
the
protection
of
human
health
of
Chromium
I11
ingested
through
water
0
from
the
toxic
properties
and
contaminated
aquatic
organisms,
the
ambient
water
criterion
is
determined
to
be
170
mg/
LJ.

For
the
protection
of
human
health
from
the
toxic
properties
of
Chromium
I11
ingested
through
contaminated
aquatic
organisms
alone,
the
ambient
water
criterion
is
determined
to
be
3433
mg/
L.

The
ambient
water
quality
criterion
for
total
Chromium
VI
is
recommended
to
be
identical
to
the
existing
drinking
water
standard
which
is
50
ug/
L.
Analysis
of
the
toxic
effects
data
resulted
in
a
calculated
level
which
is
protective
of
human
health
against
the
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms.
The
calculated
value
is
comparable
to
the
present
standard.
For
this
reason
a
selective
criterion
based
on
exposure
solely
from
consumption
of
6.5
grams
of
aquatic
organisms
was
not
derived.

(
45
F.
R.
79318
Nov.
28,1980)
(
50
F.
R.
30784,
July
29,
1985)
SEE
APPENDIX
A
FOR
METHODOLOGY
2­
CHLOROPHENOL
Aquatic
Life
The
available
data
for
2­
chlorophenol
indicate
that
acute
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
4,380
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
definitive
data
are
available
concerning
the
chronic
toxicity
of
2­

chlorophenol
to
sensitive
freshwater
aquatic
life,
but
flavor
impairment
occurs
in
one
species
of
fish
at
concentrations
as
low
as
2,000
ug/
L.

No
saltwater
organisms
have
been
tested
with
2­

I
chlorophenol
and
therefore,
no
statement
can
be
made
concerning
acute
or
chronic
toxicity.

Human
Health
Sufficient
data
are
not
available
for
2­
chlorophenol
to
derive
a
level
which
would
protect
against
the
potential
toxicity
of
this
compound.
Using
avai
1
able
organol
eptic
data,
to
control
undesirable
taste
and
odor
qualities
of
ambient
water
the
estimated
level
is
0.1
ug/
L.
It
should
be
recognized
that
organoleptic
data
have
limitations
as
a
basis
for
establishing
a
water
quality
criterion,
and
have
no
demonstrated
relationship
to
potential
adverse
human
health
effects.

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
COLOR
Waters
shall
be
virtually
free
from
substances
producing
objectionable
color
for
aesthetic
purposes;

the
source
of
supply
should
not
exceed
75
color
units
on
the
platinum­
cobalt
scale
for
domestic
water
supplies;
and
not
reduce
the
depth
of
the
compensation
point
for
photosynthetic
activity
by
more
than
10
percent
from
the
seasonally
established
norm
for
aquatic
life.
increased
color
(
in
combination
with
turbidity)
should
INTRODUCTION:

Color
in
water
principally
results
from
degradation
processes
in
the
natural
environment.
Although
colloidal
forms
of
iron
and
manganese
occasionally
are
the
cause
of
color
in
water,
the
most
common
causes
are
complex
organic
compounds
originating
from
the
decomposition
of
naturally
occurring
organic
matter
(
AWWA,

1971).
Sources
of
organic
material
include
human
materials
from
the
soil
such
as
tannins,
human
acid
and
humates:
decaying
plankton:
and
other
decaying
aquatic
plants.
Industrial
discharges
may
contribute
similar
compounds:
for
example,
those
from
the
pulp
and
paper
and
tanning
industries.
Other
industrial
discharges
may
contain
brightly
colored
substances
such
as
those
from
certain
processes
in
textile
and
chemical
industries.

Surface
waters
may
appear
colored
because
of
suspended
matter
which
comprises
turbidity.
Such
color
is
referred
to
as
apparent
color
and
is
differentiated
from
true
color
caused
by
colloidal
human
materials
(
Sawyer,
1960).
Natural
color
is
reported
in
,
color
9mits"
which
generally
are
determined
by
use
of
the
platinum­
cobalt
method
(
Standard
Methods,
1971).

There
is
no
general
agreement
as
to
the
chemical
composition
of
natural
color,
and
in
fact
the
composition
may
vary
chemically
from
place
to
place
(
AWWA,
1971).
Black
and
Christman
(
1963a)

characterized
color­
causing
colloids
examined
as
aromatic,

polyhydroxy,
methoxy
carboxyl
ic
acids.
Shapiro
(
1964)

characterized
color­
causing
constituents
as
being
dialyzable
and
composed
of
aliphatic,
polyhydroxyl
carboxyl
ic
acids
with
molecular
weights
varying
from
less
than
200
to
approximately
400.
The
colloidal
fraction
of
color
exists
in
the
3.5
to
10
mu
diameter
range
(
Black
and
Christman,
1963b).
These
same
authors
summarized
other
characteristics
of
color
observed
in
laboratory
studies
of
natural
waters:
color
is
caused
by
light
scattering
and
fluorescence
rather
than
absorption
of
light
energy,
and
pH
affects
both
particle
size
of
the
color­
causing
colloids
and
the
intensity
of
color
itself.

RATIONALE
:

Color
in
water
is
an
important
constituent
in
terms
of
aesthetic
considerations.
To
be
aesthetically
pleasing,
water
should
be
virtually
free
from
substances
introduced
by
man's
activities
which
produce
objectionable
color.
"
Objectionable
color18
is
defined
to
be
a
significant
increase
over
natural
background
levels.
Non­
natural
colors
such
as
dyes
should
not
be
perceptible
by
the
human
eye
as
such
colors
are
especially
objectionable
to
those
who
receive
pleasure
by
viewing
water
in
its
natural
state.
Because
of
the
extreme
variations
in
the
natural
background
amount
of
color,
it
is
meaningless
to
attempt
numerical
limits.
The
aesthetic
attributes
of
water
depend
on
one's
appreciation
of
the
water
setting.
0
The
effects
of
col'or
on
public
water
supplies
also
are
principally
aesthetic.
The
1962
Drinking
Water
Standards
(
PHS,

1962)
recommended
that
color
in
finished
waters
should
not
exceed
15
units
on
the
platinum­
cobalt
scale.
Water
consistently
can
be
treated
using
standard
coagulation,
sedimentation
and
filtration
processes
to
reduce
color
to
substantially
less
than
15
color
units
when
the
source
water
does
not
exceed
7
5
color
units
AWWA,

1971;
NAS,
1974).

The
effects
of
color
in
water
on
aquatic
life
principally
are
to
reduce
light
penetration
and
thereby
generally
reduce
photosynthesis
by
phytoplankton
and
to
restrict
the
zone
for
aquatic
vascular
plant
growth.
a
The
light
supply
necessary
to
support
plant
life
is
dependent
on
both
intensity
and
effective
wave
lengths
(
Welch,
1952).
In
general,
the
rate
of
photosynthesis
increases
with
the
intensity
of
the
incident
light.
Photosynthetic
rates
are
most
affected
in
the
red
region
and
least
affected
in
the
blue­
violet
region
of
incident
light
(
Welch,
1952).
It
has
been
found
that
in
colored
waters
the
red
spectrum
is
not
a
region
of
high
absorption
so
that
the
effective
penetration,
and
therefore
the
intensity
for
photosynthesis,
is
not
as
restricted
as
are
other
wave
lengths.

It
should
be
emphasized
that
transmission
of
all
parts
of
the
spectrum
is
affected
by
color,
but
the
greatest
effect
is
on
the
0
standard
or
blue
end
of
the
spectrum
(
Birge
and
Juday,
1930).
In
TABLE
2.

Maximum
color
of
surface
waters
that
have
been
used
as
sources
f
o
r
industrial
water
supplies.

Industry
or
Industrial
Use
Color
units
Boiler
make
up
1,200
cooling
water
1,200
Pulp
and
paper
Chemical
and
allied
products
Petroleum
360
500
25
highly
colored
waters
(
45
to
132
color
units)
Birge
and
Juday
(
1930)
measured
the
light
transmission
as
a
percentage
of
the
incident
level
and
found
very
little
blue,
50
percent
or
less
yellow,
and
100
to
120
percent
red.
0
The
light
intensity
required
for
some
aquatic
vascular
plants
to
photosynthetically
balance
the
oxygen
used
in
respiration
may
be
5
percent
of
full
sunlight
during
maximum
summer
illumination
periods
(
NTAC,
1968).
As
much
as
10
percent
of
the
incident
light
may
be
required
for
plankton
to
likewise
photosynthetically
produce
sufficient
oxygen
to
balance
their
respiration
requirements
(
NTAC,
1968).
The
depth
at
which
such
a
compensation
point
is
reached,
calledthe
compensation
depth,

delineates
the
zone
of
ef
fe,
ctive
photosynthetic
oxygen
production.
To
maintain
satisfactory
biological
conditions,
this
depth
cannot
be
substantially
reduced.
0
Industrial
requirements
as
related
to
water
color
have
been
standardized
(
NAS,
1974).
Table
2
lists
the
maximum
value
used
as
a
source
of
water
for
various
industries
and
industrial
uses.

Through
treatment,
such
waters
can
be
made
to
meet
almost
any
industrial
requirement.

(
QUALITY
CRITERIA
FOR
WATER,
JULY
1976)
PB­
263943
SEE
APPENDIX
C
FOR
METHODOLOGY
0
AQUATIC
LIFE
SUMMARY:
*
COPPER
Acute
toxicity
data
freshwater
animals.
At
are
available
for
species
in
41
genera
of
a
hardness
of
50
mg/
L
the
genera
range
in
sensitivity
from
16.74
ug/
L
for
Ptychocheilus
­
to
10,240
ug/
L
for
Acroneuria.
Data
for
eight
species
indicate
that
acute
toxicity
decreases
as
hardness
increases.
Additional
data
for
several
species
indicate
that
toxicity
also
decreases
with
increases
in
alkalinity
and
total
organic
carbon.

Chronic
values
are
available
 or
15
freshwater
species
and
range
from
3.873
ug/
L
 or
brook
trout
to
60.36
ug/+
for
northern
pike.
Fish
and
invertebrate
species
seem
to
be
about
equally
sensitive
to
the
chronic
toxicity
of
copper.

Toxicity
tests
have
been
conducted
on
copper
with
a
wide
range
of
freshwater
plants
and
the
sensitivities
are
similar
to
those
of
animals.
Complexing
effects
of
the
test
media
and
a
lack
of
good
analytical
data
make
interpretation
and
application
of
these
results
difficult.
Protection
of
animal
species,

however,
appears
to
offer
adequate
protection
of
plants.
Copper
does
not
appear
to
bioconcentrate
very
much
in
the
edible
portion
of
freshwater
aquatic
species.

The
acute
sensitivities
of
saltwater
animals
to
copper
range
from
5.8
ug/
L
for
the
blue
mussel
to
600
ug/
L
for
the
green
crab.

A
chronic
life­
cycle
test
has
been
conducted
with
a
mysid,
and
adverse
effects
were
observed
at
77
ug/
L
but
not
at
38
ug/
L,

which
resulted
in
an
acute­
chronic
ratio
of
3.346.
Several
*
Indicates
susDended.
canceled
or
restricted
bv
U.
S.
EPA
Office
of
Pesticides
ahd
Toxic
Substances
­
saltwater
algal
species
have
been
tested,
and
effects
were
observed
between
5
and
100
ug/
L.
Oysters
can
bioaccumulate
copper
up
to
28,200
times,
and
become
bluish­
green,
apparently
without
significant
mortality.
In
long­
term
exposures,
the
bay
scallop
was
killed
at
5
ug/
L.

NATIONAL
CRITERIA:

The
procedures
described
in
the
Guidelines
for
Deriving
Numerical
National
Water
Quality
Criteria
for
the
Protection
of
Aquatic
Organisms
and
Uses
indicate
that,
except
possibly
where
a
locally
important
species
is
very
sensitive,
freshwater
aquatic
organisms
and
their
uses
should
not
be
affected
unacceptably
if
the
4­
day
average
concentration
(
in
ug/
L)
of
copper
does
not
exceed
the
numerical
value
given,
by
.(
0.8545[
ln(
hardness)
3
­
1.465)

more
than
once
every
3
years
on
the
average
and
if
the
1­
hour
average
concentration
(
in
ug/
L)
does
not
exceed
the
numerical
value
given
by
,(
0.9422
[
ln(
hardness)
1­
1.464)
more
than
once
every
3
years
on
the
average.
For
example,
at
hardnesses
of
50,
100,

and
200
mg/
L
as
CaC03
the
4­
day
average
concentrations
of
copper
are
6.5,
12,
and
21
ug/
L,
respectively,
and
the
1­
hour
average
concentrations
are
9.2,
18,
and
34
ug/
L.

The
procedures
described
in
the
Guidelines
indicate
that,

except
possibly
where
a
locally
important
species
is
very
sensitive,
saltwater
aquatic
organisms
and
their
uses
should
not
be
affected
unacceptably
if
the
1­
hour
average
concentration
of
copper
does
not
exceed
2.9
ug/
L
more
than
once
every
3
years
on
the
average.

EPA
believes
that
a
measurement
such
as
tsacid­
solublets
would
p
r
o
v
i
d
e
a
more
s
c
i
e
n
t
i
f
i
c
a
l
l
y
c
o
r
r
e
c
t
b
a
s
i
s
upon
which
t
o
e
s
t
a
b
l
i
s
h
c
r
i
t
e
r
i
a
f
o
r
metals.
The
c
r
i
t
e
r
i
a
were
developed
on
t
h
i
s
basis.
However,
a
t
t
h
i
s
t
i
m
e
,
no
EPA
approved
methods
f
o
r
such
a
measurement
a
r
e
a
v
a
i
l
a
b
l
e
t
o
implement
t
h
e
c
r
i
t
e
r
i
a
through
t
h
e
r
e
g
u
l
a
t
o
r
y
programs
of
t
h
e
Agency
and
t
h
e
States.

The
Agency
is
considering
development
and
approval
of
methods
f
o
r
a
measurement
such
a
s
acid­
s
o
l
u
b
l
e
.
U
n
t
i
l
a
v
a
i
l
a
b
l
e
,

however,
EPA
recommends
applying
the
c
r
i
t
e
r
i
a
using
t
h
e
t
o
t
a
l
recoverable
method.
This
has
two
impacts:
(
1)
c
e
r
t
a
i
n
species
of
some
metals
cannot
be
analyzed
d
i
r
e
c
t
l
y
because
t
h
e
t
o
t
a
l
r
e
c
o
v
e
r
a
b
l
e
method
does
n
o
t
d
i
s
t
i
n
g
u
i
s
h
between
i
n
d
i
v
i
d
u
a
l
oxidation
s
t
a
t
e
s
,
and
(
2)
these
c
r
i
t
e
r
i
a
may
be
overly
protective
when
based
on
t
h
e
t
o
t
a
l
recoverable
method.

The
recommended
exceedence
frequency
of
3
years
is
t
h
e
Agency's
best
s
c
i
e
n
t
i
f
i
c
judgment
of
t
h
e
average
amount
of
t
i
m
e
it
w
i
l
l
t
a
k
e
an
unstressed
system
t
o
r
e
c
o
v
e
r
from
a
p
o
l
l
u
t
i
o
n
e
v
e
n
t
i
n
which
exposure
t
o
copper
exceeds
t
h
e
c
r
i
t
e
r
i
o
n
.
A
stressed
system,
f
o
r
example,
one
i
n
which
several
o
u
t
f
a
l
l
s
occur
i
n
a
l
i
m
i
t
e
d
a
r
e
a
,
would
be
expected
t
o
r
e
q
u
i
r
e
more
t
i
m
e
f
o
r
recovery.
The
r
e
s
i
l
i
e
n
c
e
of
ecosystems
and
t
h
e
i
r
a
b
i
l
i
t
y
t
o
recover
d
i
f
f
e
r
g
r
e
a
t
l
y
,
however,
and
s
i
t
e­
s
p
e
c
i
f
i
c
c
r
i
t
e
r
i
a
may
be
established
i
f
adequate
j
u
s
t
i
f
i
c
a
t
i
o
n
is
provided.

The
use
of
criteria
i
n
developing
waste
treatment
f
a
c
i
l
i
t
i
e
s
r
e
q
u
i
r
e
s
t
h
e
s
e
l
e
c
t
i
o
n
of
an
a
p
p
r
o
p
r
i
a
t
e
wasteload
a
l
l
o
c
a
t
i
o
n
model.
Dynamic
models
a
r
e
preferred
f
o
r
t
h
e
application
of
these
c
r
i
t
e
r
i
a
.
L
i
m
i
t
e
d
d
a
t
a
o
r
o
t
h
e
r
f
a
c
t
o
r
s
may
make
t
h
e
i
r
use
i
m
p
r
a
c
t
i
c
a
l
,
i
n
which
case
one
should
r
e
l
y
on
a
steady­
s
t
a
t
e
model.
The
Agency
recommends
t
h
e
interim
use
of
1Q5
o
r
l
Q
l
O
f
o
r
0
Criterion
Maximum
Concentration
design
flow
and
745
or
7410
for
the
Criterion
Continuous
Concentration
(
CCC)
design
flow
in
steady­
state
models
for
unstressed
and
stressed
systems
respectively.
These
matters
are
discussed
in
more
detail
in
the
Technical
Support
Document
for
Water
Quality­
Based
Toxics
Control
(
U.
S.
EPA,
1985).

HUMAN
HEALTH
CRITERIA:

Sufficient
data
is
not
available
for
copper
to
derive
a
level
which
would
protect
against
the
potential
toxicity
of
this
compound.
Using
available
organoleptic
data,
for
controlling
undesirable
taste
and
odor
quality
of
ambient
water,
the
estimated
level
is
1
mg/
L.
It
should
be
recognized
that
organoleptic
data
as
a
basis
for
establishing
a
water
quality
criteria
have
1
imitations
and
have
no
demonstrated
relationship
to
potential
adverse
human
health
effects.
I
(
45
F.
R.
79318
Nov.
28,1980)
(
50
F.
R.
30784,
July
29,
1985)
SEE
APPENDIX
A
FOR
METHODOLOGY
CYANIDE
AQUATIC
­
LIFE
SUMMARY:

D
a
t
a
on
t
h
e
a
c
u
t
e
t
o
x
i
c
i
t
y
of
free
cyanide
(
t
h
e
sum
of
cyanide
p
r
e
s
e
n
t
a
s
HCN
and
CN­,
expressed
a
s
CN)
are
a
v
a
i
l
a
b
l
e
f
o
r
a
w
i
d
e
v
a
r
i
e
t
y
of
freshwater
s
p
e
c
i
e
s
t
h
a
t
are
i
n
v
o
l
v
e
d
i
n
d
i
v
e
r
s
e
community
functions.
The
acute
s
e
n
s
i
t
i
v
i
t
i
e
s
ranged
from
44.73
ug/
L
t
o
2,490
ug/
L,
b
u
t
a
l
l
of
t
h
e
s
p
e
c
i
e
s
w
i
t
h
a
c
u
t
e
s
e
n
s
i
t
i
v
i
t
i
e
s
above
400
ug/
L
were
i
n
v
e
r
t
e
b
r
a
t
e
s
.
A
long­
term
s
u
r
v
i
v
a
l
,
and
a
p
a
r
t
i
a
l
and
l
i
f
e­
c
y
c
l
e
t
e
s
t
with
f
i
s
h
gave
c
h
r
o
n
i
c
v
a
l
u
e
s
o
f
13.57,
7.849,
and
16.39
ug/
L,
r
e
s
p
e
c
t
i
v
e
l
y
.

Chronic
v
a
l
u
e
s
f
o
r
two
f
r
e
s
h
w
a
t
e
r
i
n
v
e
r
t
e
b
r
a
t
e
s
p
e
c
i
e
s
were
18.33
and
34.06
ug/
L.
Freshwater
p
l
a
n
t
s
were
a
f
f
e
c
t
e
d
a
t
cyanide
concentrations
ranging
from
30
ug/
L
t
o
26,000
ug/
L.

The
acute
t
o
x
i
c
i
t
y
of
f
r
e
e
cyanide
t
o
s
a
l
t
w
a
t
e
r
species
ranged
from
4.893
ug/
L
t
o
>
10,000
ug/
L
and
i
n
v
e
r
t
e
b
r
a
t
e
s
w
e
r
e
both
t
h
e
most
and
l
e
a
s
t
s
e
n
s
i
t
i
v
e
species.
Long­
term
s
u
r
v
i
v
a
l
i
n
an
e
a
r
l
y
l
i
f
e­
s
t
a
g
e
test
w
i
t
h
the
sheepshead
minnow
gave
a
chronic
v
a
l
u
e
of
36.12
ug/
L.
Long­
term
s
u
r
v
i
v
a
l
i
n
a
mysid
l
i
f
e­
c
y
c
l
e
t
e
s
t
r
e
s
u
l
t
e
d
i
n
a
c
h
r
o
n
i
c
v
a
l
u
e
of
69.71
ug/
L.
T
e
s
t
s
w
i
t
h
t
h
e
red
macroalga,
Champia
parvula
­­
I
showed
cyanide
t
o
x
i
c
i
t
y
a
t
11
t
o
25
ug/
L,
b
u
t
o
t
h
e
r
s
p
e
c
i
e
s
were
a
f
f
e
c
t
e
d
a
t
c
o
n
c
e
n
t
r
a
t
i
o
n
s
up
t
o
3,000
ug/
L.

NATIONAL
CRITERIA:

The
procedures
described
i
n
t
h
e
G
u
i
d
e
l
i
n
e
s
f
o
r
Deriving
Numerical
National
Water
Q
u
a
l
i
t
y
Criteria
f
o
r
t
h
e
Protection
of
Aquatic
Organisms
and
Their
Uses
indicate
t
h
a
t
,
except
possibly
where
a
l
o
c
a
l
l
y
important
s
p
e
c
i
e
s
is
very
s
e
n
s
i
t
i
v
e
,
freshwater
a
q
u
a
t
i
c
o
r
g
a
n
i
s
m
s
a
n
d
t
h
e
i
r
u
s
e
s
s
h
o
u
l
d
n
o
t
be
a
f
f
e
c
t
e
d
unacceptably
if
the
4­
day
average
concentration
of
cyanide
does
not
exceed
5.2
ug/
L
more
than
once
every
3
years
on
the
average
and
if
the
1­
hour
average
concentration
does
not
exceed
22
ug/
L
more
than
once
every
3
years
on
the
average.

The
procedures
described
in
the
Guidelines
indicate
that,

except
possibly
where
a
locally
important
species
is
very
sensitive,
saltwater
aquatic
organisms
and
their
uses
should
not
be
affected
unacceptably
if
the
1­
hour
average
concentration
of
cyanide
does
not
exceed
1.0
ug/
L
more
than
once
every
3
years
on
the
average.

EPA
believes
that
a
measurement
such
as
"
acid
soluble"
would
provide
a
more
scientifically
correct
basis
upon
which
to
establish
criteria
for
cyanide.
The
criteria
were
developed
on
this
basis.
However,
at
this
time,
no
EPA­
approved
methods
for
such
a
measurement
are
available
to
implement
the
criteria
through
the
regulatory
programs
of
the
Agency
and
the
States.

The
Agency
is
considering
development
and
approval
of
methods
for
a
measurement
such
as
acid
soluble.
Until
available,

however,
EPA
recommends
applying
the
criteria
using
the
total
recoverable
method.
These
criteria
may
be
overly
protective
when
based
on
the
total
recoverable
method.

The
recommended
exceedence
frequency
of
3
years
is
the
Agency's
best
scientific
judgment
of
the
average
amount
of
time
it
will
take
an
unstressed
system
to
recover
from
a
pollution
event
in
which
exposure
to
cyanide
exceeds
the
criterion.
A
stressed
system,
for
example,
one
in
which
several
outfalls
occur
in
a
limited
area,
would
be
expected
to
require
more
time
for
recovery.
The
resilience
of
ecosystems
and
their
ability
to
recover
differ
greatly,
however,
and
site­
specif
ic
criteria
may
be
established
if
adequate
justification
is
provided.

The
use
of
criteria
in
designing
waste
treatment
facilities
requires
the
selection
of
an
appropriate
wasteload
a1
location
model.
Dynamic
models
are
preferred
for
the
application
of
these
criteria.
Limited
data
or
other
factors
may
make
their
use
impractical,
in
which
case
one
should
rely
on
a
steady­
state
model.
The
Agency
recommends
the
interim
use
of
1Q5
or
lQlO
for
Criterion
Maximum
Concentration
design
flow
and
745
or
7410
f
o
r
the
Criterion
Continuous
Concentration
design
flow
in
steady­

state
models
for
unstressed
and
stressed
systems
respectively.

These
matters
are
discussed
in
more
detail
in
the
Technical
Support
Document
for
Water
Quality­
Based
Toxics
Control
(
U.
S.

EPA,
1985).

H"
HEALTH
CRITERIA
The
ambient
water
quality
criterion
for
cyanide
is
recommended
to
be
identical
to
the
existing
drinking
water
standard
which
is
200
ug/
L.
Analysis
of
the
toxic
effects
data
resulted
in
a
calculated
level
which
is
protective
of
human
health
against
the
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms.
The
calculated
value
is
comparable
to
the
present
standard.
For
this
reason
a
selective
criterion
based
on
exposure
solely
from
consumption
of
6.5
grams
of
aquatic
organisms
was
not
derived.

NOTE:
The
U.
S.
EPA
is
currently
developing
Acceptable
Daily
Intake
(
ADI)
or
Verified
Reference
Dose
(
RfD)
values
for
Agency­
wide
use
for
this
chemical.
The
new
value
should
be
substituted
when
it
becomes
available.
The
January,
1986,
draft
Verified
Reference
Dose
document
cites
an
RfD
of
.02
mg/
kg/
day
for
free
cyanide.
­,
0
CRITERIA:

DDT
_.
­­
DDT
AND
METABOLITES
Aquatic
Life
For
DDT
and
its
metabolites
the
criterion
to
protect
freshwater
aquatic
life
as
derived
using
the
Guidelines
is
0.0010
ug/
L
as
a
24­
hour
average
and
the
concentration
should
not
exceed
1.1
ug/
L
at
any
time.

For
DDT
and
its
metabolites
the
criterion
to
protect'

saltwater
aquatic
life
as
derived
using
the
Guidelines
is
0.0010
ug/
L
as
a
24­
hour
average
and
the
concentration
should
not
exceed
0.13
ug/
L
at
any
time.

TDE
­
The
available
data
for
TDE
indicate
that
acute
toxicity
to
freswater
aquatic
life
occurs
at
concentrations
as
low
as
0.6
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
TDE
to
sensitive
freshwater
aquatic
life.

The
available
data
for
TDE
indicate
that
acute
toxicity
to
saltwater
aquatic
life
occurs
at
concentrations
as
low
as
3.6
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
TDE
to
sensitive
saltwater
aquatic
life.

DDE
­
The
available
data
for
DDE
indicate
that
acute
toxicity
'./

to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
1,050
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
DDE
to
sensitive
freshwater
aquatic
life.

The
available
data
for
DDE
indicate
that
acute
toxicity
to
saltwater
aquatic
life
occurs
in
concentrations
as
low
as
14
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
DDE
to
sensitive
saltwater
aquatic
life.

Human
Health
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
DDT
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentration
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
and
The
corresponding
recommended
criteria
are
0.24
nq/
L,
0.024
ng/
L,
and
0.0024
ng/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
0.24
ng/
L,
0.024
ng/
L,
and
0.0024
ng/
L,
respectively.

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
a
CRITERION:
DEMETON
0.1
ug/
L
for
freshwater
and
marine
aquatic
life
RATIONALE:

Static
LC50
bioassays
yielded
toxicity
values
for
the
organo­

phosphorus
pesticide
demeton
for
carp,
goldfish,
fathead
minnow,

channel
catfish,
guppy,
rainbow,
trout
and
bluegill,
ranging
from
70
ug/
L
to
15,000
ug/
L
(
Henderson
and
Pickering,
1958;
Ludemann
and
Neumann,
1982;
Macek
and
McAllister,
1970;
McCann
and
Jasper,

1972;
Pickering
et
al.
1962).
Results
of
these
tests
demonstrate
an
apparent
sharp
division
in
species
sensitivity,
with
bluegill
(
Lepomis
macrochirusr,
rainbow
trout
(
Salmo
gairdneri)
and
guppy,
(
Poecilia
­­­
­­­­­­­­­
reticulatg,'
being
susceptible
to
lower
concentrations
while
the
remaining
species
were
comparatively
resistant.
In
the
96­
hOUr
exposures
toxicity
did
not
increase
significantly
with
time,
indicating
that
concentrations
close
to
nominal
may
not
have
been
maintained
for
more
than
a
few
hours,

Bluegills
with
a
24­
hour
LC50
of
70
ug/
L
were
the
most
sensitive
fish
(
Mccann
and
Jasper,
1972).

When
fish
were
exposed
to
acutely
toxic
levels
of
demeton
 or
12
hours
by
Weiss
(
1959,
1961)
the
maximum
inhibition
of
brain
acetiylcholinesterase
(
AChE)
was
not
reached.
The
lowest
levels
of
AChE
occurred
after
24
to
48
hOUrS.
It
was
demonstrated
that
maximum
inhibition
could
last
as
long
as
two
weeks
after
exposure,
and
subsequent
recovery
to
levels
approaching
normal
took
many
more
weeks.
Weiss
(
1958)
reported
a
significant
increase
in
mortality
of
fathead
minnows
exposed
for
a
second
time
to
the
organophosphate,
Sarin,
before
the
fish
had
recovered
normal
brain
AChE
levels.
The
resistance
of
fully
recovered
fish
was
equal
to
that
of
previously
unexposed
controls.
Weiss
and
Gakstatter
(
1964a)
reported
no
significant
inhibition
of
brain
AChE
in
bluegills,
goldfish
and
shiners
(
Notemigonus
__
crysoleucasl,
__
­
­­­­­
following
15­
day
exposures
to
demeton
at
continuously
replenished,
nominal
concentrations
of
1
ug/
L.

Acute
toxicity
values
reported
for
invertebrates
range
from
10
to
100,000
ug/
L
(
Ludemann
and
Neumann,
1962;
Sanders,
1972).

In
general,
molluscs
and
tubifex
worms
were
very
resistant
while
the
smaller
crustaceans
and
insect
larvae
were
susceptible.

Ludemann
and
Neumann
(
1962)
reported
that
Chironomus
plumosus
larvae
were
the
most
sensitive
species
they
tested.
A
24­
hour
exposure
at
10
ug/
L
produced
undefined
effects
while
100
percent
were
killed
at
1000
ug/
L.
Calculated
LC50
data
for
invertebrates
apparently
are
limited
to
a
single,
nominal
concentration
static
exposure
of
Gammarus
______
fasciatus
(
Sanders,
1972).
These
24­
and
96­
hourLC50
valuesarereportedas
500
and27
ug/
L,
indicatinga
time­
related
effect
not
observed
in
the
bioassays
with
fishes.

As
only
a
fewofthe
sensitive
s
p
e
c
i
e
s
h
a
v
e
b
e
e
n
t
e
s
t
e
d
a
n
d
g
r
e
a
t
variance
in
response
can
result
with
different
test
methods,

caution
must
be
exercised
in
estimating
the
sub­
acute
concentration
for
aquatic
fauna
in
general.
It
appears
that
no
study
has
been
made
of
possible
residual
effects
other
than
AChE
inhibition,
which
might
result
from
short
exposures
to
subacute
concentrations
of
organophosphates.

There
are
few
data
on
the
toxicity
of
demeton
to
marine
organisms.
Butler
(
1964)
reported
a
48­
hour
EC50
of
63
ug/
L
for
the
pink
shrimp,
Peneaus
duorarum,
and
a
24­
hour
LC50
of
550
ug/
L
for
the
spot,
Leiostomus
xanthurus.

Chronic
demeton
toxicity
data
for
freshwater
organism
are
not
currently
available.
Since
no
data
are
available
at
this
time
to
indicate
long­
term
no­
effect
levels
for
aquatic
organisms,
a
criterion
must
be
derived
based
partly
on
the
fact
that
all
organophosphates
inhibit
the
production
of
the
AChE
enzyme.

Demeton
is
unique,
however,
in
that
the
persistence
of
its
AChE­

inhibiting
ability
is
greater
than
that
of
10
other
common
organophosphates,
even
though
its
acute
toxicity
is
apparently
less.
The
effective
"
half­
life"
of
AChE
inhibition
for
demeton
is
greater
than
one
year
(
Weiss
and
Gakstatter,
1964b).
Because
such
inhibition
may
be
additive
with
repeated
exposures
and
may
be
compounded
by
any
of
the
organophosphates,
it
is
recommended
that
a
criterion
for
demeton
be
based
primarily
on
its
enzyme­

inhibiting
potential.
A
criterion
of
0.1
ug/
L
demeton
for
freshwater
and
marine
aquatic
life
is
recommended
since
it
will
not
be
expected
to
significantly
inhibit
AChE
over
a
prolonged
period
of
time.
In
addition,
the
criteria
recommendation
is
in
close
agreement
with
the
criteria
for
the
other
organophosphates.

(
QUALITY
CRITERIA
FOR
WATER,
JULY
1976)
PB­
263943
SEE
APPENDIX
C
FOR
METHODOLOGY
DICHLOROBENZENES
Aquatic
­
L
i
f
e
The
available
data
for
dichlorobenzenes
indicate
that
acute
and
chronic
toxicity
to
freshwater
aquatic
life
occur
at
concentrations
as
low
as
1,120
and
763
ug/
L,
respectively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.

The
available
data
for
dichlorobenzenes
indicate
that
acute
toxicity
to
saltwater
aquatic
life
occurs
at
concentrations
as
low
as
1,970
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic'
toxicity
of
dichlorobenzenes
to
sensitive
saltwater
aquatic
life.
0
Human
Health
For
the
protection
of
human
health
from
the
toxic
properties
of
dichlorobenzene
ingested
through
water
and
contaminated
aquatic
organisms,
the
ambient
water
criterion
is
determined
to
be
4
0
0
ug/
L.

For
the
protection
of
human
health
from
the
toxic
properties
of
dichlorobenzenes
ingested
through
contaminated
aquatic
organisms
alone,
the
ambient
water
criterion
is
determined
to
be
2.6
mg/
L.

(
4
5
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B.
FOR
METHODOLOGY
0
.,
DICHLOROBENZIDINE
CRITERIA:

Aquatic
Life
The
data
base
available
for
dichlorobenzidines
and
freshwater
organisms
is
limited
to
one
test
on
bioconcentration
of
3,3­

dichlorobenzidine,
and
therefore,
no
statement
can
be
made
concerning
acute
or
chronic
toxicity.

No
saltwater
organisms
have
been
tested
with
any
dichlorobenzidine,
and
therefore,
no
statement
can
be
made
concerning
acute
or
chronic
toxicity.

Human
Health
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
dichlorobenzidine
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentrations
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
and
lo­'.

The
corresponding
recommended
criteria
are
0.103
ug/
L,
0.010
ug/
L,
and
0.001
ug/
L,
respectively.
If
these
estimates
are
made
for
consumption'of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
0.204
ug/
L,
0.020
ug/
L,
and
0.002
ug/
L,
respectively.
0
(
45
F.
R.
79318,
November
28,
1980)

_.*
SEE
APPENDIX
B
FOR
METHODOLOGY
DICHLOROETHYLENES
Aquatic
Life
The
available
data
for
dichloroethylenes
indicate
that
acute
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
11,600
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
definitive
data
are
available
concerning
the
chronic
toxicity
of
dichloroethylenes
to
sensitive
freshwater
aquatic
life.

The
available
data
for
dichloroethylenes
indicate
that
acute
and
chronic
toxicity
to
saltwater
aquatic
life
occurs
at
concentrations
as
low
as
224,000
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
dichloroethylenes
to
sensitive
saltwater
aquatic
life.

Human
Health
1,1­
Dichloroethylene
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
1,
l
dichloroethylene
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentrations
should
be
zero,

based
on
the
non
threshold
assumption
for
this
chemical.

However,
zero
level
may
not
be
attainable
at
the
present
time.

Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
10­
5e
10­

0
.­
6,
and
The
corresponding
recommended
criteria
are
0.33
ug/
L,
0.033
ug/
L,
and
0.003
ug/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
18.5
ug/
L,
1.85
ug/
L,
and
0.185
ug/
L,
respectively.

1,2­
Dichloroethylene
Using
the
present
guidelines,
a
satisfactory
criterion
cannot
be
derived
at
this
tine
because
of
insufficient
available
data
for
1,2­
dichloroethylene.

(
4
5
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
2,4­
DICHLOROPHENOL
CRITERIA:

Aquatic
Life
The
available
data
for
2,4­
dichlorophenol
indicate
that
acute
and
chronic
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
2,020
and
365
ug/
L,
respectively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
Mortality
to
early
life
stages
of
one
species
of
fish
occurs
at
concentrations
as
low
as
70
ug/
L.

Only
one
test
has
been
conducted
with
saltwater
organisms
and
2,4­
dichlorophenol
and
therefore,
no
statement
can
be
made
concerning
acute
or
chronic
toxicity.

Human
Health
For
comparison
purposes,
two
approaches
were
used
to
derive
criterion
levels
for
2,4­
dichlorophenol.
Based
on
available
toxicity
data,
to
protect
public
health
the
derived
level
is
3.09
mg/
L.
Using
available
organoleptic
data,
to
control
undesirable
taste
and
odor
qualities
of
ambient
water
the
estimated
level
is
0.3
ug/
L.
It
should
be
recognized
that
organoleptic
data
have
limitations
as
a
basis
for
establishing
a
water
quality
criterion,
and
have
no
demonstrated
relationship
to
potential
adverse
human
health
effects.

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
DICHLOROPROPANES/
DICHLOROPROPROPENES
CRITERIA:

Aquatic
Life
The
available
data
for
dichloropropanes
indicate
that
acute
and
chronic
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
23,000
and
5,700
ug/
L,
respectively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.

The
available
data
for
dichloropropene
indicate
that
acute
and
chronic
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
6,060
and
2
4
4
ug/
L,
respeptively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.

The
available
data
for
dichloropropane
indicate
that
acute
and
chronic
toxicity
to
saltwater
aquatic
life
occur
at
concentrations
as
l
o
w
as
10,300
and
3,040
ug/
L,
respectively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.

The
available
data
for
dichloropropene
indicate
that
acute
toxicity
to
saltwater
aquatic
life
occurs
at
concentrations
as
low
as
790
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
dichloropropene
to
sensitive
saltwater
aquatic
life.
Human
Health
Using
the
present
guidelines,
a
satisfactory
criterion
cannot
be
derived
at
this
time
because
of
insufficient
available
data
for
dichloropropanes.

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
Aquatic
Life
The
available
data
for
2,4­
dimethylphenol
indicate
that
acute
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
2,120
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
dimethylphenol
to
sensitive
freshwater
aquatic
life.

No
saltwater
organisms
have
been
tested
with
2,4­

dimethyl­
phenol
and
therefore,
no
statement
can
be
made
concerning
acute
or
chronic
toxicity.

Human
Health
Sufficient
data
are
not
available
for
2,4­
dimethylphenol
to
derive
a
level
which
would
protect
against
the
potential
toxicity
of
this
compound.
Using
available
organoleptic
data,
to
control
undesirable
taste
and
odor
quality
of
ambient
water
the
estimated
level
is
400
ug/
L.
It
should
be
recognized
that
organoleptic
data
have
limitations
as
a
basis
for
establishing
a
water
quality
criterion,
and
have
no
demonstrated
relationship
to
potential
adverse
human
health
effects.

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
CRITERIA:
DINITROTOLUENE
Aquatic
Life
The
available
data
for
dinitrotoluenes
indicate
that
acute
and
chronic
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
330
and
230
ug/
L,
respectively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.

The
available
data
for
dinitrotoluenes
indicate
that
acute
toxicity
to
saltwater
aquatic
life
occurs
at
concentrations
as
low
as
5
9
0
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
dinitrotoluenes
to
sensitive
saltwater
aquatic
life
but
a
decrease
in
algal
cell
numbers
occurs
at
concentrations
as
low
as
370
ug/
L.

Human
Health
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
2,4­
dinitrotoluene
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentration
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.
However,

zero
level
may
not
be
attainable
at
the
present
time.
Therefore,

the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
and
The
corresponding
recommended
criteria
are
1.1
ug/
L,
0.11
ug/
L,
and
0.011
ug/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organ­
isms
only,
excluding
consumption
of
water,
the
levels
are
91
ug/
L,
9.1
ug/
L,
and
0.91
ug/
L,

respectively.

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
DIPHENYLHYDRAZINE
Aquatic
Life
The
available
data
for
1,2­
diphenylhydrazine
indicate
that
acute
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
270
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
1,2­
diphenylhydrazine
to
sensitive
freshwater
aquatic
life.

NO
saltwater
organisms
have
been
tested
with
1,2­

diphenylhydrazine
and
therefore,
no
statement
.
can
be
made
concerning
acute
or
chronic
toxicity.

Human
Health
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
diphenylhydrazine
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentrations
shopld
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
and
The
corresponding
recommended
criteria
are
422
ng/
L,
42
ng/
L,
and
4
ng/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
5.6
ug/
L,
0.56
ug/
L,
and
0.056
ug/
L,
0
respectively.

0
(
4
5
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
ENDOSULFAN
Aquatic
Life
For
endosulfan
the
criterion
to
protect
freshwater
aquatic
life
as
derived
using
the
Guidelines
is
0.056
ug/
L
as
a
24­
hour
average
and
the
concentration
should
not
exceed
0.22
ug/
L
at
any
time.

For
endosulfan
the
criterion
to
protect
saltwater
aquatic
life
as
derived
using
the
Guidelines
is
0.0087
ug/
L
as
a
24­
hour
average
and
the
concentration
should
not
exceed
0.034
ug/
L
at
any
time.

Human
Health
For
the
protection
of
human
'
health
from
the
toxic
properties
of
endosulfan
ingested
through
water
and
contaminated
aquatic
organisms,
the
ambient
water
criterion
is
determined
to
be
74
ug/
L.

For
the
protection
of
human
health
from
the
toxic
properties
of
endosulfan
ingested
through
contaminated
aquatic
organisms
alone,
the
ambient
water
criterion
is
determined
to
be
159
ug/
L.

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
CRITERIA:
*
ENDRIN
Aquatic
Life
For
endrin
the
criterion
to
protect
freshwater
aquatic
life
as
derived
using
the
Guidelines
is
0.0023
ug/
L
as
a
24­
hour
average,
and
the
concentration
should
not
exceed
0.18
ug/
L
at
any
time.

For
endrin
the
criterion
to
protect
saltwater
aquatic
life
as
derived
using
the
Guidelines
is
0.0023
ug/
L
as
a
24­
hour
average,

and
the
concentration
should
not
exceed
0.037
ug/
L
at
any
time.

Human
Health
The
ambient
water
quality
criterion
for
endrin
is
recommended
to
be
identical
to
the
existing
water
standard
which
is
1.0
ug/
L.

Analysis
of
the
toxic
effects
data
resulted
in
a
calculated
level
which
is
protective
of
human
health
against
the
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms.
The
calculated
value
is
comparable
to
the
present
standard.
For
this
reason
a
selective
criterion
based
on
exposure
solely
from
assumption
of
6.5
g
of
aquatic
organisms
was
not
derived.

*
Indicates
suspended,
canceled
or
restricted
by
W.
S.
EPA
Office
of
Pesticides
and
Toxic
Substances
(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
0
CRITERIA:

The
avai
3
B
da
ETHYLBENZENE
A
q
u
a
t
i
c
Life
3
for
ethylbenzene
idicate
t
at
acute
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
32,000
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
definitive
data
are
available
concerning
the
chronic
toxicity
of
ethylbenzene
to
sensitive
freshwater
aquatic
life.

The
available
data
for
ethylbenzene
indicate
that
acute
toxicity
to
saltwater
aquatic
life
occurs
at
concentrations
as
low
as
430
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
ethylbenzene
to
sensitive
saltwater
aquatic
life.

Human
Health
For
the
protection
of
human
health
from
the
toxic
properties
of
ethylbenzene
ingested
through
water
and
contaminated
aquatic
organisms,
the
ambient
water
criterion
is
determined
to
be
1.4
W/
L.

For
the
protection
of
human
health
from
the
toxic
properties
of
ethylbenzene
ingested
through
contaminated
aquatic
organisms
alone,
the
ambient
water
criterion
is
determined
to
be
3.28
mg/
L.

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
NOTE:
The
U.
S.
EPA
is
currently
developing
Acceptable
Daily
Intake
(
ADI)
or
Verified
Reference
Dose
(
RfD)
values
for
Agency­
wide
use
for
this
chemical.
The
new
value
should
be
substituted
when
it
becomes
available.
The
January,
1986,
draft
Verified
Reference
Dose
document
cites
an
RfD
of
0.1
mg/
kg/
day
for
ethylbenzene.
FLUORANTHENE
Aquatic
Life
The
available
data
for
fluoranthene
indicate
that
acute
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
l
o
w
as
3,980
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
f
luoranthene
to
sensitive
­
freshwater
aquatic
life.

The
available
data
 or
fluoranthene
indicate
that
acute
and
chronic
toxicity
to
saltwater
aquatic
life
occur
at
concentrations
as
low
as
40
and
16
ug/
L,
respectively,
and
would
occur
at
lower
concentrations'
among
species
that
are
more
sensitive
than
those
tested.
a
Human
Health
For
the
protection
of
human
health
from
the
toxic
properties
of
fluoranthene
ingested
through
water
and
contaminated
aquatic
organisms,
the
ambient
water
criterion
is
determined
to
be
42
ug/
L.

For
the
protection
of
human
health
from
the
toxic
properties
of
fluoranthene
ingested
through
contaminated
aquatic
organisms
alone,
the
ambient
water
criterion
is
determined
to
be
54
ug/
L.

(
45
F.
R.
79318,
Novembe­
r
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
GASES,
TOTAL
DISSOLVED
CRITERION:

To
protect
freshwater
and
marine
aquatic
life,
the
total
dissolved
gas
concentrations
in
water
should
not
exceed
110
percent
of
the
saturation
value
for
gases
at
the
existing
atmospheric
and
hydrostatic
pressures.

RATIONALE:

Fish
in
water
containing
excessive
dissolved
gas
pressure
or
tension
are
killed
when
dissolved
gases
in
their
circulatory
system
come
out
of
solution
to
form
bubbles
(
emboli)
which
block
the
flow
of
blood
through
the
capillary
vessels.
In
aquatic
organisms
this
is
commonly
referred
to
as
"
gas
bubble
diseasett.

External
bubbles
(
emphysema)
also
appear
in
the
fins,
on
the
opercula,
in
the
skin
and
in
other
body
tissues.
Aquatic
invertebrates
are
also
affected
by
gas
bubble
disease,
but
usua1l.
y
at
supersaturation
levels
higher
than
those
lethal
to
fish.
0
The
standard
method
of
analyzing
for
gases
in
solutions
has
been
the
Van
Slyke
method
(
Van
Slyke
et
al.
1934);
now,
gas
chromatography
also
is
used
f
o
r
determination
of
individual
and
total
gases.
For
determination
of
total
gas
pressure,
Weiss
has
developed
the
saturometer,
a
device
based
upon
a
thin­
wall
silicone
rubber
tube
that
is
permeable
to
gases
but
impermeable
to
water.
Gases
pass
from
the
water
through
the
tube,
thus
0
raising
the
internal
gas
pressure
which
is
measured
by
a
manometer
or
pressure
gauge
connected
to
the
tube
(
NAS,
1974).

This
method
alone
does
not
separate
the
total
gas
pressure
into
the
separate
components,
but
Winkler
oxygen
determinations
can
be
run
simultaneously,
and
gas
concentrations
can
be
calculated.

Total
dissolved
gas
concentrations
must
be
determined
because
analysis
of
individual
gases
may
not
determine
with
certainty
that
gas
supersaturation
exists.
For
example,
water
could
be
highly
supersaturated
with
oxygen,
but
if
nitrogen
were
at
less
than
saturation,
the
saturation
as
measured
by
total
gas
pressure
might
not
exceed
100
percent.
Also,
if
the
water
was
highly
supersaturated
with
dissolved
oxygen,
the
oxygen
alone
might
be
sufficient
to
create
gas
pressures
or
tensions
greater
than
the
Criterion
limits,
but
one
would
not
know
the
total
gas
pressure
or
tension,
or
by
how
much
the
criterion
was
exceeded.
The
rare
and
inert
gases
such
as
argon,
neon
and
helium
are
not
usually
involved
in
causing
gas
bubble
disease
as
their
contribution
to
total
gas
pressures
is
very
low.
Dissolved
nitrogen
(
NZ),
which
comprises
roughly
80
percent
of
the
earth's
atmosphere,
is
nearly
inert
biologically
and
is
the
most
significant
cause
of
gas
bubble
disease
in
aquatic
animals.
Dissolved
oxygen,
which
is
extremely
bioactive,
is
consumed
by
the
metabolic
processes
of
the
organism
and
is
less
important
in
causing
serious
gas
bubble
disease
though
it
may
be
involved
in
initiating
emboli
formation
in
the
blood
(
Nebeker
et
al.
1976a).

Percent
saturation
of
water
containing
a
given
amount
of
gas
varies
with
the
absolute
temperature
and
with
the
pressure.

Because
of
the
pressure
changes,
percent
saturation
with
a
given
amount
of
gas
changes
with
depth
of
the
water.
Gas
supersaturation
decreases
by
10
percent
per
meter
of
increase
in
water
depth
because
of
hydrostatic
pressure;
a
gas
that
is
at
130
percent
saturation
at
the
surface
would
be
at
100
percent
saturation
at
3
meters'
depth.
Compensation
 or
altitude
may
be
needed
because
a
reduction
in
atmospheric
pressure
changes
the
water/
gas
equilibria,
resulting
in
changes
in
solubility
of
dissolved
gases.

There
are
several
ways
that
total
dissolved
gas
supersaturation
can
occur:

1.
Excessive
biological
activity­­
dissolved
oxygen
concentrations
often
reach
supersaturation
because
of
excessive
algal
photosynthesis.
Renfro
(
1963)
reported
gas
bubble
disease
in
fishes
resulting,
in
part,
from
algal
blooms.
Algal
blooms
often
accompany
an
increase
in
water
temperature
and
this
higher
temperature
further
contributes
to
supersaturation.
I
2.
Lindroff
(
1957)
reported
that
water
spillage
at
hydropower
dams
caused
supersaturation.
When
excess
water
is
spilled
over
the
face
of
a
dam
it
entrains
air
as
it
plunges
to
the
stilling
or
plunge
pool
atthebaseofthedam.
Themomentum
of
the
fall
carries
the
water
and
entrained
gases
to
great
depths
in
the
pool;
and,
under
increased
hydrostatic
pressure,
the
entrained
gases
are
driven
into
solution,
causing
supersaturation
of
dissolved
gases.

3.
Gas
bubble
disease
may
be
induced
by
discharges
from
power­
generating
and
other
thermal
sources
(
Marcello
et
al.
./
1975).
Cool,
gas­
saturated
water
is
heated
as
it
passes
through
the
condenser
or
heat
exchanger.
As
the
temperature
of
the
water
rises,
percent
saturation
increases
because
of
the
reduced
solubility
of
gases
at
higher
temperatures.
Thus,
the
discharged
water
becomes
supersaturated
with
gases
and
fish
or
other
organisms
living
in
the
heated
water
may
exhibit
gas
bubble
disease
(
DeMont
and
Miller,
1972:
Malouf
et
al.
1972;
Keup,

1975).
!
In
recent
years,
gas
bubble
disease
has
been
identified
as
a
major
problem
affecting
valuable
stocks
of
salmon
and
trout
in
the
Columbia
River
system
(
Rulifson
and
Abel,
1971).
The
disease
is
caused
by
high
concentrations
of
dissolved
atmospheric
gas
which
enter
the
river's
water
during
heavy
spillinq
at
hydroelectric
dams.
A
report
by
Ebel
et
al.
(
1975)
presents
results
from
field
and
laboratory
studies
on
the
lethal,

sublethal
and
physiological
effects
of
gas
on
fish,
depth
distribution
of
fish
in
the
river
(
fish
can
compensate
for
some
high
concentrations
of
gas
by
moving
deeper
into
the
water
column),
detection
and
avoidance
of
gas
concentrations
by
fish,

intermittent
exposure
of
fish
to
gas
concentrations,
and
bioassays
of
many
species
of
fish
exposed
to
different
concentrations
of
gas.
Several
conclusions
resulting
from
these
studies
are:

1.
When
either
juvenile
or
adult
salmonids
are
confined
to
shallow
water
(
1
m),
substantial
mortality
occurs
at
and
above
115
percent
total
dissolved
gas
saturation.

2.
When
either
juvenile
or
adult
salmonids
are
free
to
sound
and
obtain
hydrostatic
compensation
either
in
the
laboratory
or
in
the
field,
substantial
mortality
still
occurs
when
saturation
,
levels
(
of
total
dissolved
gases)
exceed
120
percent
saturation.

3.
On
the
basis
of
survival
estimates
made
in
the
Snake
River
from
1966
to
1975,
it
is
concluded
that
juvenile
fish
losses
ranging
from
40to
95
p
e
r
c
e
n
t
d
o
o
c
c
u
r
a
n
d
a
m
a
j
o
r
p
o
r
t
i
o
n
of
this
mortality
can
be
attributed
to
fish
exposure
to
supersaturation
by
atmospheric
gases
during
years
of
high
flow.

4.
Juvenile
salmonids
subjected
to
sublethal
periods
of
exposure
to
supersaturation
can
recover
when
returned
to
normally
saturated
water,
but
adults
do
not
recover
and
generally
die
from
direct
and
indirect
effects
of
the
exposure.

5.
Some
species
of
salmon
and
trout
can
detect
and
avoid
supersaturated
water;
others
may
not.

6.
Higher
survival
was
'
observed
during
periods
of
intermittent
exposure
than
during
continuous
exposure.

7.
In
general,
in
acute
bioassays,
salmon
and
trout
were
less
tolerant
than
the
nonsalmonids.

Dawley
and
Ebel
(
1975)
found
that
exposure
of
juvenile
spring
chinook
salmon,
Oncorhynchus
tshawytscha,
­­
and
steelhead
trout,

Salmo
qairdneri,
to
120
percent
saturation
for
1.5
days
resulted
in
over
50
percent
mortality;
100
percent
mortality
occurred
in
less
than
3
days.
They
also
determined
that
the
threshold
level
where
significant
mortalities
begin
occurring
is
at
115
percent
nitrogen
saturation
(
111
percent
total
gas
saturation
in
this
test).

Rucker
(
1974),
using
juvenile
coho
salmon,
Oncorhynchus
kisutch,
determined
the
effect
of
individual
ratios
of
oxygen
and
nitrogen
and
established
that
a
decrease
in
lethal
effect
­.,

occurred
when
the
nitrogen
content
fell
below
109
percent
saturation
even
though
total
gas
saturation
remained
at
119
percent
saturation,
indicating
the
importance
of
determining
the
concentration
of
the
individual
components
(
02
and
N
2
)
of
the
atmospheric
supersaturation.
Nebeker
et
al.
(
1976a),
using
juvenile
sockeye
salmon,
Oncorhynchus
nerka,
also
showed
that
there
was
a
significant
increase
in
fish
mortality
when
the
nitrogen
concentration
was
increased
while
holding
the
total
percent
saturation
constant.
They
also
showed
that
there
was
no
significant
difference
in
fish
mortality
at
different
C02
concentrations.

Research
collected
by
Bouck
et
al.
(
1975)
showed
that
gas
supersaturated
water
at
and
above
115
percent
total
gas
saturation
is
acutely
lethal
to
most
species
of
salmonids,
with
120
percent
saturation
and
above
rapidly
lethal
to
all
salmonids
tested.
Levels
as
low
as
110
percent
will
produce
emphysema
in
most
species.
Steelhead
trout
were
most
sensitive
to
gas­

supersaturatea
water
followed
by
sockeye
salmon,
Oncorhyncnus
nerka.
Chinook
salmon,
______
Oncorhynchus
_____
­_­­_
tshawytscha
­­­­­
I
were
intermediate
in
sensitivity.
Coho
salmon,
Oncorhyncnus
kisutch,

were
significantly
the
more
tolerant
of
the
salmonids
though
­­­­­

still
much
more
susceptible
than
non­
salmonids
like
bass
or
carp.

­
DaFnia
­­
rmqna
__
exhibited
a
sensitivity
to
supersaturation
similar
to
that
of
the
salmonids
(
Nebeker
et
al.
1975),
with
115
percent
saturation
lethal
within
a
few
days.
Stoneflies
exhibited
an
intermediate
sensitivity
similar
to
bass
with
mortality
at
130
percent
saturation.
Crayfish
were
very
tolerant,
with
levels
near
140
percent
total
gas
saturation
resulting
in
mortality.
No
d
i
f
f
e
r
e
n
c
e
s
are
proposed
i
n
t
h
e
c
r
i
t
e
r
i
a
 
o
r
freshwater
and
marine
aquatic
l
i
f
e
a
s
t
h
e
data
a
v
a
i
l
a
b
l
e
indicate
t
h
a
t
there
probably
is
l
i
t
t
l
e
d
i
f
f
e
r
e
n
c
e
i
n
o
v
e
r
a
l
l
t
o
l
e
r
a
n
c
e
s
between
marine
and
freshwater
species.

The
development
of
gas
bubble
disease
i
n
menhaden,
S
o
o
r
t
i
a
sp.,
and
t
h
e
i
r
t
o
l
e
r
a
n
c
e
t
o
g
a
s
s
a
t
u
r
a
t
i
o
n
i
n
l
a
b
o
r
a
t
o
r
y
bioassays
and
i
n
t
h
e
f
i
e
l
d
(
P
i
l
g
r
i
m
Nuclear
Power
S
t
a
t
i
o
n
Discharge
Canal)
a
r
e
discussed
by
Clay
e
t
a
l
.
(
1975)
and
Marcello
e
t
a
l
.
(
1
9
7
5
)
.
A
t
1
0
0
p
e
r
c
e
n
t
and
1
0
5
p
e
r
c
e
n
t
n
i
t
r
o
g
e
n
saturation,
no
gas
bubbles
developed
e
x
t
e
r
n
a
l
l
y
o
r
i
n
any
of
t
h
e
i
n
t
e
r
n
a
l
organs
of
menhaden.
A
t
105
percent
nitrogen
saturation,

however,
c
e
r
t
a
i
n
b
e
h
a
v
i
o
r
a
l
changes
became
apparent.
Fish
sloughed
o
f
 
mucus,
s
w
a
m
e
r
r
a
t
i
c
a
l
l
y
,
were
more
e
x
c
i
t
a
b
l
e
,
and
became
darker
i
n
color.
Menhaden
behavioral
changes
observed
a
t
110
p
e
r
c
e
n
t
n
i
t
r
o
g
e
n
s
a
t
u
r
a
t
i
o
n
were
s
i
m
i
l
a
r
t
o
t
h
o
s
e
noted
a
t
3.05
percent.
I
n
a
d
d
i
t
i
o
n
,
a
t
110
p
e
r
c
e
n
t
gas
emboli
were
found
i
n
t
h
e
i
n
t
e
s
t
i
n
e
s
,
t
h
e
p
y
l
o
r
i
c
caeca,
and
o
c
c
a
s
i
o
n
a
l
l
y
t
h
e
operculum.
The
behavioral
changes
described
w
e
r
e
a
l
s
o
observed
a
t
115
p
e
r
c
e
n
t
,
and
c
l
e
a
r
l
y
d
e
f
i
n
e
d
subcutaneous
emphysema
was
observed
i
n
t
h
e
f
i
n
s
and
o
c
c
a
s
i
o
n
a
l
l
y
i
n
t
h
e
eye.
A
t
1
2
0
p
e
r
c
e
n
t
and
130
percent
nitrogen
saturation,
menhaden
developed
within
a
f
e
w
hours
c
l
a
s
s
i
c
symptoms
of
g
a
s
bubble
d
i
s
e
a
s
e
.
E
x
t
e
r
n
a
l
l
y
,

emboli
were
e
v
i
d
e
n
t
i
n
a
l
l
f
i
n
s
,
t
h
e
operculum
and
w
i
t
h
i
n
the
o
r
a
l
cavity.

Exophthalmia
a
l
s
o
occurred
and
emboli
developed
i
n
i
n
t
e
r
n
a
l
The
bulbous
a
r
t
e
r
i
o
s
i
s
and
s
w
i
m
b
l
a
d
d
e
r
w
e
r
e
s
e
v
e
r
e
l
y
J
distended,
and
emboli
were
found
along
t
h
e
l
e
n
g
t
h
of
t
h
e
g
i
l
l
a
r
t
e
r
i
o
l
e
s
,
r
e
s
u
l
t
i
n
g
i
n
hemostasis.
A
t
water
temperatures
of
30
organs.

0
OC,
menhaden
did
not
survive,
regardless
of
gas
saturation
level.

At
water
temperatures
of
15
,
22
,
and
25
OC
100
percent
of
the
menhaden
died
within
24
hours
at
120
percent
and
130
percent
gas
saturation.
Fifty
percent
died
after
96
hours
at
115
percent
(
22
OC)
Menhaden
survival
after
96
hours
at
110
percent
nitrogen
saturation
ranged
from
92
percent
at
22O
and
25'
to
83
percent
at
15
OC.
Observations
on
the
relationship
between
the
mortality
rate
of
menhaden
and
gas
saturation
levels
at
Pilgrim
Station
during
the
April
1975,
incident
suggest
that
the
fish
may
tolerate
somewhat
higher
gas
saturation
levels
in
nature.

It
has
been
shown
by
Bouck
et
al.
(
1975)
and
Dawley
et
al.

(
1975)
that
survival
of
salmon
and
steelhead
smolts
in
seawater
is
not
affected
by
prior
exposure
to
gas
supersaturation
while
in
fresh
water.
No
significant
mortality
of
juvenile
coho
and
sockeye
salmon
occurred
when
they
were
exposed
to
sublethal
concentrations
of
supersaturated
water
and
then
transferred
to
seawater
(
Nebeker
et
al.
197633).

(
QUALITY
CRITERIA
FOR
WATER,
JULY
1976)
PB­
263943
SEE
APPENDIX
C
FOR
METHODOLLIGY
GUTHION
CRITERION:

.01
ug/
L
for
freshwater
and
marine
aquatic
life.

RATIONALE
:

Ninety­
six­
hour
LC50
values
for
fish
exposed
to
the
organophosphorus
pesticide
guthion
range
from
4
to
4270
ug/
L
(
Katz,
1961:
Pickering
et
al.
1962;
Lahav
and
Sarig,
1969;
Macek
et
al.
1969;
Macek
and
McAllister,
1970).
The
only
long­
term
fish
exposure
data
available
are
those
obtained
recently
by
Adelman
and
Smith
(
unpublished
data)
Decreased
spawning
(
eggs
produced
per
female)
was
observed
in
fathead
minnows,
PimephalE
prrelas
­­­
I
exposed
during
a
complete
life
cycle.
An
estimated
"
safe"
long­
term
exposure
concentration
f
o
r
fathead
minnows
lies
between
0.3
and
0.5
ug/
L.
survival
of
larvae
was
reduced
at
approximately
0.7
ug/
L.
0
An
investigation
of
the
persistence
of
guthion
in
fish
revealed
that
50
percent
of
the
chemical
was
lost
in
less
than
one
week
(
Meyer,
1965).
Analysis
of
plankton
and
pond
water
in
the
same
study
indicated
a
50
percent
loss
of
guthioninabout48
hours.
Flint
et
al.
(
1970)
determined
the
half­
life
of
guthion
at
30C
in
pond
water
and
in
a
phosphate
buffer
protected
from
light
in
the
laboratory.
The
half­
life
in
pond
water
was
1.2
days
whereas
that
in
the
laboratory
solution
was
10
days.
The
more
rapid
degradation
in
pond
water
was
attributed
to
the
effect
of
sunlight
and
microorganisms.

Organophosphate
pesticides
are
toxic
because
they
inhibit
the
enzyme
acetylcholinesterase
(
AChE)
which
is
essential
to
nerve
impulse
conduction
and
transmission
(
Holland
et
al.
1967).
Weiss
(
1958,
1959,
1961)
demonstrated
that
a
40
to
70
percent
inhibition
of
fish
brain
AChE
usually
is
lethal.
Centrarchids
generally
are
considered
one
of
the
more
sensitive
groups
of
fish
to
guthion
(
Pickering
et
al.
1962;
Weiss
and
Gakstatter,
1964;

Meyer,
1965).
Weiss
and
Gakskatter
(
1964)
found
that
over
a
15­

day
period
bluegills,
­_
Lepomis
__
macrochirus
1
exhibited
AChE
inhibition
at
1.0
ug/
L
guthion
but
not
at
0.1
ug/
L.
Exposure
at
0.05
ug/
L
for
30
days
also
failed
to
produce
inhibition
below
the
range
of
normal
variation,
but
the
authors
stated
that
it
appeared
there
was
a
downward
trend
in
brain
enzyme
activity
and
that
if
exposure
was
continued
a
definite
reduction
might
develop.
Weiss
(
1961)
found
that
about
30
days
were
required
for
fathead
minnow
and
bluegill
brain
AChE
levels
to
recover
after
8
to
24
hours
exposure
to
10
ug/
L
guthion.

Benke
and
Murphy
(
1974)
showed
that
repetitive
injection
of
fish
with
guthion
caused
cumulative
inhibition
of
brain
AChE
and
mortality.
After
substantial
inhibition
by
guthion
exposure,
it
takes
several
weeks
for
brain
AChE
of
fishes
to
return
to
normal
even
though
exposure
is
discontinued
(
Weiss,
1959,
1960;
Carter,

1971).
Inhibition
of
brain
AChE
of
fishes
by
46
percent
or
more
has
been
associated
with
harmful
effects
in
exposures
to
there
organophosphate
pesticides
for
a
life
cycle
(
Eaton,
1970)
and
for
shorter
periods
(
Carter,
1971;
Coppage
and
Duke,
1971;
Coppage,

1972;
Coppage
and
Matthews,
1974:
Post
and
Leasure,
1974;
Coppage
et
al.
in
press).
In
static
tests,
similar
inhibition
of
AChE
and
mortality
were
caused
in
the
sheepshead
minnow,
Cyprinodon
variegatus,
in
2,
24,
48
and
72
hours
at
concentrations
of
50,
7,

3.5
and
3
ug/
L,
respectively
(
Coppage,
1972).
These
data
indicate
that
reduction
of
brain
AChE
activity
of
marine
fishes
by
70
to
80
percent
or
more
in
short­
term
exposures
to
guthion
may
be
associated
with
some
deaths.

There
is
no
evidence
to
indicate
that
guthion
would
cause
adverse
effects
through
the
food
chain.
Tissue
residue
accumulation
for
whole
fish
calculated
from
the
data
of
Meyer
(
1965)
indicate
no
more
than
a
twentyfold
accumulation.
LC50
toxicity
values
for
birds
are
relatively
high
and
range
from
70
to
2,000
mg/
kg
(
Tucker
and
Crabtree,
1970).

Ninety­
six­
hour
LC50
values
for
aquatic
invertebrates
range
from
0.10
to
22.0
ug/
L
(
Nebeker
and
Gaufin,
1964;
Gaufin
et
al.

1965:
Jensen
and
Gaufin,
1966:
Sanders
and
Cope,
1968:
Sanders,

1969,
1972).
Sanders
(
1972)
exposed
the
grass
shrimp,

Paleomonetes
kadiakensis
­­_­__
r
to
guthion
in
a
continuous
flow
bioassay
for
up
to
20
days
and
found
that
the
5­
and
20­
day
LC50
values
were
1.2
and
0.16
ug/
L,
respectively.
He
found
that
the
amphipod,
­­­­
Gammarus
­
I­­_­
fasciatus
1
was
the
most
sensitive
aquatic
organism
tested,
with
a
96­
hour
LC50
of
0.10
ug/
L.
Jensen
and
Gaufin
(
1966),
also
using
a
continuous
flow
system,
exposed
two
species
of
stonefly
naiads
in
4­
and
30­
day
studies.
They
observed
96­
hour
and
30­
day
LC50
values
for
Acroneuria
pacifica
of
2.0
and
0.24
ug/
L,
respectively,
whereas
for
Pteronarcys
­
­
californica
the
values
were
4.6
and
1.3
ug/
L,
respectively.

Results
of
other
toxicity
studies
on
marine
organisms
have
1
.
3
been
reported.
The
24­
hour
LC50
for
the
white
mullet,
_­
M
U
G
1
_
0
curema,
was
found
to
be
5.5
ug/
L
guthion
(
Butler,
1963).
The
96­

hour
LC50
for
the
striped
mullet,
Mugil
cephalus,
was
determined
by
Lahav
and
Sarig
(
1969)
to
be
8
ug/
L
guthion.
Portman
(
1972)

reported
the
48­
hour
LC50
for
the
fish,
­
Pleuronectes
­
limanda,
to
be
10
to
30
ug/
L.
The
48­
hour
LC50
for
the
European
shrimp,

Cranqon
crangon,
was
found
to
be
0.33
ug/
L
guthion
(
Portman,

1972).
Butler
(
1963)
found
that
the
24­
hour
EC50
 or
blue
crab,

­­­­­­­
Callinectes
­
sapidus,
was
550
ug/
L
and
the
48­
hour
EC50
 or
pink
shrimp,
Penaeus
duorarum,
was
4.4
ug/
L
guthion.
The
48­
hour
TLm
was
estimated
to
be
6
2
0
ug/
L
 
o
r
fertilized
oyster
eggs,

Crassostrea
­­
vixinica
­­­
I
and
860
ug/
L
for
fertilized
clam
eggs,

Mercenaria
­
mercenaria
(
Davis
and
Hidu,
1969).

A
criterion
level
of
.01
ug/
L
for
guthion
is
based
upon
use
of
an
0.1
application
factor
applied
to
the
96­
hour
LC50
of
0.1
ug/
L
for
­­­­­
Gammarus
and
a
similar
value
of
0.3
ug/
L
for
the
European
shrimp.

(
QUALITY
CRITERIA
FOR
WATER,
JULY
1976)
PB­
263943
SEE
APPENDIX
C
FOR
METHODOLOGY
HALOETHERS
CRITERIA:

Aquatic
Life
The
available
data
for
haloethers
indicate
that
acute
and
chronic
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
360
and
122
ug/
L,
respectively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.

No
saltwater
organisms
have
been
tested
with
any
haloether
and
therefore,
no
statement
can
be
made
concerning
acute
or
chronic
toxicity.

Human
Health
Using
the
present
guidelines,
a
satisfactory
criterion
cannot
be
derived
at
this
time
because
of
insufficient
available
data
for
ha
1
oethers.
0
(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
HALOMETHANES
CRITERIA:

Aquatic
Life
The
available
data
for
halomethanes
indicate
that
acute
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
a
11,000
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chroni­
c
toxicity
of
halomethanes
to
sensitive
freshwater
aquatic
life.

The
available
data
for
halomethanes
indicate
that
acute
and
chronic
toxicity
to
saltwater
aquatic
life
occurs
at
concentrations
as
low
as
12,000
and
6,400
ug/
L,
respectively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
A
decrease
in
algal
cell
numbers
occurs
at
concentrations
as
low
as
11,500
ug/
L.
0
Human
Health
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
chloromethane,

bromomethane,
dichloromethane,
bromodichloromethane,

tribromomethane,
dichlorodifluoromethane,
trichlorofluoromethane,

or
combinations
of
these
chemicals
through
ingestion
of
contaminated
water
and
aquatic
organisms,
the
ambient
water
concentration
should
be
zero,
based
on
the
non
threshold
assumption
for
this
chemical,
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
0
­
are
estimated
at
loe6
and
The
corresponding
recommended
criteria
are
1.9
ug/
L,
0.19
ug/
L,
and
0.019
uq/
L,

respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
157
ug/
L,
15.7
ug/
L,
and
1.57
ug/
L,
respectively.

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
HARDNESS
I"
RODUC"
ION:

Water
hardness
is
caused
by
the
polyvalent
metallic
ions
disolved
in
water.
In
fresh
water
these
are
primarily
calcium
and
magnesium
although
other
metals
such
as
iron,
strontium
and
manganese
contribute
to
the
extent
that
appreciable
concentrations
are
present.
Hardness
commonly
is
reported
as
an
equivalent
concentration
of
calcium
carborate
(
CaC03).

The
concept
of
hardness
comes
from
water
supply
practice.
It
is
measured
by
soap
requirements
for
adequate
lather
formation
and
as
an
indicator
of
the
rate
of
scale
formation
in
hot
water
heaters
and
low
pressure
boilers.
A
commonly
used
classification
is
given
in
the
following
table
(
Sawyer,
1960).

TABLE
3.

Classification
of
Water
by
Hardness
Content
Conc.
mg/
L
CaC03
Description
0
­
75
75
­
150
150
­
300
300
and
up
soft
moderately
hard
hard
very
hard
Natural
sources
of
hardness
principally
are
limestones
which
are
dissolved
by
percolating
rainwater
made
acid
by
dissolved
carbon
dioxide.
Industrial
and
industrially
related
sources
include
the
inorganic
chemical
industry
and
discharges
from
operating
and
abandoned
mines.

Hardness
in
fresh
water
frequently
is
distinguished
in
carbonate
and
non­
carbonate
fractions.
The
carbonate
fraction
is
chemically
equivalent
to
the
bicarbonates
present
in
water.
I
Since
bicarbonates
generally
are
measured
as
alkalinity,
the
carbonate
hardness
usually
is
considered
equal
to
the
alkalinity.

RATIONALE:

The
determination
of
hardness
in
raw
waters
subsequently
treated
and
used
for
domestic
water
supplies
is
useful
as
a
parameter
to
characterize
the
total
dissolved
solids
present
and
for
calculating
dosages
where
lime­
soda
softening
is
practiced.

Because
hardness
concentrations
in
water
have
not
been
proven
health
related,
the
final
level
achieved
principally
is
a
function
of
economics.
Since
hardness
in
water
can
be
removed
with
treatment
by
such
processes
as
lime­
soda
softening
and
zeolite
or
ion
exchange
systems,
a
criterion
for
raw
waters
used
for
public
water
supply
is
not
practical.

The
effects
of
hardness
on
freshwater
fish
and
other
aquatic
life
appear
to
be
related
to
the
ions
causing
the
hardness
rather
than
hardness.
Both
the
NTAC
(
1968)
and
NAS
(
1974)
panels
have
recommended
against
the
use
of
the
term
hardness
but
suggest
the
inclusion
of
the
concentrations
of
the
specific
ions.
This
procedure
should
avoid
confusion
in
future
studies
but
is
not
helpful
in
evaluating
previous
studies.
For
most
existing
data,

it
is
difficult
to
determine
whether
toxicity
o
f
various
metal
ions
is
reduced
because
of
the
formation
of
metallic
hydroxides
and
carbonates
caused
by
the
associated
increases
in
alkalinity,

or
because
of
an
antagonistic
effect
of
one
of
the
principal
cations
contributing
to
hardness,
e.
g.,
calcium,
or
a
combination
of
both
effects.
Stiff
(
1971)
presented
a
theory
(
without
proof)
t
h
a
t
i
f
cupric
ions
were
t
h
e
t
o
x
i
c
form
of
copper
whereas
copper
carbonate
complexes
were
r
e
l
a
t
i
v
e
l
y
non­
toxic,
then
t
h
e
observed
difference
i
n
t
o
x
i
c
i
t
y
of
copper
between
hard
and
s
o
f
t
waters
can
be
e
x
p
l
a
i
n
e
d
by
t
h
e
d
i
f
f
e
r
e
n
c
e
i
n
a
l
k
a
l
i
n
i
t
y
r
a
t
h
e
r
t
h
a
n
hardness.
Doudoroff
and
K
a
t
z
(
1953),
i
n
t
h
e
i
r
review
of
t
h
e
l
i
t
e
r
a
t
u
r
e
on
t
o
x
i
c
i
t
y
,
presented
d
a
t
a
showing
t
h
a
t
i
n
c
r
e
a
s
i
n
g
calcium
i
n
p
a
r
t
i
c
u
l
a
r
reduced
t
h
e
t
o
x
i
c
i
t
y
of
other
heavy
metals.

Under
u
s
u
a
l
conditions
i
n
fresh
water
and
assuming
t
h
a
t
o
t
h
e
r
b
i
v
a
l
e
n
t
metals
behave
s
i
m
i
l
a
r
l
y
t
o
copper,
it
is
reasonable
t
o
assume
t
h
a
t
both
effects
occur
simultaneously
and
e
x
p
l
a
i
n
t
h
e
observed
reduction
of
t
o
x
i
c
i
t
y
of
m
e
t
a
l
s
i
n
waters
containing
carbonate
hardness.
T
h
e
amount
of
reduced
t
o
x
i
c
i
t
y
related
t
o
hardness,
as
measured
by
a
40­
hour
LC50
f
o
r
rainbow
t
r
o
u
t
,
has
been
estimated
t
o
be
about
f
o
u
r
t
i
m
e
s
f
o
r
copper
and
z
i
n
c
when
t
h
e
hardness
was
increased
from
1
0
t
o
1
0
0
mg/
L
a
s
CaC03
(
NAS,

1974)
­

L
i
m
i
t
s
on
hardness
f
o
r
i
n
d
u
s
t
r
i
a
l
u
s
e
s
a
r
e
q
u
i
t
e
v
a
r
i
a
b
l
e
.

Table
4
lists
maximum
values
t
h
a
t
have
been
accepted
by
various
i
n
d
u
s
t
r
i
e
s
a
s
a
source
of
raw
w
a
t
e
r
(
NAS,
1
9
7
4
)
.
Subsequent
treatment
g
e
n
e
r
a
l
l
y
can
reduce
hardness
t
o
t
o
l
e
r
a
b
l
e
l
i
m
i
t
s
although
c
o
s
t
s
of
such
t
r
e
a
t
m
e
n
t
a
r
e
a
n
important
f
a
c
t
o
r
i
n
determining
its
d
e
s
i
r
a
b
i
l
i
t
y
 or
a
p
a
r
t
i
c
u
l
a
r
water
source.

Hardness
is
not
a
determination
of
concern
f
o
r
i
r
r
i
g
a
t
i
o
n
use
o
f
water.
The
c
o
n
c
e
n
t
r
a
t
i
o
n
s
o
f
t
h
e
c
a
t
i
o
n
s
c
a
l
c
i
u
m
and
magnesium,
which
comprise
hardness,
are
important
i
n
determining
t
h
e
exchangeable
sodium
i
n
a
g
i
v
e
n
water.
T
h
i
s
p
a
r
t
i
c
u
l
a
r
calculation
w
i
l
l
be
discussed
under
t
o
t
a
l
dissolved
s
o
l
i
d
s
rather
TABLE
4.

Maximum
Hardness
Levels
Accepted
By
Industry
as
a
Raw
Water
Source*

Industry
Electric
utilities
Maximum
Concentration
m&
L
as
CaC03
­
5,000
Textile
120
Pulp
and
paper
475
Chemical
1,000
Petroleum
900
Primary
metals
1,000
*
Requirements
for
final
use
within
a
process
may
be
essential11
zero,
which
requires
treatment
f
o
r
concentration
reductions.
than
hardness.

a
(
QUALITY
CRITERIA
FOR
WATER,
JULY
1976)
PB­
263943
SEE
APPENDIX
C
FOR
METHODOLOGY
HEFTACHLOR
CRITERIA:

Aquatic
Life
For
heptachlor
the
criterion
to
protect
freshwater
aquatic
life
as
derived
using
the
Guidelines
is
0.0038
ug/
L
as
a
24­
hour
average,
and
the
concentration
should
not
exceed
0.52
ug/
L
at
any
time.

For
heptachlor
the
criterion
to
protect
saltwater
aquatic
life
as
derived
using
the
Guidelines
is
0.0036
ug/
L
as
a
24­
hour
average,
and
the
concentration
should
not
exceed
0.053
ug/
L
at
any
time.

Human
Health
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
heptachlor
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentration
should
be
zero,
based
on
the
non
threshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
and
The
corresponding
recommended
criteria
are
2.00
ng/
L,
0.20
ng/
L,
and
0.020
ng/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
2.04
ng/
L,
0.20
ng/
L,
and
0.020.
ng/
L,
0
respectively.

0
(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
HEXACHLOROBUTADIENE
CRITERIA:

Aquatic
Life
The
available
data
for
hexachlorobutadiene
indicate
that
acute
and
chronic
toxicity
to
freshwater
aquatic
life
occur
at
concentrations
as
low
as
90
and
9.3
ug/
L,
respectively,

and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.

The
available
data
for
hexachlorobutadiene
indicate
that
acute
toxicity
to
saltwater
aquatic
life
occurs
at
concentrations
as
low
as
?
2
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
I
available
concerning
the
chronic
toxicity
of
hexachlorobutadiene
to
sensitive
saltwater
aquatic
life.
0
Human
Health
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
hexachlorobutadiene
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentrations
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
'
are
estimated
at
and
The
corresponding
recommended
criteria
are
4.47
ug/
L,
0.45
ug/
L,
and
0.045
ug/
L,
respectively.
If
these
estimates
are
0
i
­.

made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
t
h
e
l
e
v
e
l
s
are
500
ug/
L,
50
ug/
L,
and
5.0
ug/
L,
respectively.

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
HEXACHUIROCYCLOHEXANE
CRITERIA:

Aquatic
Life
Lindane
For
lindane
the
criterion
to
protect
freshwater
aquatic
life
as
derived
using
the
Guidelines
is
0.080
ug/
L
as
a
24­
hour
average
and
the
concentration
should
not
exceed
2.0
ug/
L
at
any
time.

For
saltwater
aquatic
life
the
concentration
of
lindane
should
not
exceed
0.16
ug/
L
at
any
time.
No
data
are
available
concerning
the
chronic
toxicity
of
lindane
to
sensitive
saltwater
aquatic
life.

BHC
­
The
available
data
for
a
mixture
of
isomers
of
BHC
indicate
that
acute
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
100
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
a
mixture
of
isomers
of
BHC
to
sensitive
freshwater
aquatic
life.

The
available
data
for
a
mixture
of
isomers
of
BHC
indicate
that
acute
toxicity
to
saltwater
aquatic
life
occurs
at
concentrations
as
low
as
0.34
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
a
mixture
of
isomers
of
BHC
to
sensitive
saltwater
aquatic
life.

0
i.
Human
Health
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
hexachlorocyclohexane
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentrations
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
,
and
10­
over
the
lifetime
are
estimated
at
loq5,

The
corresponding
recommended
criteria
are
2
2
ng/
L,
2.2
7
ng/
L,
and
.
2
2
ng/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
74
ng/
L,
7.4
ng/
L,
and
.74
ng/
L,

respectively.
10­
6
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
hexachlorocyclohexane
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentrations
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
and
The
corresponding
recommended
criteria
are
134
ng/
L,
13.4
ng/
L,

and
1.34
ng/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
450
ng/
L,
45.0
ng/
L,
and
4.50
ng/
L,

respectively.
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
due
to
exposure
of
r­

hexachlorocyclohexane
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentrations
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
and
The
corresponding
recommended
criteria
are
186
ng/
L,
18.6
ng/
L,
and
1.86
ng/
L,

respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
625
ng/
L,
62.5
ng/
L,
and
6.25
ng/
L,
respectively.

For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
technical­

hexachlorocyclohexane
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentrations
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
10­
5,
loe6,
and
The
corresponding
recommended
criteria
are
52
ng/
L,
5.2
ng/
L,
and
­
52
ng/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
174
ng/
L,
17.4
ng/
L,
and
1.74
ng/
L,
respectively.

,
/
.*
,
Using
the
present
guidelines,
s
a
t
i
s
f
a
c
t
o
r
y
c
r
i
t
e
r
i
a
cannot
be
derived
a
t
t
h
i
s
t
i
m
e
f
o
r
d­
and
e­
hexachlorocyclohexane
because
of
i
n
s
u
f
f
i
c
i
e
n
t
a
v
a
i
l
a
b
l
e
data.

(
4
5
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
HEXACHLOROCYCLOPENTADIENE
CRITERIA:

Aquatic
Life
The
available
data
for
hexachlorocyclopentadiene
indicate
that
acute
and
chronic
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
7.0
and
5.2
ug/
L,
respectively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.

The
available
data
for
hexachlorocyclopentadiene
indicate
that
acute
toxicity
to
saltwater
aquatic
life
occurs
at
concentrations
as
low
as
7.0
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
NO
data
are
available
concerning
the
chronic
toxicity
of
hexachlorocyclopentadiene
to
sensitive
saltwater
aquatic
life.
0
Human
Health
For
comparison
purposes,
two
approaches
were
used
to
derive
criterion
levels
for
hexachlorocyclopentadiene.
Based
on
available
toxicity
data,
to
protect
public
health
the
derived
level
is
2
0
6
ug/
L.
Using
available
organoleptic
data,
to
control
undersirable
taste
and
odor
quality
of
ambient
water
the
estimated
level
is
1
ug/
L.
It
should
be
recognized
that
organoleptic
data
have
limitations
as
a
basis
for
establishing
water
quality
criteria,
and
have
no
demonstrated
relationship
to
potential
adverse
human
health
effects.

­.
,
(
4
5
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
0.3
mg/
L
f
o
r
domestic
water
supplies
(
welfare).

1.0
mg/
L
f
o
r
freshwater
aquatic
l
i
f
e
.

INTRODUCTION:

Iron
is
t
h
e
fourth
most
abundant,
by
weight,
of
t
h
e
elements
t
h
a
t
make
up
t
h
e
e
a
r
t
h
'
s
c
r
u
s
t
.
Common
i
n
many
r
o
c
k
s
,
it
is
an
important
component
of
many
s
o
i
l
s
,
e
s
p
e
c
i
a
l
l
y
t
h
e
c
l
a
y
s
o
i
l
s
where
u
s
u
a
l
l
y
it
i
s
a
major
c
o
n
s
t
i
t
u
e
n
t
.
I
r
o
n
i
n
water
may
be
present
i
n
varying
q
u
a
n
t
i
t
i
e
s
dependent
upon
t
h
e
geology
of
t
h
e
area
and
o
t
h
e
r
chemical
components
of
the
waterway.

I
r
o
n
is
an
essential
trace
element
r
e
q
u
i
r
e
d
by
b
o
t
h
p
l
a
n
t
s
and
animals.
I
n
some
waters
it
may
be
a
l
i
m
i
t
i
n
g
f
a
c
t
o
r
f
o
r
the
growth
of
a
l
g
a
e
and
o
t
h
e
r
plants;
t
h
i
s
is
t
r
u
e
e
s
p
e
c
i
a
l
l
y
i
n
some
marl
l
a
k
e
s
where
it
is
p
r
e
c
i
p
i
t
a
t
e
d
by
t
h
e
h
i
g
h
l
y
a
l
k
a
l
i
n
e
c
o
n
d
i
t
i
o
n
s
.
It
is
a
v
i
t
a
l
oxygen
t
r
a
n
s
p
o
r
t
mechanism
i
n
t
h
e
blood
of
all
vertebrate
and
some
i
n
v
e
r
t
e
b
r
a
t
e
animals.

The
ferrous,
o
r
b
i
v
a
l
e
n
t
(
Fe++),
and
t
h
e
ferric,
o
r
t
r
i
v
a
l
e
n
t
(
Fe+++)
i
r
o
n
s
,
are
t
h
e
primary
forms
o
f
concern
i
n
t
h
e
a
q
u
a
t
i
c
environment,
although
other
forms
may
be
i
n
organic
and
inorganic
wastewater
streams.
The
f
e
r
r
o
u
s
(
Fe++)
form
c
a
n
p
e
r
s
i
s
t
i
n
waters
v
o
i
d
of
d
i
s
s
o
l
v
e
d
oxygen
and
o
r
i
g
i
n
a
t
e
s
u
s
u
a
l
l
y
from
groundwaters
o
r
mines
when
these
a
r
e
pumped
o
r
d
r
a
i
n
e
d
.
For
p
r
a
c
t
i
c
a
l
purposes
t
h
e
f
e
r
r
i
c
(
Fe
+++)
form
i
s
i
n
s
o
l
u
b
l
e
.
I
r
o
n
can
e
x
i
s
t
i
n
n
a
t
u
r
a
l
o
r
g
a
n
o
m
e
t
a
l
l
i
c
o
r
humic
compounds
and
c
o
l
l
o
i
d
a
l
forms.
B
l
a
c
k
o
r
brown
swamp
w
a
t
e
r
s
may
c
o
n
t
a
i
n
i
r
o
n
c
o
n
c
e
n
t
r
a
t
i
o
n
s
o
f
s
e
v
e
r
a
l
mg/
L
i
n
t
h
e
p
r
e
s
e
n
c
e
o
r
absence
o
f
..
.
.,_.
,~
.,.,.

...,,

~~

dissolved
oxygen,
but
t
h
i
s
i
r
o
n
form
has
l
i
t
t
l
e
effect
on
aquatic
l
i
f
e
.

(
QUALITY
CRITERIA
FOR
WATER,
JULY
1976)
PB­
263943
SEE
APPENDIX
C
FOR
METHODOLOGY
a
..

.

i
.
:/
ISOPHORONE
Aquatic
Life
The
available
data
for
isophorone
indicate
that
acute
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
117,000
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
isophorone
to
sensitive
freshwater
aquatic
life.

The
available
data
for
isophorone
indicate
that
acute
toxicity
to
saltwater
aquatic
life
occurs
at
concentrations
as
low
as
12,900
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
isophorone
to
sensitive
saltwater
aquatic
life.
I
Human
Health
For
the
protection
of
human
health
from
the
toxic
properties
of
isophorone
ingested
through
water
and
contaminated
aquatic
organisms,
the
ambient
water
criterion
is
determined
to
be
5.2
W/
L.

For
the
protection
of
human
health
from
the
toxic
properties
of
isophorone
ingested
through
contaminated
aquatic
organisms
alone,
the
ambient
water
criterion
is
determined
to
be
5
2
0
mg/
L.

(
45
F.
R.
79318,
November
28,
1980)

0
;;.
:
SEE
APPENDIX
B
FOR
METHODOLOGY
.
.
­...*.
LEAD
AQUATIC
LIFE
SUMMARY:

The
acute
toxicity
of
lead
to
several
species
of
freshwater
animals
has
been
shown
to
decrease
as
the
hardness
of
water
increases.
At
a
hardness
of
50
mg/
L
the
acute
sensitivities
of
10
species
range
from
142.5
ug/
L
for
an
amphipod
to
235,900
ug/
L
for
a
midge.
Data
on
the
chronic
effects
of
lead
on
freshwater
animals
are
available
for
two
fish
and
two
invertebrate
species.

The
chronic
toxicity
of
lead
also
decreases
as
hardness
increases
and
the
lowest
and
highest
available
chronic
values
(
12.26
and
128.1
ug/
L)
are
both
for
a
cladoceran,
but
in
soft
and
hard
water,
respectively.
Acute­
chronic
ratios
are
available
for
three
species
and
range
from
18
to
62.
Freshwater
algae
are
affected
by
concentrations
of
lead
above
500
ug/
L,
based
on
data
for
four
species.
Bioconcentration
factors
are
available
for
four
invertebrate
and
two
fish
species
and
range
from
42
to
1,700.
0
Acute
values
are
available
for
13
saltwater
animal
species
and
range
from
315
ug/
L
for
the
mummichog
to
27,000
ug/
L
for
the
soft
shell
clam.
A
chronic
toxicity
test
was
conducted
with
a
mysid;
unacceptable
effects
were
observed
at
37
ug/
L
but
not
at
17
ug/
L
and
the
acute­
chronic
ratio
for
this
species
is
124.8.
A
species
of
macroalgae
was
affected
at
20
ug/
L.

Available
bioconcentration
factors
range
from
17.5
to
2,570.

NATIONAL
CRITERIA:

The
procedures
described
in
the
Guidelines
for
Deriving
0
­
2
Numerical
National
Water
Quality
Criteria
for
the
Protection
of
Aquatic
Organisms
and
Their
Uses
indicate
that,
except
possibly
where
a
locally
important
species
is
very
sensitive,
freshwater
aquatic
organisms
and
their
uses
should
not
be
affected
unacceptably
if
the
4­
day
average
concentration
(
in
ug/
L)
of
lead
does
not
exceed
the
numerical
value
given
by
(
1.273
[
ln(
hardness)
3
­
4.705)
more
than
once
every
3
years
on
the
e
average
and
if
the
1­
hour
average
concentration
(
in
ug/
L)
does
not
exceed
the
numerical
value
given.
by
.(
1.273
[
ln(
hardness)
3
­

1.460)
more
than
once
every
3
years
on
the
average.
For
example,

at
hardnesses
of
50,
100,
and
200
mg/
L
as
CaC03
the
4­
day
average
concentrations
of
lead
are
1.3,
3.2,
and
7.7
ug/
L,
respectively,

and
the
1­
hour
average
concentrations
are
34,
82,
and
200
ug/
L.

The
procedures
described
in,
the
Guidelines
indicate
that,

except
possibly
where
a
locally
important
species
is
very
sensitive,
saltwater
aquatic
organisms
and
their
uses
should
not
be
affected
unacceptably
if
the
4­
day
average
concentration
of
lead
does
not
exceed
5.6
ug/
L
more
than
once
every
3
years
on
the
average
and
if
the
1­
hour
average
concentration
does
not
exceed
140
ug/
L
more
than
once
every
three
years
on
the
average.

EPA
believes
that
a
measurement
such
as
tfacid­
solubletl
would
provide
a
more
scientifically
correct
basis
upon
which
to
establish
criteria
for
metals.
The
criteria
were
developed
on
this
basis.
However,
at
this
time,
no
EPA­
approved
methods
for
such
a
measurement
are
available
to
implement
the
criteria
through
the
regulatory
programs
of
the
Agency
and
the
States.

a
The
Agency
is
considering
development
and
approval
of
methods
for
a
measurement
such
as
acid­
soluble.
Until
available,
however,

EPA
recommends
applying
the
criteria
using
the
total
recoverable
method.
T
h
i
s
has
two
impacts:
(
1)
Certain
species
of
some
metals
cannot
be
analyzed
d
i
r
e
c
t
l
y
because
t
h
e
t
o
t
a
l
recoverable
method
does
not
distinguish
between
individual
oxidation
s
t
a
t
e
s
,
and
(
2)

these
c
r
i
t
e
r
i
a
may
be
overly
protective
when
based
on
the
t
o
t
a
l
recoverable
method.

The
recommended
exceedence
frequency
of
3
y
e
a
r
s
is
t
h
e
Agency's
b
e
s
t
s
c
i
e
n
t
i
f
i
c
judgment
of
t
h
e
average
amount
of
time
it
w
i
l
l
t
a
k
e
an
unstressed
system
t
o
recover
from
a
p
o
l
l
u
t
i
o
n
e
v
e
n
t
i
n
which
exposure
t
o
lead
exceeds
t
h
e
c
r
i
t
e
r
i
o
n
.
A
stressed
system,
f
o
r
example,
one
i
n
which
several
o
u
t
f
a
l
l
s
occur
i
n
a
l
i
m
i
t
e
d
a
r
e
a
,
would
be
expected
t
o
r
e
q
u
i
r
e
more
t
i
m
e
f
o
r
recovery.
The
r
e
s
i
l
i
e
n
c
e
of
ecosystems
and
t
h
e
i
r
a
b
i
l
i
t
y
t
o
recover
d
i
f
f
e
r
g
r
e
a
t
l
y
,
however,
'
and
s
i
t
e­
s
p
e
c
i
f
i
c
c
r
i
t
e
r
i
a
may
be
established
i
f
adequate
j
u
s
t
i
f
i
c
a
t
i
o
n
is
provided.

The
use
of
c
r
i
t
e
r
i
a
i
n
designing
waste
treatment
f
a
c
i
l
i
t
i
e
s
r
e
q
u
i
r
e
s
t
h
e
s
e
l
e
c
t
i
o
n
of
an
a
p
p
r
o
p
r
i
a
t
e
wasteload
a
l
l
o
c
a
t
i
o
n
model.
Dynamic
models
a
r
e
preferred
f
o
r
the
application
of
these
c
r
i
t
e
r
i
a
.
L
i
m
i
t
e
d
d
a
t
a
o
r
o
t
h
e
r
f
a
c
t
o
r
s
may
make
t
h
e
i
r
use
i
m
p
r
a
c
t
i
c
a
l
,
i
n
which
c
a
s
e
one
should
r
e
l
y
on
a
steady­
s
t
a
t
e
model.
The
Agency
recommends
t
h
e
interim
use
of
lQ5
or
lQl0
f
o
r
Criterion
Maximum
Concentration
design
flow
and
745
o
r
7Q1Q
f
o
r
t
h
e
C
r
i
t
e
r
i
o
n
Continuous
Concentration
d
e
s
i
g
n
flow
i
n
steady­

s
t
a
t
e
models
f
o
r
unstressed
and
stressed
systems,
respectively.

These
m
a
t
t
e
r
s
a
r
e
discussed
i
n
more
d
e
t
a
i
l
i
n
t
h
e
Technical
Support
Document
f
o
r
Water
Quality­
Based
Toxics
Control
(
U.
S.
0
HUMAN
HEALTH
CRITERIA:

The
ambient
water
quality
criterion
for
lead
is
recommended
to
be
identical
to
the
existing
drinking
water
standard
which
is
50
ug/
L.
Analysis
of
the
toxic
effects
data
resulted
in
a
calculated
level
whic
is
protective
to
human
health
against
the
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms.
The
calculated
value
is
comparable
to
the
present
standard.
For
this
reason
a
selective
criterion
based
on
expoeure
Soley
from
consumption
of
6.5
grams
of
aquatic
organisms
was
not
derived.

(
45
F.
R.
79318
Nov.
28,1980)
(
50
F.
R.
30784,
July
29,
1985)
SEE
APPENDIX
A
FOR
METHODOLOGY
e
CRITERION:
MALATHION
0.1
ug/
L
for
freshwater
and
marine
aquatic
life.

RATIONALE:

The
freshwater
fish
most
sensitive
to
malathion,
an
organophosphorus
pesticide,
appear
to
be
the
salmonids
and
centrarchids.
Post
and
Schroeder
(
1971)
report
a
96­
hour
LC50
between
120
and
265
ug/
L
for
4
species
of
salmonids.
Macek
and
McAllister
(
1970)
found
a
96­
hour
LC50
range
between
101
and
285
ug/
L
for
3
species
of
centrarchids
and
3
species
of
salmonids.

Other
96­
hour
LCSO's
are:
rainbow
trout,
­­
Salmo
gairdrleri,
68
ug/
L
(
Cope,
1965)
i
largemouth
bass,
MicroEerus
­­­­­­­
I
salmoides
50
ug/
L
(
Pickering
et
al.
1962);
and
chinook
salmon,
Oncorhynchus
I­­

tshawytscha
­­­
I
23
ug/
L
(
Katz,
1961).
All
of
the
above
tests
were
in
static
systems.
Eaton
(
1970)
determined
a
96­
hour
LC50
for
bluegill,
­
LeEmis
­
­
­
­
­
­
­
­
I
macrochirus
in
a
flow­
through
system
at110
ug/
L.
Macek
and
McAllister
(
1970)
reported
a
similar
96­
hOUr
LC50
for
the
bluegill
in
a
static
exposure.
Static
96­
hour
LC50s
of
120
and
160
ug/
L
were
reported
by
Post
and
Schroeder
(
1971)

for
brook
trout,
Salvelinus
fontha&.
Bender
(
1969)
indicated
that
the
acute
toxicity
to
fathead
minnows,
Pimephales
­­
promelas,
­

is
slightly
greater
(
about
2.0
times)
in
a
static
system
than
in
a
f
low­
through
system.
The
f
low­
through
acute
toxicity
to
fathead
minnows
reported
by
Mount
and
Stephan
(
1967)
approximated
the
static
acute
toxicity
reported
by
Henderson
and
Pickering
0
(
1958)
and
Bender
(
1969).

0
x
i
Many
aquatic
invertebrates
appear
t
o
be
more
s
e
n
s
i
t
i
v
e
than
f
i
s
h
to
malathion.
The
96­
hour
LC50
f
o
r
Gammarus
l
a
c
u
s
t
r
i
s
w
a
s
1.0
ug/
L
(
Sanders,
1
9
6
9
)
;
f
o
r
P
t
e
r
o
n
a
r
c
e
l
l
a
badia
1
1.1
ug/
L
(
Sanders
and
Cope,
1
9
6
8
)
;
and
f
o
r
­­­­
Gammarus
­­­­­
f
a
s
c
i
a
t
u
s
1
0.76
ug/
L
(
Sanders,
1
9
7
2
)
.
T
h
e
4
8
­
h
O
U
r
LC50
f
o
r
Simocephalus
­­­­­
s
e
r
r
u
l
a
t
u
s
w
a
s
3.5
ug/
L
and
f
o
r
­
Daphnia
­
pulex,
1.8
ug/
L
(
Sanders
and
Cope,

1
9
6
6
)
.
­­
Daphnia
­­­­
were
immobilized
i
n
5
0
h
o
u
r
s
i
n
0.9
ug/
L
(
Anderson,
1960).
The
24­
hour
LC50s
f
o
r
two
s
p
e
c
i
e
s
of
midge
larvae
were
2.1
ug/
L
(
Mulla
and
Xhasawinah,
1969)
and
2.0
ug/
L
(
Karnak
and
Collins,
1974).

S
a
f
e
l
i
f
e
c
y
c
l
e
e
x
p
o
s
u
r
e
c
o
n
c
e
n
t
r
a
t
i
o
n
s
f
o
r
t
h
e
more
s
e
n
s
i
t
i
v
e
i
n
v
e
r
t
e
b
r
a
t
e
s
are
n
o
t
known.
T
h
e
most
s
e
n
s
i
t
i
v
e
a
q
u
a
t
i
c
organisms
probably
have
n
o
t
y
e
t
been
tested;
safe
c
o
n
c
e
n
t
r
a
t
i
o
n
s
f
o
r
t
h
e
m
o
s
t
s
'
e
n
s
i
t
i
v
e
i
n
v
e
r
t
e
b
r
a
t
e
s
exposed
through
a
complete
l
i
f
e
c
y
c
l
e
have
n
o
t
been
determined;
and
effects
of
low
c
o
n
c
e
n
t
r
a
t
i
o
n
s
on
i
n
v
e
r
t
e
b
r
a
t
e
behavior
a
r
e
unknown.

The
s
t
a
b
i
l
i
t
y
of
malathion
i
n
water
is
dependent
on
t
h
e
chemical
and
b
i
o
l
o
g
i
c
a
l
c
o
n
d
i
t
i
o
n
s
of
t
h
e
water
(
P
a
r
i
s
e
t
a
l
.

1975).
Weiss
and
Gakstatter
(
1964)
have
shown
t
h
a
t
t
h
e
h
a
l
f­
l
i
f
e
of
malathion
w
a
s
reduced
from
about
5
months
a
t
pH
6
to
1
t
o
2
weeks
a
t
pH
8.
Eichelberger
and
Lichtenberg
(
1971)
found
t
h
a
t
o
n
l
y
1
0
p
e
r
c
e
n
t
remained
i
n
t
h
e
L
i
t
t
l
e
M
i
a
m
i
R
i
v
e
r
(
pH
7.3­
8.0)

a
f
t
e
r
2
weeks.
Bender
(
1969)
s
t
a
t
e
s
t
h
a
t
one
of
t
h
e
malathion
breakdown
products
may
be
more
t
o
x
i
c
than
t
h
e
parent
compound.

It
h
a
s
been
shown
t
h
a
t
a
measured
c
o
n
c
e
n
t
r
a
t
i
o
n
of
575
ug/
L
malathion
i
n
flowing
seawater
k
i
l
l
s
4
0
to
60
p
e
r
c
e
n
t
of
t
h
e
marine
fish,
Lagodon
rhomboides,
i
n
3.5
hours
and
causes
about
75
percent
brain
a
c
e
t
y
l
c
h
o
l
i
n
e
s
t
e
r
a
s
e
(
AChE)
i
n
h
i
b
i
t
i
o
n
(
Coppage
e
t
a
l
.
1975).
S
i
m
i
l
a
r
i
n
h
i
b
i
t
i
o
n
of
AChE
and
m
o
r
t
a
l
i
t
y
were
caused
i
n
p
i
n
f
i
s
h
i
n
24,
48,
and
72
h
o
u
r
s
'
a
t
measured
concentrations
of
1
4
2
,
9
2
and
58
ug/
L,
r
e
s
p
e
c
t
i
v
e
l
y
.
A
c
o
n
c
e
n
t
r
a
t
i
o
n
o
f
3
1
ug/
L
caused
34
percent
AChE
i
n
h
i
b
i
t
i
o
n
i
n
p
i
n
f
i
s
h
but
no
deaths
i
n
72
hours.
Coppage
and
Matthews
(
1974)
demonstrated
t
h
a
t
death
may
be
a
s
s
o
c
i
a
t
e
d
w
i
t
h
r
e
d
u
c
t
i
o
n
s
of
b
r
a
i
n
AChE
a
c
t
i
v
i
t
y
of
f
o
u
r
marine
f
i
s
h
e
s
by
70
t
o
80
percent
or
more
i
n
short­
term
exposures
t
o
malathion.
Coppage
and
Duke
(
1971)
found
t
h
a
t
moribund
m
u
l
l
e
t
,
Mugil
cephalus,
i
n
an
estuary
sprayed
w
i
t
h
malathion
(
3
oz./
acre)
d
u
r
i
n
g
a
large­
scale
mosquito
c
o
n
t
r
o
l
o
p
e
r
a
t
i
o
n
had
about
98
percent
i
n
h
i
b
i
t
i
o
n
of
b
r
a
i
n
AChE.
T
h
i
s
is
i
n
agreement
w
i
t
h
70
t
o
80
percent
o
r
more
i
n
h
i
b
i
t
i
o
n
of
b
r
a
i
n
AChE
l
e
v
e
l
s
a
t
and
below
which
some
d
e
a
t
h
s
are
l
i
k
e
l
y
t
o
occur
i
n
short­
term
exposure.
S
p
o
t
,
Leiostomus
xanthurus,
and
A
t
l
a
n
t
i
c
c
r
o
a
k
e
r
,

Micropogon
­
u
n
d
u
l
e
,
a
l
s
o
had
s
u
b
s
t
a
n
t
i
a
l
i
n
h
i
b
i
t
i
o
n
of
brain
during
t
h
e
spray
operation
(
70
percent
o
r
more
i
n
h
i
b
i
t
i
o
n
)
.
0
T
o
x
i
c
i
t
y
s
t
u
d
i
e
s
h
a
v
e
been
made
on
a
number
of
marine
animals.
E
i
s
l
e
r
(
1970)
s
t
u
d
i
e
d
t
h
e
96­
hour
LC50
f
o
r
s
e
v
e
r
a
l
marine
fishes
a
t
20
OC
i
n
static,
aerated
seawater.
The
96­
hour
LC50
v
a
l
u
e
s
(
i
n
ug/
L)
were:
Menidia
menidia,
125:
­
Mugil
cephalus,

550;
F
u
n
d
u
l
u
s
rnnialis
_­­­
1
250;
F
u
n
d
u
l
u
s
h
e
t
e
r
o
c
l
i
t
u
s
,
2
4
0
;

­
s
p
h
a
e
r
o
i
d
e
s
­­_­__­­­
­­­_­­­­­
I
m
a
c
u
l
a
t
u
s
3250:
A
n
q
u
i
l
l
a
__
_­___
­­___­__
I
r
o
s
t
r
a
t
a
8
2
:
a
n
d
Thalassoma
bifasciatum,
27.
Katz
(
1961)
reported
t
h
e
s
t
a
t
i
c
24­

hour
LC50
f
o
r
Gasterosteus
a
c
u
l
e
a
t
u
s
i
n
2
5
o/
oo
saltwater
as
76.9
ug/
L
a
c
t
i
v
e
i
n
g
r
e
d
i
e
n
t
.
T
h
e
96­
hour
LC50
f
o
r
s
t
r
i
p
e
d
b
a
s
s
,
Morone
_­­_
­­.­­­­­­­
I
s
a
x
a
t
i
l
i
s
i
n
i
n
t
e
r
m
i
t
t
e
n
t
flowing
seawater
has
been
r
e
p
o
r
t
e
d
a
s
1
4
ug/
L
(
U.
S.
BSFW,
1
9
7
0
)
.

Reporting
on
s
t
u
d
i
e
s
of
t
h
e
t
o
x
i
c
i
t
y
of
malathion
on
marine
i
n
v
e
r
t
e
b
r
a
t
e
s
,
E
i
s
l
e
r
(
1969)
found
t
h
e
96­
hour
LC50
(
s
t
a
t
i
c
,
2
4
o/
oo
s
a
l
i
n
i
t
y
a
e
r
a
t
e
d
)
t
o
be
33
ug/
L
f
o
r
sand
shrimp,
Crangon
septemspinosa;
8
2
ug/
L
f
o
r
grass
shrimp,
S
a
e
m
o
n
e
t
e
s
v
u
l
g
a
r
i
s
;
­­

and
83
ug/
L
f
o
r
hermit
crab,
Pagurus
longicarpus.
Growth
of
oyster,
Crassostrea
virginica,
was
reduced
32
percent
by
96­
hour
e
x
p
o
s
u
r
e
t
o
1
mg/
L
(
B
u
t
l
e
r
,
1
9
6
3
)
.
T
h
e
48­
hour
L
C
5
0
f
o
r
f
e
r
t
i
l
i
z
e
d
eggs
of
oysters
was
estimated
by
Davis
and
Hidu
(
1969)

t
o
be
9.07
mg/
L
and
the
14­
day
LC50
f
o
r
larvae,
2.66
mg/
L.

Malathion
e
n
t
e
r
s
t
h
e
a
q
u
a
t
i
c
environment
p
r
i
m
a
r
i
l
y
as
a
r
e
s
u
l
t
of
its
application
as
an
insecticide.
Because
it
degrades
quite
rapidly
i
n
most
waters,
depending
on
pH,
its
occurrence
is
sporadic
rather
than
continuous.
Because
t
h
e
t
o
x
i
c
i
t
y
is
exerted
through
i
n
h
i
b
i
t
i
o
n
of
AChE
and
because
such
i
n
h
i
b
i
t
i
o
n
may
be
additive
w
i
t
h
repeated
exposures
and
may
be
caused
by
any
of
t
h
e
organophosphorus
insecticides,
inhibition
of
AChE
by
more
than
35
percent
may
be
expected
t
o
r
e
s
u
l
t
i
n
damage
t
o
aquatic
organisms.

An
a
p
p
l
i
c
a
t
i
o
n
f
a
c
t
o
r
of
0.1
i
s
a
p
p
l
i
e
d
t
o
t
h
e
9
6
­
h
O
U
r
LC50
data
f
o
r
Gammarus
l
a
c
u
s
t
r
i
s
,
­
G.
f
a
s
c
i
a
t
i
s
and
Daphnia,
which
a
r
e
a
l
l
approximately
1.0
ug/
L,
yielding
a
c
r
i
t
e
r
i
o
n
of
0.1
ug/
L.

(
QUALITY
CRITERIA
FOR
WATER,
JULY
1976)
PB­
263943
SEE
APPENDIX
C
FOR
METHODOLOGY
MANGANESE
50
ug/
L
for
domestic
water
supplies
(
welfare):

100
ug/
L
for
protection
of
consumers
of
marine
molluscs.

INTRODUCTION:

Manganese
does
not
occur
naturally
as
a
metal
but
is
found
in
various
salts
and
minerals,
frequently
in
association
with
iron
compounds.
The
principal
manganese­
containing
substances
are
manganese
dioxide
(
MnOZ)
,
pyrolusite,
manganese
carbonate
(
rhodocrosite)
and
manganese
silicate
(
rhodonite)
~

The
oxides
are
the
only
important
minerals
mined.
Manganese
is
not
mined
in
the
United
States
except
when
manganese
is
contained
in
iron
ores
that
are
deliberately
used
to
f
o
h
ferro­
manganese
alloys.

The
primary
uses
of
manganese
are
in
metal
alloys,
dry
cell
batteries,
micro­
nutrient
fertilizer
additives,
organic
compounds
used
in
paint
driers
and
as
chemical
reagents.
Permanganates
are
very
strong
oxidizing
agents
of
organic
materials.

Manganese
is
a
vital
micro­
nutrient
for
both
plants
and
animals.
When
manganese
is
not
present
in
sufficient
quantities,

plants
exhibit
chlorosis
(
a
yellowing
of
the
leaves)
or
failure
of
the
leaves
to
develop
properly.
Inadequate
quantities
of
manganese
in
domestic
animal
food
results
in
reduced
reproductive
capabilities
and
deformed
or
poorly
maturing
young.
Livestock
feeds
usually
high
corn
diet
have
sufficient
manganese,
but
beef
cattle
on
a
may
require
a
supplement.
RATIONALE:

Although
inhaled
manganese
dusts
have
been
reported
to
be
toxic
to
humans,
manganese
normally
is
ingested
as
a
trace
nutrient
in
food.
The
average
human
intake
is
approximately
10
mg/
day
(
Sollman,
1957).
Very
large
doses
of
ingested
manganese
can
cause
some
disease
and
liver
damage
but
these
are
not
known
to
occur
in
the
United
States.
Only
a
few
manganese
toxicity
problems
have
been
found
throughout
the
world
and
these
have
occurred
under
unique
circumstances,
i.
e.,
a
well
in
Japan
near
a
deposit
of
buried
batteries
(
McKee
and
Wolf,
1963).

It
is
possible
to
partially
sequester
manganese
with
special
treatment
but
manganese
is
not
removed
in
the
conventional
treatment
of
domestic
waters
(
Riddick
et
al.
1958:
Illig,
1960).

Consumer
complaints
arise
when
manganese
exceeds
a
concentration
of
150
ug/
L
in
water
supplies
(
Griffin,
1960).
These
complaints
are
concerned
primarily
with
the
brownish
staining
of
laundry
and
objectionable
tastes
in
beverages.
It
is
possible
that
the
presence
of
low
concentrations
of
iron
may
intensify
the
adverse
effects
of
manganese.
Manganese
at
concentrations
of
about
10
to
2
0
ug/
L
is
acceptable
to
most
consumers.
A
criterion
for
domestic
water
supplies
of
50
ug/
L
should
minimize
the
objectionable
qua1
ities.

McKee
and
Wolf
(
1963)
summarized
data
on
toxicity
of
manganese
to
freshwater
aquatic
life.
Ions
of
manganese
are
found
rarely
at
concentrations
above
1
mg/
L.
The
tolerance
values
reported
range
from
1.5
mg/
L
to
over
1000
mg/
L.
Thus,

manganese
is
not
considered
to
be
a
problem
in
fresh
waters.

Permanganates
havebeen
reportedtokill
fish
in
8to
18
hours
at
concentrations
of
2.2
to
4.1
mg/
L,
but
permanganates
are
not
persistent
because
they
rapidly
oxidize
organic
materials
and
are
thereby
reduced
and
rendered
nontoxic.

Few
data
are
available
on
the
toxicity
of
manganese
to
marine
organisms.
The
ambient
concentration
of
manganese
is
about
2
ug/
L
(
Fairbridge,
1966).
The
material
is
rapidly
assimilated
and
bioconcentrated
into
nodules
that
are
deposited
on
the
sea
floor.

The
major
problem
with
manganese
may
be
concentration
in
the
edible
portions
of
molluscs,
as
bioaccumulation
factors
as
high
as
12,000
have
been
reported
(
NAS,
1974).
In
order
to
protect
against
a
possible
health
hazard
to
humans
by
manganese
accumulation
in
shellfish,
a
criterion
of
100
ug/
L
is
recommended
for
marine
water.

Manganese
is
not
known
to
be
a
problem
i
q
water
consumed
by
livestock.
At
concentrations
of
slightly
less
than
1
mg/
L
to
a
few
milligrams
per
liter,
manganese
nay
be
toxic
to
plants
from
irrigation
water
applied
to
soils
with
pH
values
lower
than
6.0.

The
problem
may
be
rectified
by
liming
soils
to
increase
the
pH.

Problems
may
develop
with
long­
term
(
20
year)
continuous
irrigation
on
other
soils
with
water
containing
about
10
mg/
L
of
manganese
(
NAS,
1974).
But,
as
stated
above,
manganese
is
rarely
found
in
surface
waters
at
concentrations
greater
than
1
mg/
L.

Thus,
no
specific
criterion
for
manganese
in
agricultural
waters
is
proposed.
In
select
areas,
and
where
acidophilic
crops
ate
cultivated
and
irrigated,
a
criterion
of
200
ug/
L
is
suggested
0
for
consideration.
i
\+./
Most
industrial
users
of
water
can
operate
successfully
where
the
criterion
proposed
for
public
water
supplies
is
observed.

Examples
of
industrial
tolerance
of
manganese
in
water
are
summarized
for
industries
such
as
dyeing,
milk
processing,
paper,

textiles,
photography
and
plastics
(
McKee
and
Wolf,
1
9
6
3
)
.
A
more
restrictive
criterion
may
be
needed
to
protect
or
ensure
product
quality.

(
QUALITY
CRITERIA
FOR
WATER,
JULY
1976)
PB­
2
6
3
9
4
3
SEE
APPENDIX
C
FOR
METHODOLOGY
*
MERCURY
AQUATIC
LIFE
Data
are
SUMMARY
:

a
v
a
i
l
a
b
l
e
on
t
h
e
acute
t
o
x
i
c
i
t
y
of
mercury(
I1)
,
to
28
genera
of
freshwater
animals.
Acute
v
a
l
u
e
s
f
o
r
i
n
v
e
r
t
e
b
r
a
t
e
species
range
from
2.2
ug/
L
f
o
r
Daphnia
pulex
t
o
2,000
ug/
L
f
o
r
t
h
r
e
e
i
n
s
e
c
t
s
.
Acute
v
a
l
u
e
s
f
o
r
f
i
s
h
e
s
range
from
3
0
ug/
L
f
o
r
the
guppy
t
o
1,000
ug/
L
f
o
r
t
h
e
­
Mozambique
tilapia.
Few
data
a
r
e
a
v
a
i
l
a
b
l
e
f
o
r
v
a
r
i
o
u
s
organomercury
compounds
and
mercurous
n
i
t
r
a
t
e
,
and
they
a
l
l
appear
t
o
be
4
t
o
3
1
t
i
m
e
s
more
a
c
u
t
e
l
y
t
o
x
i
c
than
mercury(
I1).

A
v
a
i
l
a
b
l
e
chronic
d
a
t
a
i
n
d
i
c
a
t
e
t
h
a
t
methylmercury
is
t
h
e
most
chronically
toxic
of
t
h
e
tested
mercury
compounds.
T
e
s
t
s
on
methylmercury
w
i
t
h
Daphnia
magna
and
brook
t
r
o
u
t
produced
chronic
v
a
l
u
e
s
less
than
0.07
ug/
L.
For
mercury(
I1)
t
h
e
chronic
v
a
l
u
e
obtained
with
­
D
a
a
n
i
a
__­
m
s
n
a
­
was
about
1.1
ug/
L
and
t
h
e
acute­

chronic
r
a
t
i
o
was
4.5.
I
n
both
a
l
i
f
e­
c
y
c
l
e
t
e
s
t
and
a
n
e
a
r
l
y
l
i
f
e­
stage
test
on
mercuric
chloride
w
i
t
h
t
h
e
fathead
minnow,
t
h
e
chronic
value
was
less
than
0.26
ug/
L
and
t
h
e
acute­
chronic
r
a
t
i
o
was
over
600.
0
Freshwater
p
l
a
n
t
s
show
a
w
i
d
e
range
of
s
e
n
s
i
t
i
v
i
t
i
e
s
t
o
mercury,
b
u
t
t
h
e
most
s
e
n
s
i
t
i
v
e
p
l
a
n
t
s
a
p
p
e
a
r
t
o
be
l
e
s
s
s
e
n
s
i
t
i
v
e
than
t
h
e
most
s
e
n
s
i
t
i
v
e
freshwater
animals
t
o
both
mercury(
I1)
and
methylmercury.
A
bioconcentration
f
a
c
t
o
r
of
4
,
9
9
4
is
a
v
a
i
l
a
b
l
e
f
o
r
m
e
r
c
u
r
y
(
I
I
)
,
b
u
t
t
h
e
bioconcentration
factors
f
o
r
methylmercury
range
f
r
o
m
4,000
t
o
85,000.

0
\..,
*
I
n
d
i
c
a
t
e
s
suspended,
canceled
o
r
r
e
s
t
r
i
c
t
e
d
by
U.
S.
EPA
Office
of
Pesticides
and
Toxic
Substances
Data
on
the
acute
toxicity
of
mercuric
chloride
are
available
for
2
9
genera
of
saltwater
animals,
including
annelids,
molluscs,

crustaceans,
echinoderms,
and
fishes.
Acute
values
range
from
3.5
ug/
L
for
a
mysid
to
1,678
ug/
L
for
winter
flounder.
Fishes
tend
to
be
more
resistant
and
molluscs
and
crustaceans
tend
to
be
more
sensitive
to
the
acute
toxic
effects
of
mercury(
I1).

Results
of
a
life­
cycle
test
with
the
mysid
show
that
mercury(
I1)

at
a
concentration
of
1.6
ug/
L
significantly
affected
time
of
first
spawn
and
productivity;
the
resulting
acute­
chronic
ratio
was
3.1.

Concentrations
of
mercury
that
affected
growth
and
photosynthetic
activity
of
one
saltwater
diatom
and
six
species
of
brown
algae
range
from
10
to
160
ug/
L.
Bioconcentration
factors
of
10,000
and
40,000
have
been
obtained
for
mercuric
chloride
and
methylmercury
with
an
oyster.

NATIONAL
CRITERIA:

Derivation
of
a
water
quality
criterion
for
mercury
is
more
complex
than
for
most
metals
because
of
methylation
of
mercury
in
sediment,
in
fish,
and
in
the
food
chain
of
fish.
Apparently
almost
all
mercury
currently
being
discharged
is
mercury(
I1).

Thus
mercury(
I1)
should
be
the
only
important
possible
cause
of
acute
toxicity
and
the
Criterion
Maximum
Concentrations
can
be
based
on
the
acute
values
for
mercury(
I1).

The
best
available
data
concerning
long­
term
exposure
of
fish
to
mercury(
I1)
indicates
that
concentrations
above
0.23
ug/
L
caused
statistically
significant
effects
on
the
fathead
minnow
and
caused
the
concentration
of
total
mercury
in
the
whole
body
t
o
exceed
1.0
mg/
kg.
Although
it
is
n
o
t
known
what
p
e
r
c
e
n
t
of
t
h
e
mercury
i
n
t
h
e
f
i
s
h
w
a
s
methylmercury,
it
is
a
l
s
o
n
o
t
known
whether
uptake
from
food
would
increase
the
concentration
i
n
t
h
e
f
i
s
h
i
n
n
a
t
u
r
a
l
s
i
t
u
a
t
i
o
n
s
.
Species
such
as
rainbow
t
r
o
u
t
,
coho
salmon,
and
e
s
p
e
c
i
a
l
l
y
t
h
e
b
l
u
e
g
i
l
l
,
might
s
u
f
f
e
r
chronic
effects
and
accumulate
h
i
g
h
residues
of
mercury
about
t
h
e
same
a
s
t
h
e
fathead
minnow.

W
i
t
h
regard
t
o
long­
term
exposure
t
o
methylmercury,
M
c
K
i
m
e
t
a
l
.
(
1976)
found
t
h
a
t
brook
t
r
o
u
t
can
exceed
the
FDA
action
l
e
v
e
l
without
s
u
f
f
e
r
i
n
g
s
t
a
t
i
s
t
i
c
a
l
l
y
s
i
g
n
i
f
i
c
a
n
t
adverse
effects
on
s
u
r
v
i
v
a
l
,
growth,
o
r
reproduction.
Thus
f
o
r
methylmercury
t
h
e
Final
Residue
Value
would
be
s
u
b
s
t
a
n
t
i
a
l
l
y
lower
than
the
Final
Chronic
Value.

Basing
a
freshwater
c
r
i
t
e
r
i
o
n
on
t
h
e
F
i
n
a
l
Residue
Value
of
0.012
ug/
L
derived
from
the
bioconcentration
f
a
c
t
o
r
of
81,700
f
o
r
methylmercury
w
i
t
h
t
h
e
fathead
minnow
(
Olson
e
t
a
l
.
1975)

e
s
s
e
n
t
i
a
l
l
y
assumes
t
h
a
t
a
l
l
discharged
mercury
is
methylmercury.

On
t
h
e
o
t
h
e
r
hand,
there
is
t
h
e
p
o
s
s
i
b
i
l
i
t
y
t
h
a
t
i
n
f
i
e
l
d
s
i
t
u
a
t
i
o
n
s
uptake
from
food
might
add
t
o
t
h
e
u
p
t
a
k
e
from
water.

S
i
m
i
l
a
r
considerations
apply
t
o
t
h
e
d
e
r
i
v
a
t
i
o
n
of
t
h
e
saltwater
c
r
i
t
e
r
i
o
n
of
0.025
ug/
L
u
s
i
n
g
t
h
e
BCF
of
4
0
,
0
0
0
o
b
t
a
i
n
e
d
f
o
r
methylmercury
w
i
t
h
t
h
e
Eastern
oyster
(
Kopfler,
1974).
Because
t
h
e
F
i
n
a
l
Residue
Values
f
o
r
methylmercury
a
r
e
s
u
b
s
t
a
n
t
i
a
l
l
y
below
t
h
e
F
i
n
a
l
Chronic
Values
f
o
r
m
e
r
c
u
r
y
(
I
I
)
,
it
is
probably
not
too
important
t
h
a
t
many
fishes,
including
t
h
e
rainbow
t
r
o
u
t
,

coho
salmon,
b
l
u
e
g
i
l
l
,
and
haddock
might
n
o
t
be
a
d
e
q
u
a
t
e
l
y
protected
by
the
freshwater
and
saltwater
F
i
n
a
l
Chronic
Values
f
o
r
mercury(
I1).
In
contrast
to
all
the
complexities
of
deriving
numerical
criteria
for
mercury,
monitoring
for
unacceptable
environmental
effects
should
be
relatively
straightforward.
The
most
sensitive
adverse
effect
will
probably
be
exceedence
of
the
FDA
action
level.
Therefore,
existing
discharges
should
be
acceptable
if
the
concentration
of
methylmercury
in
the
edible
portion
of
exposed
consumed
species
does
not
exceed
the
FDA
action
level.

The
procedures
described
in
the
Guidelines
for
Deriving
Numerical
National
Water
Quality
Criteria
for
the
Protection
of
Aquatic
Organisms
and
Their
Uses
indicate
that,
except
possibly
where
a
locally
important
species
is
very
sensitive,
freshwater
aquatic
organisms
and
their
uses
should
not
be
affected
unacceptably
if
the
4­
day
average
concentration
of
mercury
does
I
not
exceed
0.012
ug/
L
more
than
once
every
3
years
on
the
average
and
if
the
1­
hour
average
concentration
does
not
exceed
2.4
ug/
L
more
than
once
every
3
years
on
the
average.
If
the
4­
day
average
concentration
exceeds
0.012
ug/
L
more
than
once
in
a
3
­

year
period,
the
edible
portion
of
consumed
species
should
be
analyzed
to
determine
whether
the
concentration
of
methylmercury
exceeds
the
FDA
action
level.

The
procedures
described
in
the
Guidelines
indicate
that,

except
possibly
where
a
localy
important
species
is
very
sensitive,
saltwater
aquatic
organisms
and
their
uses
should
not
be
affected
unacceptably
if
the
4­
day
average
concentration
of
mercury
does
not
exceed
0.025
ug/
L
more
than
once
every
3
years
on
the
average
and
if
the
1­
hour
average
concentration
does
not
exceed
2.1
ug/
L
more
than
once
every
3
years
on
the
average.
If
the
4­
day
average
concentration
exceeds
0.025
ug/
L
more
than
once
i
n
a
3­
year
period,
t
h
e
e
d
i
b
l
e
protion
of
consumed
species
should
be
a
n
a
l
y
z
e
d
t
o
d
e
t
e
r
m
i
n
e
w
h
e
t
h
e
r
t
h
e
c
o
n
c
e
n
t
r
a
t
i
o
n
of
mathylmercury
exceeds
t
h
e
FDA
action
l
e
v
e
l
.
0
EPA
b
e
l
i
e
v
e
s
t
h
a
t
a
measurement
such
as
"
acid­
soluble"
would
p
r
o
v
i
d
e
a
more
s
c
i
e
n
t
i
f
i
c
a
l
l
y
c
o
r
r
e
c
t
b
a
s
i
s
upon
w
h
i
c
h
t
o
e
s
t
a
b
l
i
s
h
c
r
i
t
e
r
i
a
f
o
r
metals.
The
c
r
i
t
e
r
i
a
w
e
r
e
developed
on
t
h
i
s
basis.
However,
a
t
t
h
i
s
t
i
m
e
,
no
EPA
approved­
methods
f
o
r
such
a
measurement
are
a
v
a
i
l
a
b
l
e
t
o
implement
t
h
e
c
r
i
t
e
r
i
a
through
t
h
e
r
e
g
u
l
a
t
o
r
y
programs
of
t
h
e
Agency
and
t
h
e
S
t
a
t
e
s
.

The
Agency
is
considering
development
and
approval
of
methods
f
o
r
a
measurement
such
as
acid­
soluble.
U
n
t
i
l
a
v
a
i
l
a
b
l
e
,
however,

EPA
recommends
applying
t
h
e
criteria
using
t
h
e
t
o
t
a
l
recoverable
method.
T
h
i
s
has
two
impacts:
(
1)
c
e
r
t
a
i
n
species
of
some
metals
cannot
be
analyzed
d
i
r
e
c
t
l
y
because
t
h
e
t
o
t
a
l
recoverable
method
does
not
d
i
s
t
i
n
g
u
i
s
h
between
i
n
d
i
v
i
d
u
a
l
oxidation
states,
and
(
2
)

these
criteria
may
be
o
v
e
r
l
y
p
r
o
t
e
c
t
i
v
e
when
based
on
t
h
e
t
o
t
a
l
recoverable
method.
0
T
h
e
recommended
exceedence
frequency
o
f
3
y
e
a
r
s
is
t
h
e
Agency's
best
s
c
i
e
n
t
i
f
i
c
judgment
of
t
h
e
a
v
e
r
a
g
e
amount
of
t
i
m
e
it
w
i
l
l
t
a
k
e
an
u
n
s
t
r
e
s
s
e
d
system
t
o
r
e
c
o
v
e
r
from
a
p
o
l
l
u
t
i
o
n
e
v
e
n
t
i
n
which
exposure
t
o
mercury
exceeds
t
h
e
c
r
i
t
e
r
i
o
n
.
A
stressed
system,
f
o
r
example,
one
i
n
which
s
e
v
e
r
a
l
o
u
t
f
a
l
l
s
occur
i
n
a
l
i
m
i
t
e
d
area,
would
be
expected
t
o
r
e
q
u
i
r
e
more
t
i
m
e
f
o
r
recovery.
T
h
e
r
e
s
i
l
i
e
n
c
e
of
ecosystems
and
t
h
e
i
r
a
b
i
l
i
t
y
t
o
recover
d
i
f
f
e
r
g
r
e
a
t
l
y
,
however,
and
s
i
t
e­
specif
ic
criteria
may
be
e
s
t
a
b
l
i
s
h
e
d
i
f
adequate
j
u
s
t
i
f
i
c
a
t
i
o
n
is
provided.
0
..,
,'
The
use
of
criteria
in
designing
waste
treatment
facilities
requires
the
selection
of
an
appropriate
wasteload
a1
location
model.
Dynamic
models
are
preferred
for
the
application
of
these
criteria.
Limited
data
or
other
factors
may
make
their
use
impractical,
in
which
case
one
should
rely
on
a
steady­
state
model.
The
Agency
recommends
the
interim
use
of
1Q5
or
lQl0
for
Criterion
Maximum
Concentration
design
flow
and
745
or
7410
for
the
Criterion
Continuous
Concentration
design
flow
in
steady­

state
models
for
unstressed
and
stressed
systems
respectively.

These
matters
are
discussed
in
more
detail
in
the
Technical
Support
Document
for
Water
Quality­
Based
Toxics
Control
(
U
.
S
EPA,
1985).

HUMAN
HEALTH
CRITERIA
I
For
the
protection
of
human
health
from
the
toxic
properties
of
mercury
ingested
through
water
and
contaminated
aquatic
organisms,
the
ambient
water
criterion
is
determined
to
be
144
nWL.

For
the
protection
of
human
health
from
the
toxic
properties
of
mercury
ingested
through
contaminated
aquatic
organisms
alone,

the
ambient
water
criterion
is
determined
to
be
146
ng/
L.

NOTE:
These
values
include
the
consumption
of
freshwater,
estuarine,
and
marine
species.

(
45
F.
R.
79318
Nov.
28,1980)
(
50
F.
R.
30784,
July
29,
1985)
SEE
APPENDIX
A
FOR
METHODOLOGY
0
CRITERIA:
METHOXYCHMR
100
ug/
L
f
o
r
domestic
water
supply
(
health);

0.03
ug/
L
f
o
r
freshwater
and
marine
aquatic
l
i
f
e
.

RATIONALE
:

The
highest
l
e
v
e
l
of
methoxychlor
found
t
o
have
minimal
o
r
no
long­
term
e
f
f
e
c
t
s
i
n
man
is
2.0
mg/
kg
of
body
weight/
day
(
Lehman,

1965).
Where
adequate
human
data
a
r
e
a
v
a
i
l
a
b
l
e
f
o
r
corroboration
of
t
h
e
animal
r
e
s
u
l
t
s
,
t
h
e
t
o
t
a
l
tfsafe't
drinking
water
i
n
t
a
k
e
level
is
assumed
t
o
be
l/
lOO
of
t
h
e
no­
effect
o
r
minimal
effect
l
e
v
e
l
r
e
p
o
r
t
e
d
f
o
r
t
h
e
most
s
e
n
s
i
t
i
v
e
animal
tested,
i
n
t
h
i
s
case,
man.

Applying
t
h
e
a
v
a
i
l
a
b
l
e
d
a
t
a
and
based
upon
t
h
e
assumptions
t
h
a
t
20
p
e
r
c
e
n
t
of
t
h
e
t
o
t
a
l
i
n
t
a
k
e
of
methoxychlor
is
from
d
r
i
n
k
i
n
g
water,
and
t
h
a
t
t
h
e
average
person
weighs
7
0
kg
and
consumes
2
l
i
t
e
r
s
of
water
per
day,
the
formula
f
o
r
calculating
a
c
r
i
t
e
r
i
o
n
is
2.0
mg/
kg
x
0.2
x
70
kg
x
1/
100
x
1/
2
=
0.14
mg/
L.

A
c
r
i
t
e
r
i
o
n
l
e
v
e
l
f
o
r
domestic
water
supply
of
1
0
0
ug/
L
is
recommended.

Few
data:
are
a
v
a
i
l
a
b
l
e
on
a
c
u
t
e
and
chronic
e
f
f
e
c
t
s
of
methoxychlor
on
freshwater
f
i
s
h
.
Merna
and
E
i
s
e
l
e
(
1973)

observed
reduced
h
a
t
c
h
a
b
i
l
i
t
y
of
f
a
t
h
e
a
d
minnow
p
i
m
z
h
a
l
e
s
­

prcelas)
­­
embryos
a
t
0.125
ug/
L
and
l
a
c
k
of
spawning
a
t
2.0
ug/
L.
Yellow
perch,
Perca
­­
flavescens,
exposed
t
o
0.6
ug/
L
f
o
r
8
months
exhibited
reduced
growth.
The
36­
hour
LC50
concentration
was
7.5
and
22
ug/
L
f
o
r
t
h
e
f
a
t
h
e
a
d
minnow
and
yellow
perch,

.­
r
e
s
p
e
c
t
i
v
e
l
y
.
Korn
and
Earnest
(
1374)
obtained
a
96­
hour
LC50
of
3.3
ug/
L
with
juvenile
stripped
bass,
Morone
­­­­­_
I
saxatilis
exposed
to
methoxychlor
in
a
flowing­
water
bioassay.

Sanders
(
1972)
determined
a
96­
hour
LC50
value
of
0.5
ug/
L
for
the
crayfish,
Orconectes
nais.
Merna
and
Eisele
(
1973)

obtained
a
96­
hour
LC50
value
of
0.61
ug/
L
for
the
scud,
Gammarus
pseudolimnaeus
and
96­
hour
LC5O's
ranging
from
1.59
to
7.05
ug/
L
for
the
crayfish,
Orconectes
­­
I
nais
and
three
aquatic
insect
larvae.
In
28­
day
exposuresI
reduction
in
emergence
of
mayflies,

­­­­­
Stenonema
sp.
and
in
pupation
of
caddisf
lies,
Cheumatospsyche
_­

sp.,
were
observed
at
0.5
and
0.25
ug/
L
concentrations,

respectively.
They
also
found
methoxychlor
to
be
degraded
in
a
few
weeks
or
less
in
natural
waters.

Eisele
(
1974)
conducted
a
study
in
which
a
section
of
a
natural
stream
was
dosed
at
0.2
ug/
L
methoxychlor
for
1
year.

The
near
extinction
of
one
species
of
scud,
wllella
­­­­
azteca,
and
reductions
in
populations
of
other
sensitive
species,
as
well
as
biomass,
were
observed.
Residue
accumulation
of
up
to
1,000
times
the
level
in
the
stream
was
observed
in
first­
year
crayfish,
Orconectes
__
nais.
Metcalf
et
al.
(
1971)
traced
the
rapid
conversion
of
methoxychlor
to
water
soluble
compounds
and
elimination
from
the
tissues
of
snails,
mosquito
larvae
and
mosquitofish.
Thus,
methoxychlor
appears
to
be
considerably
less
bioaccumlative
in
aquatic
organisms
than
some
of
the
other
chlorinated
pesticides.

Methoxyhlor
has
a
very
low
accumulation
rate
in
birds
and
mammals
(
Stickel,
1973),
and
relatively
low
avian
(
Heath
et
al.

1972)
and
mammalian
(
Hodge
et
al.
1950)
toxicities.
No
administrative
guidelines
for
acceptable
levels
in
edible
fish
tissues
have
been
established
by
the
U.
S.
Food
and
Drug
Administration.

The
above
data
indicate
that
0.1
ug/
L
methoxychlor
would
be
just
below
chronic
effect
level
for
the
fathead
minnow
and
one­

fifth
the
acute
toxicity
level
in
a
crayfish
species.
Therefore,

a
criterion
level
of
0.03
ug/
L
is
recommended.
This
criterion
should
protect
fish
as
sensitive
as
striped
bass
and
is
10
times
lower
than
the
level
causing
effects
on
some
invertebrate
populations
in
a
1­
year
dosing
of
a
natural
stream.

Bahner
and
Nimmo
(
1974)
found
the
96­
hour
LC50
of
methoxychlor
for
the
pink
shrimp,
Penaeus
­­­­
I
duorarum
to
be
3.5
ug/
L
and
the
30­
day
LC50
to
be
1.3
ug/
L.
Using
an
application
factor
of
0.01
with
the
pink
shrimp's
acute
toxicity
of
3.5
ug/
L,
the
recommended
criterion
for
the
marine
environment
is
0.03
ug/
L.

Butler
(
1971)
found
accumulation
factors
of
470
and
1,500
for
the
molluscs,
Mercenaria
_­_____­_­
______­___
mercenaria
and
gyg
arenaria,

respectively,
when
exposed
to
1
ug/
L
methoxychlor
for
5
days.

Using
the
1,500
accumulation
factor
as
a
basis,
a
water
concentration
of
0.2
ug/
L
would
be
required
to
meet
the
U.
S.
Food
and
Drug
Administration's
guideline
for
methoxychlor
in
meat
products.
Thus,
the
recommended
marine
criterion
of
0.03
ug/
L
is
an
order
of
magnitude
lower
than
this
concentration.

(
QUALITY
CRITERIA
FOR
WATER,
JULY
1976)
PB­
263943
SEE
APPENDIX
C
FOR
METHODOLOGY
­
.,
MIREX
CRITERION:

0.001
ug/
L
f
o
r
freshwater
and
marine
aquatic
l
i
f
e
.

RATIONAL??.:

Mirex
is
used
t
o
c
o
n
t
r
o
l
t
h
e
imported
f
i
r
e
a
n
t
S
o
l
e
n
z
s
i
s
­

saevissima
richteri
i
n
t
h
e
southeastern
United
States.
Its
use
is
e
s
s
e
n
t
i
a
l
l
y
limited
t
o
the
control
of
t
h
i
s
insect
and
it
is
always
presented
i
n
b
a
i
t
.
I
n
t
h
e
most
common
formulation,

technical
grade
mirex
i
s
dissolved
i
n
soybean
o
i
l
and
sprayed
on
corncob
g
r
i
t
s
.
The
b
a
i
t
produced
i
n
t
h
i
s
manner
consists
of
0.3
p
e
r
c
e
n
t
mirex,
14.1
percent
soybean
o
i
l
and
85
percent
corncob
g
r
i
t
s
.
b
a
i
t
o
f
t
e
n
is
a
p
p
l
i
e
d
a
t
a
rate
of
1.4
kg/
ha,
The
mirex
I
equivalent
t
o
4.2
grams
of
toxicant
per
hectare.

R
e
l
a
t
i
v
e
l
y
few
studies
have
been
made
of
t
h
e
e
f
f
e
c
t
s
of
mirex
on
freshwater
i
n
v
e
r
t
e
b
r
a
t
e
s
og
yhrdr,
o
n
l
y
Dudke
e
t
a
l
.
(
1971)

r
e
p
o
r
t
chemical
analyses
of
mirex
i
n
t
h
e
water.
Their
study
r
e
p
o
r
t
e
d
effects
on
two
c
r
a
y
f
i
s
h
s
p
e
c
i
e
s
exposed
t
o
mirex
by
three
techniques.
F
i
r
s
t
,
f
ield­
col
l
e
c
t
e
d
crayfish
were
exposed
t
o
s
e
v
e
r
a
l
s
u
b
l
e
t
h
a
l
c
o
n
c
e
n
t
r
a
t
i
o
n
s
of
t
e
c
h
n
i
c
a
l
grade
mirex
s
o
l
u
t
i
o
n
s
 
o
r
v
a
r
i
o
u
s
p
e
r
i
o
d
s
o
f
t
i
m
e
;
second,
c
r
a
y
f
i
s
h
w
e
r
e
exposed
t
o
m
i
r
e
x
l
e
a
c
h
e
d
from
b
a
i
t
(
0.3
p
e
r
c
e
n
t
a
c
t
i
v
e
ingredient)
;
and
t
h
i
r
d
,
the
crayfish
w
e
r
e
fed
mirex
bait.

Procambarus
u
a
n
d
i
n
g
i
j
u
v
e
n
i
l
e
s
were
exposed
t
o
1
o
r
5
ug/
L
f
o
r
6
t
o
144
hours,
t
r
a
n
s
f
e
r
r
e
d
t
o
c
l
e
a
n
water
and
observed
f
o
r
10
days.
A
f
t
e
r
5
days
i
n
clean
w
a
t
e
r
,
95
percent
of
t
h
e
animals
exposed
t
o
1
ug/
L
f
o
r
14
hours
w
e
r
e
dead.
Exposure
t
o
5
ug/
L
f
o
r
6,
24,
and
58
hours
r
e
s
u
l
t
e
d
i
n
26,
50,
and
98
percent
m
o
r
t
a
l
i
t
y
10
days
a
f
t
e
r
t
r
a
n
s
f
e
r
t
o
c
l
e
a
n
water.
Crayfish,
Procambarus
0
'..

­
hayi,
were
exposed
to
0.1
and
0.5
ug/
L
for
48
hours.
Four
days
after
transfer
to
clean
water,
65
percent
of
the
animals
exposed
to
0.1
ug/
L
were
dead.
At
the
0.5
ug/
L
concentration,
71
percent
of
the
animals
were
dead
after
4
days
in
clean
water.

Tissue
residue
accumulations
(
wet
weight
basis)
ranged
from
940­

to
27,210­
fold
above
water
concentrations.
In
leached
bait
experiments,
10
bait
particles
were
placed
in
2
liters
of
water
but
isolated
from
20
juvenile
crayfish.
Thirty
percent
of
the
crayfish
were
dead
in
4
days
and
95
percent
were
dead
in
7
days.

Water
analysis
indicated
mirex
concentrations
of
0.86
ug/~.
In
feeding
Qxperiments,
108
crayfish
each
were
fed
one
bait
particle.
Mortality
was
noticed
on
the
first
day
after
feeding,

j8
and
by
the
sixth
day
77
percent
were
dead.
In
another
experiment,

all
crayfish
were
dead
4
days
after
having
been
fed
2
bait
particles
each.
From
this
report
it
is
obvious
that
mirex
is
extremely
toxic
to
these
species
of
crayfish.
Mortality
and
accumulation
increases
with
time
of
exposure
to
the
insecticide.

Concentrations
as
low
as
0.1
ug/
L
or
the
ingestion
of
one
particle
resulted
in
death.
$

Research
to
determine
effects
of
mirex
on
fish
has
been
concentrated
on
species
which
have
economic
and
sport
fishery
importance.
Hyde
et
al.
(
1974)
applied
mirex
bait
(
0.3
percent
mirex)
at
the
standard
rate
(
1.4
kg/
ha)
in
four
ponds
containing
channel
catfish,
Ictalurus
punctatus.
Three
applications
were
made
over
an
8~
011th
period
with
the
first
application
8
days
after
fingerling
(
average
weight
18.4
g
)
catfish
were
placed
in
the
ponds.
Fish
were
collected
at
each
subsequent
application
(
approximately
4­
month
i
n
t
e
r
v
a
l
s
)
.
Two
and
one
h
a
l
f
months
after
t
h
e
f
i
n
a
l
a
p
p
l
i
c
a
t
i
o
n
,
t
h
e
ponds
were
drained,
a
l
l
f
i
s
h
were
measured
and
weighed,
and
t
h
e
p
e
r
c
e
n
t
s
u
r
v
i
v
a
l
was
c
a
l
c
u
l
a
t
e
d
.

Mirex
r
e
s
i
d
u
e
s
i
n
t
h
e
f
i
s
h
a
t
termination
of
t
h
e
experiment
ranged
from
0.015
ug/
g
(
ppm)
i
n
t
h
e
f
i
l
l
e
t
t
o
0.255
uq/
g
i
n
t
h
e
f
a
t
.
0
I
n
another
study,
Van
Valin
e
t
a
l
.
(
1988)
exposed
b
l
u
e
g
i
l
l
s
,

­
Lepomis
macrochirus,
and
t
h
e
g
o
l
d
f
i
s
h
,
Carassius
a
u
r
a
t
u
s
,
t
o
mirex
by
feeding
a
mirex­
treated
d
i
e
t
(
1,
3,
and
5
mg
mirex
p
e
r
kg
body
weight)
o
r
by
t
r
e
a
t
i
n
g
h
o
l
d
i
n
g
ponds
w
i
t
h
mirex
b
a
i
t
(
1.3,
1
0
0
,
and
1000
ug/
L
computed
water
concentration).
They
reported
no
m
o
r
t
a
l
i
t
y
or
t
i
s
s
u
e
pathology
f
o
r
t
h
e
b
l
u
e
g
i
l
l
s
;

however,
a
f
t
e
r
58
days
of
exposure,
g
i
l
l
breakdown
i
n
g
o
l
d
f
i
s
h
was
found
i
n
t
h
e
100
and
1
0
0
0
ug/
L
c
o
n
t
a
c
t
exposure
ponds,
and
kidney
breakdown
was
occurring
i
n
the
1000
ug/
L
ponds.
Mortality
i
n
t
h
e
feeding
experiments
was
n
o
t
r
e
l
a
t
e
d
t
o
t
h
e
l
e
v
e
l
of
exposure,
although
growth
of
t
h
e
b
l
u
e
g
i
l
l
s
fed
5
ug/
L
mirex
w
a
s
reduced.

I
n
laboratory
and
f
i
e
l
d
test
systems,
reported
concentrations
of
mirex
u
s
u
a
l
l
y
are
between
0
.
5
and
1.0
ug/
L
(
Van
Valin
e
t
a
l
.

1968:
Ludke
e
t
a
l
.
1971).
Although
mirex
seldom
is
found
above
1
ug/
L
i
n
t
h
e
aquatic
environment,
several
f
i
e
l
d
studies
have
shown
t
h
a
t
t
h
e
i
n
s
e
c
t
i
c
i
d
e
is
accumulated
through
t
h
e
food
chain.

Borthwick
e
t
a
l
.
(
1973)
r
e
p
o
r
t
e
d
t
h
e
accumulation
of
mirex
i
n
South
Carolina
e
s
t
u
a
r
i
e
s
.
T
h
e
i
r
d
a
t
a
r
e
v
e
a
l
e
d
t
h
a
t
mirex
was
transported
from
treated
land
and
marsh
t
o
t
h
e
estuary
animals
and
t
h
a
t
accumulation,
e
s
p
e
c
i
a
l
l
y
i
n
predators,
occurred.
I
n
t
h
e
test
area,
water
supplies
consistently
w
e
r
e
less
than
0.01
ug/
L.
Residues
i
n
f
i
s
h
v
a
r
i
e
d
from
non­
detectable
percent
of
t
h
e
samples
containing
and
the
percent
of
samples
containing
mirex
residues.
t
o
0.8
ug/
g
w
i
t
h
15
The
amount
of
mirex
increased
a
t
higher
t
r
o
p
h
i
c
l
e
v
e
l
s
.
F
i
f
t
y­
f
o
u
r
p
e
r
c
e
n
t
of
t
h
e
raccoons
sampled
contained
mirex
r
e
s
i
d
u
e
s
up
t
o
4.4
ug/
g
and
78
p
e
r
c
e
n
t
of
t
h
e
b
i
r
d
s
contained
r
e
s
i
d
u
e
s
up
t
o
1
7
ug/
g.
Navgi
and
de
l
a
Cruz
(
1973)
r
e
p
o
r
t
e
d
average
r
e
s
i
d
u
e
s
f
o
r
molluscs
(
0.15
ug/
g),
f
i
s
h
(
0.26
ug/
g),
i
n
s
e
c
t
s
(
0.29
ug/
g),
c
r
u
s
t
a
c
e
a
n
s
(
0
.
4
4
ug/
g)
and
annelids
(
0.63
ug/
g.
They
a
l
s
o
reported
t
h
a
t
mirex
was
found
i
n
areas
n
o
t
t
r
e
a
t
e
d
w
i
t
h
mirex
which
s
u
g
g
e
s
t
s
movement
of
t
h
e
p
e
s
t
i
c
i
d
e
i
n
t
h
e
environment.
Wol
fe
and
Norment
(
1973)
sampled
an
a
r
e
a
f
o
r
one
y
e
a
r
f
o
l
l
o
w
i
n
g
an
a
e
r
i
a
l
a
p
p
l
i
c
a
t
i
o
n
of
mirex
b
a
i
t
(
2.1
g
mirex/
ha).
Crayfish
r
e
s
i
d
u
e
s
ranged
from
0.04
t
o
0.16
ug/
g.
F
i
s
h
r
e
s
i
d
u
e
s
w
e
r
e
about
2
t
o
2
0
t
i
m
e
s
g
r
e
a
t
e
r
t
h
a
n
t
h
e
c
o
n
t
r
o
l
s
and
averaged
from
0.01
t
o
0.78
ug/
g.
Kaiser
(
1974),

r
e
p
o
r
t
e
d
t
h
e
presence
of
rnirex
i
n
f
i
s
h
from
t
h
e
Bay
of
Q
u
i
n
t
e
,

Lake
Ontario,
Canada.
Concentrations
range
from
0.02
ug/
g
i
n
the
gonads
of
t
h
e
northern
long
nose
gar,
Lepistosteus
osseus,
t
o
0.05
ug/
g
i
n
t
h
e
a
r
e
a
l
f
i
n
of
t
h
e
n
o
r
t
h
e
r
n
p
i
k
e
,
___­
Esox
l
u
c
i
u
s
.

Mirex
has
never
been
registered
f
o
r
u
s
e
i
n
Canada.

Mirex
does
n
o
t
appear
t
o
be
g
r
e
a
t
l
y
t
o
x
i
c
t
o
b
i
r
d
s
,
w
i
t
h
LCSO's
f
o
r
t
h
e
young
of
four
species
ranging
from
547
t
o
g
r
e
a
t
e
r
than
1667
ug/
g
(
Heath
e
t
a
l
.
1
9
7
2
)
.
Long­
term
d
i
e
t
a
r
y
dosages
caused
no
adverse
effect
a
t
3
u
g
/
g
w
i
t
h
m
a
l
l
a
r
d
s
a
n
d
1
3
ug/
gwith
pheasants
(
Heath
and
Spann,
1973).
However,
it
has
been
reported
(
Stoke
e
t
a
l
.
1978)
t
h
a
t
the
persistence
of
mirex
i
n
b
i
r
d
t
i
s
s
u
e
exceeds
t
h
a
t
of
a
l
l
organochlorine
compounds
tested
except
f
o
r
DDE.
Delayed
mortality
occurred
among
birds
subjected
to
doses
above
expected
environmental
concentration.

A
summary
examination
of
the
data
available
at
this
time
shows
a
mosaic
of
effects.
Crayfish
and
channel
catfish
survival
is
affected
by
mirex
in
the
water
or
by
ingestion
of
the
bait
particles.
Bioaccumulation
is
well
established
for
a
wide
variety
of
organisms
but
the
effect
of
this
bioaccumulation
on
the
aquatic
ecosystem
is
unknown.
There
is
evidence
that
mirex
is
very
persistent
in
bird
tissue.
Considering
the
extreme
toxicity
and
potential
for
bioaccumulation,
every
effort
should
be
made
to
keep
mirex
bait
particles
out
of
water
containing
aquatic
organisms
and
water
concentrations
should
not
exceed
0.001
ug/
L
mirex.
This
value
is
based
upon
an
application
factor
of
0.01
applied
to
the
lowest
levels
at
which
effects
on
crayfish
have
been
observed.

Data
upon
which
to
base
a
marine
criterion
involve
several
estuarine
and
marine
crustaceans.
A
concentration
of
0.1
ug/
L
technical
grade
mirex
in
flowing
seawater
was
lethal
to
juvenile
pink
shrimp,
Penaeus
durorarum,
in
a
3­
week
exposure
(
Lowe
et
al.

1971).
In
static
tests
with
larval
stages
(
megalopal)
of
the
mud
crab,
Rhithropanopeus
harrisii,
reduced
survival
was
observed
in
0.1
ug/
L
mirex
(
Bookhout
et
al.
1972).
In
three
of
four
28­
day
seasonal
f
low­
through
experiments,
Tagatz
et
al.
(
1975)
found
reduced
survival
of
__­
Callinectes
sapidus,
Penaeus
durorarum,
and
grass
shrimp,
­_­_­­_­­_
Palaemonetes
~
us&,
at
levels
of
0.12
ug/
L
in
summer,
0.06
ug/
L
in
fall
and
0.09
ug/
L
in
winter.

Since
two
reports,
Lowe
et
al.
(
1971)
and
Bookhout
et
al.
­

(
1972),
stated
that
effects
of
mirex
on
estuarine
and
marine
c
r
u
s
t
a
c
e
a
n
s
w
e
r
e
observed
o
n
l
y
a
f
t
e
r
c
o
n
s
i
d
e
r
a
b
l
e
t
i
m
e
had
e
l
a
p
s
e
d
,
it
seems
reasonable
t
h
a
t
l
e
n
g
t
h
of
exposure
is
an
important
consideration
 or
t
h
i
s
chemical.
This
may
not
be
t
h
e
case
i
n
f
r
e
s
h
water
s
i
n
c
e
t
h
e
c
r
a
y
f
i
s
h
were
a
f
f
e
c
t
e
d
w
i
t
h
i
n
4
8
hours.
Therefore,
a
3­
t
o
4­
week
exposure
might
be
considered
88acute81
and
by
applying
an
a
p
p
l
i
c
a
t
i
o
n
f
a
c
t
o
r
of
0.01
t
o
a
reasonable
average
of
toxic­
effect
l
e
v
e
l
s
as
summarized
above,
a
recommended
marine
c
r
i
t
e
r
i
o
n
of
0.001
ug/
L
r
e
s
u
l
t
s
.

(
QUALITY
CRITERIA
FOR
WATER,
JULY
1976)
PB­
263943
SEE
APPENDIX
C
FOR
METHODOLOGY
N?
QrnLENE
Aquatic
Life
The
available
data
for
naphthalene
indicate
that
acute
and
chronic
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
2,300
and
620
ug/
L,
respectively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.

The
available
data
for
naphthalene
indicate
that
acute
toxicity
to
saltwater
aquatic
1
ife
occurs
at
concentrations
as
low
as
2,350
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
naphthalene
to
sensitive
saltwater
aquatic
life.
I
Human
Health
Using
the
present
guidelines,
a
satisfactory
criterion
cannot
be
derived
at
this
time
because
of
insufficient
available
data
for
naphthalene.

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
NICKEL
Aquatic
Life
FOI
total
recoverable
nickel
the
criterion
(
in
ug/
L)
to
protect
freshwater
aquatic
life
as
derived
using
the
Guidelines
is
the
numerical
value
given
by
e(
0.76[
1n(
hardness)]+
l.
06)
as
a
24­
hour
average,
and
the
concentration
(
in
ug/
L)
should
not
exceed
the
numerical
value
given
by
.(
0.76[
ln
(
hardness)]+
4.02)

at
any
time.
For
example,
at
hardnesses
of
5
0
,
100,
and
200
mg/
L
as
CaC03
the
criteria
are
56,
96,
and
160
ug/
L,
respectively,
as
24­
hour
averages,
and
the
concentrations
should
not,
exceed
1,100,

1,800,
and
3,100
ug/
L,
respectively,
at
any
time.

For
total
recoverable
nickel
the
criterion
to
protect
saltwater
aquatic
life
as
derived
using
the
Guidelines
is
7.1
ug/
L
as
a
24­
hour
average,
and
the
concentration
should
not
exceed
140
ug/
L
at
any
time.
0
Human
Health
For
the
protection
of
human
health
from
the
toxic
properties
of
nickel
ingested
through
water
and
contaminated
aquatic
organisms,
the
ambient
water
criterion
is
determined
to
be
632
wl/
L.

For
the
protection
of
human
health
from
the
toxic
properties
of
nickel
ingested
through
contaminated
aquatic
organisms
alone,
the
ambient
water
criterion
is
determined
to
be
4.77
W/
L.

(
45
F.
R.
79318,
November
28,
1980)
0
SEE
APPENDIX
B
FOR
METHODOLOGY
0
CRITERION:
NITF?
ATES/
NITRITES
10
mg/
L
n
i
t
r
a
t
e
nitrogen
(
N)
f
o
r
domestic
water
supply
(
h
e
a
l
t
h
)
.

INTRODUCTION:

Two
gases
(
m
o
l
e
c
u
l
a
r
n
i
t
r
o
g
e
n
and
n
i
t
r
o
u
s
oxide)
and
f
i
v
e
forms
o
f
nongaseous,
combined
n
i
t
r
o
g
e
n
(
amino
and
amide
groups,

ammonium,
n
i
t
r
i
t
e
,
and
n
i
t
r
a
t
e
)
are
important
i
n
t
h
e
n
i
t
r
o
g
e
n
cycle.
The
amino
and
amide
groups
are
found
i
n
s
o
i
l
o
r
g
a
n
i
c
matter
and
as
c
o
n
s
t
i
t
u
e
n
t
s
of
p
l
a
n
t
and
animal
p
r
o
t
e
i
n
.
The
ammonium
ion
either
is
released
from
proteinaceous
organic
matter
and
urea,
o
r
is
s
y
n
t
h
e
s
i
z
e
d
i
n
i
n
d
u
s
t
r
i
a
l
p
r
o
c
e
s
s
e
s
i
n
v
o
l
v
i
n
g
atmospheric
n
i
t
r
o
g
e
n
f
i
x
a
t
i
o
n
.
The
n
i
t
r
i
t
e
i
o
n
i
s
formed
from
t
h
e
n
i
t
r
a
t
e
or
t
h
e
ammonium
ions
by
c
e
r
t
a
i
n
microorganisms
found
i
n
s
o
i
l
,
water,
sewage,
and
t
h
e
d
i
g
e
s
t
i
v
e
tract.
The
n
i
t
r
a
t
e
ion
is
formed
by
t
h
e
complete
o
x
i
d
a
t
i
o
n
of
ammonium
i
o
n
s
by
s
o
i
l
o
r
water
microorganisms;
n
i
t
r
i
t
e
is
an
intermediate
product
of
t
h
i
s
n
i
t
r
i
f
i
c
a
t
i
o
n
process.
I
n
oxygenated
n
a
t
u
r
a
l
water
systems
n
i
t
r
i
t
e
is
r
a
p
i
d
l
y
o
x
i
d
i
z
e
d
t
o
n
i
t
r
a
t
e
.
Growing
p
l
a
n
t
s
assimilate
n
i
t
r
a
t
e
o
r
ammonium
ions
and
convert
them
t
o
protein.

A
process
known
as
d
e
n
i
t
r
i
f
i
c
a
t
i
o
n
takes
p
l
a
c
e
when
n
i
t
r
a
t
e
­

containing
s
o
i
l
s
become
anaerobic
and
t
h
e
conversion
t
o
n
i
t
r
i
t
e
,

molecular
nitrogen,
or
n
i
t
r
o
u
s
oxide
occurs.
Ammonium
ions
may
a
l
s
o
be
produced
i
n
some
circumstances.
0
Among
t
h
e
major
p
o
i
n
t
s
o
u
r
c
e
s
of
n
i
t
r
o
g
e
n
e
n
t
r
y
i
n
t
o
water
bodies
are
municipal
and
i
n
d
u
s
t
r
i
a
l
wastewaters,
s
e
p
t
i
c
t
a
n
k
s
,

and
feed
l
o
t
discharges.
D
i
f
f
u
s
e
s
o
u
r
c
e
s
o
f
n
i
t
r
o
g
e
n
i
n
c
l
u
d
e
farm­
site
f
e
r
t
i
l
i
z
e
r
and
animal
wastes,
lawn
f
e
r
t
i
l
i
z
e
r
,
leachate
a
from
waste
disposal
in
dumps
or
sanitary
landfills,
atmospheric
fallout,
nitric
oxide
and
nitrite
discharges
from
automobile
exhausts
and
other
combustion
processes,
and
losses
from
natural
sources
such
as
mineralization
of
soil
organic
matter
(
NAS,

1972).
Water
reuse
systems
in
some
fish
hatcheries
employ
a
nitrification
process
for
ammonia
reduction;
this
may
result
in
exposure
of
the
hatchery
fish
to
elevated
levels
of
nitrite
(
Russo
et
al.
1974).

RATIONALE
:

In
quantities
normally
found
in
food
or
feed,
nitrates
become
toxic
only
under
conditions
in
which
they
are,
or
may
be,
reduced
to
nitrites.
Otherwise,
at
"
reasonable"
concentration
nitrates
are
rapidly
excreted
in
the
urine.
High
intake
of
nitrates
constitutes
a
hazard
primarily
to
warmblooded
animals
under
conditions
that
are
favorable
to
reduction
to
nitrite.
Under
certain
circumstances,
nitrate
can
be
reduced
to
nitrite
in
the
gastrointestinal
tract
which
then
reaches
the
bloodstream
and
reacts
directly
with
hemoglobin
to
produce
methemoglobin,

consequently
impairing
transport.
I
The
reaction
of
nitrite
with
hemoglobin
can
be
hazardous
in
infants
under
3
months
of
age.
Serious
and
occasionally
fatal
poisonings
in
infants
have
occurred
following
ingestion
of
untreated
well
waters
shown
to
contain
nitrate
at
concentrations
greater
than
10
mg/
L
nitrate
nitrogen
(
N)
(
NAS,
1974).
High
nitrate
concentrations
frequently
are
found
in
shallow
farm
and
rural
community
wells,
often
as
the
result
o
f
inadequate
protection
from
barnyard
drainage
or
from
septic
tanks
(
USPHS,
1961;
Stewart
et
al.
1967).
Increased
concentrations
of
nitrates
also
have
been
found
in
streams
from
farm
tile
drainage
in
areas
of
intense
fertilization
and
farm
crop
production
(
Harmeson
et
al.
1971).
Approximately
2,000
cases
of
infant
methemoglobinemia
have
been
reported
in
Europe
and
North
America
since
1945;
7
to
8
percent
of
the
affected
infants
died
(
Walton,
1951;

Sattelmacher,
1962).
Many
infants
have
drunk
water
in
which
the
nitrate
nitrogen
content
was
greater
than
10
mg/
L
without
developing
methemoglobinemia.
Many
pub1
ic
water
supplies
in
the
United
States
contain
levels
that
routinely
exceed
this
amount,

but
only
one
U.
S.
case
of
infant
methemoglobinemia
associated
with
a
public
water
supply
has
ever
been
reported
(
Virgil
et
al.

1965).
The
differences
in
susceptibility
to
methemoglobinemia
are
not
yet
understood
but
appear
to
be
related
to
a
combination
of
factors
including
nitrate
concentration,
enteric
bacteria,
and
the
lower
acidity
characteristic
of
the
digestive
systems
of
baby
mammals.
Methemoglobinemia
systems
and
other
toxic
effects
were
observed
when
high
nitrate
well
waters
containing
pathogenic
bacteria
were
fed
to
laboratory
mammals
(
Wolff
et
al.
1972).

Conventional
water
treatment
has
no
significant
effect
on
nitrate
removal
from
water
(
NAS,
1974).

Because
of
the
potential
risk
of
methemoglobinemia
to
bottle­

fed
infants,
and
in
view
of
the
absence
of
substantiated
physiological
effects
at
nitrate
concentrations
below
10
mg/
L
nitrate
nitrogen,
this
level
is
the
criterion
for
domestic
water
supplies.
Waters
with
nitrite
nitrogen
concentrations
over
1
0
~
mg/
L
should
not
be
used
for
infant
feeding.
Waters
with
a
significant
nitrite
concentration
usually
would
be
heavily
polluted
and
probably
bacteriologically
unacceptable.

Westin
(
1974)
determined
that
the
respective
96­
hOUr
and
7
­

day
LC50
values
for
chinook
salmon,
Oncorhynchus
tshawytscha,

were
1,310
and
1,080
mg/
L
nitrate
nitrogen
in
fresh
water
and
990
and
900
mg/
L
nitrate
nitrogen
in
15
o/
oo
saline
water.
For
fingerling
rainbow
trout,
Salmo
qairdneri,
the
respective
96­
hour
and
7­
day
LC50
values
were
1,360
and
1,060
mg/
L
nitrate
nitrogen
in
fresh
water,
and
1,050
and
900
mg/
L
nitrate
nitrogen
in
15
o/
oo
saline
water.
Trama
(
1954)
reported
that
the
96­
hOUr
LC50
for
bluegills,
Asomis
­­
macrochirus
­­­­­­
I
at
20
°
C
was
2
,
0
0
0
mg/
L
nitrate
nitrogen
(
sodium
nitrate)
and
420
mg/
L
nitrate
nitrogen
(
potassium
nitrate).
Knepp
and
Arkin
(
1973)
observed
that
largemouth
bass,
Micropterus
salmoides,
and
channel
catfish,

Ictalurus
punctatus,
could
be
maintained
at
concentrations
up
to
400
mg/
L
nitrate
(
90
mg/
L
nitrate
nitrogen)
without
significant
effect
upon
their
growth
and
feeding
activities.

The
96­
hour
and
7­
day
LC50
values
for
chinook
salmon,

­___
Qncorhynchus
tshawytscha,
were
found
to
be
0.9
and
0.7
mg/
L
nitrite
nitrogen
in
fresh
water
(
Westin,
1974).
Smith
and
Williams
(
1974)
tested
the
effects
of
nitrite
nitrogen
and
observed
that
yearling
rainbow
trout,
Salmo
gairdneri,
suffered
a
55
percent
mortality
after
24
hours
at
0.55
mg/
L;
fingerling
rainbow
trout
suffered
a
50
percent
mortality
after
24
hours
of
exposure
at
1.6
mg/
L;
and
chinook
salmon,
Oncorhynchus
tshawytscha,
suffered
a
40
percent
mortality
within
24
hours
at
0.5
mg/
L.
There
w
e
r
e
no
m
o
r
t
a
l
i
t
i
e
s
among
rainbow
t
r
o
u
t
exposed
t
o
0.15
mg/
L
n
i
t
r
i
t
e
nitrogen
f
o
r
48
hours.
These
data
indicate
t
h
a
t
salmonids
are
more
s
e
n
s
i
t
i
v
e
t
o
n
i
t
r
i
t
e
t
o
x
i
c
i
t
y
than
a
r
e
other
f
i
s
h
s
p
e
c
i
e
s
,
e.
g.,
minnows,
­_­­
Phoxinus
l
a
e
v
i
s
t
h
a
t
suffered
a
50
percent
mortality
within
1.5
hours
of
exposure
t
o
2,030
mg/
L
n
i
t
r
i
t
e
n
i
t
r
o
g
e
n
,
b
u
t
required
1
4
days
of
exposure
 or
m
o
r
t
a
l
i
t
y
t
o
occur
a
t
1
0
mg/
L
(
K
l
i
n
g
l
e
r
,
1
9
5
7
)
,
and
carp,

Cyprinus
­­
c
a
r
e
,
when
r
a
i
s
e
d
i
n
a
water
reuse
system,
t
o
l
e
r
a
t
e
d
up
t
o
1.8
mg/
L
n
i
t
r
i
t
e
nitrogen
(
Saeki,
1965).

G
i
l
b
e
t
t
e
,
e
t
a
l
.
(
1952)
observed
t
h
a
t
t
h
e
critical
range
f
o
r
creek
chub,
Semotilus
­
atromaculatus,
was
80
t
o
400
mg/
L
n
i
t
r
i
t
e
nitrogen.
Wallen
e
t
a
l
.
(
1957)
reported
a
24­
hour
LC50
of
1.6
mg/
L
n
i
t
r
i
t
e
n
i
t
r
o
g
e
n
,
and
48­
and
96­
hour
LC50
v
a
l
u
e
s
of
1.5
mg/
L
n
i
t
r
i
t
e
nitrogen
f
o
r
mosquitofish,
Gambusia
a
f
f
i
n
i
s
.
McCoy
(
1972)
tested
t
h
e
n
i
t
r
i
t
e
s
u
s
c
e
p
t
i
b
i
l
i
t
y
of
13
f
i
s
h
species
and
found
t
h
a
t
logperch,
Percina
­­
caprodes
­
_
I
were
t
h
e
most
s
e
n
s
i
t
i
v
e
s
p
e
c
i
e
s
tested
(
m
o
r
t
a
l
i
t
y
a
t
5
mg/
L
n
i
t
r
i
t
e
n
i
t
r
o
g
e
n
i
n
less
than
3
hours
of
exposure)
whereas
carp,
9
p
r
i
n
u
s
c
a
r
p
i
o
,
and
b
l
a
c
k
bullheads,
­­­­
I
c
t
a
l
u
r
u
s
­­­­
melas
8
s
u
r
v
i
v
e
d
40
mg/
L
n
i
t
r
i
t
e
nitrogen
 or
a
48­
hour
exposure
period;
t
h
e
common
white
sucker,

Catostomus
­­..­­­­­­­
I
commersoni
and
t
h
e
q
u
i
l
lback,
Carpiodes
cyprinus,

survived
100
mg/
L
f
o
r
48
and
36
hours,
respectively.
e
Russo
e
t
a
l
.
(
1974)
performed
f
low­
through
n
i
t
r
i
t
e
bioassays
i
n
hard
water
(
hardness
=
1
9
9
mg/
L
CaC03;
a
l
k
a
l
i
n
i
t
y
=
176
mg/
L
CaCQ3;
pH
=
7.9)
on
rainbow
t
r
o
u
t
,
Salmo
g
a
i
r
d
n
e
r
i
,
of
f
o
u
r
d
i
f
f
e
r
e
n
t
s
i
z
e
s
,
and
obtained
9
6
­
h
O
U
r
LC50
values
ranging
from
0.19
t
o
0.39
mg/
L
n
i
t
r
i
t
e
nitrogen.
Duplicate
bioassays
on
1
2
­

gram
rainbow
t
r
o
u
t
w
e
r
e
continued
long
enough
f
o
r
t
h
e
i
r
t
o
x
i
c
i
t
y
c
u
r
v
e
s
t
o
l
e
v
e
l
o
f
f
,
and
asymptotic
LC50
concentrations
of
0.14
and
0.15
mg/
L
were
reached
i
n
8
days;
on
day
1
9
,
a
d
d
i
t
i
o
n
a
l
m
o
r
t
a
l
i
t
i
e
s
occurred.
For
2­
gram
rainbow
t
r
o
u
t
,
t
h
e
minimum
tested
l
e
v
e
l
of
n
i
t
r
i
t
e
n
i
t
r
o
g
e
n
a
t
which
no
m
o
r
t
a
l
i
t
i
e
s
w
e
r
e
observed
after
10
days
was
0.14
mg/
L;
f
o
r
the
235­
gram
t
r
o
u
t
,
t
h
e
minimum
l
e
v
e
l
w
i
t
h
no
mortality
a
f
t
e
r
1
0
days
w
a
s
0.06
mg/
L.

It
is
concluded
t
h
a
t
(
1)
l
e
v
e
l
s
of
n
i
t
r
a
t
e
nitrogen
a
t
o
r
below
90
mg/
L
would
have
no
adverse
effects
on
warmwater
f
i
s
h
(
Knepp
and
A
r
k
i
n
,
1
9
7
3
)
;
(
2
)
n
i
t
r
i
t
e
n
i
t
r
o
g
e
n
a
t
o
r
below
5
mg/
L
should
be
p
r
o
t
e
c
t
i
v
e
of
most
warmwater
f
i
s
h
(
McCoy,
1
9
7
2
)
;
and
(
3)
n
i
t
r
i
t
e
nitrogen
a
t
o
r
below
0.06
mg/
L
should
be
protective
of
salmonid
f
i
s
h
e
s
(
Russo
e
t
a
l
.
1
9
7
4
;
Russo
and
Thurston,

1975).
These
l
e
v
e
l
s
e
i
t
h
e
r
are
n
o
t
known
t
o
occur
o
r
would
be
unlikely
t
o
occur
i
n
natural
surface
waters.
I
Recognizing
t
h
a
t
c
o
n
c
e
n
t
r
a
t
i
o
n
s
of
n
i
t
r
a
t
e
o
r
n
i
t
r
i
t
e
t
h
a
t
would
e
x
h
i
b
i
t
t
o
x
i
c
effects
on
warm­
o
r
coldwater
f
i
s
h
could
r
a
r
e
l
y
occur
i
n
nature,
r
e
s
t
r
i
c
t
i
v
e
c
r
i
t
e
r
i
a
are
not
recommended.

(
QUALITY
CRITERIA
FOR
WATER,
JULY
1976)
PB­
263943
SEE
APPENDIX
C
FOR
METHODOLOGY
NITROBENZENE
CRITERIA:
Aquatic
Life
The
available
data
for
nitrobenzene
indicate
that
acute
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
27,000
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
definitive
data
are
available
concerning
the
chronic
toxicity
of
nitrobenzene
to
sensitive
freshwater
aquatic
life.

The
available
data
for
nitrobenzene
indicate
that
acute
toxicity
to
saltwater
aquatic
life
occurs
at
concentrations
as
low
as
6,680
ug/
L
and
would
occur
at
lower
concentrations
among
species
thaf;
are
more
sensitive
than
those
tested.
No
definitive
data
are
available
concerning
the
chronic
toxicity
of
nitrobenzene
to
sensitive
saltwater
aquatic
life.

Human
Health
For
comparison
purposes,
two
approaches
were
used
to
derive
criterion
levels
for
nitrobenzene.
Based
on
available
toxicity
data,
to
protect
public
health
the
derived
level
is
19.8
mg/
L.

using
available
organoleptic
data,
to
control
undesirable
taste
and
odor
qualities
of
ambient
water
the
estimated
level
is
30
ug/
L.
It
should
be
recognized
that
organoleptic
data
have
limitations
as
a
basis
for
establishing
a
water
quality
criterion,
and
have
no
demonstrated
relationship
to
potential
adverse
human
health
effects.

(
4
5
F.
R.
79318,
November
2
8
,
1980)

NOTE:
The
U.
S.
EPA
is
currently
developing
Acceptable
Daily
Intake
(
ADI)
or
Verified
Reference
Dose
(
RID)
values
for
Agency­
wide
use
for
this
chemical.
The
new
value
should
SEE
APPENDIX
B
FOR
METHODOLOGY
be
substituted
when
it
becomes
available.
The
January,
1986,
draft
Verified
Reference
Dose
document
cites
an
RfD
of
.0005
mg/
kg/
day
for
nitrobenzene.
NITROPHENOLS
CRITERIA:

Aquatic
Life
The
available
data
for
nitrophenols
indicate
that
acute
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
230
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
nitrophenols
to
sensitive
freshwater
aquatic
life
but
toxicity
to
one
species
of
algae
occurs
at
concentrations
as
low
as
150
ug/
L.

The
available
data
for
nitrophenols
indicate
that
acute
toxicity
to
saltwater
aquatic
life
occurs
at
concentrations
as
low
as
4,850
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
nitrophenols
to
sensitive
saltwater
aquatic
life.
1
Human
Health
Because
of
insufficient
available
data
for
mono­
and
trinitrophenols,
satisfactory
criteria
cannot
be
derived
at
this
time,
using
the
present
guidelines.

For
the
protection
of
human
health
from
the
toxic
properties
of
dinitrophenols
and
2,4­
dinitro­
o­
cresol
ingested
through
water
and
contaminated
aquatic
organisms,
the
ambient
water
criteria
are
determined
to
be
70
ug/
L
and
13.4
ug/
L,
respectively.

For
the
protection
of
human
health
from
the
toxic
properties
of
dinitrophenols
and
2,4­
dinitro­
o­
cresol
ingested
through
contaminated
aquatic
organisms
alone,
the
ambient
water
criteria
­
are
determined
t
o
be
14.3
mg/
L
and
765
ug/
L,
r
e
s
p
e
c
t
i
v
e
l
y
.

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
NITROSAMINES
Aquatic
Life
The
available
data
for
nitrosamines
indicate
that
acute
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
5,850
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
nitrosamines
to
sensitive
freshwater
aquatic
life.

The
available
data
for
nitrosamines
indicate
that
acute
toxicity
to
saltwater
aquatic
life
occurs
at
concentrations
as
low
as
3,300,000
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
nitrosamines
to
sensitive
saltwater
aquatic
life.

Human
Health
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
N­
nitrosodiethylamine
and
all
other
nitrosamines
except
those
listed
below,
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentrations
should
be
zero,
based
on
the
non
threshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
lo"*,
and
1OpcPs
The
corresponding
recommended
criteria
are
8.0
ng/
L,
0.8
ng/
L,
and
0
.
0
8
ng/
L,
respectively.
If
these
estimates
are
made
for
L
0
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
12,400
ng/
L,
1,240
ng/
L,
and
124
ng/
L,

respectively.

For
the
maximum
protection
of
human
health
from
the
potential
carcinoqenic
effects
of
exposure
to
N­
nitrosodimethylamine
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentrations
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.

However,
zero
level
may
not
be
attainable
at
the
present
time.

Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
10­

6,
and
The
corresponding
recommended
criteria
are
14
ng/
L,
1.4
ng/
L,
and
0.14
ng/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
160,000
ng/
L,
16,000
ng/
L,

and
1,600
ng/
L,
respectively.

For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
N­
nitrosodibutylamine
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentrations
should
be
zero,

based
on
the
non
threshold
assumption
for
this
chemical.

However,
zero
level
may
not
be
attainable
at
the
present
time.

Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
10­

6,
and
The
corresponding
recommended
criteria
are
64
ng/
L,
6.4
ng/
L,
and
0.64
ng/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,

excluding
consumption
of
water,
the
levels
are
5,868
ng/
L,
587
ng/
L,
and
58.7
ng/
L,
respectively.

0
For
the
maximum
protection
o'f
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
N­
nitrosopyrrolidine
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentrations
should
be
zero
based
on
the
nonthreshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
.
over
the
lifetime
are
estimated
at
and
The
corresponding
recommended
criteria
are
160
ng/
L,
16
ng/
L,
and
1.6
ng/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
919,000
ng/
L,
91,900
ng/
L,
and
9,190
ng/
L,

respectively.

For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
N­
nitrosodiphenylamine
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentrations
should
be
zero,

based
on
the
non
threshold
assumption
f
o
r
this
chemical.

However,
zero
level
may
not
be
attainable
at
the
present
time.

Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
10­
5,

The
corresponding
recommended
criteria
are
49,000
ng/
L,
4,900
ng/
L,
and
490
ngfL,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,

excluding
consumption
of
water,
the
levels
are
161,000
ng/
L,
16,100
ng/
L,
and
1,610
ng/
L,
respectively.
and
L,
(
4
5
F.
R.
79318,
November
2
8
,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
OIL
AND
GREASE
­­

For
domestic
water
supply:
Virtually
free
from
oil
and
grease,
particularly
from
the
tastes
and
odors
that
emanate
from
petroleum
products.

For
aquatic
life:

(
1)
0.01
of
the
lowest
continuous
flow
96­
hour
LC50
to
several
important
freshwater
and
marine
species,
each
having
a
demonstrated
high
susceptibility
to
oils
and
petrochemicals.

(
2)
Levels
of
oils
or
petrochemicals
in
the
sediment
which
cause
deleterious
effects
to
the
biota
should
not
be
allowed.

(
3
)
Surface
waters
shall
be
virtually
free
from
floating
nonpetroleum
oils
of
vegetable
or
animal
origin,
as
well
as
petroleum­
derived
oils.
1
INTRODUCTION:

It
has
been
estimated
that
between
5
and
10
million
metric
tons
of
oil
enter
the
marine
environment
annually
(
Blumer,
1970).

A
major
difficulty
encountered
in
the
setting
of
criteria
for
oil
and
grease
is
that
these
are
not
definitive
chemical
categories,
but
include
thousands
of
organic
compounds
with
varying
physical,
chemical,
and
toxicological
properties.
They
may
be
volatile
or
nonvolatile,
soluble
or
insoluble,
persistent
or
easily
degraded.

RATIONALE
:

Field
and
laboratory
evidence
have
demonstrated
both
acute
lethal
toxicity
and
long­
term
sublethal
toxicity
of
o
i
l
s
to
aquatic
organisms.
Events
such
as
the
Tampico
Maru
wreck
of
1957
in
Baja,
California,
(
Diaz­
Piferrer,
1962),
and
the
No.
2
fuel
oil
spill
in
West
Falmouth,
Massachusetts,
in
1969
..
I
(
Hampson
and
Sanders,
1
9
6
9
)
,
both
of
which
caused
immediate
death
t
o
a
wide
v
a
r
i
e
t
y
o
f
organisms,
a
r
e
i
l
l
u
s
t
r
a
t
i
v
e
o
f
t
h
e
l
e
t
h
a
l
t
o
x
i
c
i
t
y
t
h
a
t
may
be
a
t
t
r
i
b
u
t
e
d
t
o
o
i
l
p
o
l
l
u
t
i
o
n
.
S
i
m
i
l
a
r
l
y
,
a
g
a
s
o
l
i
n
e
s
p
i
l
l
i
n
South
Dakota
i
n
November
1969
(
Bugbee
and
Walter,
1973)
was
reported
t
o
have
caused
immediate
death
t
o
t
h
e
majority
of
freshwater
i
n
v
e
r
t
e
b
r
a
t
e
s
and
2,500
f
i
s
h
,
30
percent
of
which
were
n
a
t
i
v
e
s
p
e
c
i
e
s
of
t
r
o
u
t
.
Because
o
f
t
h
e
wide
range
o
f
compounds
i
n
c
l
u
d
e
d
i
n
t
h
e
c
a
t
e
g
o
r
y
of
o
i
l
,
it
is
impossible
t
o
establish
meaningful
96­
hour
LC50
v
a
l
u
e
s
f
o
r
o
i
l
and
g
r
e
a
s
e
w
i
t
h
o
u
t
s
p
e
c
i
f
y
i
n
g
t
h
e
p
r
o
d
u
c
t
i
n
v
o
l
v
e
d
.

However,
a
s
t
h
e
d
a
t
a
i
n
T
a
b
l
e
6
show,
t
h
e
most
s
u
s
c
e
p
t
i
b
l
e
category
of
organisms,
t
h
e
marine
l
a
r
v
a
e
,
appear
t
o
be
i
n
t
o
l
e
r
a
n
t
o
f
p
e
t
r
o
l
e
u
m
p
o
l
l
u
t
a
n
t
s
,
p
a
r
t
i
c
u
l
a
r
l
y
t
h
e
w
a
t
e
r
s
o
l
u
b
l
e
compounds,
a
t
concentrations
as
low
as
0.1
mg/
L.

T
h
e
long­
term
s
u
b
l
e
t
h
a
l
e
f
f
e
c
t
s
of
o
i
l
p
o
l
l
u
t
i
o
n
refer
t
o
i
n
t
e
r
f
e
r
e
n
c
e
s
with
c
e
l
l
u
l
a
r
and
physiological
processes
such
as
f
e
e
d
i
n
g
and
r
e
p
r
o
d
u
c
t
i
o
n
and
do
n
o
t
lead
t
o
i
m
m
e
d
i
a
t
e
death
of
t
h
e
organism.
Disruption
of
such
behavior
apparently
can
r
e
s
u
l
t
from
petroleum
p
r
o
d
u
c
t
c
o
n
c
e
n
t
r
a
t
i
o
n
s
a
s
low
a
s
1
0
t
o
100
ug/
L
(
see
T
a
b
l
e
7).

T
a
b
l
e
7
summarizes
some
of
t
h
e
s
u
b
l
e
t
h
a
l
t
o
x
i
c
i
t
i
e
s
f
o
r
various
petroleum
p
o
l
l
u
t
a
n
t
s
and
aquatic
species.
I
n
a
d
d
i
t
i
o
n
t
o
s
u
b
l
e
t
h
a
l
effects
reported
a
t
t
h
e
10
t
o
100
ug/
L
l
e
v
e
l
,
it
has
been
shown
t
h
a
t
petroleum
p
r
o
d
u
c
t
s
can
harm
a
q
u
a
t
i
c
l
i
f
e
a
t
concentrations
as
low
a
s
1
ug/
L
(
Jacobson
and
Boylan,
1973).

Bioaccumulation
of
petroleum
products
presents
two
e
s
p
e
c
i
a
l
l
y
important
p
u
b
l
i
c
h
e
a
l
t
h
problems:
(
1)
t
h
e
t
a
i
n
t
i
n
g
o
f
edible,
aquatic
species,
and
(
2
)
the
possibility
of
edible
marine
organisms
incorporating
the
high
boiling,
carcinogenic
polycyclic
aromatics
in
their
tissues.
Nelson­
Smith
(
1971)
reported
that
0.01
mg/
L
of
crude
oil
caused
tainting
in
oysters.
Moore
et
al.

(
1973)
reported
that
concentrations
as
low
as
1
to
10
ug/
L
could
lead
to
tainting
within
very
short
periods
of
time.
It
has
been
shown
that
chemicals
responsible
for
cancer
in
animals
and
man
(
such
as
3,4­
benzopyrene)
occur
in
crude
oil
(
Blumer,
1970).
It
also
has
been
shown
that
marine
organisms
are
capable
of
incorporating
potentially
carcinogenic
compounds
into
their
body
fat
where
the
compounds
remain
unchanged
(
Blumer,
1970).

Oil
pollutants
may
also
be
incorporated
into
sediments.

There
is
evidence
that
once
this
occurs
in
the
sediments
below
the
aerobic
surface
layer,
petroleum
oil
can
remain
unchanged
and
toxic
for
long
periods,
since
its
rate
of
bacterial
degradation
is
slow.
For
example,
Blumer
(
1970)
reported
that
No.
2
fuel
oil
incorporated
into
the
sediments
after
the
West
Falmouth
spill
persisted
for
over
a
year,
and
even
began
spreading
in
the
form
of
oil­
laden
sediments
to
more
distant
areas
that
had
remained
unpolluted
immediately
after
the
spill.
The
persistence
of
unweathered
oil
within
the
sediment
could
have
a
long­
term
effect
on
the
structure
of
the
benthic
community
or
cause
the
demise
of
specific
sensitive
important
species.
Moore
et
al.
(
1973)

reported
concentrations
of
5
mg/
L
for
the
carcinogen
3,
4­

benzopyrene
in
marine
sediments.

Mironov
(
1967)
reported
that
0­
01
mg/
L
oil
produced
deformed
and
inactive
flatfish
larvae.
Mironov
(
1970)
also
reported
inhibition
or
delay
of
cellular
division
in
algae
by
oil
­,
concentrations
of
to
10­
1
mg/
L.
Jacobson
and
Boylan
(
1973)

reported
a
reduction
in
the
chemotactic
perception
of
food
by
the
snail,
Nassarius
­­
obsoletus
at
kerosene
concentrations
of
0.001
to
0.004
mg/
L.
Bellen
et
al.
(
1972)
reported
decreased
survival
and
fecundity
in
worms
at
concentrations
of
0.01
to
10
mg/
L
of
detergent.

Because
of
the
great
variability
in
the
toxic
properties
of
oil,
it
is
difficult
to
establish
a
numerical
criterion
which
would
be
applicable
to
all
t­
jpes
of
oil.
Thus,
an
application
factor
of
0.01
of
the
96­
hour
LC50
as
determined
by
using
continuous
flow
with
a
sensitive
resident
species
should
be
employed
for
individual
petrochemical
components.

There
is
a
paucity
of
toxicological
data
on
the
ingestion
of
the
components
of
refinery
wastewaters
by
humans
or
by
test
animals.
It
is
apparent
that
any
tolerable
health
concentrations
for
petroleum­
derived
substances
far
exceed
the
limits
of
taste
and
odor.
Since
petroleum
derivatives
become
organoleptically
objectionable
at
concentrations
far
below
the
human
chronic
toxicity,
it
appears
that
hazards
to
humans
will
not
arise
from
drinking
oil­
polluted
waters
(
Johns
Hopkins
Univ.,
1956;
Mckee
and
Wolf,
1963).
Oils
of
animal
or
vegetable
origin
generally
are
nontoxic
to
humans
and
aquatic
life.

In
view
of
the
problem
of
petroleum
oil
incorporation
in
sediments,
its
persistence
and
chronic
toxic
potential,
and
the
present
lack
of
sufficient
toxicity
data
to
support
specific
criteria,
concentrations
of
oils
in
sediments
should
not
approach
levels
that
cause
deleterious
effects
to
important
species
or
the
bottom
community
as
a
whole.

Petroleum
and
nonpetroleum
o
i
l
s
share
some
similar
physical
and
chemical
p
r
o
p
e
r
t
i
e
s
.
Because
they
share
common
p
r
o
p
e
r
t
i
e
s
,

t
h
e
y
may
c
a
u
s
e
s
i
m
i
l
a
r
h
a
r
m
f
u
l
e
f
f
e
c
t
s
i
n
t
h
e
a
q
u
a
t
i
c
environment
by
forming
a
sheen,
f
i
l
m
,
o
r
d
i
s
c
o
l
o
r
a
t
i
o
n
on
t
h
e
s
u
r
f
a
c
e
of
t
h
e
water.
L
i
k
e
petroleum
o
i
l
s
,
nonpetroleum
o
i
l
s
may
occur
a
t
four
levels
of
t
h
e
aquatic
environment:
(
a)
f
l
o
a
t
i
n
g
on
t
h
e
s
u
r
f
a
c
e
,
(
b)
e
m
u
l
s
i
f
i
e
d
i
n
t
h
e
water
column,
(
c)

s
o
l
u
b
i
l
i
z
e
d
,
and
(
d)
s
e
t
t
l
e
d
on
t
h
e
bottom
a
s
a
sludge.
Analogous
t
o
t
h
e
g
r
e
a
s
e
b
a
l
l
s
from
v
e
g
e
t
a
b
l
e
o
i
l
and
animal
f
a
t
s
are
t
h
e
t
a
r
b
a
l
l
s
of
petroleum
o
r
i
g
i
n
which
have
been
found
i
n
the
marine
environment
o
r
washed
ashore
on
beaches.

O
i
l
s
of
any
kind
can
cause
(
a)
drowning
of
waterfowl
because
of
loss
of
buoyancy,
exposure
because
of
loss
of
i
n
s
u
l
a
t
i
n
g
c
a
p
a
c
i
t
y
of
f
e
a
t
h
e
r
s
,
and
s
t
a
r
v
a
t
i
o
n
and
v
u
l
n
e
r
a
b
i
l
i
t
y
t
o
0
predators
because
of
lack
of
mobility;
(
b)
l
e
t
h
a
l
effects
on
f
i
s
h
by
c
o
a
t
i
n
g
e
p
i
t
h
e
l
i
a
l
s
u
r
f
a
c
e
s
of
g
i
l
l
s
,
t
h
u
s
p
r
e
v
e
n
t
i
n
g
r
e
s
p
i
r
a
t
i
o
n
;
(
c)
p
o
t
e
n
t
i
a
l
f
i
s
h
k
i
l
l
s
r
e
s
u
l
t
i
n
g
from
biochemical
oxygen
demand:
(
d)
a
s
p
h
y
x
i
a
t
i
o
n
of
b
e
n
t
h
i
c
l
i
f
e
forms
when
f
l
o
a
t
i
n
g
masses
become
engaged
w
i
t
h
surface
debris
and
s
e
t
t
l
e
on
t
h
e
bottom:
a
n
d
(
e
)
a
d
v
e
r
s
e
a
e
s
t
h
e
t
i
c
e
f
f
e
c
t
s
of
f
o
u
l
e
d
s
h
o
r
e
l
i
n
e
s
and
beaches.
These
and
o
t
h
e
r
e
f
f
e
c
t
s
have
been
documented
i
n
t
h
e
U.
S.
Department
of
H
e
a
l
t
h
,
Education
and
Welfare
r
e
p
o
r
t
on
Oil
S
p
i
l
l
s
A
f
f
e
c
t
i
n
g
t
h
e
Minnesota
and
M
i
s
s
i
s
s
i
p
p
i
R
i
v
e
r
s
and
t
h
e
1975
P
r
o
c
e
e
d
i
n
g
s
o
f
t
h
e
J
o
i
n
t
Conference
on
Prevention
and
Control
of
Oil
S
p
i
l
l
s
.

e
O
i
l
s
of
animal
o
r
vegetable
o
r
i
g
i
n
g
e
n
e
r
a
l
l
y
are
chemically
_
,

nontoxic
t
o
humans
o
r
aquatic
l
i
f
e
;
however,
f
l
o
a
t
i
n
g
sheens
of
such
oils
result
in
deleterious
environmental
effects
described
in
this
criterion.
Thus,
it
is
recommended
that
surface
waters
shall
be
virtually
free
from
floating
nonpetroleum
oils
of
vegetable
or
animal
origin.
This
same
recommendation
applies
to
floating
oils
of
petroleum
origin
since
they
too
may
produce
similar
effects.

(
QUALITY
CRITERIA
FOR
WATER,
JULY
1976)
PB­
263943
SEE
APPENDIX
C
FOR
METHODOLOGY
DISSOLVED
OXYGEN
0
NATIONAL
CRITERIA:

The
national
criteria
for
ambient
dissolved
oxygen
concentra­

tions
for
the
protection
of
freshwater
aquatic
life
are
presented
in
Table
1.
The
criteria
are
derived
from
the
production
impair­

ment
estimates
which
are
based
primarily
upon
growth
data
and
information
on
temperature,
disease,
and
pollutant
stresses.
The
average
dissolved
oxygen
concentrations
selected
are
values
0.5
mg/
L
above
the
slight
production
impairment
values
and
repre­

sent
values
between
no
production
impairment
and
slight
production
impairment.
Each
criterion
may
thus
be
viewed
as
an
estimate
of
the
threshold
concentration
below
which
detrimental
effects
are
expected.

Criteria
for
coldwater
fish
are
intended
to
apply
to
waters
containing
a
population
of
one
or
more
species
in
the
family
Salmonidae
{
Bailey
et
al.,
1970)
or
to
waters
containing
other
coldwater
or
coolwater
fish
deemed
by
the
user
to
be
closer
to
salmonids
in
sensitivity
than
to
most
warmwater
species.

Although
the
acute
lethal
limit
 or
salmonids
is
at
or
below
3
mg/
L,
the
coldwater
minimum
has
been
established
at
4
mg/
L
because
a
significant
proportion
of
the
insect
species
common
to
salmonid
habitats
are
less
tolerant
of
acute
exposures
to
low
dissolved
oxygen
than
are
salmonids.
Some
coolwater
species
may
require
more
protection
than
that
afforded
by
the
other
life
stage
criteria
for
warmwater
fish
and
it
may
be
desirable
to
protect
sensitive
coolwater
species
with
the
coldwater
criteria.
Many
states
have
more
stringent
dissolved
oxygen
standards
for
cooler
waters,
waters
that
contain
either
._
a
salmonids,
nonsalmonid
coolwater
fish,
or
the
sensitive
centra­

chid,
the
smallmouth
bas5
The
warmwater
criteria
are
necessary
to
protect
early
life
stages
of
warmwater
fish
as
sensitive
as
as
channel
catfish
and
to
protect
other
life
stages
of
fish
as
sensitive
as
largemouth
bass.
Criteria
for
early
life
stages
are
intended
to
apply
only
where
and
when
these
stages
occur.
These
criteria
represent
dissolved
oxygen
concentrations
which
EPA
believes
provide
a
reasonable
and
adequate
degree
of
protection
for
freshwater
aquatic
life.

The
criteria
do
not
represent
assured
no­
effect
levels.

However,
because
the
criteria
represent
worst
case
conditions
(
i.
e.
for
wasteload
allocation
and
waste
treatment
plant
design)
~

conditions
will
be
better
than
the
criteria
nearly
all
of
the
time
at
most
sites.
In
situations
where
criteria
conditions
are
just
maintained
for
considerable
periods
the
proposed
criteria
represent
some
risk
of
production
impairment.
This
impairment
would
depend
on
innumerable
other
factors.
If
slight
production
impairment
or
a
small
but
undefinable
risk
of
moderate
impairment
is
unacceptable,
than
one
should
use
the
"
no
production
impair­

ment"
values
given
in
the
document
as
means
and
the
"
slight
production
impairment')
values
as
minima.
The
table
which
pre­

sents
these
concentrations
is
reproduced
here
as
table
2
.

The
criteria
do
represent
dissolved
oxygen
concentrations
believed
to
protect
the
more
sensitive
populations
of
organisms
against
potentially
damaging
production
impairment.
The
dissolved
oxygen
concentrations
in
the
criteria
are
intended
to
be
protective
at
typically
high
seasonal
environmental
tempera­

tures
for
the
appropriate
taxonomic
and
life
stage
classi­
Table
1.
Water
quality
criteria
for
ambient
dissolved
oxygen
0
concentration.

Coldwater
Criteria
Warmwater
Criteria
Stages
Ue
Stages
Stages
Stages
Other
Life
Early
kife
Other
Life
30
Day
Mean
NA3
6.5
NA
5.5
7
Day
Mean
9.5
(
6.5)
NA
6.0
NA
7
Day
Mean
NA
Minimum
5.0
NA
4.0
1
Day
8.0
(
5.0)
4.0
5.0
3.0
Minimum
415
These
are
water
column
concentrations
recommended
to
achieve
the
required
intergravel
dissolved
oxygen
concentrations
shown
in
parentheses.
The
3
mg/
L
differential
is
discussed
in
the
criteria
document.
For
species
that
have
early
life
stages
exposed
directly
to
the
water
column,
the
figures
in
parentheses
apply.

Includes
all
embryonic
and
larval
stages
and
all
juvenile
forms
to
30­
days
following
hatching.
0
3
NA
(
not
applicable).

For
highly
manipulatable
discharges,
further
restrictions
apply
(
see
page
37)

A
l
l
minima
should
be
considered
as
instantaneous
concentrations
to
be
achieved
at
all
times.

fications,
temperatures
which
are
often
higher
than
those
used
in
the
research
from
which
the
criteria
were
generated,
especially
for
other
than
early
life
stages.

Where
natural
conditions
a
l
o
n
e
c
r
e
a
t
e
d
i
s
s
o
l
v
e
d
oxygen
concentrations
less
than
110
percent
of
the
applicable
criteria
means
or
minima
or
both,
the
minimum
acceptable
concentration
is
90
percent
of
the
natural
concentration.
These
values
are
similar
to
those
presented
graphically
by
Doudoroff
and
Shumway
(
1970)
and
those
calculated
from
Water
Quality
Criteria
1972
(
NAS/
NAE,
1973).
Absolutely
no
anthropogenic
dissolved
oxygen
depression
in
the
potentially
lethal
area
below
the
1­
day
minima
should
be
allowed
unless
special
care
is
taken
to
ascertain
the
tolerance
of
resident
species
to
low
dissolved
oxygen.

.
If
daily
cycles
of
dissolved
oxygen
are
essentially
sinusoidal,
a
reasonable
daily
average
is
calculated
from
the
day's
high
and
low
dissolved
oxygen
values.
A
time­
weighted
average
may
be
required
if
the
dissolved
oxygen
cycles
are
decidedly
non­
sinusoidal.
Determining
the
magnitude
of
daily
dissolved
oxygen
cycles
requires
at
least
two
appropriate­

ly
timed
measurements
daily,
and
characterizing
the
shape
of
the
cycle
requires
several
more
appropriately
spaced
measurements.

Once
a
series
of
daily
mean
dissolved
oxygen
concentrations
are
calculated,
an
average
of
these
daily
means
can
be
calcu­

lated
(
Table
3).
For
embryonic,
larval,
and
early
life
stages,

the
averaging
period
should
not
exceed
7
days.
This
short
time
is
needed
to
adequately
protect
these
often
short
duration,
most
sensitive
life
stages.
Other
life
stages
can
probably
be
adequately
protected
by
30­
day
averages.
Regardless
of
the
averaging
period,
the
average
should
be
considered
a
moving
average
rather
than
a
calendar­
week
or
calendar­
month
average.
Table
2.
Dissolved
Oxygen
Concentrations
(
mg/
L)
Versus
0
Quantitative
Level
of
Effect.

1.
Salmonid
Waters
a.
Embryo
and
Larval
Stages
No
Production
Impairment
=
11*
(
8
)
Slight
Production
Impairment
=
9*
(
6)
Moderate
Production
Impairment
=
8*
(
5)
Severe
Production
Impairment
=
7*
(
4
)
Limit
to
Avoid
Acute
Mortality
=
6*
(
3)

(*
Note:
These
are
water
column
concentrations
recommended
to
achieve
the
required
intergravel
dissolved
oxygen
concentrations
shown
in
parentheses.
The
3
mg/
L
difference
is
discussed
in
the
criteria
document.)

b.
Other
Life
Stages
No
Production
Impairment
=
8
light
Production
Impairment
=
6
Moderate
Production
Impairment
=
5
Severe
Production
Impairment
=
4
Limit
to
Avoid
Acute
Mortality
=
3
2.
Nonsalmonid
Waters
a.
Early
Life
Stages
No
Production
Impairment
=
6
.
5
Slight
Production
Impairment
=
5.5
Moderate
Production
Impairment
=
5
Severe
Production
Impairment
=
4
,
5
Limit
to
Avoid
Acute
Mortality
=
4
b.
Other
Life
Stages
No
Production
Impairment
=
6
Slight
Production
Impairment
=
5
Moderate
Production
Impairment
=
4
Severe
Production
Impairment
=
3.5
Limit
to
Avoid
Acute
Mortality
=
3
3
.
Invertebrates
No
Production
Impairment
=
8
Some
Production
Impairment
=
5
Acute
Mortality
Limit
=
4
Table
3
.
Sample
calculations
for
determining
daily
means
and
7­
day
mean
dissolved
oxygen
concentrations
(
30­
day
averages
are
calculated
in
a
similar
fashion
using
30
days
data).
Dissolved
Oxygen
(
mg/
L)

Day
Daily
Max.
Daily
Min.
Daily
Mean
9.0
7.0
10.0
7.0
11.0
8.0
12.
oa
8.0
10.0
8.0
11.0
9
0
12.03
10.0
8.0
8.5
9.5
9.0
10.0
1Q.
5c
9.5b
57.0
65.0
1­
day
Minimum
7.0
?­
day
Mean
Minimum
8.1
?­
day
Mean
9.3
a
Above
air
saturation
concentration
(
assumed
to
be
11.0
mg/
L
for
this
example).

b
(
11.0
­
b
8.0)
2.

c
(
11
0
+
10.0)
2.

The
criteria
have
been
established
on
the
basis
that
the
maximum
dissolved
oxygen
value
actually
used
in
calculating
any
daily
mean
should
not
exceed
the
air
saturation
value.
This
consideration
is
based
primarily
on
analysis
of
studies
of
c
y
c
l
i
n
g
d
i
s
s
o
l
v
e
d
oxygen
and
the
growth
of
largemouth
bass
(
Stewart
et
al.,
1967),
which
indicated
that
high
dissolved
oxygen
levels
(>
6
mg/
L)
had
no
beneficial
effect
on
growth.

During
periodic
cycles
of
dissolved
oxygen
concentrations,

minima
lower
than
acceptable
constant
exposure
levels
are
toler­

able
so
long
as:
1.
the
average
concentration
attained
meets
or
exceeds
the
2
.
the
average
dissolved
oxygen
concentration
is
calculated
as
3
.
criterion;

recommended
in
Table
3
;
and
the
minima
are
not
unduly
stressful
and
clearly
are
not
lethal.

A
daily
minimum
has
been
included
to
make
certain
that
no
acute
mortality
of
sensitive
species
occurs
as
a
result
of
lack
of
oxygen.
Because
repeated
exposure
to
dissolved
oxygen
concentrations
at
or
near
the
acute
lethal
threshold
will
be
stressful
and
because
stress
can
indirectly
produce
mortality
or
other
adverse
effects
(
e­
g.,
through
disease),
the
criteria
are
designed
to
prevent
significant
episodes
of
continuous
or
regularly
recurring
exposures
to
dissolved
oxygen
concentrations
at
or
near
the
lethal
threshold.
This
protection
has
been
achieved
by
setting
the
daily
minimum
for
early
life
stages
at
the
subacute
lethality
threshold,
by
the
use
of
a
7­
day
averaging
period
for
early
life
stages,
by
stipulating
a
7­
day
mean
minimum
value
for
other
life
stages,
and
by
recommending
additional
limits
for
manipulatable
discharges.

The
previous
EPA
criterion
for
dissolved
oxygen
published
in
Quality
Criteria
for
Water
(
USEPA,
1976)
was
a
minimum
of
5
mg/
L
(
usually
applied
as
a
7410)
which
is
similar
to
the
current
criterion
minimum
except
for
other
life
stages
of
warmwater
fish
which
now
allows
a
criteria
are
similar
of
the
Federal
Water
7­
day
mean
minimum
of
4
mg/
L.
The
new
to
those
contained
in
the
1968
"
Green
Book"

Pollution
Control
Federation
(
FWPCA,
1968).
A.
The
Criteria
and
Monitoring
and
Design
Conditions
The
acceptable
mean
concentrations
should
be
attained
most
of
the
time,
but
some
deviation
below
these
values
would
probably
not
cause
significant
harm.
Deviations
below
the
mean
will
probably
be
serially
correlated
and
hence
apt
to
occur
on
consecutive
days.
The
significance
of
deviations
below
the
mean
will
depend
on
whether
they
occur
continuously
or
in
daily
cycles,
the
former
being
more
adverse
than
the
latter.
Current
knowledge
regarding
such
deviations
is
limited
primarily
to
labo­

ratory
growth
experiments
and
by
extrapolation
to
other
activity­

related
phenomena.

Under
conditions
where
large
daily
cycles
of
dissolved
oxygen
occur,
it
is
possible
to
meet
the
criteria
mean
values
and
consistently
violate
the
mean
minimum
criteria.
Under
these
conditions
the
mean
minimum
criteria
will
clearly
be
the
limiting
regulation
unless
alternatives
such
as
nutrient
control
can
dampen
the
daily
cycles.

The
significance
of
conditions
which
fail
to
meet
the
recommended
dissolved
oxygen
criteria
depend
largely
upon
five
factors:
(
1)
the
duration
of
the
event;
(
2)
the
magnitude
of
the
dissolved
oxygen
depression;
(
3
)
the
frequency
of
recurrence;
(
4
)

the
proportional
area
of
the
site
failing
to
meet
the
criteria,

and
(
5)
the
biological
significance
of
the
site
where
the
event
occurs.
Evaluation
of
an
event's
significance
must
be
largely
case­
and
site­
specific.
Common
sense
would
dictate
that
the
magnitude
of
the
depression
would
be
the
single
most
important
factor
in
general,
especially
if
the
acute
value
is
violated.
A
logical
extension
of
these
considerations
is
that
the
event
must
be
considered
in
the
context
of
the
level
of
resolution
of
the
monitoring
or
modeling
effort.
Evaluating
the
extent,
duration,

and
magnitude
of
an
event
must
be
a
function
of
the
spatial
and
temporal
frequency
of
the
data.
Thus,
a
single
deviation
below
the
criterion
takes
on
considerably
less
significance
where
continuous
monitoring
occurs
than
where
sampling
is
com­

prised
of
once­
a­
week
grab
samples.
This
is
so
because
based
on
continuous
monitoring
the
event
is
provably
small,
but
with
the
much
less
frequent
sampling
the
event
is
not
provably
small
and
can
be
considerably
worse
than
indicated
by
the
sample.
The
frequency
of
recurrence
is
of
considerable
interest
to
those
modeling
dissolved
oxygen
concentrations
because
the
return
period,
or
period
between
recurrences,
is
a
primary
modeling
consideration
contingent
upon
probabilities
of
receiving
water
volumes,
waste
loads,
temperatures,
etc.
It
should
be
apparent
that
return
period
cannot
be
isolated
from
the
other
four
factors
discussed
above.
Ultimately,
the
question
of
return
period
may
be
decided
on
a
site­
specific
basis
taking
into
account
the
other
factors
(
duration,
magnitude,
areal
extent,
and
biologi­

cal
significance)
mentioned
above.
Future
studies
of
temporal
patterns
of
dissolved
oxygen
concentrations,
both
within
and
between
years,
must
be
conducted
to
provide
a
better
basis
for
selection
of
the
appropriate
return
period.

In
conducting
wasteload
a1
location
and
treatment
plant
design
computations,
the
choice
of
temperature
in
the
models
will
be
important.
Probably
the
best
option
would
be
to
use
temperatures
consistent
with
those
expected
in
the
receiving
water
over
the
critical
dissolved
oxygen
period
for
t
h
e
biota.

B.
The
C
r
i
t
e
r
i
a
and
Manipulatable
Discharges
If
d
a
i
l
y
minimum
D
O
s
are
p
e
r
f
e
c
t
l
y
s
e
r
i
a
l
l
y
c
o
r
r
e
l
a
t
e
d
,

i.
e,
i
f
t
h
e
annual
lowest
d
a
i
l
y
minimum
dissolved
oxygen
concen­

t
r
a
t
i
o
n
is
a
d
j
a
c
e
n
t
i
n
t
i
m
e
t
o
t
h
e
n
e
x
t
lower
d
a
i
l
y
minimum
d
i
s
s
o
l
v
e
d
oxygen
c
o
n
c
e
n
t
r
a
t
i
o
n
and
one
of
these
two
minima
is
a
d
j
a
c
e
n
t
t
o
t
h
e
t
h
i
r
d
l
o
w
e
s
t
d
a
i
l
y
minimum
d
i
s
s
o
l
v
e
d
oxygen
c
o
n
c
e
n
t
r
a
t
i
o
n
,
etc.,
t
h
e
n
i
n
o
r
d
e
r
t
o
meet
t
h
e
7­
day
mean
minimum
c
r
i
t
e
r
i
o
n
it
is
u
n
l
i
k
e
l
y
t
h
a
t
t
h
e
r
e
w
i
l
l
be
more
t
h
a
n
three
o
r
four
consecutive
d
a
i
l
y
minimum
v
a
l
u
e
s
below
t
h
e
accept­

able
7­
day
mean
minimum.
Unless
t
h
e
d
i
s
s
o
l
v
e
d
oxygen
p
a
t
t
e
r
n
is
e
x
t
r
e
m
e
l
y
e
r
r
a
t
i
c
,
it
i
s
a
l
s
o
u
n
l
i
k
e
l
y
t
h
a
t
t
h
e
l
o
w
e
s
t
d
i
s
s
o
l
v
e
d
oxygen
concentration
w
i
l
l
be
appreciably
below
t
h
e
a
c
c
e
p
t
a
b
l
e
7­
day
mean
minimum
o
r
t
h
a
t
d
a
i
l
y
minimum
v
a
l
u
e
s
below
t
h
e
7­
day
mean
minimum
w
i
l
l
occur
i
n
more
t
h
a
n
one
o
r
two
weeks
each
y
e
a
r
.
For
some
d
i
s
c
h
a
r
g
e
s
,
t
h
e
d
i
s
t
r
i
b
u
t
i
o
n
of
d
i
s
s
o
l
v
e
d
oxygen
concentrations
can
be
manipulated
t
o
varying
degrees.
Applying
t
h
e
d
a
i
l
y
minimum
t
o
manipulatable
discharges
would
a
l
l
o
w
repeated
weekly
c
y
c
l
e
s
of
minimum
a
c
u
t
e
l
y
acceptable
d
i
s
s
o
l
v
e
d
oxygen
v
a
l
u
e
s
,
a
condition
of
unacceptable
stress
and
p
o
s
s
i
b
l
e
adverse
b
i
o
l
o
g
i
c
a
l
effect.
For
t
h
i
s
r
e
a
s
o
n
,
t
h
e
a
p
p
l
i
c
a
t
i
o
n
of
t
h
e
one
day
minimum
c
r
i
t
e
r
i
o
n
t
o
m
a
n
i
p
u
l
a
t
a
b
l
e
d
i
s
c
h
a
r
g
e
s
must
l
i
m
i
t
e
i
t
h
e
r
t
h
e
frequency
of
o
c
c
u
r
r
e
n
c
e
o
f
v
a
l
u
e
s
below
t
h
e
a
c
c
e
p
t
a
b
l
e
7­
day
mean
minimum
o
r
must
impose
f
u
r
t
h
e
r
l
i
m
i
t
s
on
t
h
e
e
x
t
e
n
t
of
excursions
below
t
h
e
7­
day
mean
minimum.
For
such
c
o
n
t
r
o
l
l
e
d
discharges,
it
is
recommended
t
h
a
t
t
h
e
o
c
c
u
r
r
e
n
c
e
of
d
a
i
l
y
minima
below
t
h
e
a
c
c
e
p
t
a
b
l
e
7­
day
mean
I
minimum
b
e
l
i
m
i
t
e
d
t
o
3
weeks
per
year
or
t
h
a
t
t
h
e
acceptable
one­
day
minimum
be
increased
t
o
4
.
5
mg/
L
f
o
r
coldwater
f
i
s
h
and
3
.
5
mg/
L
f
o
r
warmwater
f
i
s
h
,
Such
d
e
c
i
s
i
o
n
s
c
o
u
l
d
be
s
i
t
e
­

s
p
e
c
i
f
i
c
based
upon
the
extent
of
control
and
s
e
r
i
a
l
correlation.
e
PARATHION
CRITERION:

0.04
ug/
L
f
o
r
freshwater
and
marine
a
q
u
a
t
i
c
l
i
f
e
.

RATIONALE:

Acute
s
t
a
t
i
c
LC50
v
a
l
u
e
s
of
t
h
e
organophosphorus
p
e
s
t
i
c
i
d
e
,

parathion,
f
o
r
freshwater
f
i
s
h
have
ranged
g
e
n
e
r
a
l
l
y
from
about
50
ug/
L
f
o
r
more
s
e
n
s
i
i
t
i
v
e
s
p
e
c
i
e
s
such
a
s
b
l
u
e
g
i
l
l
s
,
Lepomis
macrochi=,
t
o
about
2.5
mg/
L
f
o
r
t
h
e
more
r
e
s
i
s
t
a
n
t
s
p
e
c
i
e
s
s
u
c
h
as
minnows
(
U.
S.
Environ.
P
r
o
t
.
Agency,
1975).
I
n
f
l
o
w
i
n
g
w
a
t
e
r
e
x
p
o
s
u
r
e
s
,
S
p
a
c
i
e
(
1975)
o
b
t
a
i
n
e
d
96­
hour
LC50
v
a
l
u
e
s
o
f
0.5
mg/
L,
1
.
6
mg/
L,
and
1
.
7
6
mg/
L
f
o
r
b
l
u
e
g
i
l
l
s
,
_­
Lepomis
____

macrochirus,
f
a
t
h
e
a
d
minnows,
­­
Pimep&
ales
p
r
o
m
e
l
a
s
,
and
brook
t
r
o
u
t
,
S
a
l
v
e
l
i
n
u
s
_
_
_
_
_
_
_
I
f
o
n
t
i
n
a
l
i
s
r
e
s
p
e
c
t
i
v
e
l
y
.
Korn
and
E
a
r
n
e
s
t
(
1974)
found
a
96­
hOUr
LC50
o
f
18
ug/
L
f
o
r
j
u
v
e
n
i
l
e
f
r
e
s
h
w
a
t
e
r
and
e
s
t
u
a
r
i
n
e
s
t
r
i
p
e
d
bass,
Morone
s
a
x
a
t
i
l
i
s
,
i
n
a
flowing
water
system,
I
Few
c
h
r
o
n
i
c
e
x
p
o
s
u
r
e
d
a
t
a
a
r
e
a
v
a
i
l
a
b
l
e
f
o
r
a
q
u
a
t
i
c
organisms.
Brown
bullheads,
f
c
t
a
l
u
r
u
s
nebulosus,
exposed
t
o
30
ug/
L
p
a
r
a
t
h
i
o
n
f
o
r
30
days
e
x
h
i
b
i
t
e
d
t
r
e
m
o
r
s
;
a
t
6
0
ug/
L
t
h
e
y
convulsed
and
were
found
to
have
developed
a
deformed
vertebral
column
(
Mount
and
Boyle,
1
9
6
9
)
.
I
n
a
23­
mOnth
exposure
o
f
b
l
u
e
g
i
l
l
s
,
S
p
a
c
i
e
(
1975)
o
b
s
e
r
v
e
d
d
e
f
o
r
m
i
t
i
e
s
(
s
c
o
l
i
o
s
i
s
and
a
characteristic
protrusion
i
n
t
h
e
t
h
r
o
a
t
region)
a
t
0.34
ug/
L,
b
u
t
n
o
t
a
t
0.16
ug/
L.
Tremors,
c
o
n
v
u
l
s
i
o
n
s
,
h
y
p
e
r
s
e
n
s
i
t
i
v
i
t
y
,
and
hemorrhages
also
were
evident
at
higher
concentrations.

Reproductive
impairment
and
d
e
f
o
r
m
i
t
i
e
s
were
observed
i
n
f
a
t
h
e
a
d
minnows
e
x
p
o
s
e
d
t
o
4
.
0
ug/
L
f
o
r
8
1
/
2
months.
Development
of
brook
trout,
S,
fontinalis
­­
embryos
exposed
to
32
ug/
L
was
abnormal
and
mortalities
associated
with
premature
hatching
were
observed.
Embryos
at
10
ug/
L
appeared
normal.
No
adverse
effects
on
juveniles
and
adults
was
evident
during
9
months'
exposure
to
7
ug/
L.

Inhibition
of
cholinesterase
enzymes
is
the
well­
established
mode
of
physiological
action
of
parathion
and
other
organic
phosphorus
pesticides
(
Weiss,
1958).
The
degree
of
inhibition
of
brain
acetylcholinesterase
(
AChE)
activity
has
been
the
most
frequently
used
measure
of
effect
of
these
pesticides.
Various
studies
(
Weiss,
1958,
1959,
1961;
Murphy
et
al.,
1968;
Gibson
et
al.
1969)
have
shown
the
degree
of
inhibition
to
be
dependent
upon
toxicant
concentration,
length
of
exposure,
and
species
sensitivity.
The
results
of
these
studies
have
also
indicated
AChE
inhibition
ranging
from
25
to
90
(
1959)
also
showed
that
susceptibility
of
recovery
of
AChE
activity
following
that
death
results
from
percent
of
normal.
Weiss
depended
upon
the
extent
prior
exposure
and
that
the
recovery
period
for
fish
exposed
to
parathion
was
relatively
long.
In
bluegills,
AChE
activity
was
only
50
percent
recovered
30
days
after
exposure
to
1
mg/
L
for
6
to
7
hours
(
Weiss,
1961).

Some
of
the
other
physiological
effects
observed
to
result
from
exposure
of
fish
to
parathion
have
been
inhibition
of
spermatogenesis
in
guppies,
_______­
Poecilia
reticulata
at
10
ug/
L
(
Billard
and
deKinkelin,
1970),
alternation
of
oxygen
consumption
rate
in
bluegills,
Lepomis
macrochirus,
at
100
ug/
L
(
Dowden,
1966),
and
liver
enlargement
associated
with
increased
pesticide­
hydrolizing
capability
in
mosquitof
ish,
Gambusia
affinis
(
Ludke,
1970).
0
Parathion
has
been
found
acutely
toxic
to
aquatic
invertebrates
at
under
1
ug/
L
e.
g.,
a
50­
hour
LC50
of
0.8
ug/
L
for
­
Daphnia
maqgg;
48­
hour
LC50
of
0.6
ug/
L
for
­­
Daphnia
­­
pulex;
­­­

48­
hour
LC50
of
0.37
for
Simocephalus
­­­_
_____­_
serrulatus
(
a
daphnid)

(
Sanders
and
Cope,
1966);
a
5­
day
LC50
of
0.93
ug/
L
for
the
larval
stonefly,
Acroneuria
pacifica
(
Jensen
and
Gaufin,
1964);

and
a
96­
hour
LC50
of
0.43
ug/
L
for
the
larval
caddisfly
­
HydroEsychc
­­­
­
­­
californica
­
­­­­­­­­
(
Gaufin
et
al.
1965).
Mulla
and
Khasawinah
(
1969)
obtained
a
24­
hour
LC50
of
0.5
ug/
L
for
4th
instar
larvae
of
the
midge
­­
Tanypus
­
qrodhausi.
Spacie
(
1975)

obtained
96­
hour
LC5O's
in
flow­
through
bioassays
of
0.62
ug/
L
for
Daphnia
magna,
0.40
ug/
L
for
the
scud,
Gammarus
fasciatus
and
31.0
ug/
L
for
4th
instar
of
Chironomous
tentans,
a
midge.

Other
invertebrates
have
been
found
acutely
sensitive
to
parathion
in
concentrations
of
from
1
to
30
ug/
L
in
water
(
U.
S.

Environ.
Prot.
Agency,
1975).

Few
longer
exposures
have
been
conducted.
Jensen
and
Gaufin
(
1964)
obtained
30­
day
LC50'
s
for
Pteronarcys
­
­­­­
I­­
californica
and
Acroneuria
­­­­_
pacifica
­­­­
of
2.2
and
0.44
ug/
L,
respectively.
Spacie
(
1975)
found
the
3­
week
LC50
for
Daphnia
mxna
­
to
be
0.14
ug/
L.

Statistically
significant
reproductive
impairment
occurred
at
concentrations
above
0.08
ug/
L.
A
43­
day
LC50
of
0.07
ug/
L
was
reported
for
Gammarus
fasciatus
and
a
concentration
of
0.04
ug/
L
produced
significantly
greater
mortality
than
among
controls.

Limited
information
is
available
on
persistence
of
parathion
~

in
water.
Eichelberger
and
Lichtenberg
(
1971)
determined
the
half­
life
in
river
water
(
pH
7.3
­
8.0)
to
be
1
week.
Using
AChE
inhibitory
capacity
as
the
indicator,
Weiss
and
Gakstatter
(
1964)
found
the
half­
life
of
parathion
or
its
active
breakdown
products
to
be
40,
35,
and
2
0
days
in
`
lnaturalll
waters
having
a
pH
of
5.1.
7.0,
and
8.4,
respectively.
The
possibility
of
breakdown
resulting
in
compounds
more
toxic
than
parathion
was
suggested
by
Burke
and
Ferguson
(
1969)
who
determined
that
the
toxicity
of
this
pesticide
to
mosquitofish
1
Gambusia
­­­_
I
affinis
was
greater
in
static
than
in
flowing
water
test
systems.

Sanders
(
1972),
in
96­
hour
bioassays
with
the
scud,
Gammarus
_­

­­­­­­­
I
fasciatus
and
glass
shrimp,
Palaemonetes
­_­
I_­
kadiakensis
I
also
observed
greater
toxicity
under
static
than
in
flow­
through
conditions.

a
Tissue
accumulations
of
parathion
by
exposed
aquatic
organisms
are
not
great
and
do
not
appear
to
be
very
persistent.

Mount
and
Boyle
(
1969)
observed
concentrations
in
the
blood
of
bullhead,
­­­­­­­­­
Ictalurus
melas
­
­
­
­
­
I
up
to
about
50
times
water
concentrations.
Spacie
(
1975)
found
muscle
concentrations
in
chronically
exposed
brook
trout,
­
S
.
___­­_
I
fontinalis
to
be
several
hundred
times
water
concentrations;
bluegills,
­­
Lepomis
­­­­

macrochirus,
had
about
25
times
water
concentrations
in
their
bodies.
Leland
(
1968)
demonstrated
a
biological
half­
life
of
parathion
in
rainbow
trout,
Salmo
Eirdneri
1
exposed
and
then
placed
in
fresh
water
to
be
only
30
to
4
0
hours.
It
is
not
expected
that
parathion
residues
in
aquatic
organisms
exposed
to
the
recommended
criterion
concentrations
will
be
a
hazard
to
consumer
organisms.
a
Weiss
and
Gakstatter
(
1964)
have
shown
that
15­
day
continuous
exposure
to
parathion
(
1.0
ug/
L)
can
produce
progressively
greater
(
i.
e­,
cumulative)
brain
AChE
inhibition
in
a
fish
species.
After
substantial
inhibition
by
parathion
exposure,
it
takes
several
weeks
for
brain
AChE
of
exposed
fishes
to
return
to
normal
even
though
exposure
is
discontinued
(
Weiss,
1959,

1961).
fishes
by
46
percent
or
more
has
been
associated
with
harmful
effects
in
exposures
to
organophosphate
pesticides
for
one
life
cycle
(
Eaton,
1970)
and
for
short
periods
(
Carter,
1971;
Coppage
and
Duke,
1971;
Coppage,

1972;
Coppage
and
Matthews,
1974;
Post
and
Leasure,
1974;
Coppage
et
al.
1975).
It
has
been
shown
that
a
concentration
of
10
ug
Inhibition
of
brain
AChE
of
parathion/
L
of
flowing
seawater
kills
40
to
60
percent
of
the
marine
fishes
&
agodon
­­­­
­­­­­­­­­­
rhomboides
(
pinfish)
and
Leostomus
­­

xanthurus
__
(
spot)
in
24
hours
and
causes
about
87
to
92
percent
brain
AChE
inhibition
(
Coppage
and
Matthews,
1974.)
Similar
inhibition
of
AChE
and
mortality
were
caused
in
sheepshead
minnows,
Cyprinodon
variegatus,
in
2,
24,
48,
and
72
hours
at
concentrations
of
5,000,
2,000,
100,
and
10
ug/
L,
respectively
in
static
tests
(
Coppage,
1972).
These
data
indicate
that
reductions
of
brain
AChE
activity
of
marine
fishes
by
70
to
80
percent
or
more
in
short­
term
exposures
to
parathion
may
be
associated
with
some
deaths.
0
Other
estimates
of
parathion
toxicity
to
marine
organisms
follow.
The
48­
hour
EC50
for
parathion
to
Penaeus
duorarum
was
found
to
be
0.2
ug/
L
(
Lowe
et
al.
1970).
Lahav
and
Sarig
(
1969)

reported
the
96­
hour
LC50
for
mullet,
Mugil
cephalus
to
be
125
ug/
L.
The
shell
growth
of
the
oyster,
Crassostrea
virginica,
was
found
by
Lowe
et
al.
(
1970)
to
be
decreased
by
22
percent
after
96
hours
in
1.0
mg/
L.

An
application
factor
of
0.1
is
applied
to
the
96­
hour
LC50
data
for
invertebrates
which
range
upward
from
0.4
ug/
L.
A
criteria
of
0.04
ug/
L
is
recommended
for
marine
and
freshwater
aquatic
life.

(
QUALITY
CRITERIA
FOR
WATER,
JULY
1976)
PB­
263943
SEE
APPENDIX
C
FOR
METHODOLOGY
CRITERIA:
0
PENTACHLOROPHENOL
Aquatic
Life
The
available
data
for
pentachlorophenol
indicate
that
acute
and
chronic
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
55
and
3.2
ug/
L,
respectively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.

The
available
data
for
pentachlorophenol
indicate
that
acute
and
chronic
toxicity
to
saltwater
aquatic
life
occur
at
concentrations
as
l
o
w
as
53
and
34
ug/
L,
respectively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.

Human
Health
For
comparison
purposes,
two
approaches
were
used
to
derive
criterion
levels
for
pentachlorophenol.
Based
on
available
toxicity
data,
to
protect
public
health
the
derived
level
is
1.01
W/
L.
Using
available
organoleptic
data,
to
control
undesirable
taste
and
odor
qualities
of
ambient
water
the
estimated
level
is
30
ug/
L.
It
should
be
recognized
that
organoleptic
data
have
limitations
as
a
basis
for
establishing
a
water
quality
criterion,
and
have
no
demonstrated
relationship
to
potential
adverse
human
health
effects.

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
CRITERIA:

Range
5
­
9
Domestic
water
supplies
(
welfare)

6.5
­
9.0
Freshwater
aquatic
l
i
f
e
6.5
­
8.5
Marine
aquatic
l
i
f
e
(
but
not
more
than
0.2
u
n
i
t
s
o
u
t
s
i
d
e
of
normal
lyoccurring
range.
)

INTRODUCTION
:

I8pH"
is
a
measure
of
t
h
e
hydrogen
i
o
n
a
c
t
i
v
i
t
y
i
n
a
water
sample.
It
i
s
mathematically
r
e
l
a
t
e
d
t
o
hydrogen
i
o
n
a
c
t
i
v
i
t
y
according
t
o
t
h
e
expression:
pH
=
­
log
10
(
H'),
where
(
H')
is
t
h
e
hydrogen
ion
a
c
t
i
v
i
t
y
.

The
pH
o
f
n
a
t
u
r
a
l
waters
is
a
measure
of
acid­
base
equilibrium
achieved
by
t
h
e
various
dissolved
compounds,
s
a
l
t
s
,

and
gases.
The
p
r
i
n
c
i
p
a
l
system
r
e
g
u
l
a
t
i
n
g
p
H
i
n
n
a
t
u
r
a
l
waters
is
t
h
e
c
a
r
b
o
n
a
t
e
system
which
i
s
composed
of
carbon
d
i
o
x
i
d
e
(
C02),
c
a
r
b
o
n
i
c
a
c
i
d
,
(
H
z
C
0
3
)
,
b
i
c
a
r
b
o
n
a
t
e
i
o
n
(
H
C
0
3
)
and
c
a
r
b
o
n
a
t
e
i
o
n
s
(
C
o
g
)
.
The
i
n
t
e
r
a
c
t
i
o
n
s
and
k
i
n
e
t
i
c
s
of
t
h
i
s
system
have
been
described
by
Stumm
and
Morgan
(
1970).

pH
is
an
important
f
a
c
t
o
r
i
n
t
h
e
chemical
and
b
i
o
l
o
g
i
c
a
l
systems
of
n
a
t
u
r
a
l
waters.
The
d
e
g
r
e
e
o
f
d
i
s
s
o
c
i
a
t
i
o
n
of
weak
a
c
i
d
s
o
r
bases
i
s
affected
by
changes
i
n
pH.
T
h
i
s
effect
is
important
because
t
h
e
t
o
x
i
c
i
t
y
of
many
compounds
is
affected
by
t
h
e
degree
of
dissociation.
One
such
example
is
hydrogen
cyanide
(
HCN).
Cyanide
t
o
x
i
c
i
t
y
t
o
f
i
s
h
i
n
c
r
e
a
s
e
s
a
s
t
h
e
p
H
i
s
lowered
­

because
t
h
e
chemical
e
q
u
l
i
b
r
i
u
m
is
s
h
i
f
t
e
d
toward
a
n
i
n
c
r
e
a
s
e
d
c
o
n
c
e
n
t
r
a
t
i
o
n
o
f
HCN.
S
i
m
i
l
a
r
r
e
s
u
l
t
s
h
a
v
e
been
shown
f
o
r
hydrogen
s
u
l
f
i
d
e
(
Has)
(
Jones,
1
9
6
4
)
.
The
solubility
of
metal
compounds
contained
in
bottom
sediments
or
as
suspended
material,
also
is
affected
by
pH.
For
example,
laboratory
equilibrium
studies
under
anaerobic
conditions
indicated
that
pH
was
an
important
parameter
involved
in
releasing
manganese
from
bottom
sediments
(
Delfino
and
Lee,

1971).

The
pH
of
a
water
does
not
indicate
ability
to
neutralize
additions
of
acids
or
bases
without
appreciable
change.
This
characteristic,
termed
"
buffering
capacity,"
is
control
led
by
the
amounts
of
alkalinity
and
acidity
present.

RATIONALE:

Knowledge
of
pH
in
the
raw
water
used
 or
public
water
supplies
is
important
because
without
adjustment
to
a
suitable
level,
such
waters
may
be
corrosive
and
adversely
affect
treatment
processes
including
coagulation
and
chlorination.

Coagulation
for
removal
of
colloidal
color
by
use
of
aluminum
o
r
iron
salts
generally
has
an
optimum
pH
range
of
5.0
to
6
­
5
(
Sawyer,
1960).
Such
optima
are
predicated
upon
the
availability
of
sufficient
alkalinity
to
complete
the
chemical
reactions.

The
effect
of
pH
on
chlorine
in
water
principally
is
on
the
equilibrium
between
hypochlorous
acid
(
HOC1)
and
the
hypochlorite
ion
(
OC1­)
according
to
the
reaction:

HOCl
=
H+
+
OC1­

Butterfield
(
1984)
has
shown
that
chlorine
disinfection
is
more
effective
at
values
less
than
pH
7.
Another
study
(
Reid
and
Carlson,
1974)
has
indicated,
however,
that
in
natural
waters
no
significant
difference
in
the
kill
rate
for
Escherichia
­­
coli
was
c
observed
between
pH
6
and
pH
8.

corrosion
of
plant
equipment
and
piping
in
the
distribution
system
can
lead
to
expensive
replacement
as
well
as
the
introduction
of
metal
ions
such
as
copper,
lead,
zinc,
and
cadmium.
Langelier
(
1936)
developed
a
method
to
calculate
and
control
water
corrosive
activity
that
employs
calcium
carbonate
saturation
theory
and
predicts
whether
the
water
would
tend
to
dissolve
or
deposit
calcium
carbonate.
By
maintaining
the
pH
at
the
proper
level,
the
distribution
system
can
be
provided
with
a
protective
calcium
carbonate
lining
which
prevents
metal
pipe
corrosion.
Generally,
this
level
is
above
pH
7
and
frequently
approaches
pH
8.3,
the
point
of
maximum
bicarbonate/
carbonate
buffering.

Since
pH
is
relatively
easily
adjusted
prior
to
and
during
water
treatment,
a
rather
wide
range
is
acceptable
for
waters
serving
as
a
source
of
public
water
supply.
A
range
of
p~
from
5.0
to
9.0
would
provide
a
water
treatable
by
typical
(
coagulation,
sedimentation,
filtration,
and
chlorination)

treatment
plant
processes.
As
the
range
is
extended,
the
cost
of
neutralizing
chemicals
increases.

A
review
of
the
effects
of
pH
on
fresh
water
fish
has
been
published
by
the
European
Inland
Fisheries
Advisory
Commission
(
1969).
The
commission
concluded:

There
is
no
definite
pH
range
within
which
a
fishery
is
unharmed
and
outside
which
it
is
damaged,
but
rather,
there
is
a
gradual
deterioration
as
the
pH
values
are
further
removed
from
the
normal
range.
The
pH
range
which
is
not
directly
lethal
to
fish
is
5
­
9;
however,
the
toxicity
of
several
common
pollutants
is
markedly
affected
by
pH
changes
within
this
range,
and
increasing
acidity
or
alkalinity
may
make
these
poisons
more
toxic.
A
l
s
o
,
an
acid
discharge
may
liberate
sufficient
C02
from
bicarbonate
in
the
water
either
to
be
directly
toxic,
or
to
cause
the
pH
range
5
­
6
to
become
lethal.

Mount
(
1973)
performed
bioassays
on
the
fathead
minnow,

Pimephal­
promelas,
for
a
13­
mQnth,
one
generation
time
period
to
determine
chronic
pH
effects.
Tests
were
run
at
pH
levels
of
4.5,
5.2.

PH
Range
Effect
on
Fish*

5.0
­
6.0
Unlikely
to
be
harmful
to
any
species
unless
either
the
concentration
of
free
C02
is
greater
than
20
ppm,
or
the
water
contains
iron
salts
which
are
precipitated
as
ferric
hydroxide,
the
toxicity
of
which
is
not
known.

is
present
in
excess
of
100
ppm.
6.0
­
6.5
Unlikely
to
be
harmful
to
fish
unless
free
carbon
dioxide
6.5
­
9.0
Harmless
to
fish,
although
the
toxicity
of
other
poisons
may
be
affected
by
changes
within
this
range.

EIFAC,
1969
5.9,
6.6,
and
a
c
o
n
t
r
o
l
o
f
7.5.
A
t
t
h
e
two
l
o
w
e
s
t
pH
v
a
l
u
e
s
(
4.5
and
5.2)
behavior
was
abnormal
and
t
h
e
f
i
s
h
were
deformed.
A
t
pH
v
a
l
u
e
s
less
t
h
a
n
6.6,
egg
p
r
o
d
u
c
t
i
o
n
and
h
a
t
c
h
a
b
i
l
i
t
y
w
e
r
e
reduced
when
compared
w
i
t
h
t
h
e
c
o
n
t
r
o
l
.
It
w
a
s
concluded
t
h
a
t
a
pH
of
6.6
w
a
s
marginal
f
o
r
v
i
t
a
l
l
i
f
e
functions.

B
e
l
l
(
1971)
performed
b
i
o
a
s
s
a
y
s
w
i
t
h
nymphs
of
c
a
d
d
i
s
f
l
i
e
s
(
t
w
o
s
p
e
c
i
e
s
)
s
t
o
n
e
f
l
i
e
s
(
f
o
u
r
s
p
e
c
i
e
s
)
,
d
r
a
g
o
n
f
l
i
e
s
(
two
s
p
e
c
i
e
s
)
,
and
m
a
y
f
l
i
e
s
(
one
s
p
e
c
i
e
s
)
.
A
l
l
a
r
e
i
m
p
o
r
t
a
n
t
f
i
s
h
food
organisms.
The
30­
day
TL50
v
a
l
u
e
s
ranged
from
2.45
t
o
5.38
w
i
t
h
t
h
e
c
a
d
d
i
s
f
l
i
e
s
being
t
h
e
most
t
o
l
e
r
a
n
t
and
t
h
e
m
a
y
f
l
i
e
s
being
t
h
e
l
e
a
s
t
t
o
l
e
r
a
n
t
.
The
pH
v
a
l
u
e
s
a
t
which
50
percent
of
t
h
e
organisms
emerged
ranged
from
4.0
t
o
6.6
w
i
t
h
i
n
c
r
e
a
s
i
n
g
p
e
r
c
e
n
t
a
g
e
emergence
o
c
c
u
r
r
i
n
g
w
i
t
h
t
h
e
i
n
c
r
e
a
s
i
n
g
pH
v
a
l
u
e
s
.

Based
on
p
r
e
s
e
n
t
e
v
i
d
e
n
c
e
,
a
pH
r
a
n
g
e
of
6.5
t
o
9.0
a
p
p
e
a
r
s
t
o
provide
adequate
p
r
o
t
e
c
t
i
o
n
f
o
r
t
h
e
l
i
f
e
of
freshwater
f
i
s
h
and
bottom
dwelling
i
n
v
e
r
t
e
b
r
a
t
e
s
f
i
s
h
food
organisms.
Outside
of
t
h
i
s
range,
f
i
s
h
s
u
f
f
e
r
adverse
p
h
y
s
i
o
l
o
g
i
c
a
l
effects
increasing
i
n
s
e
v
e
r
i
t
y
a
s
t
h
e
d
e
g
r
e
e
of
d
e
v
i
a
t
i
o
n
i
n
c
r
e
a
s
e
s
u
n
t
i
l
l
e
t
h
a
l
l
e
v
e
l
s
are
reached.

Conversely,
r
a
p
i
d
i
n
c
r
e
a
s
e
s
i
n
pH
c
a
n
c
a
u
s
e
i
n
c
r
e
a
s
e
d
NH3
concentrations
t
h
a
t
a
r
e
a
l
s
o
t
o
x
i
c
.
Ammonia
has
been
shown
t
o
be
1
0
t
i
m
e
s
as
t
o
x
i
c
a
t
pH
8.0
as
a
t
pH
7.0
(
EIFAC,
1969).

T
h
e
c
h
e
m
i
s
t
r
y
of
marine
w
a
t
e
r
s
d
i
f
f
e
r
s
f
r
o
m
that
of
fresh
water
because
of
t
h
e
l
a
r
g
e
c
o
n
c
e
n
t
r
a
t
i
o
n
of
s
a
l
t
s
p
r
e
s
e
n
t
.
I
n
a
d
d
i
t
i
o
n
t
o
a
l
k
a
l
i
n
i
t
y
based
on
t
h
e
c
a
r
b
o
n
a
t
e
system,
there
i
s
a
l
s
o
a
l
k
a
l
i
n
i
t
y
from
o
t
h
e
r
weak
a
c
i
d
s
a
l
t
s
such
a
s
b
o
r
a
t
e
.

Because
of
t
h
e
b
u
f
f
e
r
i
n
g
system
p
r
e
s
e
n
t
i
n
seawater,
t
h
e
n
a
t
u
r
a
l
l
y
o
c
c
u
r
r
i
n
g
v
a
r
i
a
b
i
l
i
t
y
o
f
pH
is
less
t
h
a
n
i
n
f
r
e
s
h
water.
Some
marine
communities
are
more
s
e
n
s
i
t
i
v
e
t
o
pH
change
than
others
(
NAS,
1974).
Normal
pH
v
a
l
u
e
s
i
n
seawater
are
8.0
t
o
8.2
a
t
t
h
e
s
u
r
f
a
c
e
,
d
e
c
r
e
a
s
i
n
g
t
o
7.7
t
o
7.8
w
i
t
h
i
n
c
r
e
a
s
i
n
g
d
e
p
t
h
(
Capurro,
1970).
The
NAS
Committee's
review
(
NAS,
1974)

indicated
t
h
a
t
plankton
and
benthic
i
n
v
e
r
t
e
b
r
a
t
e
s
are
probably
more
s
e
n
s
i
t
i
v
e
t
h
a
n
f
i
s
h
t
o
changes
i
n
pH
and
t
h
a
t
mature
forms
and
larvae
of
o
y
s
t
e
r
s
are
adversely
affected
a
t
t
h
e
extremes
of
t
h
e
pH
r
a
n
g
e
o
f
6.5
t
o
9.0.
However,
i
n
t
h
e
s
h
a
l
l
o
w
,

b
i
o
l
o
g
i
c
a
l
l
y
a
c
t
i
v
e
w
a
t
e
r
s
i
n
t
r
o
p
i
c
a
l
o
r
s
u
b
t
r
o
p
i
c
a
l
areas,

l
a
r
g
e
d
i
u
r
n
a
l
pH
c
h
a
n
g
e
s
o
c
c
u
r
n
a
t
u
r
a
l
l
y
b
e
c
a
u
s
e
o
f
p
h
o
t
o
s
y
n
t
h
e
s
i
s
.
pH
v
a
l
u
e
s
may
r
a
n
g
e
from
9.5
i
n
t
h
e
daytime
t
o
7.3
i
n
t
h
e
e
a
r
l
y
morning
b
e
f
o
r
e
dawn.
A
p
p
a
r
e
n
t
l
y
,
t
h
e
s
e
communities
are
adapted
t
o
such
v
a
r
i
a
t
i
o
n
s
or
i
n
t
o
l
e
r
a
n
t
s
p
e
c
i
e
s
are
a
b
l
e
t
o
a
v
o
i
d
extremes
bymoving
o
u
t
o
f
t
h
e
a
r
e
a
.

For
open
ocean
waters
where
t
h
e
d
e
p
t
h
i
s
s
u
b
s
t
a
n
t
i
a
l
l
y
g
r
e
a
t
e
r
than
t
h
e
euphotic
zone,
t
h
e
pH
should
not
be
changed
more
than
0.2
u
n
i
t
s
o
u
t
s
i
d
e
of
t
h
e
n
a
t
u
r
a
l
l
y
occurring
v
a
r
i
a
t
i
o
n
o
r
i
n
any
c
a
s
e
o
u
t
s
i
d
e
t
h
e
range
of
6.5
t
o
8.5.
For
s
h
a
l
l
o
w
,
h
i
g
h
l
y
productive
c
o
a
s
t
a
l
and
e
s
t
u
a
r
i
n
e
a
r
e
a
s
where
n
a
t
u
r
a
l
l
y
occurring
v
a
r
i
a
t
i
o
n
s
approach
t
h
e
l
e
t
h
a
l
l
i
m
i
t
s
f
o
r
some
species,
changes
i
n
pH
s
h
o
u
l
d
be
avoided,
b
u
t
i
n
any
case
n
o
t
exceed
t
h
e
l
i
m
i
t
s
e
s
t
a
b
l
i
s
h
e
d
f
o
r
fresh
water,
i.
e.,
pH
of
6.5
t
o
9.0.
A
s
w
i
t
h
f
r
e
s
h
w
a
t
e
r
c
r
i
t
e
r
i
a
,
r
a
p
i
d
pH
f
l
u
c
t
u
a
t
i
o
n
s
t
h
a
t
are
caused
by
waste
discharges
should
be
avoided.
Additional
support
f
o
r
these
l
i
m
i
t
s
i
s
p
r
o
v
i
d
e
d
by
Z
i
r
i
n
o
a
n
d
Yamamoto
(
1
9
7
2
)
.
These
i
n
v
e
s
t
i
g
a
t
o
r
s
developed
a
model
which
i
l
l
u
s
t
r
a
t
e
s
t
h
e
effects
of
0
%.

variable
pH
on
copper,
zinc,
cadmium,
and
lead;
small
changes
i
n
p~
cause
large
shifts
in
these
metallic
complexes.
Such
changes
may
affect
toxicity
of
these
metals.

For
the
industrial
classifications
considered,
the
NAS
report
(
NAs,
1974)
tabulated
the
range
of
pH
values
used
by
industry
for
various
process
and
cooling
purposes.
In
general,
process
waters
used
varied
from
pH
3.0
to
11.7,
while
cooling
waters
used
varied
from
5.0
to
8.9.
Desirable
pH
values
are
undoubtedly
closer
to
neutral
to
avoid
corrosion
and
other
deleterious
chemical
reactions.
Waters
with
pH
values
outside
these
ranges
are
considered
unusable
for
industrial
purposes.

The
pH
of
water
applied
for
irrigation
purposes
is
not
normally
a
critical
parameter.
Compared
with
the
large
buffering
capacity
of
the
soil
matrix,
the
pH
of
applied
water
is
rapidly
changed
to
approximately
that
of
the
soil.
The
greatest
danger
in
acid
soils
is
that
metallic
ions
such
as
iron,
manganese,
or
aluminum
may
be
dissolved
in
concentrations
which
are
subsequently
directly
toxic
to
plants.
Under
alkaline
conditions,

the
danger
to
plants
is
the
toxicity
of
sodium
carbonates
and
bicarbonates
either
directly
or
indirectly
(
NAS,
1974).

To
avoid
undesirable
effects
in
irrigation
waters,
the
pH
should
not
exceed
a
range
of
4.5
to
9.0.

(
QUALITY
CRITERIA
FOR
WATER,
JULY
1976)
PB­
263943
SEE
APPENDIX
C
FOR
METHODOLOGY
CRITERIA:
0
PHENOL
Aquatic
Life
The
available
data
for
phenol
indicate
that
acute
and
chronic
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
l
o
w
as
10,200
and
2,560
ug/
L,
respectively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.

The
available
data
for
phenol
indicate
that
toxicity
to
saltwater
aquatic
life
occurs
at
concentrations
as
low
as
5,800
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
NO
data
are
available
concerning
the
chronic
toxicity
of
phenol
to
sensitive
saltwater
aquatic
life.

0
Human
Health
For
comparison
purposes,
two
approaches
were
used
to
derive
criterion
levels
for
phenol.
Based
on
available
toxicity
data,

to
protect
public
health
the
derived
level
is
3.5
mg/
L.

Using
available
organoleptic
data,
to
control
undesirable
taste
and
odor
qualities
of
ambient
water
the
estimated
level
is
0.3
mg/
L.
It
should
be
recognized
that
organoleptic
data
have
limitations
as
a
basis
for
establishing
a
water
quality
criterion,
and
have
no
demonstrated
relationship
to
potential
adverse
human
health
effects.

NOTE:
The
U.
S.
EPA
is
currently
developing
Acceptable
Daily
Intake
(
ADI)
or
Verified
Reference
Dose
(
RfD)
values
for
Agency­
wide
use
for
this
chemical.
The
new
value
should
be
s
u
b
s
t
i
t
u
t
e
d
when
it
becomes
available.
The
January,
1986,
draft
Verified
Reference
Dose
document
cites
an
RfD
of
0.1
mg/
kg/
day
for
phenol.
.'/
(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
0
PHOSPHORUS
CRITERION::

0.10
ug/
L
yellow
(
elemental)
phosphorus
f
o
r
marine
o
r
e
s
t
u
a
r
i
n
e
water.

INTRODUCTION:

Phosphorus
i
n
t
h
e
elemental
form
is
p
a
r
t
i
c
u
l
a
r
l
y
t
o
x
i
c
and
is
s
u
b
j
e
c
t
t
o
bioaccumulation
i
n
much
t
h
e
same
way
a
s
mercury.

Phosphorus
a
s
phosphate
i
s
one
bf
t
h
e
major
n
u
t
r
i
e
n
t
s
r
e
q
u
i
r
e
d
f
o
r
p
l
a
n
t
n
u
t
r
i
t
i
o
n
and
is
e
s
s
e
n
t
i
a
l
f
o
r
l
i
f
e
.
I
n
excess
of
a
cr
it
ica
1
concent
r
a
t
i
o
n
,
phosphates
stirnu
l
a
t
e
p
1
a
n
t
growths.

During
t
h
e
p
a
s
t
30
years,
a
formidable
case
has
developed
f
o
r
t
h
e
b
e
l
i
e
f
t
h
a
t
i
n
c
r
e
a
s
i
n
g
s
t
a
n
d
i
n
g
c
r
o
p
s
of
a
q
u
a
t
i
c
p
l
a
n
t
s
,
which
o
f
t
e
n
i
n
t
e
r
f
e
r
e
w
i
t
h
water
u
s
e
s
and
are
n
u
i
s
a
n
c
e
s
t
o
man,

frequently
are
caused
by
increasing
s
u
p
p
l
i
e
s
of
phosphorus.
Such
phenomena
a
r
e
a
s
s
o
c
i
a
t
e
d
w
i
t
h
a
c
o
n
d
i
t
i
o
n
o
f
accelerated
eutrophication
o
r
aging
of
waters.
Generally,
it
is
recognized
t
h
a
t
phosphorus
is
not
t
h
e
s
o
l
e
cause
of
eutrophication
but
there
is
s
u
b
s
t
a
n
t
i
a
t
i
n
g
evidence
t
h
a
t
frequently
it
is
t
h
e
key
element
of
a
l
l
of
t
h
e
e
l
e
m
e
n
t
s
r
e
q
u
i
r
e
d
by
f
r
e
s
h
w
a
t
e
r
p
l
a
n
t
s
,
and
g
e
n
e
r
a
l
l
y
,
it
is
p
r
e
s
e
n
t
i
n
t
h
e
l
e
a
s
t
amount
r
e
l
a
t
i
v
e
t
o
need.

Therefore,
an
increase
i
n
phosphorus
allows
use
of
other
already
p
r
e
s
e
n
t
n
u
t
r
i
e
n
t
s
for
p
l
a
n
t
growth.
F
u
r
t
h
e
r
,
of
a
l
l
of
t
h
e
elements
r
e
q
u
i
r
e
d
f
o
r
p
l
a
n
t
growth
i
n
t
h
e
water
environment,

phosphorus
is
t
h
e
most
e
a
s
i
l
y
c
o
n
t
r
o
l
l
e
d
by
man.

Large
d
e
p
o
s
i
t
s
of
phosphate
rock
are
found
n
e
a
r
t
h
e
western
shore
of
C
e
n
t
r
a
l
F
l
o
r
i
d
a
,
a
s
w
e
l
l
a
s
i
n
a
number
of
o
t
h
e
r
States.

Deposits
i
n
F
l
o
r
i
d
a
are
found
i
n
t
h
e
form
of
p
e
b
b
l
e
s
which
v
a
r
y
0
.
.,
in
size
from
fine
sand
to
about
the
size
of
a
human
foot.
These
pebbles
are
embedded
in
a
matrix
of
clay
and
sand.
The
phosphate
rock
beds
lie
within
a
few
feet
of
the
surface
and
mining
is
accomplished
by
using
hydraulic
water
jets
and
a
washing
operation
that
separates
the
phosphates
from
waste
materials.
The
process
is
similar
to
that
of
strip­
mining.

Florida,
Idaho,
Montana,
North
Carolina,
South
Carolina,

Tennessee,
Utah,
Virginia,
and
Wyoming
share
phosphate
mining
activities.

Phosphates
enter
waterways
from
several
different
sources.

The
human
body
excretes
about
one
pound
per
year
of
phosphorus
expressed
as
I1Pt1.
The
use
of
phosphate
detergents
and
other
domestic
phosphates
increases
the
per
capita
contribution
to
about
3
.
5
pounds
per
year
of
phosphorus
as
P.
Some
industries,

such
as
potato
processing,
have
wastewaters
high
in
phosphates.

Crop,
forest,
idle,
and
urban
land
contribute
varying
amounts
of
phosphorus­
diffused
sources
in
drainage
to
watercourses.
This
drainage
may
be
surface
runoff
of
rainfall,
effluent
from
tile
lines,
or
return
flow
from
irrigation.
Cattle
feedlots,

concentrations
of
domestic
duck
or
wild
duck
populations,
tree
leaves,
and
fallout
from
the
atmosphere
all
are
contributing
sources.

Evidence
indicates
that:
(
1)
high
phosphorus
concentrations
are
associated
with
accelerated
eutrophication
of
waters,
when
other
growth­
promoting
factors
are
present:
(
2)
aquatic
plant
problems
develop
in
reservoirs
and
other
standing
waters
at
phosphorus
values
lower
than
those
critical
in
flowing
streams:

(
3
)
reservoirs
and
lakes
collect
phosphates
from
influent
streams
and
s
t
o
r
e
a
portion
of
them
within
consolidated
sediments,
thus
s
e
r
v
i
n
g
a
s
a
phosphate
s
i
n
k
;
and
(
4
)
phosphorus
c
o
n
c
e
n
t
r
a
t
i
o
n
s
c
r
i
t
i
c
a
l
t
o
noxious
p
l
a
n
t
growth
v
a
r
y
and
n
u
i
s
a
n
c
e
growths
may
r
e
s
u
l
t
from
a
p
a
r
t
i
c
u
l
a
r
c
o
n
c
e
n
t
r
a
t
i
o
n
of
phosphate
i
n
one
g
e
o
g
r
a
p
h
i
c
a
l
a
r
e
a
b
u
t
n
o
t
i
n
a
n
o
t
h
e
r
.
The
amount
o
r
p
e
r
c
e
n
t
a
g
e
of
i
n
f
l
o
w
i
n
g
n
u
t
r
i
e
n
t
s
t
h
a
t
may
be
r
e
t
a
i
n
e
d
by
a
l
a
k
e
o
r
reservoir
is
v
a
r
i
a
b
l
e
and
w
i
l
l
depend
upon:
(
1)
t
h
e
n
u
t
r
i
e
n
t
loading
t
o
t
h
e
l
a
k
e
o
r
r
e
s
e
v
o
i
r
;
(
2
)
t
h
e
volume
of
t
h
e
euphotic
zone;
(
3
)
t
h
e
extent
of
b
i
o
l
o
g
i
c
a
l
activities;
(
4
)
t
h
e
detention
time
w
i
t
h
i
n
a
l
a
k
e
b
a
s
i
n
o
r
t
h
e
t
i
m
e
a
v
a
i
l
a
b
l
e
f
o
r
b
i
o
l
o
g
i
c
a
l
a
c
t
i
v
i
t
i
e
s
;

t
h
e
penstock
from
t
h
e
r
e
s
e
r
v
o
i
r
.
and
(
5
)
t
h
e
l
e
v
e
l
of
d
i
s
c
h
a
r
g
e
from
t
h
e
l
a
k
e
o
r
of
Once
n
u
t
r
i
e
n
t
s
are
combined
w
i
t
h
i
n
t
h
e
a
q
u
a
t
i
c
ecosystem,

t
h
e
i
r
removal
is
tedious
and
expensive.
Phosphates
a
r
e
used
by
a
l
g
a
e
and
h
i
g
h
e
r
a
q
u
a
t
i
c
p
l
a
n
t
s
and
may
b
e
s
t
o
r
e
d
i
n
excess
of
use
within
t
h
e
p
l
a
n
t
c
e
l
l
.
With
decomposition
of
t
h
e
p
l
a
n
t
c
e
l
l
,

some
phosphorus
may
be
released
immediately
through
b
a
c
t
e
r
i
a
l
a
c
t
i
o
n
 
o
r
r
e
c
y
c
l
i
n
g
w
i
t
h
i
n
t
h
e
b
i
o
t
i
c
community,
w
h
i
l
e
t
h
e
remainder
may
be
deposited
with
sediments.
Much
of
t
h
e
material
t
h
a
t
combines
with
t
h
e
c
o
n
s
o
l
i
d
a
t
e
d
sediments
w
i
t
h
i
n
t
h
e
l
a
k
e
bottom
is
bound
permanently
and
w
i
l
l
n
o
t
be
r
e
c
y
c
l
e
d
i
n
t
o
t
h
e
s
y
s
tern.
0
RATIONAm
:

Elemental
Phosphorus
I
s
o
m
(
1960)
reported
an
LC50
of
0.105
mg/
L
a
t
4
8
h
o
u
r
s
and
0.025
mg/
L
a
t
1
6
0
h
o
u
r
s
f
o
r
b
l
u
e
g
i
l
l
s
u
n
f
i
s
h
,
__
Lepomis
_­­­

y.
~

macrochirus,
exposed
t
o
ye1
low
phosphorus
i
n
d
i
s
t
i
l
l
e
d
water
a
t
~
~

­
~~

__

­

26
OC
and
pH
7.
The
125­
and
195­
hour
LC50'
s
of
yellow
phosphorus
to
Atlantic
cod,
Gadus
morhua,
and
Atlantic
salmon,
Salmo
salar,

smolts
in
continuous­
exposure
experiments
were
1.89
and
0.79
ug/
L,
respectively
(
Fletcher
and
Hoyle,
1972).
No
evidence
of
an
incipient
lethal
level
was
observed
since
the
lowest
concentration
of
p4
tested
was
0.79
ug/
L.
Salmon
that
were
exposed
to
elemental
phosphorus
concentrations
of
40
ug/
L
or
less
developed
a
distinct
external
red
color
and
showed
signs
of
extensive
hemolysis.
The
predominant
features
of
p4
poisoning
in
salmon
were
external
redness,
hemolysis,
and
reduced
hematocrits.

Following
the
opening
of
an
elemental
phosphorus
production
plant
in
Long
Harbour,
Placentia
Bay,
Newfoundland,
divers
observed
dead
fish
upon
the
bottom
throughout
the
Harbour
(
Peer,

1972).
Mortalities
were
confined
to
a
water
depth
of
less
than
18
meters.
There
was
visual
evidence
of
selective
mortality
among
benthos.
Live
mussels
were
found
within
300
meters
of
the
effluent
pipe,
while
all
scallops
within
this
area
were
dead.

Fish
will
concentrate
elemental
phosphorus
from
water
containing
as
little
as
1
ug/
L
(
Idler,
1969).
In
one
set
of
experiments,
a
cod
swimming
in
water
containing
1
ug/
L
elemental
phosphorus
for
18
hours
concentrated
phosphorus
to
5
0
ug/
kg
in
muscle,
150
ug/
kg
in
fatty
tissue,
and
25,000
ug/
kg
in
the
liver
(
Idler,
1969;
Jangaard,
1970).
The
experimental
findings
showed
that
phosphorus
is
quite
stable
in
the
fish
tissues.
The
criterion
of
0.10
ug/
L
elemental
phosphorus
for
marine
or
estuarine
waters
is
.1
of
demonstrated
lethal
levels
to
important
marine
organisms
and
of
levels
that
have
been
found
to
result
in
significant
bioaccumulation.
0
Phosphate
Phosphorus
Although
a
t
o
t
a
l
phosphorus
c
r
i
t
e
r
i
o
n
t
o
c
o
n
t
r
o
l
nuisance
a
q
u
a
t
i
c
growths
i
s
n
o
t
p
r
e
s
e
n
t
e
d
,
it
is
b
e
l
i
e
v
e
d
t
h
a
t
t
h
e
following
r
a
t
i
o
n
a
l
e
t
o
support
such
a
c
r
i
t
e
r
i
o
n
,
which
c
u
r
r
e
n
t
l
y
is
evolving,
should
be
considered.

T
o
t
a
l
phosphate
phosphorus
c
o
n
c
e
n
t
r
a
t
i
o
n
s
i
n
excess
of
1
0
0
ug/
L
P
may
interfere
with
coagulation
i
n
water
treatment
plants.

When
such
concentrations
exceed
25
ug/
L
a
t
the
t
i
m
e
of
the
spring
turnover
on
a
volume­
weighted
b
a
s
i
s
i
n
l
a
k
e
s
o
r
r
e
s
e
r
v
o
i
r
s
,
they
may
occasionally
s
t
i
m
u
l
a
t
e
excessive
o
r
nuisance
growths
of
algae
and
o
t
h
e
r
a
q
u
a
t
i
c
p
l
a
n
t
s
.
A
l
g
a
l
growths
i
n
p
a
r
t
'
u
n
d
e
s
i
r
a
b
l
e
tastes
and
odors
t
o
water,
i
n
t
e
r
f
e
r
e
with
water
treatment,
become
a
e
s
t
h
e
t
i
c
a
l
l
y
u
n
p
l
e
a
s
a
n
t
,
and
a
l
t
e
r
t
h
e
chemistry
of
t
h
e
water
s
u
p
p
l
y
.
They
c
o
n
t
r
i
b
u
t
e
t
o
t
h
e
phenomenon
o
f
c
u
l
t
u
r
a
l
0
eutrophication.

To
p
r
e
v
e
n
t
t
h
e
development
o
f
b
i
o
l
o
g
i
c
a
l
n
u
i
s
a
n
c
e
s
and
t
o
c
o
n
t
r
o
l
accelerated
o
r
c
u
l
t
u
r
a
l
eutrophication,
t
o
t
a
l
phosphates
as
phosphorus
(
P)
should
not
exceed
50
ug/
L
i
n
any
stream
a
t
t
h
e
p
o
i
n
t
where
it
e
n
t
e
r
s
any
l
a
k
e
o
r
r
e
s
e
r
v
o
i
r
,
n
o
r
2
5
ug/
L
w
i
t
h
i
n
t
h
e
l
a
k
e
o
r
r
e
s
e
r
v
o
i
r
.
A
desired
g
o
a
l
f
o
r
t
h
e
p
r
e
v
e
n
t
i
o
n
of
p
l
a
n
t
n
u
i
s
a
n
c
e
s
i
n
streams
o
r
o
t
h
e
r
f
l
o
w
i
n
g
w
a
t
e
r
s
n
o
t
discharging
d
i
r
e
c
t
l
y
t
o
l
a
k
e
s
o
r
impoundments
is
100
ug/
L
t
o
t
a
l
P
(
Mackenthun,
1973)
Most
r
e
l
a
t
i
v
e
l
y
uncontaminated
l
a
k
e
d
i
s
t
r
i
c
t
s
axe
known
t
o
have
s
u
r
f
a
c
e
w
a
t
e
r
s
t
h
a
t
c
o
n
t
a
i
n
from
10
t
o
30
ug/
L
t
o
t
a
l
phosphorus
as
P
(
Hutchinson,
1957).

T
h
e
m
a
j
o
r
i
t
y
of
t
h
e
N
a
t
i
o
n
'
s
e
u
t
r
o
p
h
i
c
a
t
i
o
n
problems
a
r
e
associated
with
l
a
k
e
s
o
r
r
e
s
e
r
v
o
i
r
s
and
c
u
r
r
e
n
t
l
y
there
are
more
0
,
..
..
data
t
o
support
t
h
e
establishment
of
a
l
i
m
i
t
i
n
g
phosphorus
l
e
v
e
l
i
n
t
h
o
s
e
waters
t
h
a
n
i
n
streams
o
r
r
i
v
e
r
s
t
h
a
t
do
n
o
t
d
i
r
e
c
t
l
y
impact
such
water.
There
a
r
e
n
a
t
u
r
a
l
c
o
n
d
i
t
i
o
n
s
,
a
l
s
o
,
t
h
a
t
would
d
i
c
t
a
t
e
t
h
e
c
o
n
s
i
d
e
r
a
t
i
o
n
of
e
i
t
h
e
r
a
more
o
r
l
e
s
s
s
t
r
i
n
g
e
n
t
phosphorus
l
e
v
e
l
.
Eutrophication
problems
may
occur
i
n
waters
where
t
h
e
phosphorus
c
o
n
c
e
n
t
r
a
t
i
o
n
is
less
than
t
h
a
t
i
n
d
i
c
a
t
e
d
above
and,
o
b
v
i
o
u
s
l
y
,
such
waters
would
need
more
s
t
r
i
n
g
e
n
t
n
u
t
r
i
e
n
t
l
i
m
i
t
s
.
L
i
k
e
w
i
s
e
,
there
a
r
e
t
h
o
s
e
waters
w
i
t
h
i
n
the
Nation
where
phosphorus
is
not
now
a
l
i
m
i
t
i
n
g
n
u
t
r
i
e
n
t
and
where
t
h
e
need
f
o
r
phosphorus
l
i
m
i
t
s
i
s
s
u
b
s
t
a
n
t
i
a
l
l
y
diminished.
Such
conditions
are
described
i
n
the
l
a
s
t
paragraph
of
t
h
i
s
r
a
t
i
o
n
a
l
e
.

There
a
r
e
two
basic
needs
i
n
e
s
t
a
b
l
i
s
h
i
n
g
a
phosphorus
c
r
i
t
e
r
i
o
n
f
o
r
flowing
waters:
ode
is
t
o
c
o
n
t
r
o
l
the
development
of
p
l
a
n
t
n
u
i
s
a
n
c
e
s
w
i
t
h
i
n
t
h
e
f
l
o
w
i
n
g
water
and,
i
n
t
u
r
n
,
t
o
c
o
n
t
r
o
l
and
prevent
animal
p
e
s
t
s
t
h
a
t
may
become
associated
w
i
t
h
such
p
l
a
n
t
s
;
t
h
e
o
t
h
e
r
is
t
o
p
r
o
t
e
c
t
t
h
e
downstream
r
e
c
e
i
v
i
n
g
waterway,
r
e
g
a
r
d
l
e
s
s
o
f
its
proximity
i
n
l
i
n
e
a
r
distance.
It
is
e
v
i
d
e
n
t
t
h
a
t
a
p
o
r
t
i
o
n
o
f
t
h
a
t
phosphorus
t
h
a
t
e
n
t
e
r
s
a
s
t
r
e
a
m
o
r
other
flowing
waterway
e
v
e
n
t
u
a
l
l
y
w
i
l
l
reach
a
r
e
c
e
i
v
i
n
g
lake
o
r
e
s
t
u
a
r
y
e
i
t
h
e
r
a
s
a
component
of
t
h
e
f
l
u
i
d
m
a
s
s
,
as
bed
l
o
a
d
sediments
t
h
a
t
are
carried
downstream,
o
r
a
s
f
l
o
a
t
i
n
g
o
r
g
a
n
i
c
materials
t
h
a
t
may
d
r
i
f
t
j
u
s
t
above
the
stream's
bed
o
r
f
l
o
a
t
on
its
water's
s
u
r
f
a
c
e
.
Superimposed
on
t
h
e
l
o
a
d
i
n
g
from
t
h
e
i
n
f
l
o
w
i
n
g
waterway,
a
l
a
k
e
o
r
e
s
t
u
a
r
y
may
r
e
c
e
i
v
e
a
d
d
i
t
i
o
n
a
l
p
h
o
s
p
h
o
r
u
s
a
s
f
a
l
l
o
u
t
from
t
h
e
a
i
r
s
h
e
d
o
r
a
s
a
d
i
r
e
c
t
introduction
from
s
h
o
r
e
l
i
n
e
areas.
A
n
o
t
h
e
r
method
t
o
c
o
n
t
r
o
l
t
h
e
i
n
f
l
o
w
o
f
n
u
t
r
i
e
n
t
s
,

p
a
r
t
i
c
u
l
a
r
l
y
phosphates,
i
n
t
o
a
l
a
k
e
is
t
h
a
t
of
p
r
e
s
c
r
i
b
i
n
g
a
n
annual
l
o
a
d
i
n
g
t
o
t
h
e
r
e
c
e
i
v
i
n
g
water.
Vollenweider
(
1973)

suggests
t
o
t
a
l
phosphorus
(
P)
loadings
i
n
grams
per
square
meter
of
s
u
r
f
a
c
e
a
r
e
a
p
e
r
y
e
a
r
t
h
a
t
w
i
l
l
be
a
c
r
i
t
i
c
a
l
l
e
v
e
l
f
o
r
e
u
t
r
o
p
h
i
c
c
o
n
d
i
t
i
o
n
s
w
i
t
h
i
n
t
h
e
r
e
c
e
i
v
i
n
g
waterway
f
o
r
a
p
a
r
t
i
c
u
l
a
r
water
volume
where
t
h
e
mean
d
e
p
t
h
of
t
h
e
l
a
k
e
i
n
meters
is
d
i
v
i
d
e
d
by
t
h
e
h
y
d
r
a
u
l
i
c
d
e
t
e
n
t
i
o
n
t
i
m
e
i
n
y
e
a
r
s
.

Vollenweider's
data
suggest
a
range
of
loading
v
a
l
u
e
s
t
h
a
t
should
r
e
s
u
l
t
i
n
o
l
i
g
o
t
r
o
p
h
i
c
l
a
k
e
water
q
u
a
l
i
t
y
.
0
Eutrophic
or
C
r
i
t
i
c
a
l
Oligotrophic
o
r
Mean
Depth/
Hydraulic
Permissible
Detention
Time
Loading
Loading
(
meters/
year)
(
grams/
meter2/
year)
(
grams/
meter2/
year)

0.5
1.0
2.5
5.0
7.5
10.0
25.0
50.0
75.0
100.0
0.07
0.10
0.16
0.22
0.27
0.32
0.50
0.71
0.87
1­
00
0­
14
0.20
0.32
0.45
0.55
0.63
1.00
1.41
1.73
2.00
There
may
be
waterways
wherein
higher
concentrations
o
r
loadings
of
t
o
t
a
l
phosphorus
do
not
produce
eutrophy,
as
w
e
l
l
a
s
those
waterways
wherein
lower
concentrations
o
r
loadings
of
t
o
t
a
l
0
L­
7
phosphorus
may
be
associated
with
populations
of
nuisance
organisms.
Waters
now
containing
less
than
the
specified
amounts
of
phosphorus
should
not
be
degraded
by
the
introduction
of
additional
phosphates.

It
should
be
recognized
that
a
number
of
specific
exceptions
can
occur
to
reduce
the
threat
of
phosphorus
as
a
contributor
to
lake
eutrophy:
1.
Naturally
occurring
phenomena
may
limit
the
development
of
plant
nuisances.
2
.
Technological
or
cost­

effective
limitations
may
help
control
introduced
pollutants.
3.

Waters
may
be
highly
laden
with
natural
silts
or
colors
which
reduce
the
penetration
of
sunlight
needed
 or
plant
photosynthesis.
4.
Some
waters
morphometric
features
of
steep
banks,
great
depth,
and
substantial
flows
contribute
to
a
history
of
no
plant
problems.
Wateks
may
be
managed
primarily
for
waterfowl
or
other
wildlife.
7
.
In
some
waters
nutrient
other
than
phosphorus
is
limiting
to
plant
growth:
the
level
and
nature
of
such
limiting
nutrient
would.
not
be
expected
to
increase
to
an
extent
that
would
influence
eutrophication.
6.
In
some
waters
phosphorus
control
cannot
be
sufficiently
effective
under
present
technology
to
make
phosphorus
the
limiting
nutrient.
5.

No
national
criterion
is
presented
for
phosphate
phosphorus
for
the
control
of
eutrophication.

(
QUALITY
CRITERIA
FOR
WATER,
JULY
1976)
PB­
263943
SEE
APPENDIX
C
FOR
METHODOLOGY
PHTHALATE
ESTERS
CRITERIA:

Aquatic
Life
The
available
data
for
phthalate
esters
indicate
that
acute
and
chronic
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
940
and
3
ug/
L,
respectively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.

The
available
data
for
phthalate
esters
indicate
that
acute
toxicity
to
saltwater
aquatic
life
occurs
at
concentrations
as
low
as
2,944
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
phthalate
esters
to
sensitive
saltwater
aquatic
life
but
toxicity
to
one
species
of
algae
occurs
at
concentrations
as
low
as
3.4
ug/
L.
,

0
Human
Health
For
the
protection
of
human
health
from
the
toxic
properties
of
dimethyl
phthalate
ingested
through
water
and
contaminated
aquatic
organisms,
the
ambient
water
criterion
is
determined
to
be
313
mg/
L.

For
the
protection
of
human
health
from
the
toxic
properties
of
dimethyl
phthalate
ingested
through
contaminated
aquatic
organisms
alone,
the
ambient
water
criterion
is
determined
to
be
2.9
g
/
l
.
For
the
protection
of
human
health
from
the
toxic
properties
of
diethyl
phthalate
ingested
through
water
and
contaminated
aquatic
organisms,
the
ambient
water
criterion
is
determined
to
be
350
mg/
L.

For
the
protection
of
human
health
from
the
toxic
properties
of
diethyl
phthalate
ingested
through
contaminated
aquatic
organisms
alone,
the
ambient
water
criterion
is
determined
to
be
1.8
g
/
l
.

For
the
protection
of
human
health
from
the
toxic
properties
of
dibutyl
phthalate
ingested
through
water
and
contaminated
aquatic
organisms,
the
ambient
water
criterion
is
determined
to
be
34
mg/
L.

For
the
protection
of
human
health
from
the
toxic
properties
of
dibutyl
phthalate
ingested
through
contaminated
aquatic
organisms
alone,
the
ambient
water
criterion
is
determined
to
be
154
mg/
L.

For
the
protection
of
human
health
from
the
toxic
properties
of
di­
2­
ethylhexyl
phthalate
ingested
through
water
and
contaminated
aquatic
organisms,
the
ambient
water
criterion
is
determined
to
be
15
mg/
L.

For
the
protection
of
human
health
from
the
toxic
properties
of
di­
2­
ethylhexyl
phthalate
ingested
through
contaminated
aquatic
organisms
alone,
the
ambient
water
criterion
is
determined
to
be
50
mg/
L.

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
POLYCHLORINATED
BIPHENYLS
Aquatic
Life
For
polychlorinated
biphenyls
the
criterion
to
protect
freshwater
aquatic
life
as
derived
using
the
Guidelines
is
0.014
ug/
L
as
a
24­
hour
average.
The
concentration
of
0.014
ug/
L
is
probably
too
high
because
it
is
based
on
bioconcentration
factors
measured
in
laboratory
studies,
but
field
studies
apparently
produce
factors
at
least
10
times
higher
for
fishes.
The
available
data
indicate
that
acute
toxicity
to
freshwater
aquatic
life
probably
will
occur
only
at
concentrations
above
2.0
ug/
L
and
that
the
24­
hour
average
should
provide
adequate
protection
against
acute
toxicity.

For
polychlorinated
biphenyls
the
criterion
to
protect
0
saltwater
aquatic
life
as
derived
using
the
Guidelines
is
0.030
ug/
L
as
a
24­
hour
average.
The
concentration
of
0.030
ug/
L
is
probably
too
high
because
it
is
based
on
bioconcentration
factors
measured
in
laboratory
studies,
but
field
studies
apparently
produce
factors
at
least
10
times
higher
for
fishes.
The
available
data
indicate
that
acute
toxicity
to
saltwater
aquatic
life
probably
will
only
occur
at
concentrations
above
10
ug/
L
and
that
the
24­
hour
average
criterion
should
provide
adequate
protection
against
acute
toxicity.

Human
Health
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
polychlorinated
biphenyls
through
ingestion
of
contaminated
water
and
contaminated
aquatic
0
.
i"
i
organisms,
the
ambient
water
concentration
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.
However,

zero
level
may
not
be
attainable
at
the
present
time.
Therefore,

the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
and
The
corresponding
recommended
criteria
are
0.79
ng/
L,
0.079
ng/
L,

and
0.0079
ng/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
0.79
ng/
L,
0.079
ng/
L,
and
0.0079
ng/
L,

respectively.

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
POLYNUCLEAR
AROMATIC
HYDROCARBONS
CRITERIA
:

Aquatic
Life
The
limited
freshwater
data
base
available
for
polynuclear
aromatic
hydrocarbons,
mostly
from
short­
term
bioconcentration
studies
with
two
compounds,
does
not
permit
a
statement
concerning
acute
or
chronic
toxicity.

The
available
data
for
polynuclear
aromatic
hydrocarbons
indicate
that
acute
toxicity
to
saltwater
aquatic
life
occurs
at
concentrations
as
low
as
300
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
polynuclear
aromatic
hydrocarbons
to
sensitive
saltwater
aquatic
life.

Human
Health
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
polynuclear
aromatic
hydrocarbons
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentration
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
lom6,
and
The
corresponding
recommended
criteria
are
28.0
ng/
L,
2.8
ng/
L,
and
0
.
2
8
ng/
L,
respectively.

If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excludinq
consumption
of
water,
the
levels
are
311.0
nq/
L.
31.1
ng/
L,
and
3.11
ng/
L,
r
e
s
p
e
c
t
i
v
e
l
y
.

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
SELENIUM
CRITERIA:

Aquatic
Life
For
total
recoverable
inorganic
selenite
the
criterion
to
protect
freshwater
aquatic
life
as
derived
using
the
Guidelines
is
35
ug/
L
as
a
24­
hour
average,
and
the
concentration
should
not
exceed
260
ug/
L
at
any
time.

For
total
recoverable
inorganic
selenite
the
criterion
to
protect
saltwater
aquatic
life
as
derived
using
the
Guidelines
is
54
ug/
L
as
a
24­
hour
average,
and
the
concentration
should
not
exceed
410
ug/
L
at
any
time.

The
available
data
for
inorganic
selenate
indicate
that
acute
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
760
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
inorganic
selenate
to
sensitive
freshwater
aquatic
life.

No
data
are
available
concerning
the
toxicity
of
inorganic
selenate
to
saltwater
aquatic
life.

Human
Health
The
ambient
water
quality
criterion
for
selenium
is
recommended
to
be
identical
to
the
existing
water
standard
which
is
10
ug/
L.
Analysis
of
the
toxic
effects
data
resulted
in
a
calculated
level
which
is
protective
of
human
health
against
the
ingestion
of
contaminated
water
and
contaminated
aquatic
(­
organisms.
The
calculated
value
is
comparable
to
the
present
0
standard.
For
t
h
i
s
reason
a
s
e
l
e
c
t
i
v
e
c
r
i
t
e
r
i
o
n
based
on
exposure
solely
from
consumption
o
f
6.5
grams
of
a
q
u
a
t
i
c
organisms
was
not
derived.

(
45
F.
R.
79318,
November
2
8
,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
SILVER
A
q
u
a
t
i
c
­
L
i
f
e
For
f
r
e
s
h
w
a
t
e
r
a
q
u
a
t
i
c
l
i
f
e
t
h
e
c
o
n
c
e
n
t
r
a
t
i
o
n
(
i
n
ug/
L)
of
t
o
t
a
l
recoverable
s
i
l
v
e
r
should
not
exceed
t
h
e
numerical
v
a
l
u
e
g
i
v
e
n
by
e
(
l
.
7
2
[
l
n
(
h
a
r
d
n
e
s
s
)
3
­
6.52)
a
t
any
t
i
m
e
.
For
example,
a
t
h
a
r
d
n
e
s
s
e
s
of
5
0
,
1
0
0
,
and
2
0
0
mg/
L
as
CaC03,
t
h
e
concentration
of
t
o
t
a
l
recoverable
s
i
l
v
e
r
should
not
exceed
1
.
2
,
4.1,
and
13
ug/
L,
r
e
s
p
e
c
t
i
v
e
l
y
,
a
t
any
t
i
m
e
.
The
a
v
a
i
l
a
b
l
e
d
a
t
a
i
n
d
i
c
a
t
e
t
h
a
t
c
h
r
o
n
i
c
t
o
x
i
c
i
t
y
t
o
freshwater
a
q
u
a
t
i
c
l
i
f
e
may
occur
a
t
concentrations
as
low
as
0.12
ug/
L.

For
saltwater
a
q
u
a
t
i
c
l
i
f
e
t
h
e
c
o
n
c
e
n
t
r
a
t
i
o
n
of
t
o
t
a
l
recoverable
s
i
l
v
e
r
s
h
o
u
l
d
n
o
t
exceed
2.3
ug/
L
a
t
any
t
i
m
e
.
N
o
data
are
a
v
a
i
l
a
b
l
e
concerning
the
chronic
t
o
x
i
c
i
t
y
of
s
i
l
v
e
r
t
o
s
e
n
s
i
t
i
v
e
saltwater
aquatic
l
i
f
e
.

Human
Health
The
a
m
b
i
e
n
t
w
a
t
e
r
q
u
a
l
i
t
y
c
r
i
t
e
r
i
o
n
f
o
r
s
i
l
v
e
r
i
s
recommended
t
o
be
i
d
e
n
t
i
c
a
l
t
o
t
h
e
e
x
i
s
t
i
n
g
water
s
t
a
n
d
a
r
d
,

which
is
50
ug/
L.
Analysis
of
t
h
e
t
o
x
i
c
e
f
f
e
c
t
s
d
a
t
a
r
e
s
u
l
t
e
d
i
n
a
c
a
l
c
u
l
a
t
e
d
l
e
v
e
l
which
is
p
r
o
t
e
c
t
i
v
e
of
human
h
e
a
l
t
h
a
g
a
i
n
s
t
t
h
e
i
n
g
e
s
t
i
o
n
of
contaminated
w
a
t
e
r
and
contaminated
a
q
u
a
t
i
c
organisms.
T
h
e
c
a
l
c
u
l
a
t
e
d
v
a
l
u
e
i
s
comparable
t
o
t
h
e
p
r
e
s
e
n
t
s
t
a
n
d
a
r
d
.
For
t
h
i
s
reason
a
s
e
l
e
c
t
i
v
e
c
r
i
t
e
r
i
o
n
based
on
e
x
p
o
s
u
r
e
s
o
l
e
l
y
from
c
o
n
s
u
m
p
t
i
o
n
o
f
6
.
5
grams
o
f
a
q
u
a
t
i
c
organisms
w
a
s
not
derived.

a
­
'.
'
(
45
F.
R.
79318,
November
2
8
,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
SOLIDS
(
DISSOLVED)
AND
­
SALINlTY
CRITERION:

250
mg/
L
f
o
r
chlorides
and
s
u
l
f
a
t
e
s
i
n
domestic
water
s
u
p
p
l
i
e
s
(
welfare).

INTRODUCTION:

D
i
s
s
o
l
v
e
d
s
o
l
i
d
s
and
t
o
t
a
l
d
i
s
s
o
l
v
e
d
s
o
l
i
d
s
a
r
e
terms
g
e
n
e
r
a
l
l
y
a
s
s
o
c
i
a
t
e
d
w
i
t
h
f
r
e
s
h
w
a
t
e
r
systems
and
c
o
n
s
i
s
t
of
inorganic
s
a
l
t
s
,
small
amounts
of
organic
matter,
and
dissolved
materials
(
Sawyer,
1960).
The
e
q
u
i
v
a
l
e
n
t
terminology
i
n
Standard
Methods
is
f
i
l
t
r
a
b
l
e
residue
(
Standard
Methods,
1971).
S
a
l
i
n
i
t
y
is
an
oceanographic
term,
and
although
n
o
t
p
r
e
c
i
s
e
l
y
e
q
u
i
v
a
l
e
n
t
t
o
t
h
e
t
o
t
a
l
dissolved
s
a
l
t
content
it
is
related
t
o
it
(
Capurro,

1970).
For
most
purposes,
t
h
e
terms
t
o
t
a
l
dissolved
s
a
l
t
content
and
s
a
l
i
n
i
t
y
a
r
e
e
q
u
i
v
a
l
e
n
t
.
T
h
e
p
r
i
n
c
i
p
a
l
i
n
o
r
g
a
n
i
c
a
n
i
o
n
s
dissolved
i
n
water
include
t
h
e
carbonates,
c
h
l
o
r
i
d
e
s
,
s
u
l
f
a
t
e
s
,

and
n
i
t
r
a
t
e
s
(
p
r
i
n
c
i
p
a
l
l
y
i
n
ground
waters)
;
t
h
e
p
r
i
n
c
i
p
a
l
cations
a
r
e
sodium,
potassium,
calcium,
and
magnesium.

RATIONALE
:

Excess
dissolved
s
o
l
i
d
s
are
objectionable
i
n
drinking
water
because
of
p
o
s
s
i
b
l
e
physiological
e
f
f
e
c
t
s
,
unpalatable
mineral
t
a
s
t
e
s
,
and
h
i
g
h
e
r
c
o
s
t
s
because
of
c
o
r
r
o
s
i
o
n
o
r
t
h
e
n
e
c
e
s
s
i
t
y
f
o
r
a
d
d
i
t
i
o
n
a
l
treatment.

The
p
h
y
s
i
o
l
o
g
i
c
a
l
effects
d
i
r
e
c
t
l
y
r
e
l
a
t
e
d
t
o
d
i
s
s
o
l
v
e
d
s
o
l
i
d
s
include
l
a
x
a
t
i
v
e
e
f
f
e
c
t
s
p
r
i
n
c
i
p
a
l
l
y
from
sodium
s
u
l
f
a
t
e
and
magnesium
s
u
l
f
a
t
e
and
t
h
e
adverse
effect
of
sodium
on
c
e
r
t
a
i
n
p
a
t
i
e
n
t
s
a
f
f
l
i
c
t
e
d
w
i
t
h
cardiac
disease
and
women
w
i
t
h
toxemia
a
s
s
o
c
i
a
t
e
d
w
i
t
h
pregnancy.
One
s
t
u
d
y
w
a
s
made
u
s
i
n
g
d
a
t
a
,.
~

e
collected
from
wells
in
North
Dakota.
Results
from
a
questionnaire
showed
that
with
wells
in
which
sulfates
ranged
from
1,000
to
1,500
mg/
L,
62
percent
of
the
respondents
indicated
laxative
effects
associated
with
consumption
of
the
water.

However,
nearly
one­
quarter
of
the
respondents
to
the
questionnaire
reported
difficulties
when
concentrations
ranged
from
200
to
500
mg/
L
(
Moore,
1952).
To
protect
transients
to
an
area,
a
sulfate
level
of
250
mg/
L
should
afford
reasonable
protection
from
laxative
effects.

As
indicated,
sodium
frequently
is
the
principal
component
of
dissolved
solids.
Persons
on
restricted
sodium
diets
may
have
an
intake
restricted
from
500
to
1,000
mg/
day
(
Nat.
Res.
Coun.,

1954).
That
portion
ingested
in
water
must
be
compensated
by
reduced
levels
in
food
ingested
so
that
the
total
does
not
exceed
the
allowable
intake.
Using
certain
assumptions
of
water
intake
(
e.
g.,
2
liters
of
water
consumed
per
day)
and
sodium
content
of
food,
it
has
been
calculated
that
for
very
restricted
sodium
diets,
20
mg/
L
in
water
would
be
the
maximum,
while
for
moderately
restricted
diets,
270
mg/
L
would
be
maximum.
Specific
sodium
levels
for
entire
water
supplies
have
not
been
recommended
but
various
restricted
sodium
intakes
are
recommended
because:

(
1)
the
general
population
is
not
adversely
affected
by
sodium,

but
various
restricted
sodium
intakes
are
recommended
by
physicians
for
a
significant
portion
of
the
population,
and
(
2)

270
mg/
L
of
sodium
is
representative
of
mineralized
waters
that
may
be
aesthetically
unacceptable,
but
many
domestic
water
supplies
exceed
this
level.
Treatment
for
removal
bf
sodium
in
water
supplies
is
costly
(
NAS,
1974).

A
study
based
on
consumer
surveys
in
29
California
water
systems
was
made
to
measure
the
taste
threshold
of
dissolved
salts
in
water
(
Bruvold
et
al.,
1969).
Systems
were
selected
to
eliminate
possible
interferences
from
other
taste­
causing
substances
than
dissolved
salts.
The
study
revealed
that
consumers
rated
waters
with
319
to
397
mg/
L
dissolved
solids
as
llexcellentlt
while
those
with
1,283
to
1,333
mg/
L
dissolved
solids
were
"
unacceptable"
depending
on
the
rating
system
used.
A
*
lgoodll
rating
was
registered
for
dissolved
solids
less
than
658
to
755
mg/
L.
The
1962
PHS
Drinking
Water
Standards
recommended
a
maximum
dissolved
solids
concentration
of
500
mg/
L
unless
more
suitable
supplies
were
unavailable.

Specific
constituents
included
in
the
dissolved
solids
in
water
may
cause
mineral
tastes
at
lower
concentrations
than
other
constituents.
Chloride
ions
have
frequently
been
cited
as
having
a
low
taste
threshold
in
water.
Data
from
Ricter
and
MacLean
(
1939)
on
a
taste
panel
of
53
adults
indicated
that
61
mg/
L
NaCl
was
the
median
level
for
detecting
a
difference
from
distilled
water.
At
a
median
concentration
of
395
mg/
L
chloride
a
salty
taste
was
distinguishable,
although
the
range
was
from
120
to
1,215
mg/
L.
Lockhart,
@
t
al.
1955)
evaluated
the
effect
of
chlorides
on
water
used
for
brewing
coffee
indicated
threshold
concentrations
for
chloride
ranging
from
210
mg/
L
to
310
mg/
L
depending
on
the
associated
cation.
These
data
indicate
that
a
level
of
250
mg/
L
chlorides
is
a
reasonable
maximum
level
to
0
\.
p
protect
consumers
of
drinking
water.
The
causation
of
corrosion
and
encrustation
of
metallic
surfaces
by
water
containing
dissolved
solids
is
well
known.
In
water
distribution
systems
corrosion
is
controlled
by
insulating
dissimilar
metal
connections
by
nonmetallic
materials,
using
pH
control
and
corrosion
inhibitors,
or
some
form
of
galvanic
or
impressed
electrical
current
systems
(
Lehmann,
1964).
In
household
systems
water
piping,
wastewater
piping,
water
heaters,

faucets,
toilet
flushing
mechanisms,
garbage
grinders
and
both
clothes
and
dishwashing
machines
incure
damage.

By
using
water
with
1,150
mg/
L
dissolved
solids
as
compared
with
250
mg/
L,
service
life
was
reduced
from
70
percent
for
toilet
flushing
mechanisms
to
30
percent
for
washing
equipment.

Such
increased
corrosion
was
calculated
in
1968
to
cost
the
consumer
an
additional
$
0.50
per
1,000
gallons
used.

All
species
of
fish
and
other
aquatic
life
must
tolerate
a
range
of
dissolved
solids
concentrations
in
order
to
survive
under
natural
conditions.
Based
on
studies
in
Saskatchewan
it
has
been
indicated
that
several
common
freshwater
species
survived
10,000
mg/
L
dissolved
solids,
that
whitefish
and
pike­

perch
survived
15,000
mg/
L,
but
only
the
stickleback
survived
20,000
mg/
L
dissolved
solids.
It
was
concluded
that
lakes
with
dissolved
solids
in
excess
of
15,000
mg/
L
were
unsuitable
for
most
freshwater
fishes
(
Rawson
and
Moore,
1944).
The
1968
NTAC
Report
also
recommended
less
than
that
caused
chloride.
maintaining
osmotic
pressure
levels
of
by
a
15,000
mg/
L
solution
of
sodium
Marine
f
i
s
h
e
s
a
l
s
o
e
x
h
i
b
i
t
v
a
r
i
a
n
c
e
i
n
a
b
i
l
i
t
y
t
o
t
o
l
e
r
a
t
e
s
a
l
i
n
i
t
y
changes.
However,
f
i
s
h
k
i
l
l
s
i
n
Laguna
Madre
o
f
f
t
h
e
Texas
coast
h
a
v
e
o
c
c
u
r
r
e
d
w
i
t
h
s
a
l
i
n
i
t
i
e
s
i
n
t
h
e
range
o
f
75
t
o
1
0
0
o/
oo.
Such
c
o
n
c
e
n
t
r
a
t
e
d
seawater
is
caused
by
e
v
a
p
o
r
a
t
i
o
n
and
l
a
c
k
of
exchange
w
i
t
h
t
h
e
Gulf
of
Mexico
(
R
o
u
n
s
a
f
e
l
l
and
Everhart,
1953).

E
s
t
u
a
r
i
n
e
s
p
e
c
i
e
s
of
f
i
s
h
a
r
e
t
o
l
e
r
a
n
t
of
s
a
l
i
n
i
t
y
changes
ranging
from
f
r
e
s
h
t
o
brackish
t
o
seawater.
Anadromous
s
p
e
c
i
e
s
likewise
are
t
o
l
e
r
a
n
t
although
evidence
i
n
d
i
c
a
t
e
s
t
h
a
t
t
h
e
young
cannot
t
o
l
e
r
a
t
e
t
h
e
change
u
n
t
i
l
t
h
e
normal
t
i
m
e
o
f
m
i
g
r
a
t
i
o
n
(
Rounsefell
and
Everhart,
1953).
Other
a
q
u
a
t
i
c
s
p
e
c
i
e
s
are
more
dependent
on
s
a
l
i
n
i
t
y
f
o
r
p
r
o
t
e
c
t
i
o
n
from
p
r
e
d
a
t
o
r
s
o
r
r
e
q
u
i
r
e
c
e
r
t
a
i
n
minimal
s
a
l
i
n
i
t
i
e
s
f
o
r
successful
hatching
of
eggs.
The
o
y
s
t
e
r
d
r
i
l
l
c
a
n
n
o
t
t
o
l
e
r
a
t
e
s
a
l
i
n
i
t
i
e
s
less
t
h
a
n
12.5
o/
oo,

Therefore,
e
s
t
u
a
r
i
n
e
segments
containing
s
a
l
i
n
i
t
i
e
s
below
about
1
2
.
5
o/
oo
p
r
o
d
u
c
e
most
of
t
h
e
seed
o
y
s
t
e
r
s
f
o
r
p
l
a
n
t
i
n
g
(
Rounsefell
and
Everhart,
1953).
Based
on
similar
examples,
the
1968
NTAC
Report
recommended
t
h
a
t
t
o
p
r
o
t
e
c
t
f
i
s
h
and
o
t
h
e
r
marine
animals
no
changes
i
n
hydrography
o
r
stream
flow
should
be
allowed
t
h
a
t
permanently
change
i
s
o
h
a
l
i
n
e
p
a
t
t
e
r
n
s
i
n
t
h
e
estuary
by
more
than
10
percent
from
n
a
t
u
r
a
l
v
a
r
i
a
t
i
o
n
.

Many
of
t
h
e
recommended
game
b
i
r
d
l
e
v
e
l
s
f
o
r
d
i
s
s
o
l
v
e
d
s
o
l
i
d
s
concentrations
i
n
drinking
water
have
been
e
x
t
r
a
p
o
l
a
t
e
d
from
d
a
t
a
c
o
l
l
e
c
t
e
d
on
domestic
species
such
a
s
chickens.
However,
young
d
u
c
k
l
i
n
g
s
were
r
e
p
o
r
t
e
d
poisoned
i
n
S
u
i
s
a
n
Marsh
by
s
a
l
t
when
maximum
summer
s
a
l
i
n
i
t
i
e
s
v
a
r
i
e
d
from
0.55
t
o
1
.
7
4
o/
oo
w
i
t
h
~

means
a
s
high
a
s
1.26
o/
oo
(
G
r
i
f
f
i
t
h
,
1963).
I
n
d
i
r
e
c
t
e
f
f
e
c
t
s
of
excess
dissolved
s
o
l
i
d
s
are
primarily
the
elimination
of
desirable
food
p
l
a
n
t
s
and
other
habitat­
forming
p
l
a
n
t
s
.
Rapid
s
a
l
i
n
i
t
y
changes
cause
plasmolysis
of
t
e
n
d
e
r
l
e
a
v
e
s
and
stems
because
of
changes
i
n
osmotic
p
r
e
s
s
u
r
e
.
The
1968
NTAC
Report
recommended
t
h
e
following
l
i
m
i
t
s
i
n
s
a
l
i
n
i
t
y
v
a
r
i
a
t
i
o
n
from
natural
t
o
protect
w
i
l
d
l
i
f
e
habitats:

Natural
S
a
l
i
n
i
t
y
Variation
Permitted
(
O
/
O
O
)
(
o/
oo)

0
t
o
3.5
1
3.5
t
o
13.5
2
13.5
t
o
35
4
A
g
r
i
c
u
l
t
u
r
a
l
uses
of
water
a
r
e
a
l
s
o
l
i
m
i
t
e
d
by
e
x
c
e
s
s
i
v
e
d
i
s
s
o
l
v
e
d
s
o
l
i
d
s
concentrations.
S
t
u
d
i
e
s
have
i
n
d
i
c
a
t
e
d
t
h
a
t
chickens,
swine,
c
a
t
t
l
e
,
and
sheep
can
survive
on
s
a
l
i
n
e
waters
up
t
o
1
5
,
0
0
0
mg/
L
of
s
a
l
t
s
of
sodium
and
calcium
combined
with
b
i
c
a
r
b
o
n
a
t
e
s
,
c
h
l
o
r
i
d
e
s
,
and
s
u
l
f
a
t
e
s
b
u
t
o
n
l
y
1
0
,
0
0
0
mg/
L
of
corresponding
s
a
l
t
s
of
potassium
and
magnesium.
The
approximate
l
i
m
i
t
f
o
r
highly
a
l
k
a
l
i
n
e
waters
containing
sodium
and
calcium
carbonates
is
5,000
mg/
L
(
NTAC,
1968).

I
r
r
i
g
a
t
i
o
n
use
of
water
depends
n
o
t
o
n
l
y
upon
t
h
e
osmotic
effect
of
dissolved
s
o
l
i
d
s
,
but
a
l
s
o
on
t
h
e
r
a
t
i
o
of
t
h
e
various
c
a
t
i
o
n
s
p
r
e
s
e
n
t
.
I
n
a
r
i
d
and
s
e
m
i
a
r
i
d
a
r
e
a
s
g
e
n
e
r
a
l
c
l
a
s
s
i
f
i
c
a
t
i
o
n
of
s
a
l
i
n
i
t
y
hazards
has
been
prepared
(
NTAC,
1968)

(
see
Table
9
)
.

Table
9.­
Dissolved
Solids
Hazard
f
o
r
I
r
r
i
g
a
t
i
o
n
Water
(
mg/
L).

water
from
which
no
detri­
mental
effects
w
i
l
l
usually
be
noticed­­­­­­­­­­­­­­­­­­­­­
500
­
.
.
..
,,
water
which
can
have
detri­
mental
e
f
f
e
c
t
s
on
s
e
n
s
i
­
500­
1,000
t
i
v
e
crops­­­­­­­­­­­­­­­­­­­­­

water
t
h
a
t
may
have
adverse
effects
on
many
c
r
o
p
s
and
r
e
q
u
i
r
e
s
c
a
r
e
f
u
l
manage­
merit
practices­­­­­­­­­­­­­­­­­
1,000­
2,000
water
t
h
a
t
can
be
used
f
o
r
t
o
l
e
r
a
n
t
p
l
a
n
t
s
on
perme­
a
b
l
e
s
o
i
l
s
w
i
t
h
c
a
r
e
f
u
l
management
practices­­­­­­­­­­­
2,000­
5,000
The
amount
of
sodium
and
t
h
e
percentage
of
sodium
i
n
r
e
l
a
t
i
o
n
t
o
o
t
h
e
r
c
a
t
i
o
n
s
a
r
e
o
f
t
e
n
i
m
p
o
r
t
a
n
t
.
I
n
a
d
d
i
t
i
o
n
t
o
c
o
n
t
r
i
b
u
t
i
n
g
t
o
osmotic
p
r
e
s
s
u
r
e
,
sodium
is
t
o
x
i
c
t
o
c
e
r
t
a
i
n
p
l
a
n
t
s
,
e
s
p
e
c
i
a
l
l
y
f
r
u
i
t
s
,
and
f
r
e
q
u
e
n
t
l
y
causes
problems
i
n
s
o
i
l
s
t
r
u
c
t
u
r
e
,
i
n
f
i
l
t
r
a
t
i
o
n
,
and
p
e
r
m
e
a
b
i
l
i
t
y
rates
(
A
g
r
i
c
u
l
t
u
r
e
Handbook
#
60,
1954).
A
high
percentage
of
exchangeable
sodium
i
n
s
o
i
l
s
c
o
n
t
a
i
n
i
n
g
c
l
a
y
s
t
h
a
t
s
w
e
l
l
when
w
e
t
c
a
n
c
a
u
s
e
a
s
o
i
l
c
o
n
d
i
t
i
o
n
adverse
t
o
water
movement
and
p
l
a
n
t
growth.
T
h
e
exchangeable­
sodium
percentage
(
ESP)
*
is
an
index
of
t
h
e
sodium
s
t
a
t
u
s
o
f
s
o
i
l
s
.
An
ESP
o
f
1
0
t
o
15
p
e
r
c
e
n
t
is
c
o
n
s
i
d
e
r
e
d
excessive
i
f
a
h
i
g
h
p
e
r
c
e
n
t
a
g
e
o
f
s
w
e
l
l
i
n
g
c
l
a
y
m
i
n
e
r
a
l
s
is
p
r
e
s
e
n
t
(
A
g
r
i
c
u
l
t
u
r
a
l
Handbook
#
60,
1954).
0
For
s
e
n
s
i
t
i
v
e
f
r
u
i
t
s
,
t
h
e
t
o
l
e
r
a
n
c
e
f
o
r
sodium
f
o
r
i
r
r
i
g
a
t
i
o
n
water
is
f
o
r
a
sodium
a
d
s
o
r
p
t
i
o
n
r
a
t
i
o
(
SAR)**
o
f
a
b
o
u
t
4
,

whereas
for
g
e
n
e
r
a
l
c
r
o
p
s
and
forages
a
r
a
n
g
e
of
8
t
o
1
8
is
g
e
n
e
r
a
l
l
y
considered
u
s
a
b
l
e
(
NTAC,
1968).
It
is
emphasized
t
h
a
t
a
p
p
l
i
c
a
t
i
o
n
of
these
f
a
c
t
o
r
s
must
be
i
n
t
e
r
p
r
e
t
e
d
i
n
r
e
l
a
t
i
o
n
t
o
s
p
e
c
i
f
i
c
s
o
i
l
c
o
n
d
i
t
i
o
n
s
e
x
i
s
t
i
n
g
i
n
a
g
i
v
e
n
l
o
c
a
l
e
and
t
h
e
r
e
f
o
r
e
f
r
e
q
u
e
n
t
l
y
r
e
q
u
i
r
e
s
f
i
e
l
d
i
n
v
e
s
t
i
g
a
t
i
o
n
.

I
n
d
u
s
t
r
i
a
l
r
e
q
u
i
r
e
m
e
n
t
s
r
e
g
a
r
d
i
n
g
t
h
e
d
i
s
s
o
l
v
e
d
s
o
l
i
d
s
.
,­

c
o
n
t
e
n
t
o
f
raw
waters
is
q
u
i
t
e
v
a
r
i
a
b
l
e
.
T
a
b
l
e
10
i
n
d
i
c
a
t
e
s
Table
10.­
Total
Dissolved
Solids
Concentrations
of
Surface
Waters
That
Have
Been
Used
as
Sources
for
Industrial
Water
Supplies
Industry/
Use
Maximum
Concentration
(
m
g
m
Textile
150
Pulp
and
Paper
1,080
Chemical
2,500
Petroleum
3,500
Primary
Metals
.
1,500
Boiler
Make­
up
35,000
maximum
values
accepted
by
various
industries
for
process
requirements
(
NAS,
1974).
Since
water
of
almost
any
dissolved
0
solids
concentration
can
be
de­
ionized
to
meet
the
most
stringent
requirements,
the
economics
of
such
treatment
are
the
1
imiting
factor
for
industry.

*
ESP
=
100
[
a
+
b(
SAR)]
1
[
a
+
b(
SAR)]

where:
a
=
intercept
respresenting
experimental
error
(
ranges
from
­
0.06
to
0.01)

from
0.014
to
0.016)
b
=
slope
of
regression
line
(
ranges
**
SAR
=
sodium
adsorption
ratio
=
Na
­
[
0.5(
Ca
+
Mg)]""

SAR
is
expressed
as
milliequivalents
(
QUALITY
CRITERIA
FOR
WATER,
JULY
1976)
PB­
263943
SEE
APPENDIX
C
FOR
METHODOLOGY
0
SOLIDS
(
SUSPENDED,
SETTLEABLE)
­
AND
TURBIDITY
CRITERIA
Freshwater
f
i
s
h
and
o
t
h
e
r
a
q
u
a
t
i
c
l
i
f
e
:

S
e
t
t
l
e
a
b
l
e
and
suspended
s
o
l
i
d
s
should
n
o
t
reduce
the
depth
of
t
h
e
compensation
p
o
i
n
t
f
o
r
photosynthetic
a
c
t
i
v
i
t
y
by
more
t
h
a
n
1
0
p
e
r
c
e
n
t
from
t
h
e
s
e
a
s
o
n
a
l
l
y
e
s
t
a
b
l
i
s
h
e
d
norm
f
o
r
aquatic
l
i
f
e
.

INTRODUCTION:

The
term
lkuspended
and
s
e
t
t
l
e
a
b
l
e
s
o
l
i
d
s
t
1
is
d
e
s
c
r
i
p
t
i
v
e
of
t
h
e
o
r
g
a
n
i
c
and
i
n
o
r
g
a
n
i
c
p
a
r
t
i
c
u
l
a
t
e
matter
i
n
water.
The
e
q
u
i
v
a
l
e
n
t
terminology
used
f
o
r
s
o
l
i
d
s
i
n
Standard
Methods
(
APHA,

1971)
is
t
o
t
a
l
suspended
m
a
t
t
e
r
f
o
r
suspended
s
o
l
i
d
s
,
settleable
matter
f
o
r
s
e
t
t
l
e
a
b
l
e
s
o
l
i
d
s
,
v
o
l
a
t
i
l
e
suspended
matter
f
o
r
v
o
l
a
t
i
l
e
s
o
l
i
d
s
and
fixed
suspended
matter
f
o
r
fixed
suspended
s
o
l
i
d
s
.
The
term
l
t
s
o
l
i
d
s
t
l
i
s
used
i
n
t
h
i
s
d
i
s
c
u
s
s
i
o
n
because
of
its
more
common
u
s
e
i
n
t
h
e
water
p
o
l
l
u
t
i
o
n
c
o
n
t
r
o
l
l
i
t
e
r
a
t
u
r
e
.
0
RAT1
ONALE
:

Suspended
s
o
l
i
d
s
and
t
u
r
b
i
d
i
t
y
a
r
e
i
m
p
o
r
t
a
n
t
parameters
i
n
b
o
t
h
m
u
n
i
c
i
p
a
l
and
i
n
d
u
s
t
r
i
a
l
w
a
t
e
r
s
u
p
p
l
y
practices.
F
i
n
i
s
h
e
d
d
r
i
n
k
i
n
g
waters
h
a
v
e
a
maximum
l
i
m
i
t
o
f
1
t
u
r
b
i
d
i
t
y
u
n
i
t
where
t
h
e
water
e
n
t
e
r
s
t
h
e
d
i
s
t
r
i
b
u
t
i
o
n
system.
T
h
i
s
l
i
m
i
t
is
based
on
h
e
a
l
t
h
c
o
n
s
i
d
e
r
a
t
i
o
n
s
a
s
it
r
e
l
a
t
e
s
t
o
e
f
f
e
c
t
i
v
e
c
h
l
o
r
i
n
e
d
i
s
i
n
f
e
c
t
i
o
n
.
S
u
s
p
e
n
d
e
d
m
a
t
t
e
r
p
r
o
v
i
d
e
s
a
r
e
a
s
where
microorganisms
d
o
n
o
t
come
i
n
t
o
c
o
n
t
a
c
t
w
i
t
h
t
h
e
c
h
l
o
r
i
n
e
d
i
s
i
n
f
e
c
t
a
n
t
(
NAS,
1
9
7
4
)
.
The
a
b
i
l
i
t
y
o
f
common
water
t
r
e
a
t
m
e
n
t
p
r
o
c
e
s
s
e
s
(
i.
e.,
c
o
a
g
u
l
a
t
i
o
n
,
s
e
d
i
m
e
n
t
a
t
i
o
n
,
f
i
l
t
r
a
t
i
o
n
,
and
c
h
l
o
r
i
n
a
t
i
o
n
)
t
o
remove
suspended
matter
t
o
a
c
h
i
e
v
e
acceptable
f
i
n
a
l
t
u
r
b
i
d
i
t
i
e
s
is
a
f
u
n
c
t
i
o
n
o
f
t
h
e
c
o
m
p
o
s
i
t
i
o
n
of
t
h
e
m
a
t
e
r
i
a
l
a
s
w
e
l
l
as
its
concentration.
Because
o
f
t
h
e
v
a
r
i
a
b
i
l
i
t
y
0
of
such
removal
efficiency,
it
is
not
possible
to
delineate
a
general
raw
water
criterion
for
these
uses.

Turbid
water
interferes
with
recreational
use
and
aesthetic
enjoyment
of
water.
Turbid
waters
can
be
dangerous
for
swimming,

especially
if
diving
facilities
are
provided,
because
ofthe
possibility
of
unseen
submerged
hazards
and
the
difficulty
in
locating
swimmers
in
danger
of
drowning
(
NAS,
1974).
The
less
turbid
the
water
the
more
desirable
it
becomes
for
swimming
and
other
water
contact
sports.
Other
recreational
pursuits
such
as
boating
and
fishing
will
be
adequately
protected
by
suspended
solids
criteria
developed
for
protection
of
fish
and
other
aquatic
life.

Fish
and
other
aquatic
life
requirements
concerning
suspended
solids
can
be
divided
into
those
whose
effect
occurs
in
the
water
column
and
those
whose
effect
occurs
following
sedimentation
to
the
bottom
of
the
water
body.
Noted
effects
are
similar
for
both
fresh
and
marine
waters.

The
effects
of
suspended
solids
on
fish
have
been
reviewed
by
the
European
Inland
Fisheries
Advisory
Commission
(
EIFAC,
1965).

This
review
in
1965
identified
four
effects
on
the
fish
and
fish
food
populations,
namely:

(
1)
by
acting
directly
on
the
fish
swimming
in
water
in
which
solids
are
suspended,
and
either
killing
them
or
reducing
their
growth
rate,
resistance
to
disease,
etc.;

(
2)
by
preventing
the
successful
development
of
fish
eggs
and
larvae;

(
3
)
by
modifying
natural
movements
and
migrations
o
f
fish;
(
4
)
by
reducing
t
h
e
abundance
o
f
food
a
v
a
i
l
a
b
l
e
to
t
h
e
f
i
s
h
;
.
.
.

S
e
t
t
l
e
a
b
l
e
materials
which
blanket
t
h
e
bottom
of
water
bodies
damage
t
h
e
invertebrate
populations,
block
g
r
a
v
e
l
spawning
beds,

and
i
f
o
r
g
a
n
i
c
,
remove
d
i
s
s
o
l
v
e
d
oxygen
from
o
v
e
r
l
y
i
n
g
waters
(
E
l
F
A
C
,
1965;
Edberg
and
Hofsten,
1973).
I
n
a
s
t
u
d
y
downstream
from
the
discharge
of
a
rock
quarry
where
i
n
e
r
t
suspended
s
o
l
i
d
s
were
i
n
c
r
e
a
s
e
d
t
o
80
mg/
L,
t
h
e
d
e
n
s
i
t
y
of
m
a
c
r
o
i
n
v
e
r
t
e
b
r
a
t
e
s
decreased
by
60
percent
while
i
n
a
r
e
a
s
of
sediment
accumulation
b
e
n
t
h
i
c
i
n
v
e
r
t
e
b
r
a
t
e
p
o
p
u
l
a
t
i
o
n
s
a
l
s
o
decreased
by
60
p
e
r
c
e
n
t
regardless
of
t
h
e
suspended
sol
i
d
concantrations
(
Gammon,
1970).

similar
e
f
f
e
c
t
s
have
been
reported
downstream
from
an
area
which
was
i
n
t
e
n
s
i
v
e
l
y
logged.
Major
i
n
c
r
e
a
s
e
s
i
n
stream
suspended
s
o
l
i
d
s
(
25
ppm
t
u
r
b
i
d
i
t
y
upstream
v
e
r
s
u
s
390
ppm
downstream)

caused
smothering
of
bottom
i
n
v
e
r
t
e
b
r
a
t
e
s
,
reducing
organism
d
e
n
s
i
t
y
t
o
o
n
l
y
7.3
p
e
r
s
q
u
a
r
e
f
o
o
t
v
e
r
s
u
s
25.5
p
e
r
s
q
u
a
r
e
f
o
o
t
upstraam
(
Tebo,
1955).

When
s
e
t
t
l
e
a
b
l
e
s
o
l
i
d
s
b
l
o
c
k
g
r
a
v
e
l
spawning
beds
which
contain
eggs,
high
m
o
r
t
a
l
i
t
i
e
s
r
e
s
u
l
t
although
there
is
evidence
t
h
a
t
some
s
p
e
c
i
e
s
of
salmonids
w
i
l
l
n
o
t
spawn
i
n
such
areas
(
EIFAC,
1965).

I
t
has
been
p
o
s
t
u
l
a
t
e
d
t
h
a
t
silt
attached
t
o
t
h
e
eggs
prevents
s
u
f
f
i
c
i
e
n
t
exchange
of
oxygen
and
carbon
dioxide
between
t
h
e
egg
and
t
h
e
o
v
e
r
l
y
i
n
g
water.
The
important
v
a
r
i
a
b
l
e
s
are
p
a
r
t
i
c
l
e
s
i
z
e
,
stream
v
e
l
o
c
i
t
y
,
and
degree
of
turbulence
(
EIFAC,

1965).
Deposition
of
organic
materials
to
the
bottom
sediments
can
cause
imbalances
in
stream
biota
by
increasing
bottom
animal
density
principally
worm
populations,
and
diversity
is
reduced
as
pol
lution­
sensitive
forms
disappear
(
Mackenthun,
1973).
Algae
1
ikewise
flourish
in
such
nutrient­
rich
areas
although
forms
may
become
less
desirable
(
Tarzwell
and
Gaufin,
1953).

Plankton
and
inorganic
suspended
materials
reduce
light
penetration
into
the
water
body,
reducing
the
depth
of
thephotic
zone.
This
reduces
primary
production
and
decreases
fish
food.

The
NAS
commitee
in
1974
recommended
that
the
depth
of
light
penetration
not
be
reduced
by
more
than
10
percent
(
NAS,
1974).

Additionally,
the
near
surface
waters
are
heated
because
of
the
greater
heat
absorbency
of
the
particulate
material
which
tends
to
stabilize
the
water
column
and
prevents
vertical
mixing
(
NAS,

1974).
Such
mixing
reductions
decrease
the
dispersion
of
dissolved
oxygen
and
nutrients
to
lower
portions
of
the
water
body.

One
partially
offsetting
benefit
of
suspended
inorganic
material
in
water
is
the
sorption
of
organic
materials
such
as
pesticides.
Following
this
sorption
process
subsequent
sedimentation
may
remove
these
materials
from
the
water
column
into
the
sediments
(
NAS,
1974).

Identifiable
effects
of
suspended
solids
on
irrigation
use
of
water
include
the
formation
of
crusts
on
top
of
the
soil
which
inhibits
water
infiltration
and
plant
emergence,
and
impedes
soil
aeration;
the
formation
of
films
on
plant
leaves
which
blocks
sunlight
and
impedes
photosynthesis
and
which
may
reduce
the
marketability
of
some
leafy
crops
like
lettuce,
and
finally
the
adverse
effect
on
irrigation
reservoir
capacity,
delivery
canals,

and
other
distribution
equipment
(
NAS,
1974).
0
The
criterion
for
freshwater
fish
and
other
aquatic
lifeare
essentially
that
proposed
by
the
National
Academy
of
Sciences
and
the
Great
Lakes
Water
Quality
Board.

(
QUALITY
CRITERIA
FOR
WATER,
JULY
1976)
PB­
263943
SEE
APPENDIX
C
FOR
METHODOLOGY
CRITERION:
SULFIDE
=
HYDROGEN
SULFIDE
2
ug/
L'undissociated
HZS
f
o
r
f
i
s
h
and
other
aquatic
l
i
f
e
,
fresh
and
marine
water.

INTRODUCTION:

Hydrogen
s
u
l
f
i
d
e
is
a
s
o
l
u
b
l
e
,
h
i
g
h
l
y
poisonous,
gaseous
compound
h
a
v
i
n
q
t
h
e
c
h
a
r
a
c
t
e
r
i
s
t
i
c
odor
o
f
r
o
t
t
e
n
eggs.
It
is
detectable
i
n
a
i
r
by
humans
a
t
a
d
i
l
u
t
i
o
n
of
0.002
ppm.
It
w
i
l
l
d
i
s
s
o
l
v
e
i
n
w
a
t
e
r
a
t
4
,
0
0
0
mg/
L
a
t
20'
C
and
one
atmosphere
of
pressure.
Hydrogen
s
u
l
f
i
d
e
b
i
o
l
o
g
i
c
a
l
l
y
is
an
active
compound
t
h
a
t
is
found
primarily
as
an
anaerobic
degradation
product
of
both
organic
s
u
l
f
u
r
compounds
and
inorganic
s
u
l
f
a
t
e
s
.
S
u
l
f
i
d
e
s
a
r
e
c
o
n
s
t
i
t
u
e
n
t
s
of
many
i
n
d
u
s
t
r
i
a
l
wastes
s
u
c
h
a
s
t
h
o
s
e
from
t
a
n
n
e
r
i
e
s
,
paper
m
i
l
l
s
,
chemical
p
l
a
n
t
s
,
and
g
a
s
works.
T
h
e
anaerobic
decomposition
of
sewage,
sludge
beds,
algae,
and
other
n
a
t
u
r
a
l
l
y
d
e
p
o
s
i
t
e
d
o
r
g
a
n
i
c
m
a
t
e
r
i
a
l
i
s
a
major
s
o
u
r
c
e
of
hydrogen
s
u
l
f
i
d
e
.
0
When
s
o
l
u
b
l
e
s
u
l
f
i
d
e
s
are
added
t
o
water
t
h
e
y
react
w
i
t
h
hydrogen
ions
t
o
form
HS
o
r
HZS,
t
h
e
proportion
of
each
depending
on
t
h
e
pH.
The
t
o
x
i
c
i
t
y
of
s
u
l
f
i
d
e
s
derives
p
r
i
m
a
r
i
l
y
from
H2S
r
a
t
h
e
r
t
h
a
n
from
t
h
e
h
y
d
r
o
s
u
l
f
i
d
e
(
HS­)
o
r
s
u
l
f
i
d
e
(
S=)
ions'

When
hydrogen
s
u
l
f
i
d
e
d
i
s
s
o
l
v
e
s
i
n
water
it
dissociates
according
to
t
h
e
reactions:

H2S
HS­
+
H+
and
HS­
S=
+
H+

A
t
pH
9
about
99
p
e
r
c
e
n
t
of
t
h
e
s
u
l
f
i
d
e
is
i
n
t
h
e
form
of
HS­

,
a
t
pH
7
t
h
e
s
u
l
f
i
d
e
is
e
q
u
a
l
l
y
d
i
v
i
d
e
d
between
HS­
and.
H2S:
and
a
t
pH
5
about
9
9
p
e
r
c
e
n
t
of
t
h
e
s
u
l
f
i
d
e
is
p
r
e
s
e
n
t
a
s
H2S
(
NAS
0
­
_­

a
1974).
The
f
a
c
t
t
h
a
t
H2S
is
o
x
i
d
i
z
e
d
i
n
w
e
l
l­
a
e
r
a
t
e
d
water
by
n
a
t
u
r
a
l
b
i
o
l
o
g
i
c
a
l
systems
t
o
s
u
l
f
a
t
e
s
o
r
is
b
i
o
l
o
g
i
c
a
l
l
y
oxidized
t
o
elemental
s
u
l
f
u
r
has
caused
i
n
v
e
s
t
i
g
a
t
o
r
s
t
o
minimize
t
h
e
t
o
x
i
c
effects
of
H2S
on
f
i
s
h
and
o
t
h
e
r
a
q
u
a
t
i
c
l
i
f
e
.

RATIONALE:

The
degree
of
hazard
exhibited
by
s
u
l
f
i
d
e
t
o
a
q
u
a
t
i
c
animal
l
i
f
e
is
dependent
on
t
h
e
t
e
m
p
e
r
a
t
u
r
e
,
pH,
and
d
i
s
s
o
l
v
e
d
oxygen.

A
t
l
o
w
e
r
pH
v
a
l
u
e
s
a
g
r
e
a
t
e
r
p
r
o
p
o
r
t
i
o
n
is
i
n
t
h
e
form
of
t
h
e
t
o
x
i
c
u
n
d
i
s
s
o
c
i
a
t
e
d
H2S.
I
n
w
i
n
t
e
r
when
t
h
e
pH
is
n
e
u
t
r
a
l
o
r
below
o
r
when
d
i
s
s
o
l
v
e
d
oxygen
l
e
v
e
l
s
are
low
b
u
t
n
o
t
l
e
t
h
a
l
t
o
f
i
s
h
,
t
h
e
hazard
from
s
u
l
f
i
d
e
s
i
s
exacerbated.
F
i
s
h
e
x
h
i
b
i
t
a
s
t
r
o
n
g
a
v
o
i
d
a
n
c
e
r
e
a
c
t
i
o
n
t
o
s
u
l
f
i
d
e
.
Based
on
data
from
experiments
with
the
s
t
i
c
k
l
e
b
a
c
k
,
Jones
(
1964)
hypothesized
t
h
a
t
i
f
f
i
s
h
e
n
c
o
u
n
t
e
r
a
l
e
t
h
a
l
c
o
n
c
e
n
t
r
a
t
i
o
n
of
s
u
l
f
i
d
e
there
is
a
r
e
a
s
o
n
a
b
l
e
chance
t
h
e
y
w
i
l
l
be
r
e
p
e
l
l
e
d
by
it
b
e
f
o
r
e
t
h
e
y
are
harmed.
T
h
i
s
,
of
course,
assumes
t
h
a
t
an
escape
r
o
u
t
e
is
open.

Many
p
a
s
t
d
a
t
a
on
t
h
e
t
o
x
i
c
i
t
y
of
hydrogen
s
u
l
f
i
d
e
t
o
f
i
s
h
and
o
t
h
e
r
a
q
u
a
t
i
c
l
i
f
e
h
a
v
e
been
based
on
e
x
t
r
e
m
e
l
y
s
h
o
r
t
exposure
periods.
Consequently,
these
e
a
r
l
y
data
have
i
n
d
i
c
a
t
e
d
t
h
a
t
c
o
n
c
e
n
t
r
a
t
i
o
n
s
between
0.3
and
0.4
mg/
L
p
e
r
m
i
t
f
i
s
h
t
o
s
u
r
v
i
v
e
(
Van
Horn
1958,
Boon
and
F
o
l
l
i
s
1967,
Theede
e
t
a
l
.
,

1969).
Recent
:
ong­
term
d
a
t
a
,
both
i
n
f
i
e
l
d
s
i
t
u
a
t
i
o
n
s
and
under
c
o
n
t
r
o
l
l
e
d
1
a
b
o
r
a
t
o
r
.
y
c
o
n
d
i
t
i
o
n
s
,
demonstrate
hydrogen
s
u
l
f
i
d
e
t
o
x
i
c
i
t
y
a
t
lower
concentrations.

Colby
and
Smiti­
i
(
1967)
found
t
h
a
t
c
o
n
c
e
n
t
r
a
t
i
o
n
s
a
s
high
a
s
0.7
mg/
L
h
a
v
e
been
found
w
i
t
h
i
n
2
0
mm
of
t
h
e
bottom
o
f
s
l
u
d
g
e
beds,
and
t
h
e
l
e
v
e
l
s
o
f
0.1
t
o
0.02
mg/
L
w
e
r
e
common
w
i
t
h
i
n
t
h
e
first
20
mm
o
f
water
above
t
h
i
s
l
a
y
e
r
.
Walleye
(
S
t
i
z
o
s
t
e
d
i
o
n
vitreum)
eggs
h
e
l
d
i
n
t
r
a
y
s
i
n
t
h
i
s
zone
d
i
d
n
o
t
hatch.
Adelman
and
Smith
(
1970)
reported
t
h
a
t
t
h
e
h
a
t
c
h
a
b
i
l
i
t
y
of
northern
pike
(
Esox
l
u
c
i
u
s
)
eggs
was
s
u
b
s
t
a
n
t
i
a
l
l
y
reduced
a
t
2
5
ug/
L
H2S:
a
t
4
1
ug/
L
m
o
r
t
a
l
i
t
y
was
almost
complete.
Northern
p
i
k
e
f
r
y
had
96­

hour
LC50
v
a
l
u
e
s
t
h
a
t
v
a
r
i
e
d
from
17
t
o
32
ug/
L
a
t
normal
oxygen
l
e
v
e
l
s
o
f
6.0
mg/
L.
T
h
e
h
i
g
h
e
s
t
c
o
n
c
e
n
t
r
a
t
i
o
n
of
hydrogen
s
u
l
f
i
d
e
t
h
a
t
had
no
observable
effect
on
eggs
and
f
r
y
w
a
s
1
4
and
4
ug/
L,
r
e
s
p
e
c
t
i
v
e
l
y
.
S
m
i
t
h
and
Oseid
(
1
9
7
2
)
,
working
on
eggs,

fry
and
j
u
v
e
n
i
l
e
s
o
f
w
a
l
l
e
y
e
s
and
w
h
i
t
e
s
u
c
k
e
r
s
(
Catostomus
commersoni)
and
Smith
(
1971),
Safe
l
e
v
e
l
s
i
n
working
on
walleyes
and
fathead
minnows,
Pimephales
promelas,
were
found
t
o
vary
from
2.9
ug/
L
t
o
1
2
ug/
L
w
i
t
h
eggs
b
e
i
n
g
t
h
e
l
e
a
s
t
s
e
n
s
i
t
i
v
e
and
j
u
v
e
n
i
l
e
s
being
t
h
e
most
s
e
n
s
i
t
i
v
e
i
n
short­
term
tests.
I
n
96­

hour
bioassays,
fathead
minnows
and
g
o
l
d
f
i
s
h
,
Carassius
a
u
r
a
t
u
s
,

varied
g
r
e
a
t
l
y
i
n
t
o
l
e
r
a
n
c
e
t
o
hydrogen
s
u
l
f
i
d
e
w
i
t
h
changes
i
n
t
e
m
p
e
r
a
t
u
r
e
.
They
were
more
t
o
l
e
r
a
n
t
a
t
low
t
e
m
p
e
r
a
t
u
r
e
s
(
6
t
o
10,
C
)
.
H
o
l
l
a
n
d
,
e
t
a
l
.
(
1960)
r
e
p
o
r
t
e
d
t
h
a
t
1.0
mg/
L
s
u
l
f
i
d
e
caused
100
percent
m
o
r
t
a
l
i
t
y
i
n
72
hours
with
P
a
c
i
f
i
c
salmon.
0
On
t
h
e
b
a
s
i
s
of
chronic
tests
e
v
a
l
u
a
t
i
n
g
growth
and
s
u
r
v
i
v
a
l
,

t
h
e
s
a
f
e
H2S
l
e
v
e
l
f
o
r
b
l
u
e
g
i
l
l
(
Lepomis
macrochirus)
j
u
v
e
n
i
l
e
s
and
a
d
u
l
t
s
was
2
ug/
L.
Egg
deposition
i
n
b
l
u
e
g
i
l
l
s
w
a
s
reduced
a
f
t
e
r
4
6
days
i
n
1.4
ug/
L
H
2
S
(
S
m
i
t
h
and
O
s
e
i
d
,
1
9
7
4
)
.
White
sucker
eggs
were
hatched
a
t
15
ug/
L,
b
u
t
j
u
v
e
n
i
l
e
s
showed
growth
r
e
d
u
c
t
i
o
n
s
a
t
1
ug/
L.
Safe
l
e
v
e
l
f
o
r
fathead
minnows
were
between
2
and
3
ug/
L.
S
t
u
d
i
e
s
showed
t
h
a
t
s
a
f
e
l
e
v
e
l
s
f
o
r
Gammarus
Pseudolimnaeus
and
Hexagenia
­
limbata
were
2
and
15
ug/
L,

r
e
s
p
e
c
t
i
v
e
l
y
(
Oseid
and
S
m
i
t
h
,
1974a,
197413).
Some
s
p
e
c
i
e
s
0
c
typical
of
normally
stressed
habitats,
Asellus
spp.,
were
much
more
resistant
(
Oseid
and
Smith,
1974~).

Sulfide
criteria
for
domestic
or
livestock
use
have
not
been
established
because
the
unpleasant
odor
and
taste
would
preclude
such
use
at
hazardous
concentrations.

It
is
recognized
that
the
hazard
from
hydrogen
sulfide
to
aquatic
life
is
often
localized
and
transient.
Available
data
indicate
that
water
containing
concentrations
of
2.0
ug/
L
undissociated
H2S
would
not
be
hazardous
to
most
fish
and
other
aquatic
wildlife,
but
concentrations
in
excess
of
2.0
ug/
L
would
constitute
a
long­
term
hazard.

I
(
QUALITY
CRITERIA
FOR
WATER,
JULY
1976)
PB­
263943
SEE
APPENDIX
C
FOR
METHODOLOGY
TAINTING
SUBSTANCES
Materials
should
not
be
present
in
concentrations
that
individually
or
in
combination
produce
undesirable
flavors
which
are
detectable
by
organoleptic
tests
performed
on
the
edible
portions
of
aquatic
organisms.

RATIONALE
:

Fish
or
shellfish
with
abnormal
flavors,
colors,
tastes
or
odors
are
either
not
marketable
or
will
result
in
consumer
complaints
and
possible
rejection
of
the
food
source
even
though
subsequent
lots
of
organisms
may
be
acceptable.
Poor
product
quality
can
and
has
seriously
affected
or
eliminated
the
commercial
fishing
industry
in
some
areas.
Recreational
fishing
also
can
be
affected
adversely
by
off­
flavored
fish.
For
the
majority
of
sport
fishermen,
the
consumption
of
their
catch
is
part
of
their
recreation
and
off­
flavored
catches
can
result
in
diversion
of
the
sportsmen
to
other
water
bodies.
This
can
have
serious
economic
impact
on
the
established
recreation
industries
such
as
tackle
and
bait
sales
and
boat
and
cottage
rental.
0
Water
Quality
Criteria,
1972
(
NAS,
1974)
lists
a
number
of
wastewaters
and
chemical
compounds
that
have
been
found
to
lower
the
palatability
of
fish
flesh.
Implicated
wastewaters
included
those
from
2,4­
D
manufacturing
plants,
kraft
and
neutral
sulfite
pulping
processes,
municipal
wastewater
treatment
plants,
oily
wastes,
refinery
wastes,
phenolic
wastes,
and
wastes
from
slaughterhouses.
The
9
ist
of
imp1
icated
chemical
compounds
is
long:
it
includes
cresol
and
phenol
compounds,
kerosene,

naphthol,
styrene,
toluene,
and
exhaust
outboard
motor
fuel.
As
little
as
0.1
ug/
L
o­
chlorophenol
was
reported
to
cause
tainting
...
a
of
fish
flesh.

Shumway
and
a
Palensky
1973)
determined
estimated
threshold
concentrations
for
22
organic
compounds.
The
values
ranged
from
0.4
ug/
L
(
2
,
4­
dichl
orophenol
)
to
9
5,000
ug/
L
(
formaldehyde)
.
An
additional
12
compounds
were
tested,
7
of
which
were
not
found
to
impair
flavor
at
or
near
lethal
levels.

Thomas
(
1973)
reviewed
the
literature
review
on
tainting
substances
revealed
serious
problems
that
have
occurred.
Detailed
studies
and
methodology
used
to
evaluate
the
palatability
of
fishes
in
the
Ohio
River
as
affected
by
various
waste
discharges
showed
that
the
susceptibility
of
fishes
to
the
accumulation
of
tainting
substances
is
variable
and
dependent
upon
the
species,

length
of
exposure,
and
the
polJutant.
As
little
as
5
ug/
L
of
gasoline
can
impart
off­
flavors
to
fish
(
Boyle,
1967).

(
QUALITY
CRITERIA
FOR
WATER,
JULY
1976)
PB­
263943
SEE
APPENDIX
C
FOR
METHODOLOGY
TEMPERATURE
Freshwater
Aquatic
Life
For
any
time
of
year,
there
are
two
upper
limiting
temperatures
for
a
location
(
based
on
the
important
sensitive
species
found
there
at
that
time):

1.
One
limit
consists
of
a
maximum
temperature
for
short
exposures
that
is
time
dependent
and
is
given
by
the
species­

specific
equation:

Temperature
=
(
l/
b)
(
log
[
time
3
­
a)
­
2,
C
(
C,)
10
(
min)

where:
loglo
=
logarithm
to
base
10
(
common
logarithm)

a
=
intercept
on
the
"
y"
or
logarithmic
axis
of
the
l'ine
fitted
to
experimental
data
and
which
is
available
for
some
species
from
Appendix
11­
C,
National
Academy
of
Sciences
1974
document.

b
=
slope
of
the
line
fitted
to
experimental
data
and
available
for
some
species
from
Appendix
11­
C,
of
the
National
Academy
of
Sciences
document.

and
2.
The
second
value
is
a
limit
on
the
weekly
average
temperature
that:

a.
In
the
cooler
months
(
mid­
October
to
mid­
April
in
the
north
and
December
to
February
in
the
south)
will
protect
against
mortality
of
importr
to
mid­
April
in
the
north
and
December
to
February
in
the
south)
will
protect
against
mortality
of
important
species
if
the
elevated
plume
temperature
is
suddenly
dropped
to
the
ambient
temperature,
with
the
limit
being
the
b.
acclimation
temperature
minus
apt0
when
the
lower
lethal
threshold
temperature
equals
the
ambient
water
temperature
(
in
some
regions
this
limitation
may
also
be
applicable
in
summer).

or
In
the
warmer
months
(
April
through
October
in
the
north
and
March
through
November
in
the
south)
is
determined
by
adding
to
the
physiological
optimum
temperature
(
usually
for
growth)
a
factor
calculated
as
one­
third
of
the
difference
between
the
ultimate
upper
incipient
lethal
temperature
and
the
optimum
temperature
 or
the
most
sensitive
important
species
(
and
appropriate
life
state)
that
normally
is
found
at
that
location
and
time.

or
c.
During
reproductive
seasons
(
generally
April
through
June
and
September
through
October
in
the
north
and
March
through
May
and
October
through
November
in
the
south)
the
limit
is
that
temperature
that
meets
site­

specific
requirements
for
successful
migration,

spawning,
egg
incubation,
fry
rearing,
and
other
reproductive
functions
of
important
species.
These
local
requirements
should
supersede
all
other
requirements
when
they
are
applicable.

or
d.
There
is
a
site­
specific
limit
that
is
found
necessary
to
preserve
normal
species
diversity
or
prevent
appearance
of
nuisance
organisms.
Marine
Aquatic
­
Life
In
order
to
assure
protection
of
the
characteristic
indigenous
marine
community
of
a
water
body
segment
from
adverse
thermal
effects:

a.
the
maximum
acceptable
increase
in
the
weekly
average
temperature
resulting
from
artificial
sources
is
1'

C
(
1.8
F)
during
all
seasonsofthe
year,
providtng
the
summer
maxima
are
not
exceeded;

and
b.
daily
temperature
cycles
characteristic
of
the
water
body
segment
should
not
be
altered
in
either
amplitude
or
frequency.

Summer
thermal
maxima,
which
define
the
upper
thermal
limits
for
the
communities
of
the
discharge
area,
should
be
established
on
a
site­
specific
basis.
Existing
studies
suggest
the
following
regional
limits:
0
Short­
term
Maximum
Maximum
True
Daily
Mean*
Sub
tropical
regions
(
south
of
Cape
Canaveral
and
Tampa
Bay,
Florida,
and
Hawaii
32.2'
C
(
90
°
F)
29.4O
C
(
85'
F)

Cape
Hatteras,
N.
C.,
to
Cape
Canaveral,
Fla.

Long
Island
(
south
shore)
3
0
.
6
O
C
(
87O
F)
27.8O
C
(
82O
F)
32.2'
C
(
90'
F)
29.4O
C
(
85O'F)

to
Cape
Hatteras,
N.
C.

(*
True
Daily
Mean
=
average
of
24
hourly
temperature
readings.)

Baseline
thermal
conditions
should
be
measured
at
a
site
where
there
is
no
unnatural
thermal
addition
from
any
source,

which
is
in
reasonable
proximity
to
the
thermal
discharge
(
within
5
miles)
and
which
has
similar
hydrography
to
that
of
.
the
receiving
waters
at
the
discharge.

INTRODUCTION:

The
uses
of
water
by
man
in
and
out
of
its
natural
situs
in
the
environment
are
affected
by
its
temperature.
Offstream
domestic
uses
and
instream
recreation
are
both
partially
temperature
dependent.
Likewise,
the
1
ife
associated
with
the
aquatic
environment
in
any
location
has
its
species
composition
and
activity
regulated
by
water
temperature.
Since
essentially
all
of
these
organisms
are
so­
called
"
cold
blooded"
or
poikilotherms,
the
temperature
of
the
water
regulates
their
metabolism
and
ability
to
survive
and
reproduce
effectively.

Industrial
uses
for
process
water
and
for
coolingare
likewise
regulated
by
the
water's
temperature.
Temperature,
therefore,
is
an
important
physical
parameter
which
to
some
extent
regulates
many
of
the
beneficial
uses
of
water.
The
Federal
Water
Pollution
Control
Administration
in
1967
called
temperature
a
catalyst,
a
depressant,
an
activator,
a
restrictor,
a
stimulator,

a
controller,
a
killer,
one
of
the
most
important
and
most
influential
water
quality
characteristics
to
life
in
water."
0
RATIONALE
:

The
suitability
of
water
for
total
body
immersion
is
greatly
affected
by
temperature.
In
temperate
climates,
dangers
from
exposure
to
low
temperatures
is
more
prevalent
than
exposure
to
elevated
water
temperatures.
Depending
on
the
amount
of
activity
by
the
swimmer,
comfortable
temperatures
range
from
20
°
C
to
30
°
e.
Short
durations
of
lower
and
higher
temperatures
can
be
tolerated
by
most
individuals.
For
example,
for
a
30­
minute
period,
temperatures
of
10'
C
or
35O
C
can
be
tolerated
without
harm
by
most
individuals
(
NAS,
1974).

Temperature
also
affects
the
self­
purification
phenomenon
in
water
bodies
and
therefore
the
aesthetic
and
sanitary
qualities
that
exist.
Increased
temperatures
accelerate
the
biodegradation
of
organic
material
both
in
the
overlying
water
and
in
bottom
deposits
which
makes
increased
demands
on
the
dissolved
oxygen
resources
of
a
given
system.
The
typical
situation
is
exacerbated
by
the
fact
that
oxygen
becomes
less
soluble
as
water
temperature
increases.
Thus,
greater
demands
are
exerted
on
an
increasingly
scarce
resource
which
may
lead
to
total
oxygen
depletion
and
obnoxious
septic
conditions.
These
effects
have
been
described
by
Phelps
(
1944)
,
Carp
(
1963),
and
Velz
(
1970).

Indicator
enteric
bacteria,
and
presumably
enteric
pathogens,

are
likewise
affected
by
temperature.
It
has
been
shown
that
­
both
total
and
fecal
coliform
bacteria
die
away
more
rapidly
in
the
environment
with
increasing
temperatures
(
Ballentine
and
Kittrell,
1968).

Temperature
effects
have
been
shown
for
water
treatment
processes.
Lower
temperatures
reduce
the
effectiveness
of
coagulation
with
alum
and
subsequent
rapid
sand
filtration.
In
one
study,
difficulty
was
especially
pronounced
below
5O
C
(
Hannah,
et
al.,
1967).
Decreased
temperature
also
decreases
the
effectiveness
of
chlorination.
Based
on
studies
relating
chlorine
dosage
to
temperature,
and
with
a
30­
minute
contact
time,
dosages
required
for
equivalent
disinfective
effect
increased
by
as
much
as
a
factor
of
3
when
temperatures
were
decreased
from
2
0
°
C
to
loo
C
(
Reid
and
Carlson,
1974).

Increased
temperature
may
increase
the
odor
of
water
because
of
the
increased
volatility
of
odor­
causing
compounds
(
Bumson,

1938).
Odor
problems
associated
with
plankton
may
also
be
aggravated.

The
effects
o
f
temperature
on
aquatic
organisms
have
been
the
subject
of
comprehensive
literature
reviews
(
Brett,
1956;
Fry,

1967;
FWPCA,
1967;
Kine,
1970)
and
annual
literature
reviews
published
by
the
Water
Pollution
Control
Federaticn
(
Coutant,

1968,
1969,
1970,
1971;
Coutant
and
Goodyear,
1972;
Coutant
and
Pfuderer,
1973,
1974).
Only
highlights
from
the
thermal
effects
on
aquatic
life
are
presented
here.

Temperature
changes
in
water
bodies
can
alter
the
existing
aquatic
community.
The
dominance
of
various
phytoplankton
groups
in
specific
temperature
ranges
has
been
shown.
For
example,
from
20
°
C
to
25'
C,
diatoms
predominated;
green
algae
predominated
from
30'
C:
to
35O
C
and
blue­
greens
predominated
above
3.5'
C
a
i
r
n
s
,
1956).
Likewise,
changes
from
a
coldwater
f
i
s
h
e
r
y
t
o
a
warm­
water
f
i
s
h
e
r
y
can
occur
because
temperature
may
be
d
i
r
e
c
t
l
y
l
e
t
h
a
l
t
o
a
d
u
l
t
s
o
r
f
r
y
c
a
u
s
e
a
r
e
d
u
c
t
i
o
n
of
a
c
t
i
v
i
t
y
o
r
l
i
m
i
t
0
(
c
reproduction
(
B
r
e
t
t
,
1960)

Upper
and
lower
l
i
m
i
t
s
f
o
r
temperature
have
been
established
f
o
r
many
a
q
u
a
t
i
c
organisms.
C
o
n
s
i
d
e
r
a
b
l
y
more
d
a
t
a
e
x
i
s
t
f
o
r
u
p
p
e
r
a
s
opposed
t
o
l
o
w
e
r
l
i
m
i
t
s
.
T
a
b
u
l
a
t
i
o
n
s
of
l
e
t
h
a
l
temperatures
f
o
r
f
i
s
h
and
o
t
h
e
r
organisms
a
r
e
a
v
a
i
l
a
b
l
e
(
Jones,

1
9
6
4
:
FWPCA,
1
9
6
7
NAS,
1
9
7
4
)
.
F
a
c
t
o
r
s
s
u
c
h
a
s
d
i
e
t
,
a
c
t
i
v
i
t
y
,

age,
g
e
n
e
r
a
l
h
e
a
l
t
h
,
osmotic
stress,
and
even
weather
c
o
n
t
r
i
b
u
t
e
t
o
t
h
e
l
e
t
h
a
l
i
t
y
of
t
e
m
p
e
r
a
t
u
r
e
.
The
a
q
u
a
t
i
c
species,
thermal
accumulation
s
t
a
t
e
and
exposure
t
i
m
e
a
r
e
considered
t
h
e
c
r
i
t
i
c
a
l
f
a
c
t
o
r
s
(
Parker
and
Xrenkel,
1969).

The
e
f
f
e
c
t
s
o
f
s
u
b
l
e
t
h
a
l
t
e
m
p
e
r
a
t
u
r
e
s
o
n
m
e
t
a
b
o
l
i
s
m
,

r
e
s
p
i
r
a
t
i
o
n
,
behavior,
d
i
s
t
r
i
b
u
t
i
o
n
and
migration,
feeding
rate,

growth,
and
reproduction
have
been
summarized
by
Be
S
y
l
v
a
(
1969).

Another
s
t
u
d
y
h
a
s
i
l
l
u
s
t
r
a
t
e
d
t
h
a
t
i
n
s
i
d
e
t
h
e
t
o
l
e
r
a
n
c
e
zone
t
h
e
r
e
is
encompassed
a
more
r
e
s
t
r
i
c
t
i
v
e
t
e
m
p
e
r
a
t
u
r
e
r
a
n
g
e
i
n
which
normal
a
c
t
i
v
i
t
y
and
growth
o
c
c
u
r
and
y
e
t
a
n
e
v
e
n
more
r
e
s
t
r
i
c
t
i
v
e
zone
i
n
s
i
d
e
t
h
a
t
i
n
which
normal
reproduction
w
i
l
l
occur
(
B
r
e
t
t
,
1960).

D
e
S
y
l
v
a
(
1969)
has
summarized
a
v
a
i
l
a
b
l
e
data
on
t
h
e
combined
effects
of
i
n
c
r
e
a
s
e
d
t
e
m
p
e
r
a
t
u
r
e
and
t
o
x
i
c
m
a
t
e
r
i
a
l
s
o
n
f
i
s
h
i
n
d
i
c
a
t
e
t
h
a
t
t
o
x
i
c
i
t
y
g
e
n
e
r
a
l
l
y
i
n
c
r
e
a
s
e
s
w
i
t
h
i
n
c
r
e
a
s
e
d
t
e
m
p
e
r
a
t
u
r
e
and
t
h
a
t
o
r
g
a
n
i
s
m
s
s
u
b
j
e
c
t
e
d
t
o
stress
from
t
o
x
i
c
m
a
t
e
r
i
a
l
s
a
r
e
less
t
o
l
e
r
a
n
t
o
f
temperature
extremes.

The
t
o
l
e
r
a
n
c
e
o
f
o
r
g
a
n
i
s
m
s
t
o
extremes
o
f
t
e
m
p
e
r
a
t
u
r
e
is
a
f
u
n
c
t
i
o
n
o
f
t
h
e
i
r
g
e
n
e
t
i
c
a
b
i
l
i
t
y
t
o
a
d
a
p
t
t
o
t
h
e
r
m
a
l
c
h
a
n
g
e
s
0
~
within
their
characteristic
temperature
range,
the
acclimation
temperature
prior
to
exposure,
and
the
time
of
exposure
to
the
elevated
temperature
(
Coutant,
1972).
The
upper
incipient
lethal
temperature
or
the
highest
temperature
that
50
percent
of
a
sample
of
organisms
can
survive
is
determined
on
the
organism
at
the
highest
sustainable
acclimation
temperature.
The
lowest
temperature
that
50
percent
of
the
warm
acclimated
organisms
can
survive
in
is
the
ultimate
lower
incipient
lethal
temperature.

True
acclimation
to
changing
temperatures
requires
several
days
(
Brett,
1941).
The
lower
end
of
the
temperature
accommodation
range
for
aquatic
life
is
0'
C
in
fresh
water
and
somewhat
less
for
saline
waters.
However,
organisms
acclimated
to
relatively
warm
water,
when
subjected
to
reduced
temperatures
that
under
other
conditions
of
acclimation
would
not
be
detrimental,
may
suffer
a
significant
mortality
caused
by
thermal
shock
(
Coutant,

1972).

Through
the
natural
changes
in
climatic
conditions,
the
temperatures
of
water
bodies
fluctuate
daily,
as
well
as
seasonally.
These
changes
do
not
eliminate
indigenous
aquatic
populations,
but
affect
the
existing
community
structure
and
the
geographic
distribution
of
species.
Such
temperature
changes
are
necessary
to
induce
the
reproductive
cycles
of
aquatic
organisms
and
to
regulate
other
life
factors
(
Mount,
1969).

Artificially
induced
changes
such
as
the
return
of
cooling
water
or
the
release
of
cool
hypolimnetic
waters
from
impoundments
may
alter
indigenous
aquatic
ecosystems
(
Coutant,

1972).
Entrained
organisms
may
be
damaged
by
temperature
increases
across
cooling
water
condensers
if
the
increase
is
sufficiently
great
or
the
exposure
period
sufficiently
long.

Impingement
upon
condenser
screens,
chlorination
for
slime
control,
or
other
physical
insults
damage
aquatic
life
(
Raney,

1969:
Patrick,
1969
(
b)).
However,
Patrick
(
1969(
a))
has
shown
that
algae
passing
through
condensers
are
not
injured
if
the
temperature
of
the
outflowing
water
does
not
exceed
345O
C.

In
open
waters
elevated
temperatures
nay
affect
periphyton,

benthic
invertebrates,
and
fish,
in
addition
to
causing
shifts
in
algal
dominance.
Trembley
(
1960)
studies
of
the
Delaware
River
downstream
from
a
power
plant
concluded
that
the
periphyton
population
was
considerably
altered
by
the
discharge.

The
number
and
distribution
of
bottom
organisms
decrease
as
water
temperatures
increase.
The
upper
tolerance
limit
for
a
balanced
benthic
population
structure
is
approximately
32O
C,
A
0
large
number
of
these
invertebrate
species
are
able
to
tolerate
higher
temperatures
than
those
required
for
reproduction
(
FWPCA,

1967).

In
order
to
define
criteria
for
fresh
waters,
Coutant
(
1972)

cited
the
following
was
cited
as
currently
definable
requirements:

1.
Maximum
sustained
temperatures
that
are
consistent
with
maintaining
desirable
levels
of
productivity,

2.
maximum
levels
of
metabolic
acclimation
to
warm
temperatures
that
will
permit
return
to
ambient
winter
temperatures
should
artificial
sources
of
heat
cease,

3.
Time­
dependent
temperature
1
imitations
f
o
r
survival
of
brief
exposures
to
temperature
extremes,
both
upper
and
lower,
4
.
Restricted
temperature
ranges
for
various
states
of
reproduction,
including
(
for
fish)
gametogenesis,
spawning
migration,
release
of
gametes,
development
of
the
embryo,
commencement
of
independent
feeding
(
and
other
activities)
by
j
uv
eni
1
es
,
and
temper
a
tur
es
re
qu
ired
for
met
amorphos
is,
emergence,
or
other
activities
of
lower
forms,

5.
Thermal
limits
for
diverse
species
compositions
of
aquatic
communities,
particularly
where
reduction
in
diversity
creates
nuisance
growths
of
certain
organisms,
or
where
important
food
sources
(
food
chains)
are
altered,

6.
Thermal
requirements
of
downstream
aquatic
life
(
in
rivers)
where
upstream
diminution
of
a
coldwater
resource
will
adversely
affect
downstream
temperature
requirements.

The
major
portion
of
such
information
that
is
available,

however,
is
for
freshwater
fish
species
rather
than
lower
forms
of
marine
aquatic
life.

The
temperature­
time
duration
for
short­
term
exposures
such
that
50
percent
of
a
given
population
will
survive
an
extreme
temperature
frequently
is
expressed
mathematically
by
fitting
experimental
data
with
a
staright
line
on
a
semi­
logarithmic
plot
with
time
on
the
logarithmic
scale
and
temperature
on
the
linear
scale
(
see
fig.
1).
In
equation
form
this
50
percent
mortality
relationship
is:

loglo
(
time
(
minutes))
=
a
+
b
(
Temperature
(
O
C
)
)

where:
loglo=
logarithm
to
base
10
(
common
logarithm)

a
=
intercept
on
the
"
y
"
or
logarithmic
axis
of
the
line
fitted
to
experimental
data
and
which
is
available
for
some
species
from
Appendix
11­
C,
of
the
National
Academy
of
Sciences
document.

b
=
slope
of
the
line
fitted
to
experimental
data
and
which
is
available
for
some
species
from
Appendix
11­
C,
of
the
National
Academy
of
Sciences
document.

To
provide
a
safety
factor
so
that
none
or
only
a
few
organisms
will
perish,
it
has
been
found
experimentally
that
a
criterion
of
2O
C
below
maximum
temperature
is
usually
sufficient
(
Black,
1953).
To
provide
safety
for
all
the
organisms,
the
temperature
causing
a
median
mortality
for
5
0
percent
of
the
population
would
be
calculated
and
reduced
by
2'

C
in
the
case
of
an
elevated
temperature.
Available
scientific
information
includes
upper
and
lower
incipient
lethal
temperatures,

coefficients
I1at1
and
llbll
for
the
thermal
resistance
equation,
and
information
of
size,
life
stage,
and
geographic
source
of
the
particular
test
species
(
Appendix
11­
C,
NAS,
1974).

Maximum
temperatures
for
an
extensive
exposure
(
e.
g.,
more
than
1
week)
must
be
divided
into
those
for
warmer
periods
and
winter.
Other
than
for
reproduction,
the
most
temperature­

sensitive
life
function
appears
to
be
growth
(
Coutant,
1972).

Coutant
(
1972)
has
suggested
that
a
satisfactory
estimate
of
a
limiting
maximum
weekly
mean
temperature
may
be
an
average
of
the
optimum
temperature
for
growth
and
the
temperature
 or
zero
net
growth.

Because
of
the
difficulty
in
determining
the
temperature
of
zero
net
growth,
essentially
the
same
temperature
can
be
derived
by
adding
to
the
optimum
essentially
to
temperature
(
for
growth
or
other
physiological
functions)
a
factor
calculated
as
one­

third
of
the
difference
between
the
ultimate
upper
incipient
lethal
temperature
and
the
optimum
temperature
(
NAS,
1974).
In
equation
form:

Maximum
weekly
(
ultimate
upper
optimum)
average
=
optimum
+
1/
3
(
incipient
lethal
­
temperature)
temperature
temperature
(
temperature)

Since
temperature
tolerance
varies
with
various
states
of
development
of
a
particular
species,
the
criterion
f
o
r
a
­
p
a
r
t
i
c
u
l
a
r
l
o
c
a
t
i
o
n
would
be
c
a
l
c
u
l
a
t
e
d
f
o
r
t
h
e
most
important
l
i
f
e
form
l
i
k
e
l
y
t
o
be
p
r
e
s
e
n
t
d
u
r
i
n
g
a
p
a
r
t
i
c
u
l
a
r
month.
One
c
a
v
e
a
t
i
n
using
t
h
e
maximum
weekly
mean
temperature
is
t
h
a
t
t
h
e
l
i
m
i
t
f
o
r
s
h
o
r
t­
t
e
r
m
exposure
must
n
o
t
be
exceeded.
Example
c
a
l
c
u
l
a
t
i
o
n
s
f
o
r
predicting
t
h
e
summer
maximum
temperatures
f
o
r
short­
term
s
u
r
v
i
v
a
l
and
f
o
r
extensive
exposure
f
o
r
various
f
i
s
h
s
p
e
c
i
e
s
a
r
e
p
r
e
s
e
n
t
e
d
i
n
T
a
b
l
e
11.
These
c
a
l
c
u
l
a
t
i
o
n
s
u
s
e
t
h
e
above
e
q
u
a
t
i
o
n
s
and
d
a
t
a
from
EPA's
Environmental
Research
Laboxatory
i
n
Duluth.

The
w
i
n
t
e
r
maximum
t
e
m
p
e
r
a
t
u
r
e
must
n
o
t
exceed
t
h
e
ambient
water
t
e
m
p
e
r
a
t
u
r
e
by
more
t
h
a
n
t
h
e
amount
o
f
change
a
specimen
acclimated
t
o
t
h
e
plume
temperature
can
t
o
l
e
r
a
t
e
.
Such
a
change
c
o
u
l
d
o
c
c
u
r
by
a
c
e
s
s
a
t
i
o
n
of
t
h
e
s
o
u
r
c
e
of
h
e
a
t
o
r
by
t
h
e
specimen
being
d
r
i
v
e
n
from
a
n
a
r
e
a
by
a
d
d
i
t
i
o
n
o
f
b
i
o
c
i
d
e
s
o
r
o
t
h
e
r
f
a
c
t
o
r
s
.
However,
there
are
inadequate
d
a
t
a
t
o
estimate
a
s
a
f
e
t
y
f
a
c
t
o
r
f
o
r
t
h
e
Isno
stress"
l
e
v
e
l
from
c
o
l
d
shocks
(
NAS,

1974).
F
i
g
u
r
e
2
was
developed
from
a
v
a
i
l
a
b
l
e
d
a
t
a
i
n
t
h
e
l
i
t
e
r
a
t
u
r
e
(
ERL­
Duluth,
1
9
7
6
)
and
can
be
used
f
o
r
e
s
t
i
m
a
t
i
n
g
a
l
l
o
w
a
b
l
e
winter
temperature
increases.

Coutant
(
1
9
7
2
)
h
a
s
reviewed
t
h
e
e
f
f
e
c
t
s
of
t
e
m
p
e
r
a
t
u
r
e
on
a
q
u
a
t
i
c
l
i
f
e
reproduction
and
development.
Reproductive
e
v
e
n
t
s
are
noted
a
s
perhaps
t
h
e
most
t
h
e
r
m
a
l
l
y
restricted
of
a
l
l
l
i
f
e
p
h
a
s
e
s
assuming
o
t
h
e
r
f
a
c
t
o
r
s
a
r
e
a
t
o
r
n
e
a
r
optimum
l
e
v
e
l
s
.

N
a
t
u
r
a
l
s
h
o
r
t­
t
e
r
m
t
e
m
p
e
r
a
t
u
r
e
f
l
u
c
t
u
a
t
i
o
n
s
a
p
p
e
a
r
t
o
c
a
u
s
e
reduced
reproduction
of
f
i
s
h
and
invertebrates.
TABLE
11.­
Example
Calculated
Values
for
Maxima
for
Survival
for
Juveniles
and
Adults
During
the
Summer
(
Centigrade
and
Fahrenheit).

Species
Growtha
Maxima
Maximum
Weekly
Average
Temperatures
for
Growth
and
Short­
Term
b
Atlantic
salmon
Bigmouth
buffalo
Black
crappie
Bluegill
Brook
trout
Carp
Channel
catfish
Coho
salmon
Emerald
shiner
Freshwater
drum
Lake
herring
(
Cisco)
Largemouth
bass
Northern
pike
Rainbow
trout
Sauger
Smallmouth
bass
Smallmouth
buffalo
Sockeye
salmon
Striped
bass
Threadfin
shad
White
bass
White
crappie
White
sucker
Yellow
perch
20
(
68)

27
(
81)
32
(
90)
19
(
66)

32
(
90)
18
(
64)
30
(
86)

17
(
63)
32
(
90)
28
(
82)
19
(
66)
25
(
77)
29
(
84)

18
(
64)

28
(
82)
28
(
82)
29
(
84)
23
(
73)

35
(
95)
24
(
75)

35
(
95)
24
(
75)

25
(
77)

30
(
86)
24
(
75)
34
(
93)

22
(
72)

a
­
Calculated
according
to
the
equation
(
using
optimum
temperature
for
growth)

maximum
weekly
average
temperature
for
growth
=
optimum
temperature
+
1/
3
(
ultimate
incipient
lethal
temperature­

optimum
temperature.

b
­
Based
on
temperature
(
OC)
=
l
/
b
(
log"
time(
min.)
­
a)

2O
C,
acclimation
at
the
maximum
weekly
average
temperature
 or
summer
growth,
and
data
in
Appendix
11­
C
of
Water
Quality
Criteria,
published
by
National
Academy
of
Sciences.

c
­
Based
on
data
for
larvae
(
ERL­
Duluth,
1976).
0
­.
,
There
are
indadequate
data
available
quantitating
the
most
temperature­
sensitive
life
stages
among
various
aquatic
species.

Uniform
elevation
of
temperature
a
few
degrees
but
still
within
the
spawning
range
may
lead
to
advanced
spawning
for
spring
spawning
species
and
delays
for
fall
spawners.
Such
changes
may
not
be
detrimental
unless
asynchrony
occurs
between
newly
hatched
juveniles
and
their
normal
food
source.
Such
asynchrony
may
be
most
pronounced
among
anadromous
species
or
other
migrants
who
pass
from
the
warmed
area
to
a
normally
chilled,
unproductive
area.
Reported
temperature
data
on
maximum
temperatures
for
spawning
and
embryo
survival
have
been
summarized
in
Table
12
(
from
ERL­
Duluth
1976).

Although
the
limiting
effects
of
thermal
addition
to
estuarine
and
marine
waters
are
not
as
conspicuous
in
the
fall,

winter,
and
spring
as
during
the
summer
season
of
maximum
heat
stress,
nonetheless
crucial
thermal
limitations
do
exist.
Hence,

it
is
important
that
the
thermal
additions
to
the
receiving
waters
be
minimized
during
all
seasons
of
the
year.
Size
of
harvestable
stocks
of
commercial
fish
and
shellfish,
particularly
near
geographic
limits
of
the
fishery,
appear
to
be
markedly
influenced
by
slight
changes
in
the
long­
term
temperature
regime
(
Dow,
1973).

Jefferies
and
Johnson
(
1974)
studied
the
relationship
between
temperature
and
annual
variation
in
7­
year
catch
data
for
winter
flounder,
Pseudopleuronectes
_­__­_­­_­
I
americanus
in
Narragansett
Bay,

Rhode
Island,
revealed
that
a
78
percent
decrease
in
annual
catch
correlated
closely
with
a
0.5OC
increase
in
the
average
temperature
over
the
30­
month
period
between
spawning
and
recruitment
into
the
fishery.
Sissenwine's
1974
model
predicts
a
68
percent
reduction
of
recruitment
in
ye1
Powtail
flounder,

Limanda
­­­
ferrugiia,
with
a
l0C
long­
term
elevation
in
southern
New
England
waters.
TABLE
12.

Summary
of
Reported
Values
for
Maxima
for
Embryo
Survival
During
the
Spawning
Season
(
Centigrade
and
Fahrenheit)
Maximum
Weekly
Average
Temperature
for
Spawning
and
Short­
Term
Species
Atlantic
Salmon
Bigmouth
Buffalo
Black
Crappie
Bluegill
Brook
Trout
carp
Channel
Catfish
Coho
Salmon
Emerald
Shiner
Freshwater
Drum
Lake
Herring
(
Cisco)
Largemouth
Bass
Northern
Pike
Rainbow
Trout
Sauger
Smallmouth
Bass
Smallmouth
Buffalo
Sockeye
Salmon
Striped
Bass
Threadfin
Shad
White
Bass
White
Crappie
White
Sucker
Yellow
Perch
Spawning,
Embryo
Survivalb
5
17
25
9
21
27
10
24
21
3
21
11
9
17
17
10
18
18
17
18
10
12
I
10
(
41)
7
77
1
34
70)
33
(
63)
27
48)
13
81)
29
50)
13
75)
28
70)
26
37)
8
70)
27
52)
19
48)
13
50)
21
63
1
21
13
63)
50)
64
1
24
64
1
34
63)
26
23
20
64)

20
50)
54)

a
­
the
optimum
or
mean
of
the
range
of
spawning
temperatures
reported
for
the
species
(
ERL­
Duluth,
1976).

b
­
the
upper
temperature
for
successful
incubation
and
hatching
reported
for
the
species
(
ERL­
Duluth,
1976)
­

c
­
upper
temperature
for
spawning.
Community
balance
can
be
influenced
strongly
by
such
temperature­
dependent
factors
as
rates
of
reproduction,

recruitment,
and
growth
of
each
component
population.
A
few
degrees
elevation
in
average
monthly
temperature
can
appreciably
alter
a
community
through
changes
in
interspecies
relationships.

A
50
percent
reduction
in
the
softshell
clam
fishery
in
Maine
by
the
green
crab,
Carcinus
maenus,
illustrates
how
an
increase
in
winter
temperatures
can
establish
new
predator­
prey
relationships.
Over
a
period
of
4
years,
there
was
a
natural
amelioration
of
temperature
and
the
monthly
mean
for
the
coldest
month
of
each
year
did
not
fall
below
2OC.
This
apparently
precluded
appreciable
ice
formation
and
winter
cold
kill
of
the
green
crab
and
permitted
a
major
expansion
of
its
population,

with
increased
predation
of
the
softshell
clam
resulting
(
Glude,

1954:
Welch,
1968).

Temperature
is
a
primary
factor
controlling
reproduction
and
can
influence
many
events
of
the
reproductive
cycle
from
gametogenesis
to
spawning.
Among
marine
invertebrates,

initiation
of
reproduction
(
gametogenesis)
is
often
triggered
during
late
winter
by
attainment
of
a
minimum
environmental
threshold
temperature.
In
some
species,
availability
of
adequate
food
is
also
a
requisite
(
Pearse,
1970;
Sastry,
1975:
devlaming,

1971).
Elevated
temperature
can
limit
gametogenesis
by
preventing
accumulation
of
nutrients
in
the
gonads.
This
problem
could
be
acute
during
the
winter
if
food
availability
and
feeding
activity
is
reduced.
Most
marine
organisms
spawn
during
the
spring
and
summer;
gametogenesis
is
usually
initiated
during
the
0
previous
fall.
It
should
also
be
noted
that
some
species
spawn
only
during
the
fall
(
herrinhg)
,
while
others
during
the
winter
and
very
early
spring.
At
the
higher
latitudes,
winter
breeders
include
such
estuarine
community
dominants
as
acorn
barnacles,

Balanus
balanus
and
B.
balanoides,
the
edible
blue
mussel
Mytilus
_
­­
­

­­­­
I
edulis
sea
urchin,
Strongylocentrotus
drobachiensis,
sculpin,

and
the
winter
flounder,
Pseudopleuronectes
­
americanus.
The
two
boreal
barnacles
require
temperatures
below
10
°
C
before
egg
production
will
be
initiated
(
Crisp,
1957).
It
is
clear
that
adaptations
for
reproduction
exist
which
are
dependent
on
temperature
conditions
close
to
the
natural
cycle.

Juvenile
and
adult
fish
usually
thermoregulate
behaviorally
by
moving
to
water
having
temperatures
closest
to
their
thermal
preference.
This
provides
a
thermal
environment
which
approximates
the
optimal
temperature
for
many
physiological
functions,
including
growth
(
Neil1
and
Magnuson.
1974).
As
a
consequence,
fishes
usually
are
attracted
to
heated
water
during
the
fall,
winter,
and
spring.
Avoidance
will
occur
as
warmer
temperature
exceeds
the
preferendum
by
1
to
3OC
(
Coutant,
1975).

This
response
precludes
problems
of
heat
stress
for
juvenile
and
adult
fishes
during
the
summer,
but
several
potential
problems
exist
during
the
other
seasons.
The
possibility
of
cold
shock
and
death
of
plume­
entrained
fish
resulting
from
winter
plant
shutdown
is
well
recognized.
Also,
increased
incidence
of
disease
and
a
deterioration
of
physiological
condition
has
been
observed
among
plume­
entrained
fishes,
perhaps
because
of
insufficient
food
(
Massengill,
1973).
A
weight
loss
of
approximately
10
percent
for
each
lo
C
rise
in
water
temperature
has
been
observed
in
fish
when
food
is
absent.
(
Phillips
et
al.,

1960)
There
may
also
be
indirect
adverse
effects
on
the
indigenous
community
because
of
increased
predation
pressure
if
thermal
addition
leads
to
a
concentration
of
fish
which
are
dependent
on
this
community
for
their
food.

Fish
migration
is
often
linked
to
natural
environmental
temperature
cycles.
In
early
spring,
fish
employ
temperature
as
their
environmental
cue
to
migrate
northward
(
e.
g.,
menhaden,

bluefish)
or
to
move
inshore
(
winter
flounder).
Likewise,
water
temperature
strongly
influences
timing
of
spawning
runs
ofan­

adromous
fish
into
rivers
(
Leggett
and
Whitney,
1972).
In
the
autumn,
a
number
of
juvenile
marine
fishes
and
shrimp
are
dependent
on
a
drop
in
temperature
to
trigger
their
migration
from
estuarine
nursery
grounds
for
oceanic
dispersal
or
southward
migration
(
Lund
and
Maltezos,
1970;
Talbot,
1966).

Thermal
discharges
should
not
alter
diurnal
and
tidal
temperature
variations
normally
experienced
by
marine
communities.
Laboratory
studies
show
thermal
tolerance
to
be
enhanced
when
animals
are
maintained
under
a
diurnally
fluctuating
temperature
regime
rather
than
at
a
constant
temperature
(
Costlow
and
Bookhout,
1971;
Furch,
1972;
Hoss,
et
al.,).
A
daily
cyclic
regime
can
be
protective
additionally
as
it
reduces
duration
of
exposure
to
extreme
temperatures
(
Pearce,

1969;
Gonzalez,
1972).
0
Summer
thermal
maxima
should
be
established
to
protect
the
various
marine
communities
within
each
biogeographic
region.

During
the
summer,
naturally
elevated
temperatures
may
be
of
­
1
sufficent
magnitude
to
cause
death
or
emigration
(
Glynn,
1968;

Vaughn,
1961).
This
more
commonly
occurs
in
tropical
and
warm
temperate
zone
waters,
but
has
been
reported
for
enclosed
bays
and
shallow
waters
in
other
regions
as
well
(
Nichols,
1918).

Summer
heat
stress
also
can
contribute
to
increased
incidence
of
disease
or
parasitism
(
Sinderman,
1965)
:
reduce
or
block
sexual
maturation
(
Thorhaug,
et
al.,
1971:
deVlaming,
1972);
inhibit
or
block
embryonic
cleavage
of
larval
development
(
Calabrese,
1969)
;

reduce
feeding
and
growth
of
juveniles
and
adults
(
011a
and
Studholme,
1971)
:
result
in
increased
predation
(
Gonzalez,
1972);

and
reduce
productivity
of
macroalgae
and
seagrasses
(
South
and
Hill,
1970;
Zieman,
1970).
The
general
ceilings
set
forth
here
are
derived
from
studies
delineating
limiting
temperatures
for
the
more
thermally
sensitive
species
or
communities
of
a
biogeographic
region.

Thermal
effects
data
are
presently
insufficient
to
set
general
temperature
limits
for
all
coastal
biogeographic
regions.

The
data
enumerated
in
the
Appendix,
plus
any
additional
data
subsequently
generated,
should
be
used
to
develop
thermal
limits
which
specifically
consider
communities
relevant
to
given
water
bodies.

(
QUALITY
CRITERIA
FOR
WATER,
JULY
1976)
PB­
263943
SEE
APPENDIX
C
FOR
METHODOLOGY
2,3,7,8­
TETRACHMRODIBENZO­
P­
DIOXIN
CRITERIA:

Aquatic
Life
Not
enough
data
are
available
concerning
the
effects
of
2,3,7,8­
TCDD
on
aquatic
life
and
its
uses
to
allow
derivation
of
national
criteria.
The
available
information
indicates
that
acute
values
for
some
freshwater
animal
species
are
>
1.0
ug/
L;

some
chronic
values
are
<
0.01
ug/
L;
and
the
chronic
value
for
rainbow
trout
is
<
0.001
ug/
L.
Because
exposures
of
some
species
of
fishes
to
0.01
ug/
L
for
<
6
days
resulted
in
substantial
mortality
several
weeks
later,
derivation
of
aquatic
life
criteria
for
2,3,7,8­
TCDD
may
require
special
consideration.
Predicted
bioconcentration
factors
(
BCFs)
for
2,3,7,8­
TCDD
range
from
3,000
to
900,000,
but
the
available
measured
BCFs
range
from
390
to
13,000.
If
the
BCF
is
5,000,
0
concentrations
>
0.00001
ug/
L
should
result
in
concentrations
in
edible
freshwater
and
saltwater
fish
and
shellfish
that
exceed
levels
identified
in
a
U.
S.
FDA
health
advisory.
If
the
BCF
is
>
5,000
or
if
uptake
in
a
field
situation
is
greater
than
that
in
laboratory
tests,
the
value
of
0.00001
ug/
L
will
be
too
high.

Human
Health
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
2,3,7,8­
TCDD
exposure
through
ingestion
of
contaminated
water
and
contaminated
aquatic
orqanisms,
the
­
.

ambient
water
concentration
should
be
zero.
This
criterion
is
based
on
the
nonthreshold
assumption
f
o
r
2,3,7,8­
TCDD.
However,

zero
may
n
o
t
be
an
a
t
t
a
i
n
a
b
l
e
l
e
v
e
l
a
t
t
h
i
s
t
i
m
e
.

(
4
9
F
.
R
.
5831,
February
15,
1984)
SEE
APPENDIX
B
FOR
METHODOLOGY
TETRACHLOROETHYLENE
Aquatic
Life
The
available
data
for
tetrachloroethylene
indicate
that
acute
and
chronic
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
5,280
and
840
ug/
L,

respectively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.

The
available
data
for
tetrachloroethylene
indicate
that
acute
and
chronic
toxicity
to
saltwater
aquatic
life
occurs
at
concentrations
as
low
as
10,200
and
450
ug/
L,

respectively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.

Human
Health
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
tetrachloroethylene
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentrations
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
and
The
corresponding
recommended
criteria
are
8.0
ug/
L,
0.80
ug/
L,

and
0.08
ug/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
88.5
ug/
L,
8.85
ug/
L,
and
0.88
ug/
L,

respectively.

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
CRITERIA:
THALLIUM
Aquatic
Life
The
available
data
for
thallium
indicate
that
acute
and
chronic
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
1,400
and
40
ug/
L,
respectively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
Toxicity
to
one
species
of
fish
occurs
at
concentrations
as
low
as
20
ug/
L
after
2,600
hours
of
exposure.

The
available
data
for
thallium
indicate
that
acute
toxicity
to
saltwater
aquatic
life
occurs
at
concentrations
as
low
as
2,130
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
thallium
to
0
sensitive
saltwater
aquatic
life.

Human
Health
For
the
protection
of
human
health
from
the
toxic
properties
of
thallium
ingested
through
water
and
contaminated
aquatic
organisms,
the
ambient
water
criterion
is
determined
to
be
13
U9/
L.

For
the
protection
of
human
health
from
the
toxic
properties
of
thallium
ingested
through
contaminated
aquatic
organisms
alone,
the
ambient
water
criterion
is
determined
to
be
48
ug/
L.

(
45
F.
R.
79318,
November
28,
1980)
0
SEE
APPENDIX
B
FOR
METHODOLOGY
m
CRITERIA:
TOLUENE
Aquatic
Life
The
availab­
2
data
for
toluene
indicate
t
at
acute
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
17,500
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
toluene
to
sensitive
freshwater
aquatic
life.

The
available
data
for
toluene
indicate
that
acute
and
chronic
toxicity
to
saltwater
aquatic
life
occurs
at
concentrations
as
low
as
6,300
and
5,000
ug/
L,
respectively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.

0
Human
Health
For
the
protection
of
human
health
from
the
toxic
properties
of
toluene
ingested
through
water
and
contaminated
aquatic
organisms,
the
ambient
water
criterion
is
determined
to
be
14.3
w
/
L
.

For
the
protection
of
human
health
from
the
toxic
properties
of
toluene
ingested
through
contaminated
aquatic
organisms
alone,
the
ambient
water
criterion
is
determined
to
be
424
W/
L­

(
45
F.
R.
79318,
November
28,
1980)

NOTE:
The
U
.
S
.
EPA
is
currently
developing
Acceptable
Daily
Intake
(
ADI)
or
Verified
Reference
Dose
(
RfD)
values
 or
Agency­
wide
use
for
this
chemical.
The
new
value
should
be
substituted
when
it
becomes
available.
The
January,
*­
1986,
draft
Verified
Reference
Dose
document
cites
an
RfD
of
0.3
mg/
kg/
day
for
toluene.
SEE
APPENDIX
B
FOR
METHODOLOGY
0
TOXAPHENE
CRITERIA:

Aquatic
­
Life
For
toxaphene
the
criterion
to
protect
freshwater
aquatic
life
as
derived
using
the
Guidelines
is
0.013
ug/
L
as
a
24­
hour
average,
and
the
concentration
should
not
exceed
1.6
ug/
L
at
any
time.

For
saltwater
aquatic
life
the
concentration
of
toxaphene
should
not
exceed
0.070
ug/
L
at
any
time.
No
data
are
available
concerning
the
chronic
toxicity
of
toxaphene
to
sensitive
saltwater
aquatic
life.

Human
Health
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
toxaphene
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,

the
ambient
water
concentration
should
be
zero,
based
on
the
non
threshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
and
The
corresponding
recommended
criteria
are
7.1
ng/
L,
0.71
ng/
L,
and
0.07
ng/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
7.3
ng/
L,
0.73
ng/
L,
and
0.01
ng/
L,

respectively.
0
­
.
\­

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
CRITERIA:
TRICHLOROETHYLENE
Aquatic
Life
The
available
data
for
trichloroethylene
indicate
that
acute
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
45,000
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
trichloroethylene
to
sensitive
freshwater
aquatic
life
but
the
behavior
of
one
species
is
adversely
affected
at
concentrations
as
low
as
21,900
ug/
L.

The
available
data
for
trichloroethylene
indicate
that
acute
toxicity
to
saltwater
aquatic
life
occurs
at
concentrations
as
low
as
2,000
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
trichloroethylene
to
sensitive
saltwater
aquatic
life.
0
Human
Health
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
trichloroethylene
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentration
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.
However,

zero
level
may
not
be
attainable
at
the
present
time.
Therefore,

the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
lom5,
loe6,
and
lo­'*

The
corresponding
recommended
criteria
are
27
ug/
L,
2.7
ug/
L,
and
0
I
_

0.27
ug/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
807
ug/
L,
80.7
ug/
L,
and
8.07
ug/
L,

respectively.

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
CRITERIA:
VINYL
CHLORIDE
Aquatic
Life
No
freshwater
organisms
have
been
tested
with
vinyl
chloride
and
no
statement
can
be
made
concerning
acute
or
chronic
toxicity.

No
saltwater
organisms
have
been
tested
with
vinyl
chloride
and
no
statement
can
be
made
concerning
acute
or
chronic
toxicity.

Human
Health
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
vinyl
chloride
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentrations
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.
However,

zero
level
may
not
be
attainable
at
the
present
time.

Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
1
0
­
5
r
and
The
corresponding
recommended
criteria
are
20
ug/
L,
2.0
ug/
L,
and
0
.
2
ug/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
5
,
2
4
6
ug/
L,
5
2
5
ug/
L,
and
5
2
.
5
ug/
L,

respectively.

(
4
5
F.
R.
79318,
November
2
8
,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
ZINC
CRITERIA:

A
q
u
a
t
i
c
­
L
i
f
e
For
t
o
t
a
l
recoverable
zinc
t
h
e
crit
rion
t
o
p
r
o
t
e
c
t
f
r
e
hwater
aquatic
l
i
f
e
a
s
derived
using
t
h
e
Guidelines
is
47
ug/
L
as
a
24­

hour
a
v
e
r
a
g
e
and
t
h
e
c
o
n
c
e
n
t
r
a
t
i
o
n
(
i
n
ug/
L)
s
h
o
u
l
d
n
o
t
e
x
c
e
e
d
t
h
e
n
u
m
e
r
i
c
a
1
v
a
l
u
e
g
i
v
e
n
b
y
0.83
[
l
n
(
h
a
r
d
n
e
s
s
)
]+
1.95)
a
t
any
t
i
m
e
.
For
example,
a
t
e
(
h
a
r
d
n
e
s
s
e
s
of
50,
1
0
0
,
and
200
mg/
L
a
s
CaC03
t
h
e
c
o
n
c
e
n
t
r
a
t
i
o
n
of
t
o
t
a
l
r
e
c
o
v
e
r
a
b
l
e
z
i
n
c
s
h
o
u
l
d
n
o
t
exceed
180,

320,
and
570
ug/
L
a
t
any
t
i
m
e
.

For
t
o
t
a
l
recoverable
z
i
n
c
t
h
e
c
r
i
t
e
r
i
o
n
t
o
p
r
o
t
e
c
t
s
a
l
t
w
a
t
e
r
aquatic
l
i
f
e
a
s
derived
using
t
h
e
Guidelines
is
58
ug/
L
a
s
a
24­

hour
average
and
t
h
e
concentration
should
not
exceed
190
ug/
L
a
t
any
t
i
m
e
.

Human
H
e
a
l
t
h
S
u
f
f
i
c
i
e
n
t
data
a
r
e
not
a
v
a
i
l
a
b
l
e
f
o
r
zinc
t
o
d
e
r
i
v
e
a
level
which
would
p
r
o
t
e
c
t
a
g
a
i
n
s
t
t
h
e
p
o
t
e
n
t
i
a
l
t
o
x
i
c
i
t
y
of
t
h
i
s
compound.
Using
a
v
a
i
l
a
b
l
e
o
r
g
a
n
o
l
e
p
t
i
c
d
a
t
a
,
t
o
c
o
n
t
r
o
l
undesirable
t
a
s
t
e
and
odor
q
u
a
l
i
t
y
of
ambient
water
t
h
e
estimated
l
e
v
e
l
is
5
mg/
L.
It
should
be
recognized
t
h
a
t
organoleptic
data
have
l
i
m
i
t
a
t
i
o
n
s
a
s
a
b
a
s
i
s
f
o
r
e
s
t
a
b
l
i
s
h
i
n
g
a
water
q
u
a
l
i
t
y
c
r
i
t
e
r
i
a
,
and
have
no
demonstrated
r
e
l
a
t
i
o
n
s
h
i
p
to
p
o
t
e
n
t
i
a
l
adverse
human
health
effects.

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
I
~

APPENDIX
A
DERIVATION
OF
THE
1985
CRITERION
­­­
Derivation
of
numerical
national
water
quality
criteria
for
the
protection
of
aquatic
organisms
and
their
uses
is
a
complex
process
that
uses
information
from
many
areas
of
aquatic
toxicology.
After
a
decision
is
made
that
a
national
criterion
is
needed
for
a
particular
material,
all
available
information
concerning
toxicity
to,
and
bioaccumulation
by,
aquatic
organisms
is
collected,
reviewed
for
acceptability,
and
sorted.
If
enough
acceptable
data
on
acute
toxicity
to
aquatic
animals
are
available,
they
are
used
to
estimate
the
highest
1­
hour
average
concentration
that
should
not
result
in
unacceptable
effects
on
aquatic
organisms
and
their
uses.
If
justified,
this
concentration
is
made
a
function
of
a
water
quality
characteristic
such
as
pH,
salinity,
or
hardness.
Similarly,

data
on
the
chronic
toxicity
of
the
material
to
aquatic
animals
are
used
to
estimate
the
highest
4­
day
average
concentration
that
should
not
cause
unacceptable
toxicity
during
a
long­
term
exposure.
If
appropriate,
this
concentration
is
also
related
to
a
water
quality
characteristic.

Data
on
toxicity
to
aquatic
plants
are
examined
to
determine
whether
plants
are
likely
to
be
unacceptably
affected
by
concentrations
that
should
not
cause
unacceptable
effects
on
animals.
Data
on
bioaccumulation
by
aquatic
organisms
are
used
to
determine
if
residues
might
subject
edible
species
to
restrictions
by
the
U.
S.
Food
and
Drug
Administration
or
if
such
residues
might
harm
some
wildlife
consumers
of
aquatic
life.
All
other
available
data
are
examined
 or
adverse
effects
that
might
be
biologically
important.

If
a
thorough
review
of
the
pertinent
information
indicates
that
enough
acceptable
data
are
available,
numerical
national
water
quality
criteria
are
derived
for
fresh
water
or
saltwater
or
both
to
protect
aquatic
organisms
and
their
uses
from
unacceptable
effects
due
to
exposures
to
high
concentrations
for
short
periods
of
time,
lower
concentrations
for
longer
periods
of
time,
and
combinations
of
the
two.

I.
Collection
of
Data
­
­­

A.
Collect
all
available
data
on
the
material
concerning
(
a)
toxicity
to,
and
biqaccumulation
by,
aquatic
animals
and
plants,
(
b)
FDA
action
levels
[
12],
and
(
c)
chronic
feeding
studies
and
long­
term
field
studies
with
wildlife
species
that
regularly
consume
aquatic
organisms.

B.
All
data
that
are
used
should
be
available
in
typed,
dated,
and
signed
hard
copy
(
pub1
ication,

manuscript,
letter,
memorandum,
etc.)
with
enough
supporting
information
to
indicate
that
acceptable
test
procedures
were
used
and
that
the
results
are
probably
reliable.
In
some
cases
it
may
be
appropriate
to
obtain
additional
written
information
from
the
investigator,
if
possible.

Information
that
is
confidential
or
privileged
or
otherwise
not
available
for
distribution
should
not
be
used.

C.
Questionable
data,
whether
published
or
unpublished,
should
not
be
used.
For
example,
data
should
usually
be
rejected
if
they
are
from
tests
that
did
not
contain
a
control
treatment,
tests
in
which
too
many
organisms
in
the
control
treatment
died
or
showed
signs
of
stress
or
disease,
and
tests
in
which
distilled
or
deionized
water
was
used
as
the
dilution
water
without
addition
of
appropriate
salts.

D.
Data
on
technical
grade
materials
may
be
used
if
appropriate,
but
data
on
formulated
mixtures
and
emulsifiable
concentrates
of
the
material
of
concern
should
not
be
used.

E.
For
some
highly
volatile,
hydrolyzable,
or
degradable
materials
it
is
probably
appropriate
to
use
only
results
of
flow­
through
tests
in
which
the
concentrations
of
test
material
in
the
test
solutions
were
measured
often
enough
using
acceptable
analytical
methods.

F.
Data
should
be
rejected
if
they
were
obtained
using:

1.
Brine
shrimp,
because
they
usually
occur
naturally
only
in
water
with
salinity
greater
than
35
g/
kg.

2
.
Species
that
do
not
have
reproducing
wild
populations
in
North
America
(
See
Appendix
1).

3
.
Organisms
that
were
previously
exposed
to
substantial
concentrations
of
the
test
0
t.

material
or
other
contaminants.
G.
Questionable
data,
data
on
formulated
mixtures
and
emulsifiable
concentrates,
and
data
obtained
with
nonresident
species
or
previously
exposed
organisms
may
be
used
to
provide
auxiliary
information
but
should
not
be
used
in
the
derivation
of
criteria.

11.
Required
­
Data
­
A.
Certain
data
should
be
available
to
help
ensure
that
each
of
the
four
major
kinds
of
possible
adverse
effects
receives
adequate
consideration.

Results
of
acute
and
chronic
toxicity
tests
with
representative
species
of
aquatic
animals
are
necessary
so
that
data
available
for
tested
species
can
be
considered
a
useful
indication
of
the
sensitivities
of
appropriate
untested
species.

Fewer
data
concerning
toxicity
to
aquatic
plants
are
required
because
procedures
for
conducting
tests
with
plants
and
interpreting
the
results
of
such
tests
are
not
as
well
developed.
Data
concerning
bioaccumulation
by
aquatic
organisms
are
required
only
if
relevant
data
are
available
concerning
the
significance
of
residues
in
aquatic
organisms.
I
B.
To
derive
a
criterion
for
freshwater
aquatic
organisms
and
their
uses,
the
following
should
be
available:

1.
Results
of
acceptable
acute
tests
(
see
Section
IV)
with
at
least
one
species
of
freshwater
animal
in
at
least
eight
different
families
such
that
all
of
the
following
are
included:

a.

b.

C.

d.

e.

f.

g.

h.
the
family
Salmonidae
in
the
class
Osteichthyes
a
second
f
a
m
i
l
y
in
t
h
e
c
l
a
s
s
Osteichthyes,
preferably
a
commercially
or
recreationally
important
warmwater
species
(
e.
g.,
bluegill,
channel
catfish,
etc.)

a
third
family
in
the
phylum
Chordata
(
may
be
in
the
class
Osteichthyes
or
may
be
an
amphibian,
etc.)

a
planktonic
crustacean
(
e.
g.,
cladoceran,

copepod,
etc.)

a
benthic
crustacean
(
e.
g.,
ostracod,

isopod,
amphipod,
crayfish,
etc.)

an
insect
(
e.
g.,
mayfly,
dragonfly,

damselfly,
stonefly,
caddis
fly,
mosquito,

midge,
etc.)

a
family
in
a
phylum
other
than
Arthropoda
or
Chordata
(
e.
g.,
Rotifera,
Annelida,

Mol
lusca,
etc.)

a
family
in
any
order
of
insect
or
any
phylum
not
already
represented.

2.
Acute­
chronic
ratios
(
see
Section
TI)
with
species
of
aquatic
animals
in
at
least
three
different
families
provided
that
of
the
three
species:
a.
a
t
least
one
is
a
f
i
s
h
b.
a
t
least
one
is
an
invertebrate
c.
a
t
l
e
a
s
t
one
i
s
a
n
a
c
u
t
e
l
y
s
e
n
s
i
t
i
v
e
freshwater
s
p
e
c
i
e
s
(
the
o
t
h
e
r
two
may
be
saltwater
species).

3.
Results
of
a
t
least
one
acceptable
test
w
i
t
h
a
freshwater
alga
o
r
vascular
p
l
a
n
t
(
see
Section
V
I
I
I
)
.
I
f
p
l
a
n
t
s
a
r
e
among
t
h
e
a
q
u
a
t
i
c
organisms
t
h
a
t
a
r
e
most
s
e
n
s
i
t
i
v
e
t
o
t
h
e
m
a
t
e
r
i
a
l
,
r
e
s
u
l
t
s
of
a
t
e
s
t
with
a
p
l
a
n
t
i
n
another
phylum
(
d
i
v
i
s
i
o
n
)
should
a
l
s
o
be
available.

4
.
A
t
l
e
a
s
t
one
acceptable
bioconcentration
f
a
c
t
o
r
dete'rmined
w
i
t
h
a
n
a
p
p
r
o
p
r
i
a
t
e
freshwater
s
p
e
c
i
e
s
,
i
f
a
maximum
permissible
t
i
s
s
u
e
concentration
is
a
v
a
i
l
a
b
l
e
(
see
Section
IX)
.

C.
To
d
e
r
i
v
e
a
c
r
i
t
e
r
i
o
n
f
o
r
s
a
l
t
w
a
t
e
r
a
q
u
a
t
i
c
organisms
and
t
h
e
i
r
uses,
the
following
should
be
available:

1.
Results
of
acceptable
acute
tests
(
see
Section
I
V
)
w
i
t
h
a
t
l
e
a
s
t
one
s
p
e
c
i
e
s
of
s
a
l
t
w
a
t
e
r
animal
i
n
a
t
l
e
a
s
t
e
i
g
h
t
d
i
f
f
e
r
e
n
t
f
a
m
i
l
i
e
s
such
t
h
a
t
a
l
l
of
the
following
are
included:

a.
two
families
i
n
t
h
e
phylum
Chordata
b.
a
family
i
n
a
phylum
other
than
Arthropoda
o
r
Chordata
c.
either
the
Mysidae
or
Penaeidae
family
d.
three
other
families
not
in
the
phylum
Chordata
(
may
include
Mysidae
or
Penaeidae,
whichever
was
not
used
above)

e.
any
other
family.

2
.
Acute­
chronic
ratios
(
see
section
VI)
with
species
of
aquatic
animals
in
at
least
three
different
families
provided
that
of
the
three
species:

a.
at
least
one
is
a
fish
b.
at
least
one
is
an
invertebrate
c.
at
least
one
is
an
acutely
sensitive
saltwater
species
(
the
other
one
may
be
a
freshwater
species).

3
.
Results
of
at
least
one
acceptable
test
with
a
saltwater
alga
or
vascular
plant
(
see
Section
VIII.
If
plants
are
among
the
aquatic
organisms
most
sensitive
to
the
material,

results
of
a
test
with
a
plant
in
another
phylum
(
division)
should
also
be
available.

4
.
At
least
one
acceptable
bioconcentration
factor
determined
with
an
appropriate
saltwater
species,
if
a
maximum
permissible
tissue
concentration
is
available
(
see
Section
IX)
*

D.
If
all
the
required
data
are
available,
a
numerical
criterion
can
usually
be
derived,
except
in
special
0
%
.

cases.
For
example,
derivation
of
a
criterion
might
n
o
t
be
p
o
s
s
i
b
l
e
i
f
t
h
e
a
v
a
i
l
a
b
l
e
acute­

c
h
r
o
n
i
c
r
a
t
i
o
s
v
a
r
y
by
more
t
h
a
n
a
f
a
c
t
o
r
of
1
0
w
i
t
h
no
apparent
pattern.
A
l
s
o
,
i
f
a
c
r
i
t
e
r
i
o
n
is
to
be
r
e
l
a
t
e
d
t
o
a
water
q
u
a
l
i
t
y
c
h
a
r
a
c
t
e
r
i
s
t
i
c
T
(
see
S
e
c
t
i
o
n
s
V
and
V
I
I
)
,
more
d
a
t
a
w
i
l
l
be
necessary.

Similarly,
i
f
a
l
l
required
data
a
r
e
not
a
v
a
i
l
a
b
l
e
,

a
numerical
c
r
i
t
e
r
i
o
n
should
not
be
derived
except
i
n
s
p
e
c
i
a
l
cases.
For
example,
even
i
f
not
enough
a
c
u
t
e
and
chronic
d
a
t
a
are
a
v
a
i
l
a
b
l
e
,
it
might
be
p
o
s
s
i
b
l
e
t
o
derive
a
c
r
i
t
e
r
i
o
n
i
f
t
h
e
a
v
a
i
l
a
b
l
e
data
c
l
e
a
r
l
y
indicate
t
h
a
t
the
Final
Residue
Value
should
be
much
lower
than
either
the
Final
Chronic
Value
or
the
Final
P
l
a
n
t
Value.

E.
Confidence
i
n
a
c
r
i
t
e
r
i
o
n
u
s
u
a
l
l
y
increases
a
s
t
h
e
amount
of
a
v
a
i
l
a
b
l
e
p
e
r
t
i
n
e
n
t
data
i
n
c
r
e
a
s
e
s
.

Thus,
additional
data
a
r
e
u
s
u
a
l
l
y
desirable.

111.
Final
A
c
u
t
e
Value
A.
Appropriate
measures
of
t
h
e
a
c
u
t
e
(
short­
term)

t
o
x
i
c
i
t
y
o
f
t
h
e
m
a
t
e
r
i
a
l
t
o
a
v
a
r
i
e
t
y
o
f
s
p
e
c
i
e
s
of
a
q
u
a
t
i
c
animals
are
used
t
o
c
a
l
c
u
l
a
t
e
t
h
e
F
i
n
a
l
Acute
Value.
The
Final
Acute
Value
is
an
estimate
'
of
t
h
e
concentration
of
t
h
e
material
corresponding
t
o
a
cumulative
p
r
o
b
a
b
i
l
i
t
y
of
0.05
i
n
t
h
e
a
c
u
t
e
t
o
x
i
c
i
t
y
v
a
l
u
e
s
f
o
r
t
h
e
g
e
n
e
r
a
w
i
t
h
which
acceptable
a
c
u
t
e
tests
have
been
conducted
on
t
h
e
material.
However,
i
n
some
cases,
i
f
t
h
e
Species
Mean
A
c
u
t
e
V
a
l
u
e
o
f
a
c
o
m
m
e
r
c
i
a
l
l
y
o
r
recreational
1
y
important
species
is
lower
than
t
h
e
c
a
l
c
u
l
a
t
e
d
F
i
n
a
l
Acute
Value,
t
h
e
n
t
h
a
t
Species
Mean
A
c
u
t
e
Value
r
e
p
l
a
c
e
s
t
h
e
c
a
l
c
u
l
a
t
e
d
F
i
n
a
l
Acute
Value
i
n
order
t
o
provide
protection
f
o
r
t
h
a
t
important
species.

B.
Acute
t
o
x
i
c
i
t
y
tests
should
have
been
conducted
using
acceptable
procedures
[
13].

C.
Except
f
o
r
tests
w
i
t
h
s
a
l
t
w
a
t
e
r
a
n
n
e
l
i
d
s
and
mysids,
r
e
s
u
l
t
s
of
a
c
u
t
e
tests
during
which
t
h
e
t
e
s
t
organisms
were
fed
should
not
be
used,
unless
data
i
n
d
i
c
a
t
e
t
h
a
t
t
h
e
food
d
i
d
n
o
t
a
f
f
e
c
t
t
h
e
t
o
x
i
c
i
t
y
of
t
h
e
t
e
s
t
material.

D.
R
e
s
u
l
t
s
o
f
a
c
u
t
e
t
e
s
t
s
conducted
i
n
unusual
d
i
l
u
t
i
o
n
water,
e.
g.,
d
i
l
u
t
i
o
n
water
i
n
which
t
o
t
a
l
0
organic
carbon
o
r
p
a
r
t
i
c
u
l
a
t
e
m
a
t
t
e
r
exceeded
5
mg/
L,
should
not
be
used,
unless
a
r
e
l
a
t
i
o
n
s
h
i
p
is
developed
between
acute
t
o
x
i
c
i
t
y
and
organic
carbon
o
r
p
a
r
t
i
c
u
l
a
t
e
matter
o
r
u
n
l
e
s
s
d
a
t
a
show
t
h
a
t
organic
carbon,
p
a
r
t
i
c
u
l
a
t
e
m
a
t
t
e
r
,
etc.,
do
not
affect
toxicity.

E.
Acute
values
should
be
based
on
endpoints
which
r
e
f
l
e
c
t
t
h
e
t
o
t
a
l
s
e
v
e
r
e
a
c
u
t
e
adverse
impact
of
t
h
e
t
e
s
t
m
a
t
e
r
i
a
l
on
t
h
e
organisms
used
i
n
the
test.
Therefore,
only
t
h
e
following
kinds
of
data
on
acute
t
o
x
i
c
i
t
y
t
o
a
q
u
a
t
i
c
animals
should
be
used:
1.
T
e
s
t
s
w
i
t
h
daphnids
and
o
t
h
e
r
cladocerans
should
be
s
t
a
r
t
e
d
with
organisms
less
than
24
hours
o
l
d
and
t
e
s
t
s
with
midges
should
be
stressed
with
second­
o
r
t
h
i
r
d­
i
n
s
t
a
r
larvae.

The
r
e
s
u
l
t
should
be
t
h
e
48­
hr
EC50
based
on
p
e
r
c
e
n
t
a
g
e
of
organisms
immobilized
p
l
u
s
percentage
of
organisms
k
i
l
l
e
d
.
I
f
such
an
EC50
i
s
n
o
t
a
v
a
i
l
a
b
l
e
from
a
t
e
s
t
,
t
h
e
48­
hr
LC50
should
be
used
i
n
p
l
a
c
e
of
t
h
e
d
e
s
i
r
e
d
48­
hr
EC50.
An
EC50
o
r
LC50
of
l
o
n
g
e
r
t
h
a
n
4
8
hours
can
be
used
as
long
as
t
h
e
animals
were
n
o
t
f
e
d
and
t
h
e
c
o
n
t
r
o
l
animals
were
acceptable
a
t
t
h
e
end
of
t
h
e
test.

a
2.
The
r
e
s
u
l
t
of
a
t
e
s
t
w
i
t
h
embryos
and
l
a
r
v
a
e
o
f
b
a
r
n
a
c
l
e
s
,
b
i
v
a
l
v
e
m
o
l
l
u
s
c
s
(
c
l
a
m
s
,

mussels,
oysters,
and
s
c
a
l
l
o
p
s
)
,
sea
urchins,

l
o
b
s
t
e
r
s
,
c
r
a
b
s
,
shrimp,
and
abalones
should
be
t
h
e
96­
hr
EC50
based
on
t
h
e
percentage
of
organisms
with
incompletely
developed
s
h
e
l
l
s
p
l
u
s
t
h
e
percentage
of
organisms
k
i
l
l
e
d
.
I
f
such
an
EC50
is
n
o
t
a
v
a
i
l
a
b
l
e
from
a
t
e
s
t
,
t
h
e
l
o
w
e
r
o
f
t
h
e
96­
h
r
EC50
based
on
t
h
e
percentage
of
organisms
w
i
t
h
incompletely
developed
shells
and
t
h
e
96­
hr
LC50
should
be
used
i
n
p
l
a
c
e
of
t
h
e
desired
96­
hr
EC50.
I
f
t
h
e
duration
of
t
h
e
t
e
s
t
w
a
s
between
48
and
96
hours,
t
h
e
EC50
o
r
LC50
a
t
t
h
e
end
of
t
h
e
test
should
be
used.
3.
T
h
e
acute
v
a
l
u
e
s
from
tests
w
i
t
h
a
l
l
o
t
h
e
r
freshwater
and
s
a
l
t
w
a
t
e
r
animal
s
p
e
c
i
e
s
and
o
l
d
e
r
l
i
f
e
s
t
a
g
e
s
o
f
b
a
r
n
a
c
l
e
s
,
b
i
v
a
l
v
e
m
o
l
l
u
s
c
s
,
s
e
a
u
r
c
h
i
n
s
,
l
o
b
s
t
e
r
s
,
c
r
a
b
s
,

shrimps,
and
abalones
should
be
t
h
e
96­
hr
EC50
b
a
s
e
d
on
t
h
e
p
e
r
c
e
n
t
a
g
e
o
f
o
r
g
a
n
i
s
m
s
e
x
h
i
b
i
t
i
n
g
loss
of
e
q
u
i
l
i
b
r
i
u
m
p
l
u
s
t
h
e
percentage
of
organisms
immobilized
p
l
u
s
the
percentage
OF
organisms
k
i
l
l
e
d
.
I
f
such
a
n
EC50
is
n
o
t
a
v
a
i
l
a
b
l
e
from
a
t
e
s
t
,
t
h
e
96­
hr
LC50
should
be
used
i
n
p
l
a
c
e
of
t
h
e
desired
96­
hr
EC50.

4
.
T
e
s
t
s
w
i
t
h
s
i
n
g
l
e­
c
e
l
l
e
d
organisms
a
r
e
not
considered
a
c
u
t
e
t
e
s
t
a
,
even
i
f
t
h
e
d
u
r
a
t
i
o
n
was
96
hours
o
r
less.
0
5.
If
t
h
e
tests
were
conducted
p
r
o
p
e
r
l
y
,
acute
v
a
l
u
e
s
reported
a
s
"
g
r
e
a
t
e
r
than"
v
a
l
u
e
s
and
t
h
o
s
e
which
a
r
e
above
t
h
e
s
o
l
u
b
i
l
i
t
y
of
t
h
e
t
e
s
t
m
a
t
e
r
i
a
l
s
h
o
u
l
d
be
u
s
e
d
,
because
r
e
j
e
c
t
i
o
n
o
f
s
u
c
h
a
c
u
t
e
v
a
l
u
e
s
would
unnecessarily
lower
t
h
e
F
i
n
a
l
Acute
Value
by
e
l
i
m
i
n
a
t
i
n
g
a
c
u
t
e
v
a
l
u
e
s
for
r
e
s
i
s
t
a
n
t
species.

F.
If
t
h
e
acute
t
o
x
i
c
i
t
y
of
t
h
e
material
t
o
aquatic
animals
apparently
has
been
shown
t
o
be
r
e
l
a
t
e
d
t
o
a
water
q
u
a
l
i
t
y
c
h
a
r
a
c
t
e
r
i
s
t
i
c
such
as
hardness
o
r
p
a
r
t
i
c
u
l
a
t
e
m
a
t
t
e
r
f
o
r
f
r
e
s
h
w
a
t
e
r
animals
o
r
s
a
l
i
n
i
t
y
or
p
a
r
t
i
c
u
l
a
t
e
m
a
t
t
e
r
f
o
r
s
a
l
t
w
a
t
e
r
animals,
a
Final
Acute
Equation
should
be
derived
based
on
t
h
a
t
water
q
u
a
l
i
t
y
characteristic.
G
o
t
o
Section
V.

G.
If
t
h
e
a
v
a
i
l
a
b
l
e
d
a
t
a
i
n
d
i
c
a
t
e
t
h
a
t
one
or
more
l
i
f
e
s
t
a
g
e
s
a
r
e
a
t
l
e
a
s
t
a
f
a
c
t
o
r
of
2
more
r
e
s
i
s
t
a
n
t
than
one
ormore
o
t
h
e
r
l
i
f
e
s
t
a
g
e
s
o
f
t
h
e
same
species,
the
data
for
the
more
r
e
s
i
s
t
a
n
t
l
i
f
e
stages
should
not
be
used
i
n
t
h
e
c
a
l
c
u
l
a
t
i
o
n
of
t
h
e
Species
Mean
Acute
Value
(
SMAV)
because
a
species
c
a
n
o
n
l
y
be
c
o
n
s
i
d
e
r
e
d
p
r
o
t
e
c
t
e
d
from
a
c
u
t
e
t
o
x
i
c
i
t
y
i
f
a
l
l
l
i
f
e
stages
are
protected.

H.
T
h
e
agreement
o
f
t
h
e
d
a
t
a
w
i
t
h
i
n
and
between
species
should
be
considered.
Acute
v
a
l
u
e
s
t
h
a
t
appear
t
o
be
questionable
i
n
comparison
w
i
t
h
other
acute
and
chronic
data
f
o
r
the
same
species
and
for
other
species
i
n
the
same
genus
probably
should
not
be
used
i
n
c
a
l
c
u
l
a
t
i
o
n
of
a
Species
Mean
Acute
Value.
For
example,
i
f
the
acute
values
a
v
a
i
l
a
b
l
e
For
a
species
o
r
genus
d
i
f
f
e
r
by
more
than
a
f
a
c
t
o
r
of
1
0
,
some
or
a
l
l
of
t
h
e
v
a
l
u
e
s
probably
should
not
be
used
i
n
calculations.

I.
For
each
species
f
o
r
which
a
t
l
e
a
s
t
one
a
c
u
t
e
v
a
l
u
e
i
s
a
v
a
i
l
a
b
l
e
,
the
Species
Mean
Acute
Value
should
be
c
a
l
c
u
l
a
t
e
d
a
s
t
h
e
geometric
mean
of
t
h
e
r
e
s
u
l
t
s
of
a
l
l
flow­
through
tests
i
n
which
t
h
e
concentrations
of
test
material
were
measured.
For
a
species
for
which
no
such
r
e
s
u
l
t
i
s
a
v
a
i
l
a
b
l
e
,
the
Species
Mean
Acute
Value
should
be
calculated
as
t
h
e
geometric
mean
of
a
l
l
a
v
a
i
l
a
b
l
e
a
c
u
t
e
v
a
l
u
e
s
,
i.
e.,
r
e
s
u
l
t
s
of
flow­
through
tests
i
n
which
t
h
e
concentrations
were
not
measured
and
r
e
s
u
l
t
s
of
s
t
a
t
i
c
and
renewal
t
e
s
t
s
based
on
i
n
i
t
i
a
l
concentrations
of
t
e
s
t
m
a
t
e
r
i
a
l
(
nominal
c
o
n
c
e
n
t
r
a
t
i
o
n
s
a
r
e
a
c
c
e
p
t
a
b
l
e
f
o
r
most
t
e
s
t
m
a
t
e
r
i
a
l
s
i
f
measured
c
o
n
c
e
n
t
r
a
t
i
o
n
s
a
r
e
n
o
t
available).

NOTE:
Data
reported
by
o
r
i
g
i
n
a
l
investigators
should
not
be
rounded
o
f
f
.
R
e
s
u
l
t
s
of
a
l
l
intermediate
c
a
l
c
u
l
a
t
i
o
n
s
s
h
o
u
l
d
b
e
rounded
[
14]
to
f
o
u
r
significant
d
i
g
i
t
s
.

The
geometric
mean
of
N
numbers
is
t
h
e
Nth
root
of
0
NOTE:

t
h
e
product
of
t
h
e
N
numbers.
Alternatively,
the
geometric
mean
can
be
c
a
l
c
u
l
a
t
e
d
by
adding
t
h
e
logarithms
of
t
h
e
N
numbers,
d
i
v
i
d
i
n
g
t
h
e
sum
by
N,
and
t
a
k
i
n
g
t
h
e
a
n
t
i
l
o
g
of
t
h
e
q
u
o
t
i
e
n
t
.
The
geometric
mean
of
t
w
o
numbers
is
t
h
e
square
root
of
t
h
e
p
r
o
d
u
c
t
of
t
h
e
two
numbers,
and
t
h
e
geometric
mean
of
one
number
is
t
h
a
t
number.

E
i
t
h
e
r
n
a
t
u
r
a
l
(
base
0
)
or
common
(
b
a
s
e
1
0
)

logarithms
can
be
used
to
c
a
l
c
u
l
a
t
e
geometric
means
as
long
a
s
they
a
r
e
used
consistently
w
i
t
h
i
n
each
set
of
d
a
t
a
,
i.
e.,
t
h
e
a
n
t
i
l
o
g
used
must
match
the
logarithm
U
s
e
d
.
NOTE:
Geometric
means,
rather
than
arithmetic
means,
are
used
here
because
t
h
e
distributions
of
sensitivities
of
individual
organisms
in
toxicity
tests
on
most
materials
and
the
distributions
of
sensitivities
of
species
within
a
genus
are
more
likely
to
be
lognormal
than
normal.
Similarly,

geometric
means
are
used
for
acute­
chronic
ratios
and
bioconcentration
factors
because
quotients
are
likely
to
be
closer
to
lognormal
than
normal
distributions.
In
addition,
division
of
the
geometric
mean
of
a
set
of
numerators
by
the
geometric
mean
of
the
set
of
corresponding
denominators
will
result
in
the
geometric
mean
of
the
set
of
corresponding
quotients.

J.
For
each
genus
 or
which
one
or
more
Species
Mean
Acute
Values
are
available,
the
Genus
Mean
Acute
Value
should
be
calculated
as
the
geometric
mean
of
the
Species
Mean
Acute
Values
available
f
o
r
the
genus.

K.
Order
the
Genus
Mean
Acute
Value
from
high
to
low.

L.
Assign
ranks,
R,
to
the
Genus
Mean
Acute
Value
from
vvlvv
for
the
lowest
to
*'
N"
 or
the
highest.

If
two
or
more
Genus
Mean
Acute
Values
are
identical,
arbitrarily
assign
them
successive
ranks.

M.
Calculate
the
cumulative
probability,
P,
 or
each
Genus
Mean
Acute
Value
as
R/
(
N+
l).
N.
Select
t
h
e
four
Genus
Mean
Acute
Value
which
have
cumulative
p
r
o
b
a
b
i
l
i
t
i
e
s
c
l
o
s
e
s
t
t
o
0.05
(
i
f
there
a
r
e
l
e
s
s
than
59
Genus
Mean
Acute
Value,
these
w
i
l
l
always
be
t
h
e
f
o
u
r
lowest
Genus
Mean
Acute
Values).

Using
t
h
e
selected
Genus
Mean
Acute
Values
and
Fs,

calculate:

S2=
E
(
l
n
GMAV)
2)­
(
(
E
l
n
GMAV))
2/
4)

(
PI
­
((
E
/
"
~
)
)
2
/
4
)
0.

L
=
(
E(
1n
GMAV)
­
S(
E(/
Ap)))/
4
A
=
S(/"
O.
OS)
+
L
FAV
=
e
A
(
See
[
113
f
o
r
development
of
t
h
e
c
a
l
c
u
l
a
t
i
o
n
procedure
and
Appendix
2
f
o
r
example
calculation
and
computer
program.)

NOTE:
Natural
logarithms
(
logarithms
t
o
base
el
denoted
a
s
I
n
)
a
r
e
used
h
e
r
e
i
n
merely
because
they
a
r
e
e
a
s
i
e
r
t
o
use
on
some
hand
c
a
l
c
u
l
a
t
o
r
s
and
computers
t
h
a
n
common
(
base
1
0
)
l
o
g
a
r
i
t
h
m
s
.

Consistent
u
s
e
of
e
i
t
h
e
r
w
i
l
l
produce
t
h
e
same
result.

P.
If
f
o
r
a
commercially
o
r
r
e
c
r
e
a
t
i
o
n
a
l
l
y
important
s
p
e
c
i
e
s
t
h
e
geometric
mean
of
t
h
e
a
c
u
t
e
v
a
l
u
e
s
f
r
o
m
f
l
o
w
­
t
h
r
o
u
g
h
t
e
s
t
s
i
n
w
h
i
c
h
t
h
e
concentrations
of
t
e
s
t
material
were
measured
i
s
lower
than
the
calculated
F
i
n
a
l
Acute
Value,
then
t
h
a
t
geometric
mean
should
be
used
a
s
t
h
e
F
i
n
a
l
Acute
Value
instead
of
the
c
a
l
c
u
l
a
t
e
d
Final
Acute
Value.

Q.
GO
to
section
VI.
­
IV.
Final
Acute
Equation
0
A.
When
enough
data
are
available
to
show
that
acute
toxicity
to
two
o
r
more
species
is
similarly
related
to
a
water
quality
characteristic,
the
relationship
should
be
taken
into
account
as
described
in
Sections
B­
G
below
or
using
analysis
of
covariance
[
15,16].
The
two
methods
are
equivalent
and
produce
identical
results.
The
manual
method
described
below
provides
an
unuerstanding
of
this
application
of
covariance
analysis,
but
computerized
versions
of
covariance
analysis
are
much
more
convenient
for
analyzing
large
data
tests.
If
two
or
more
factors
affect
toxicity,
multiple
regression
analysis
should
be
used.

B.
For
each
species
for
which
comparable
acute
toxicity
values
are
available
at
two
or
more
different
values
of
the
water
quality
characteristic,
perform
a
least
squares
regression
of
the
acute
toxicity
values
on
the
corresponding
values
of
the
water
quality
characteristic
to
obtain
the
slope
and
its
95
percent
confidence
limits
 or
each
species.

NOTE:
Because
the
best
documented
relationship
fitting
these
data
is
that
between
hardness
and
acute
toxicity
of
metals
in
fresh
water
and
a
log­
log
relationship,
geometric
means
and
natural
logarithms
of
both
toxicity
and
water
quality
are
used
in
the
rest
of
this
section.
For
relationships
based
on
other
water
quality
characteristics
such
as
pH,
temperature,
or
salinity,
no
transformation
or
a
different
transformation
might
fit
the
data
better,
and
appropriate
changes
will
be
necessary
throughout
this
section.

C
.
Decide
whether
the
data
for
eachspecies
are
useful,
taking
into
account
the
range
and
number
of
the
tested
values
of
the
water
quality
characteristic
and
the
degree
of
agreement
within
and
between
species.
For
example,
a
slope
based
on
six
data
points
might
be
of
limited
value
if
it
is
based
only
on
data
for
a
very
narrow
range
of
values
of
the
water
quality
characteristic.
A
slope
based
on
only
two
data
points,
however,

might
be
useful
if
it
is
consistent
with
other
information
and
if
the
two
points
cover
a
broad
enough
range
of
the
water
quality
characteristic.

In
addition,
acute
values
that
appear
to
be
questionable
in
comparison
with
other
acute
and
chronic
data
available
for
the
same
species
and
for
other
species
in
the
same
genus
probably
should
not
be
used.
For
example,
if
after
adjustment
for
the
water
quality
characteristic,

the
acute
values
available
for
a
species
or
genus
d
i
f
f
e
r
by
more
than
a
f
a
c
t
o
r
of
1
0
,
probably
some
o
r
a
l
l
of
the
v
a
l
u
e
s
should
be
rejected.
I
f
useful
slopes
a
r
e
not
a
v
a
i
l
a
b
l
e
f
o
r
a
t
least
one
f
i
s
h
and
one
i
n
v
e
r
t
e
b
r
a
t
e
o
r
i
f
t
h
e
a
v
a
i
l
a
b
l
e
s
l
o
p
e
s
a
r
e
t
o
o
d
i
s
s
i
m
i
l
a
r
o
r
i
f
t
o
o
f
e
w
d
a
t
a
a
r
e
a
v
a
i
l
a
b
l
e
t
o
adequately
d
e
f
i
n
e
t
h
e
r
e
l
a
t
i
o
n
s
h
i
p
between
a
c
u
t
e
t
o
x
i
c
i
t
y
and
t
h
e
water
q
u
a
l
i
t
y
characteristic,
return
t
o
Section
IV.
G,
using
t
h
e
r
e
s
u
l
t
s
of
tests
conducted
under
conditions
and
i
n
waters
s
i
m
i
l
a
r
t
o
those
commonly
used
f
o
r
t
o
x
i
c
i
t
y
tests
with
the
species.

D.
I
n
d
i
v
i
d
u
a
l
l
y
f
o
r
each
species
c
a
l
c
u
l
a
t
e
the
geometric
mean
of
t
h
e
a
v
a
i
l
a
b
l
e
acute
values
and
t
h
e
n
d
i
v
i
d
e
each
of
t
h
e
a
c
u
t
e
v
a
l
u
e
s
f
o
r
species
by
the
mean
f
o
r
the
species.
T
h
i
s
normalizes
t
h
e
v
a
l
u
e
s
so
t
h
a
t
t
h
e
g
e
o
m
e
t
r
i
c
mean
o
f
t
h
e
normalized
values
f
o
r
each
species
individual1
y
and
f
o
r
any
combination
of
species
is
1.0.

E.
S
i
m
i
l
a
r
l
y
normalize
t
h
e
v
a
l
u
e
s
of
t
h
e
water
q
u
a
l
i
t
y
c
h
a
r
a
c
t
e
r
i
s
t
i
c
f
o
r
e
a
c
h
s
p
e
c
i
e
s
individually.

F.
I
n
d
i
v
i
d
u
a
l
l
y
f
o
r
each
species
perform
a
l
e
a
s
t
squares
regression
of
t
h
e
normalized
acute
t
o
x
i
c
i
t
y
v
a
l
u
e
s
on
t
h
e
corresponding
normalized
values
of
the
water
q
u
a
l
i
t
y
characteristic.
The
r
e
s
u
l
t
i
n
g
slopes
and
95
percent
confidence
l
i
m
i
t
s
~.
w
i
l
l
be
i
d
e
n
t
i
c
a
l
t
o
those
obtained
i
n
Section
B.
NOW,
however,
i
f
t
h
e
d
a
t
a
a
r
e
a
c
t
u
a
l
l
y
p
l
o
t
t
e
d
,

t
h
e
l
i
n
e
of
best
f
i
t
f
o
r
each
i
n
d
i
v
i
d
u
a
l
s
p
e
c
i
e
s
w
i
l
l
go
through
t
h
e
p
o
i
n
t
1,
l
i
n
t
h
e
center
of
t
h
e
graph.

G.
Treat
a
l
l
the
normalized
data
as
i
f
they
were
a
l
l
f
o
r
t
h
e
same
species
and
perform
a
l
e
a
s
t
squares
regression
of
a
l
l
the
normalized
acute
values
on
the
corresponding
normalized
values
of
the
water
q
u
a
l
i
t
y
c
h
a
r
a
c
t
e
r
i
s
t
i
c
t
o
obtain
t
h
e
pooled
acute
s
l
o
p
e
,
V
,
and
its
95
p
e
r
c
e
n
t
confidence
l
i
m
i
t
s
.

If
a
l
l
the
normalized
data
a
r
e
a
c
t
u
a
l
l
y
plotted,

t
h
e
l
i
n
e
of
best
f
i
t
w
i
l
l
go
through
t
h
e
p
o
i
n
t
1,
l
i
n
t
h
e
center
of
t
h
e
graph.

H.
For
each
species
aalculate
t
h
e
geometric
mean,
W,

of
the
a
c
u
t
e
t
o
x
i
c
i
t
y
v
a
l
u
e
s
and
t
h
e
geometric
mean,
X
I
of
t
h
e
v
a
l
u
e
s
of
t
h
e
w
a
t
e
r
q
u
a
l
i
t
y
characteristic.
(
These
were
calculated
i
n
s
t
e
p
s
D
and
E.)

I.
For
each
s
p
e
c
i
e
s
c
a
l
c
u
l
a
t
e
t
h
e
logarithm,
Y
,
of
t
h
e
Species
Mean
Acute
Value
a
t
a
selected
value,

2,
of
t
h
e
water
q
u
a
l
i
t
y
c
h
a
r
a
c
t
e
r
i
s
t
i
c
using
t
h
e
equation:

Y
=
I
n
W
­
v
(
l
n
X
­
In
2
)
.

J.
For
each
species
c
a
l
c
u
l
a
t
e
t
h
e
SMAV
a
t
Z
using
t
h
e
equation:
SMAV
=
eY.

NOTE:
Alternatively,
the
Species
Mean
Acute
Values
a
t
Z
can
be
obtained
by
skipping
s
t
e
p
H
u
s
i
n
g
t
h
e
equations
i
n
steps
I
and
J
t
o
a
d
j
u
s
t
each
acute
value
individually
t
o
2,
and
then
calculating
the
geometric
mean
of
t
h
e
a
d
j
u
s
t
e
d
v
a
l
u
e
s
 or
each
species
individually.
T
h
i
s
a
l
t
e
r
n
a
t
i
v
e
procedure
allows
an
examination
of
the
range
of
t
h
e
adjusted
acute
values
f
o
r
each
species.

K.
Obtain
the
F
i
n
a
l
Acute
Value
a
t
Z
by
using
the
procedure
described
i
n
Section
1V.
J­
0.

L.
I
f
t
h
e
S
p
e
c
i
e
s
Mean
Acute
V
a
l
u
e
a
t
Z
of
a
commercially
o
r
r
e
c
r
e
a
t
i
o
n
a
l
l
y
important
species
is
lower
than
the
calculated
Final
Acute
Value
a
t
Z,
t
h
e
n
t
h
a
t
Species
Mean
Acute
Value
should
be
used
as
t
h
e
Final
Acute
Value
a
t
Z
instead
of
the
calculated
Final
Acute
Value.

M.
The
F
i
n
a
l
Acute
Equation
is
w
r
i
t
t
e
n
as:
F
i
n
a
l
A
c
u
t
e
V
a
l
u
e
=
.
(
V
[
l
n
(
w
a
t
e
r
q
u
a
l
i
t
y
c
h
a
r
a
c
t
e
r
i
s
t
i
c
)
]
+
I
n
A
­
V
[
l
n
Z]),
where
V
=

pooled
acute
s
l
o
p
e
a
n
d
A
=
F
i
n
a
l
A
c
u
t
e
v
a
l
u
e
a
t
2.

Because
V,
A,
and
2
are
known,
t
h
e
F
i
n
a
l
Acute
Value
can
be
calculated
f
o
r
any
selected
value
of
the
water
q
u
a
l
i
t
y
c
h
a
r
a
c
t
e
r
i
s
t
i
c
.

V.
­
Final
Chronic
Value
A.
Depending
on
t
h
e
d
a
t
a
t
h
a
t
a
r
e
a
v
a
i
l
a
b
l
e
concerning
chronic
t
o
x
i
c
i
t
y
t
o
a
q
u
a
t
i
c
animals,

the
Final
Chronic
Value
might
be
calculated
i
n
the
same
manner
a
s
t
h
e
F
i
n
a
l
Acute
Value
o
r
by
dividing
t
h
e
Final
Acute
Value
by
the
Final
Acute­
NOTE
:

B.
Chronic
R
a
t
i
o
.
I
n
some
c
a
s
e
s
it
may
n
o
t
be
possible
to
calculate
a
F
i
n
a
l
Chronic
Value.

As
t
h
e
name
implies,
the
acute­
chronic
r
a
t
i
o
is
a
way
of
r
e
l
a
t
i
n
g
acute
and
chronic
t
o
x
i
c
i
t
i
e
s
.
The
acute­
chronic
r
a
t
i
o
is
b
a
s
i
c
a
l
l
y
t
h
e
i
n
v
e
r
s
e
of
t
h
e
a
p
p
l
i
c
a
t
i
o
n
f
a
c
t
o
r
,
b
u
t
t
h
i
s
new
name
i
s
better
because
it
is
more
d
e
s
c
r
i
p
t
i
v
e
and
should
h
e
l
p
p
r
e
v
e
n
t
c
o
n
f
u
s
i
o
n
between
'
a
p
p
l
i
c
a
t
i
o
n
f
a
c
t
o
r
s
"
and
"
s
a
f
e
t
y
f
a
c
t
o
r
s
.
"
Acute­
chronic
r
a
t
i
o
s
and
a
p
p
l
i
c
a
t
i
o
n
f
a
c
t
o
r
s
a
r
e
ways
of
r
e
l
a
t
i
n
g
t
h
e
a
c
u
t
e
and
c
h
r
o
n
i
c
t
o
x
i
c
i
t
i
e
s
of
a
material
t
o
aquatic
organisms.
Safety
f
a
c
t
o
r
s
are
used
to
p
r
o
v
i
d
e
an
e
x
t
r
a
margin
of
s
a
f
e
t
y
beyond
the
known
or
e
s
t
i
m
a
t
e
d
s
e
n
s
i
t
i
v
i
t
i
e
s
of
a
q
u
a
t
i
c
organisms.
Another
advantage
of
t
h
e
acute­
chronic
r
a
t
i
o
i
s
t
h
a
t
it
w
i
l
l
u
s
u
a
l
l
y
be
g
r
e
a
t
e
r
than
1;

t
h
i
s
should
avoid
t
h
e
confusion
as
to
whether
a
l
a
r
g
e
a
p
p
l
i
c
a
t
i
o
n
f
a
c
t
o
r
is
one
t
h
a
t
is
c
l
o
s
e
to
u
n
i
t
y
or
one
t
h
a
t
has
a
denominator
t
h
a
t
is
much
greater
than
t
h
e
numerator.

Chronic
values
should
be
based
on
r
e
s
u
l
t
s
of
flow­

t
h
r
o
u
g
h
(
except
r
e
n
e
w
a
l
is
a
c
c
e
p
t
a
b
l
e
f
o
r
d
a
p
h
n
i
d
s
)
c
h
r
o
n
i
c
t
e
s
t
s
i
n
w
h
i
c
h
t
h
e
c
o
n
c
e
n
t
r
a
t
i
o
n
s
of
t
e
s
t
m
a
t
e
r
i
a
l
i
n
t
h
e
t
e
s
t
s
o
l
u
t
i
o
n
s
w
e
r
e
p
r
o
p
e
r
l
y
measured
a
t
a
p
p
r
o
p
r
i
a
t
e
I
t
i
m
e
s
during
the
test.
.
*
C
.
R
e
s
u
l
t
s
of
chronic
tests
i
n
which
s
u
r
v
i
v
a
l
,
growth,
o
r
reproduction
i
n
t
h
e
c
o
n
t
r
o
l
treatment
was
unacceptably
low
should
n
o
t
be
used.
The
l
i
m
i
t
s
of
a
c
c
e
p
t
a
b
i
l
i
t
y
w
i
l
l
depend
on
t
h
e
species.

D.
R
e
s
u
l
t
s
of
chronic
t
e
s
t
s
conducted
i
n
unusual
d
i
l
u
t
i
o
n
water,
e.
g.,
d
i
l
u
t
i
o
n
water
i
n
which
t
o
t
a
l
o
r
g
a
n
i
c
carbon
o
r
p
a
r
t
i
c
u
l
a
t
e
m
a
t
t
e
r
exceeded
5
mg/
L,
should
n
o
t
be
used,
u
n
l
e
s
s
a
relationship
is
developed
between
chronic
t
o
x
i
c
i
t
y
and
organic
carbon
o
r
p
a
r
t
i
c
u
l
a
t
e
matter
o
r
unless
data
show
t
h
a
t
organic
carbon,
p
a
r
t
i
c
u
l
a
t
e
matter,
etc.,
do
not
a
f
f
e
c
t
toxicity.

E.
Chronic
v
a
l
u
e
s
should
be
based
on
endpoints
and
l
e
n
g
t
h
s
of
exposure
a
p
p
r
o
p
r
i
a
t
e
t
o
t
h
e
species.

Therefore,
only
r
e
s
u
l
t
s
of
t
h
e
following
kinds
of
chronic
t
o
x
i
c
i
t
y
tests
should
be
used:

1.
L
i
f
e­
c
y
c
l
e
t
o
x
i
c
i
t
y
tests
c
o
n
s
i
s
t
i
n
g
of
exposures
of
each
of
t
w
o
or
more
groups
of
i
n
d
i
v
i
d
u
a
l
s
o
f
a
s
p
e
c
i
e
s
to
a
d
i
f
f
e
r
e
n
t
concentration
of
t
h
e
t
e
s
t
material
throughout
a
l
i
f
e
c
y
c
l
e
.
To
ensure
t
h
a
t
a
l
l
l
i
f
e
s
t
a
g
e
s
and
l
i
f
e
p
r
o
c
e
s
s
e
s
a
r
e
exposed,
t
e
s
t
s
w
i
t
h
f
i
s
h
s
h
o
u
l
d
b
e
g
i
n
w
i
t
h
embryos
o
r
n
e
w
l
y
hatched
young
less
than
48
hours
o
l
d
,
continue
t
h
r
o
u
g
h
m
a
t
u
r
a
t
i
o
n
and
r
e
p
r
o
d
u
c
t
i
o
n
,
and
should
end
n
o
t
l
e
s
s
than
2
4
days
(
90
days
f
o
r
salmonids)
a
f
t
e
r
t
h
e
hatching
of
the
next
generation.
T
e
s
t
s
w
i
t
h
daphnids
should
begin
w
i
t
h
young
less
t
h
a
n
2
4
hours
o
l
d
a
n
d
l
a
s
t
f
o
r
n
o
t
less
t
h
a
n
2
1
days.
T
e
s
t
s
w
i
t
h
mysids
should
begin
w
i
t
h
young
less
than
24
hours
o
l
d
and
c
o
n
t
i
n
u
e
u
n
t
i
l
7
days
p
a
s
t
themediantime
of
first
brood
r
e
l
e
a
s
e
i
n
t
h
e
c
o
n
t
r
o
l
s
.
For
f
i
s
h
,
data
should
be
obtained
and
analyzed
on
s
u
r
v
i
v
a
l
and
growth
of
a
d
u
l
t
s
and
young,

maturation
of
males
and
females,
eggs
spawned
per
female,
embryo
v
i
a
b
i
l
i
t
y
(
salmonids
only)
,

and
h
a
t
c
h
a
b
i
l
i
t
y
.
For
daphnids,
d
a
t
a
should
be
obtained
and
analyzed
on
s
u
r
v
i
v
a
l
and
young
p
e
r
f
e
m
a
l
e
.
For
mysids,
d
a
t
a
s
h
o
u
l
d
be
obtained
and
analyzed
on
s
u
r
v
i
v
a
l
,
growth,
and
young
per
female.

2
.
P
a
r
t
i
a
l
l
i
f
e­
c
y
c
l
e
t
o
x
i
c
i
t
y
tests
consisting
of
exposures
of
each
of
two
or
more
groups
of
i
n
d
i
v
i
d
u
a
l
s
o
f
a
s
p
e
c
i
e
s
o
f
f
i
s
h
to
a
c
o
n
c
e
n
t
r
a
t
i
o
n
of
t
h
e
t
e
s
t
material
through
most
portions
of
a
l
i
f
e
cycle.
P
a
r
t
i
a
l
l
i
f
e
­

cycle
tests
are
allowed
w
i
t
h
f
i
s
h
species
that
r
e
q
u
i
r
e
more
than
a
y
e
a
r
to
reach
s
e
x
u
a
l
maturity,
so
t
h
a
t
a
l
l
major
l
i
f
e
stages
can
be
exposed
t
o
t
h
e
t
e
s
t
m
a
t
e
r
i
a
l
i
n
less
t
h
a
n
15
months.
Exposure
to
t
h
e
test
material
should
begin
with
immature
j
u
v
e
n
i
l
e
s
a
t
l
e
a
s
t
2
months
p
r
i
o
r
to
a
c
t
i
v
e
gonad
development,

continue
through
maturation
and
reproduction,
and
end
not
less
than
2
4
days
(
90
days
for
salmonids)
after
the
hatching
of
the
next
generation.
Data
should
be
obtained
and
analyzed
on
survival
and
growth
of
adults
and
young,
maturation
of
males
and
females,
eggs
spawned
per
female,
embryo
viability
(
salmonids
only),
and
hatchability.

3.
Early
life­
stage
toxicity
tests
consisting
of
28­
to
32­
day
(
6
0
days.
post
hatch
for
salmonids)
exposures
of
the
early
life
stages
of
a
species
of
fish
from
shortly
a,
fter
fertilization
through
embryonic,
larval,
and
early
juvenile
development.
Data
should
be
obtained
and
analyzed
on
survival
and
growth.

NOTE:
Results
of
an
early
life­
stage
test
are
used
as
predictions
of
results
of
life­
cycle
and
partial
life­
cycle
tests
with
the
same
species.

Therefore,
when
results
of
a
life­
cycle
or
partial
life­
cycle
test
are
available,
results
of
an
early
life­
stage
test
with
the
same
species
should
not
be
used.
Also,
results
of
early
life­
stage
tests
in
which
the
incidence
of
mortalities
or
abnormalities
increased
substantially
near
the
end
of
the
test
should
not
be
used
because
results
of
such
tests
are
possibly
not
good
predictions
of
the
results
of
comparable
life­
cycle
or
partial
1
if
e­
cycle
tests.
0
F.
A
chronic
v
a
l
u
e
may
be
obtained
by
c
a
l
c
u
l
a
t
i
n
g
t
h
e
geometric
mean
of
t
h
e
lower
and
upper
chronic
l
i
m
i
t
s
from
a
chronic
t
e
s
t
o
r
by
analyzing
chronic
d
a
t
a
using
r
e
g
r
e
s
s
i
o
n
a
n
a
l
y
s
i
s
.
A
l
o
w
e
r
chronic
l
i
m
i
t
is
t
h
e
h
i
g
h
e
s
t
tested
c
o
n
c
e
n
t
r
a
t
i
o
n
(
a
)
i
n
an
acceptable
c
h
r
o
n
i
c
t
e
s
t
,
(
b)
which
d
i
d
n
o
t
cause
an
unacceptable
amount
of
adverse
effect
on
any
of
the
specified
biological
measurements,
and
(
c)
below
which
no
tested
concentration
caused
an
unacceptable
effect.
An
upper
c
h
r
o
n
i
c
l
i
m
i
t
is
t
h
e
l
o
w
e
s
t
t
e
s
t
e
d
c
o
n
c
e
n
t
r
a
t
i
o
n
(
a
)
i
n
a
n
a
c
c
e
p
t
a
b
l
e
c
h
r
o
n
i
c
t
e
s
t
,
(
b)
which
d
i
d
cause
an
unacceptable
amount
of
adverse
e
f
f
e
c
t
on
one
or
more
of
the
specified
biological
measurements,
and
(
c)
above
which
a
l
l
tested
c
o
n
c
e
n
t
r
a
t
i
o
n
s
a
l
s
o
caused
such
an
e
f
f
e
c
t
.

NOTE:
Because
v
a
r
i
o
u
s
a
u
t
h
o
r
s
have
used
a
v
a
r
i
e
t
y
of
terms
and
d
e
f
i
n
i
t
i
o
n
s
t
o
i
n
t
e
r
p
r
e
t
and
r
e
p
o
r
t
r
e
s
u
l
t
s
of
chronic
tests,
reported
r
e
s
u
l
t
s
should
be
reviewed
c
a
r
e
f
u
l
l
y
.
The
amount
of
effect
t
h
a
t
is
considered
unacceptable
is
o
f
t
e
n
based
on
a
s
t
a
t
i
s
t
i
c
a
l
hypothesis
t
e
s
t
,
b
u
t
might
a
l
s
o
be
defined
i
n
terms
of
a
specified
percent
reduction
from
t
h
e
c
o
n
t
r
o
l
s
.
A
small
p
e
r
c
e
n
t
r
e
d
u
c
t
i
o
n
(
e.
g.,
3
p
e
r
c
e
n
t
)
might
be
considered
acceptable
even
i
f
it
i
s
s
t
a
t
i
s
t
i
c
a
l
l
y
s
i
g
n
i
f
i
c
a
n
t
l
y
d
i
f
f
e
r
e
n
t
from
t
h
e
c
o
n
t
r
o
l
,
whereas
a
l
a
r
g
e
p
e
r
c
e
n
t
reduction
(
e.
g.,
30
percent)
might
be
c
o
n
s
i
d
e
r
e
d
u
n
a
c
c
e
p
t
a
b
l
e
e
v
e
n
i
f
it
i
s
n
o
t
s
t
a
t
i
s
t
i
c
a
l
l
y
significant.

G
.
I
f
t
h
e
chronic
toxicity
of
t
h
e
material
t
o
aquatic
animals
apparently
has
been
shown
t
o
be
r
e
l
a
t
e
d
t
o
a
water
quality
c
h
a
r
a
c
t
e
r
i
s
t
i
c
such
a
s
hardness
o
r
p
a
r
t
i
c
u
l
a
t
e
matter
f
o
r
freshwater
animals
o
r
s
a
l
i
n
i
t
y
o
r
p
a
r
t
i
c
u
l
a
t
e
matter
f
o
r
s
a
l
t
w
a
t
e
r
a
n
i
m
a
l
s
,
a
F
i
n
a
l
Chronic
Equation
s
h
o
u
l
d
be
d
e
r
i
v
e
d
b
a
s
e
d
o
n
t
h
a
t
w
a
t
e
r
q
u
a
l
i
t
y
characteristic.
G
o
t
o
Section
V
I
I
.

H.
I
f
chronic
v
a
l
u
e
s
a
r
e
a
v
a
i
l
a
b
l
e
f
o
r
s
p
e
c
i
e
s
i
n
eight
families
a
s
described
i
n
Sections
I
I
I
.
B
.
1
or
I
I
I
.
C
.
1
,
a
Species
Mean
Chronic
Value
(
SMCV)

should
be
calculated
f
o
r
each
species
f
o
r
which
a
t
l
e
a
s
t
one
c
h
r
o
n
i
c
v
a
l
u
e
i
s
a
v
a
i
l
a
b
l
e
by
c
a
l
c
u
l
a
t
i
n
g
t
h
e
geometric
mean
of
a
l
l
chronic
values
a
v
a
i
l
a
b
l
e
f
o
r
t
h
e
species,
and
appropriate
Genus
Mean
Chronic
Values
should
be
c
a
l
c
u
l
a
t
e
d
.

T
h
e
F
i
n
a
l
Chronic
Value
should
t
h
e
n
be
obtained
using
t
h
e
procedure
described
i
n
S
e
c
t
i
o
n
1V.
J­
0
.

Then
go
t
o
Section
V1.
M.

I.
For
each
chronic
v
a
l
u
e
f
o
r
which
a
t
l
e
a
s
t
one
c
o
r
r
e
s
p
o
n
d
i
n
g
a
p
p
r
o
p
r
i
a
t
e
a
c
u
t
e
v
a
l
u
e
i
s
available,
c
a
l
c
u
l
a
t
e
an
acute­
chronic
r
a
t
i
o
,
using
for
t
h
e
numerator
t
h
e
g
e
o
m
e
t
r
i
c
mean
of
t
h
e
r
e
s
u
l
t
s
of
a
l
l
a
c
c
e
p
t
a
b
l
e
f
low­
through
(
except
0
s
t
a
t
i
c
is
acceptable
f
o
r
daphnids)
acute
tests
i
n
the
same
dilution
water
and
in
which
the
concentrations
were
measured.
For
fish,
the
acute
test(
s)
should
have
been
conducted
with
juveniles.

The
acute
test(
s)
should
have
been
part
of
the
same
study
as
the
chronic
test.
If
acute
tests
were
not
conducted
as
part
of
the
same
study,

acute
tests
conducted
in
the
same
laboratory
and
dilution
water,
but
in
a
different
study,
may
be
used.
If
no
such
acute
tests
are
available,

results
of
acute
tests
conducted
in
the
same
dilution
water
in
a
different
laboratory
may
be
used.
If
no
such
acute
tests
are
available,
an
acute­
chronic
ratio
should
not
be
calculated.

J.
For
each
species,
calculate
the
species
mean
acute­
chronic
ratio
as
the
geometric
mean
of
all
acute­
chronic
ratios
available
for
that
species.

K.
For
some
materials
the
acute­
chronic
ratio
seems
to
be
the
same
for
all
species,
but
for
other
materials
the
ratio
seems
to
increase
or
decrease
as
the
Species
Mean
Acute
Value
(
SMAV)
increases.

Thus
the
Final
Acute­
Chronic
Ratio
can
be
obtained
in
four
ways,
depending
on
the
data
available:

1.
If
the
Species
Mean
Acute­
Chronic
ratio
Seems
to
increase
or
decrease
as
the
Species
Mean
Acute
Value
increases,
the
Final
Acute­
Chronic
Ratio
should
be
calculated
as
the
geometric
mean
of
the
acute­
chronic
ratios
for
species
whose
Species
Mean
Acute
Values
are
c
l
o
s
e
to
the
Final
Acute
Value.

2
.
If
no
major
t
r
e
n
d
is
apparent
and
t
h
e
acute­

c
h
r
o
n
i
c
r
a
t
i
o
s
f
o
r
a
number
of
s
p
e
c
i
e
s
a
r
e
w
i
t
h
i
n
a
f
a
c
t
o
r
of
1
0
,
t
h
e
F
i
n
a
l
Acute­

Chronic
Ratio
should
be
c
a
l
c
u
l
a
t
e
d
a
s
t
h
e
geometric
mean
of
a
l
l
the
Species
Mean
Acute­

Chronic
Ratios
a
v
a
i
l
a
b
l
e
for
both
freshwater
and
saltwater
species.

3
.
For
a
c
u
t
e
tests
conducted
on
metals
and
p
o
s
s
i
b
l
y
o
t
h
e
r
substances
w
i
t
h
embryos
and
l
a
r
v
a
e
of
barnacles,
b
i
v
a
l
v
e
m
o
l
l
u
s
c
s
,
sea
urchins,
lobsters,
crabs,
shrimp,
and
abalones
(
see
S
e
c
t
i
o
n
I
V
.
E
.
2
)
,
it
i
s
p
r
o
b
a
b
l
y
a
p
p
r
o
p
r
i
a
t
e
to
assume
t
h
a
t
t
h
e
acute­
chronic
0
r
a
t
i
o
is
2.
Chronic
tests
are
very
d
i
f
f
i
c
u
l
t
t
o
conduct
w
i
t
h
most
such
s
p
e
c
i
e
s
,
b
u
t
it
is
l
i
k
e
l
y
t
h
a
t
the
s
e
n
s
i
t
i
v
i
t
i
e
s
of
embryos
and
l
a
r
v
a
e
would
determine
t
h
e
r
e
s
u
l
t
s
of
l
i
f
e
­

c
y
c
l
e
tests.
Thus,
i
f
t
h
e
l
o
w
e
s
t
a
v
a
i
l
a
b
l
e
Species
Mean
Acute
Values
were
determined
with
embryos
and
larvae
of
such
species,
t
h
e
Final
Acute­
Chronic
Ratio
should
probably
be
assumed
to
be
2
,
so
t
h
a
t
t
h
e
F
i
n
a
l
Chronic
Value
is
equal
to
t
h
e
Criterion
Maximum
Concentration
(
see
Section
X1.
B)

a
*
.
,
~
..?
a.
...,..
4
.
If
the
most
appropriate
Species
Mean
Acute­

Chronic
Ratios
are
less
than
2
.
0
,
and
especially
if
they
are
less
than
1.0,

acclimation
has
probably
occurred
during
the
chronic
test.
Because
continuous
exposure
and
acclimation
cannot
be
assured
to
provide
adequate
protection
in
field
situations,
the
Final
Acute­
Chronic
Ratio
should
be
assumed
to
be
2
,
so
that
the
Final
Chronic
Value
is
equal
to
the
Criterion
Maximum
Concentration
(
see
Section
X1.
B).

If
the
available
Species
Mean
Acute­
Chronic
Ratios
do
not
fit
one
of
these
cases,
a
Final
Acute­
Chronic
Ratio
probably
cannot
be
obtained,
and
a
Final
Chronic
Value
probably
cannot
be
calculated.

L.
Calculate
the
Final
Chronic
Value
by
dividing
the
Final
Acute
Value
by
the
Final
Acute­
Chronic
Ratio.
If
there
was
a
Final
Acute
Equation
rather
than
a
Final
Acute
Value,
see
also
Section
VI1.
A.

M.
If
the
Species
Mean
Chronic
Value
of
a
commercially
or
recreational
ly
important
species
is
lower
than
the
calculated
Final
Chronic
Value,

then
that
species
Mean
Chronic
Value
should
be
used
as
the
Final
Chronic
Value
instead
of
the
calculated
Final
Chronic
Value.

N.
Go
to
Section
VIII.
VI.
Final
Chronic
Equation
0
­
A.
A
Final
Chronic
Equation
can
be
derived
in
two
ways.
The
procedure
described
here
in
Section
A
will
result
in
the
chronic
slope
being
the
same
as
the
acute
slope.
The
procedure
described
in
Sections
B­
N
usually
will
result
in
the
chronic
slope
being
different
from
the
acute
slope.

1.
If
acute­
chronic
ratios
are
available
 or
enough
species
at
enough
values
of
the
water
quality
characteristic
to
indicate
that
the
acute­
chronic
ratio
is
probably
the
same
for
all
species
and
is
probably
independent
of
the
water
quality
characteristic,
calculate
the
Final
Acute­
Chronic
Ratio
as
the
geometric
mean
of
the
available
Species
Mean
Acute­

Chronic
Ratios.
'

2
.
Calculate
the
Final
Chronic
Value
at
the
selected
value
Z
of
the
water
quality
characteristic
by
dividing
the
Final
Acute
Value
at
Z
(
see
Section
V.
M)
by
the
Final
Acute­
Chronic
Ratio.

3
.
Use
V
=
pooled
acute
slope
(
see
section
V.
M)

as
L
=
pooled
chronic
slope.

4
.
Go
to
Section
VI1.
M.

B.
When
enough
data
are
available
to
show
that
chronic
toxicity
to
at
least
one
species
is
related
to
a
water
quality
characteristic,
the
relationship
should
be
taken
into
account
as
described
in
Sections
B­
G
or
using
analysis
of
covariance
[
15,16].
The
two
methods
are
equivalent
and
produce
identical
results.
The
manual
method
described
below
provides
an
understanding
of
this
application
of
covariance
analysis,
but
computerized
versions
of
covariance
analysis
are
much
more
convenient
for
analyzing
large
data
sets.
If
two
or
more
factors
affect
toxicity,
multiple
regression
analysis
should
be
used.

For
each
species
for
which
comparable
chronic
toxicity
values
are
available
at
two
or
more
different
values
of
the
water
quality
characteristic,
perform
a
least
squares
regression
of
the
chronic
toxicity
values
on
the
corresponding
values
of
the
water
quality
characteristic
to
obtain
the
slope
and
its
95
percent
confidence
limits
for
each
species.

NOTE:
Because
the
best
documented
relationship
fitting
these
data
is
that
between
hardness
and
acute
toxicity
of
metals
in
freshwater
and
a
log­
log
relationship,
geometric
means
and
natural
logarithms
of
both
toxicity
and
water
quality
are
used
in
the
rest
of
this
section.
For
relationships
based
on
other
water
quality
characteristics
such
as
pH,
temperature,
or
s
a
l
i
n
i
t
y
,
no
t
r
a
n
s
f
o
r
m
a
t
i
o
n
or
a
d
i
f
f
e
r
e
n
t
transformation
might
f
i
t
t
h
e
d
a
t
a
better,
and
appropriate
changes
w
i
l
l
be
necessary
throughout
t
h
i
s
section.
f
t
is
probably
preferable,
but
not
necessary,
to
use
t
h
e
same
transformation
t
h
a
t
was
u
s
e
d
w
i
t
h
t
h
e
a
c
u
t
e
v
a
l
u
e
s
i
n
s
e
c
t
i
o
n
v
.

D.
Decide
whether
t
h
e
d
a
t
a
f
o
r
each
s
p
e
c
i
e
s
a
r
e
u
s
e
f
u
l
,
t
a
k
i
n
g
i
n
t
o
account
t
h
e
range
and
number
of
t
h
e
t
e
s
t
e
d
v
a
l
u
e
s
of
t
h
e
w
a
t
e
r
q
u
a
l
i
t
y
c
h
a
r
a
c
t
e
r
i
s
t
i
c
and
the
degree
of
agreement
w
i
t
h
i
n
and
between
species.
For
example,
a
s
l
o
p
e
based
on
s
i
x
d
a
t
a
p
o
i
n
t
s
m
i
g
h
t
b
e
o
f
l
i
m
i
t
e
d
v
a
l
u
e
i
f
it
is
based
only
on
data
 or
a
v
e
r
y
narrow
range
of
v
a
l
u
e
s
of
t
h
e
water
q
u
a
l
i
t
y
c
h
a
r
a
c
t
e
r
i
s
t
i
c
,
A
s
l
o
p
e
based
on
o
n
l
y
two
d
a
t
a
p
o
i
n
t
s
,
however,

might
be
u
s
e
f
u
l
i
f
it
is
c
o
n
s
i
s
t
e
n
t
w
i
t
h
o
t
h
e
r
information
and
i
f
t
h
e
two
p
o
i
n
t
s
cover
a
broad
enough
range
of
the
water
q
u
a
l
i
t
y
characteristic.

I
n
a
d
d
i
t
i
o
n
,
chronic
v
a
l
u
e
s
t
h
a
t
appear
to
be
questionable
i
n
comparison
w
i
t
h
o
t
h
e
r
a
c
u
t
e
and
chronic
d
a
t
a
a
v
a
i
l
a
b
l
e
 or
t
h
e
same
species
and
f
o
r
o
t
h
e
r
s
p
e
c
i
e
s
i
n
t
h
e
same
genus
probably
s
h
o
u
l
d
n
o
t
be
used.
For
example,
i
f
a
f
t
e
r
adjustment
f
o
r
the
water
qua1
i
t
y
c
h
a
r
a
c
t
e
r
i
s
t
i
c
,

t
h
e
c
h
r
o
n
i
c
v
a
l
u
e
s
a
v
a
i
l
a
b
l
e
f
o
r
a
s
p
e
c
i
e
s
or
genus
d
i
f
f
e
r
by
more
than
a
factor
of
10,
probably
some
or
a
l
l
of
the
values
should
be
rejected.
I
f
a
useful
chronic
slope
is
not
available
for
at
least
one
species
or
if
the
available
slopes
are
too
dissimilar
or
if
too
few
data
are
available
to
adequately
define
the
relationship
between
chronic
toxicity
and
the
water
quality
characteristic,
it
might
be
appropriate
to
assume
that
the
chronic
slope
is
the
same
as
the
acute
slope,
which
is
equivalent
to
assuming
that
the
acute­
chronic
ratio
is
independent
of
the
water
quality
characteristic.
Alternatively,
return
to
Section
VI.
H,
using
the
results
of
tests
conducted
under
conditions
and
in
waters
similar
to
those
commonly
used
for
toxicity
tests
with
the
species.

E.
Individually
for
each
species
calculate
the
geometric
mean
of
the
available
chronic
values
and
then
divide
each
chronic
value
for
a
species
by
the
mean
for
the
species.
This
normalizes
the
chronic
values
so
that
the
geometric
mean
of
the
normalized
values
for
each
species
individually
and
for
any
combination
of
species
is
1.0.

F.
Similarly
normalize
the
values
of
the
water
quality
characteristic
for
each
species
individually.

G
.
Individually
for
each
species
perform
a
least
squares
regression
of
the
normalized
chronic
toxicity
values
on
the
corresponding
normalized
values
of
the
water
quality
characteristic.
The
resulting
slopes
and
the
95
percent
confidence
limits
will
be
identical
to
those
obtained
in
Section
B.
Now,
however,
if
the
data
are
actually
plotted,
the
line
of
best
fit
for
each
individual
species
will
go
through
the
point
1,1
in
the
center
of
the
graph.

K.
Treat
all
the
normalized
data
as
if
they
were
all
for
the
same
species
and
perform
a
least
squares
regression
of
all
the
normalized
chronic
values
on
the
corresponding
normalized
values
of
the
water
quality
characteristic
to
obtain
the
pooled
chronic
slope,
L,
and
its
95
percent
confidence
limits,
If
all
the
normalized
data
are
actually
plotted,
the
line
of
best
fit
will
go
through
the
point
1,
l
in
the
center
of
the
graph.

I.
For
each
species
calculate
the
geometric
mean,
M
I
0
of
the
toxicity
values
and
the
geometric
mean,
PI
of
the
values
of
the
water
quality
characteristic.

(
These
were
calculated
in
steps
E
and
F.)
J.

NOTE
:

K.

NOTE
:

L.

M.
For
each
s
p
e
c
i
e
s
c
a
l
c
u
l
a
t
e
t
h
e
logarithm,
Q,
of
t
h
e
Species
Mean
Chronic
Value
a
t
a
selected
v
a
l
u
e
,
Z,
of
t
h
e
water
q
u
a
l
i
t
y
characteristic
using
t
h
e
equation:
Q
=
I
n
M
­
L(
ln
P
­
I
n
Z).

Although
it
is
n
o
t
necessary,
it
w
i
l
l
u
s
u
a
l
l
y
be
best
t
o
use
t
h
e
same
v
a
l
u
e
of
t
h
e
water
q
u
a
l
i
t
y
characteristic
here
as
was
used
i
n
section
V.
I.

For
each
species
c
a
l
c
u
l
a
t
e
a
Species
Mean
Chronic
Value
a
t
z
using
t
h
e
equation:
SMCV
=
eQ.

Alternatively,
t
h
e
Species
Mean
Chronic
Values
a
t
Z
can
be
obtained
by
skipping
s
t
e
p
J,
u
s
i
n
g
t
h
e
equations
i
n
s
t
e
p
s
J
and
K
t
o
a
d
j
u
s
t
each
a
c
u
t
e
value
individually
t
o
2
,
and
then
c
a
l
c
u
l
a
t
i
n
g
the
geometric
means
of
t
h
e
a
d
j
u
s
t
e
d
v
a
l
u
e
s
f
o
r
each
species
individually.
T
h
i
s
a
l
t
e
r
n
a
t
i
v
e
procedure
allows
an
examination
of
the
range
of
the
adjusted
chronic
values
for
each
species.

Obtain
the
Final
Chronic
Value
a
t
2
by
using
the
procedure
described
i
n
Section
1V.
J­
0.

I
f
t
h
e
Species
Mean
Chronic
Value
a
t
Z
of
a
commercially
o
r
r
e
c
r
e
a
t
i
o
n
a
l
l
y
important
species
is
lower
than
t
h
e
calculated
Final
Chronic
Value
a
t
Z,
then
t
h
a
t
Species
Mean
Chronic
Value
should
b
e
u
s
e
d
as
t
h
e
F
i
n
a
l
C
h
r
o
n
i
c
V
a
l
u
e
a
t
Z
i
n
s
t
e
a
d
o
f
the
calculated
Final
Chronic
Value.
N
.
The
Final
Chronic
Equation
i
S
written
as:
Final
C
h
r
o
n
i
c
V
a
l
u
e
=
e
(
L
[
l
n
(
w
a
t
e
r
q
u
a
l
i
t
y
c
h
a
r
a
c
t
e
r
i
s
t
i
c
)
]
+
I
n
S
­
L
[
l
n
Z
]
)
,
where
L
=

pooled
chronic
slope
and
S
=
Final
Chronic
Value
a
t
2.
Because
L,
S
and
Z
a
r
e
known,
t
h
e
F
i
n
a
l
Chronic
Value
can
be
calculated
f
o
r
any
selected
value
of
the
water
q
u
a
l
i
t
y
characteristic.

Final
Plant
Value
A.
Appropriate
measures
of
t
h
e
t
o
x
i
c
i
t
y
of
t
h
e
material
to
aquatic
p
l
a
n
t
s
a
r
e
used
to
compare
the
r
e
l
a
t
i
v
e
s
e
n
s
i
t
i
v
i
t
i
e
s
of
a
q
u
a
t
i
c
p
l
a
n
t
s
and
animals.
Although
procedures
for
conducting
and
i
n
t
e
r
p
r
e
t
i
n
g
the
r
e
s
u
l
t
s
of
t
o
x
i
c
i
t
y
tests
w
i
t
h
p
l
a
n
t
s
are
n
o
t
w
e
l
l
developed,
r
e
s
u
l
t
s
of
tests
with
p
l
a
n
t
s
usually
indicate
t
h
a
t
c
r
i
t
e
r
i
a
which
adequately
protect
aquatic
animals
and
their
uses
w
i
l
l
probably
a
l
s
o
p
r
o
t
e
c
t
a
q
u
a
t
i
c
p
l
a
n
t
s
and
their
uses.

B.
A
p
l
a
n
t
v
a
l
u
e
i
s
the
r
e
s
u
l
t
of
a
96­
hr
t
e
s
t
conducted
w
i
t
h
an
a
l
g
a
o
r
a
chronic
t
e
s
t
conducted
w
i
t
h
an
aquatic
vascular
p
l
a
n
t
.

NOTE:
A
t
e
s
t
of
t
h
e
t
o
x
i
c
i
t
y
of
a
metal
to
a
p
l
a
n
t
u
s
u
a
l
l
y
should
not
be
used
i
f
the
medium
contained
an
excessive
amount
of
a
complexing
agent,
such
as
EDTA,
t
h
a
t
might
a
f
f
e
c
t
the
t
o
x
i
c
i
t
y
of
t
h
e
metal.

Concentrations
of
EDTA
above
about
200
ug/
L
should
probably
be
considered
excessive.
C.
T
h
e
F
i
n
a
l
P
l
a
n
t
Value
should
be
obtained
by
s
e
l
e
c
t
i
n
g
t
h
e
lowest
r
e
s
u
l
t
from
a
t
e
s
t
w
i
t
h
an
important
a
q
u
a
t
i
c
p
l
a
n
t
species
i
n
which
t
h
e
concentrations
of
t
e
s
t
material
were
measured
and
the
endpoint
was
b
i
o
l
o
g
i
c
a
l
l
y
important.

V
I
I
I
.
Final
Residue
Value
A.
The
Final
Residue
Value
is
intended
t
o
(
a)
prevent
concentrations
i
n
commercially
or
r
e
c
r
e
a
t
i
o
n
a
l
l
y
i
m
p
o
r
t
a
n
t
a
q
u
a
t
i
c
s
p
e
c
i
e
s
f
r
o
m
a
f
f
e
c
t
i
n
g
marketability
because
of
exceedence
of
applicable
FDA
a
c
t
i
o
n
l
e
v
e
l
s
and
(
b)
p
r
o
t
e
c
t
w
i
l
d
l
i
f
e
,

i
n
c
l
u
d
i
n
g
f
i
s
h
e
s
and
b
i
r
d
s
,
t
h
a
t
consume
a
q
u
a
t
i
c
organisms
from
demonstrated
unacceptable
effects.

The
F
i
n
a
l
Residue
Value
i
s
t
h
e
l
o
w
e
s
t
of
t
h
e
r
e
s
i
d
u
e
v
a
l
u
e
s
t
h
a
t
are
obtained
by
d
i
v
i
d
i
n
g
maximum
permissible
t
i
s
s
u
e
c
o
n
c
e
n
t
r
a
t
i
o
n
s
by
a
p
p
r
o
p
r
i
a
t
e
b
i
o
c
o
n
c
a
n
t
r
a
t
i
o
n
or
bioaccumulation
f
a
c
t
o
r
s
.
A
maximum
p
e
r
m
i
s
s
i
b
l
e
t
i
s
s
u
e
concentration
i
s
either
(
a)
an
FDA
a
c
t
i
o
n
l
e
v
e
l
[
12]
f
o
r
f
i
s
h
o
i
l
o
r
f
o
r
t
h
e
e
d
i
b
l
e
p
o
r
t
i
o
n
of
f
i
s
h
or
s
h
e
l
l
f
i
s
h
,
or
(
b)
a
maximum
a
c
c
e
p
t
a
b
l
e
dietary
intake
based
on
observations
on
s
u
r
v
i
v
a
l
,

growth,
or
reproduction
i
n
a
c
h
r
o
n
i
c
w
i
l
d
l
i
f
e
feeding
study
or
a
long­
term
w
i
l
d
l
i
f
e
f
i
e
l
d
study.

I
f
no
maximum
permissible
t
i
s
s
u
e
concentration
is
a
v
a
i
l
a
b
l
e
,
go
to
S
e
c
t
i
o
n
X
b
e
c
a
u
s
e
no
F
i
n
a
l
Residue
Value
can
be
derived.
B.
~
i
o
c
o
n
c
e
n
t
r
a
t
i
o
n
actors
(
B
C
F
S
)
a
n
d
bioaccumulation
f
a
c
t
o
r
s
(
BAFs)
a
r
e
q
u
o
t
i
e
n
t
s
of
t
h
e
concentration
of
a
m
a
t
e
r
i
a
l
i
n
one
o
r
more
t
i
s
s
u
e
s
of
an
a
q
u
a
t
i
c
organism
d
i
v
i
d
e
d
by
t
h
e
average
concentration
i
n
the
solution
i
n
which
t
h
e
organism
had
been
l
i
v
i
n
g
.
A
BCF
i
s
intended
to
account
Only
f
o
r
n
e
t
uptake
d
i
r
e
c
t
l
y
from
water,

and
t
h
u
s
almosthastobemeasured
i
n
a
l
a
b
o
r
a
t
o
r
y
test.
Some
uptake
during
t
h
e
bioconcentration
test
might
n
o
t
b
e
d
i
r
e
c
t
l
y
fromwater
if
t
h
e
food
sorbs
some
of
t
h
e
t
e
s
t
material
before
it
is
e
a
t
e
n
by
t
h
e
t
e
s
t
organisms.
A
BAF
i
s
intended
to
account
f
o
r
n
e
t
uptake
from
both
food
and
water
i
n
a
r
e
a
l
­

world
situation.
A
BAF
almost
has
to
be
measured
i
n
a
f
i
e
l
d
situation
i
n
which
predators
accumulate
the
material
d
i
r
e
c
t
l
y
from
water
and
by
consuming
prey
that
itself
could
have
accumulated
t
h
e
m
a
t
e
r
i
a
l
from
both
food
and
water.
The
BCF
and
BAF
a
r
e
probably
similar
f
o
r
a
material
w
i
t
h
a
low
BCF,
b
u
t
t
h
e
BAF
is
probably
higher
t
h
a
n
t
h
e
BCF
f
o
r
m
a
t
e
r
i
a
l
s
w
i
t
h
high
BCFs.
Although
BCFs
are
not
too
d
i
f
f
i
c
u
l
t
to
determine,
very
f
e
w
BAFs
have
been
measured
acceptably
because
it
is
necessary
to
make
enough
measurements
of
t
h
e
concentration
of
t
h
e
m
a
t
e
r
i
a
l
i
n
water
to
show
t
h
a
t
it
was
reasonably
constant
f
o
r
a
l
o
n
g
enough
period
of
t
i
m
e
over
the
range
of
t
e
r
r
i
t
o
r
y
inhabited
by
the
organisms.
Because
so
few
acceptable
BAFs
are
available,
only
BCFs
will
be
discussed
further.

However,
if
an
acceptable
BAF
is
available
for
a
material,
it
should
be
used
instead
of
any
available
BCFs.

C.
If
a
maximum
permissible
tissue
concentration
is
available
for
a
substance
(
e.
g.,
parent
material,

parent
material
plus
metabolites,
etc.)
,
the
tissue
concentration
used
in
the
calculation
of
the
BCF
should
be
for
the
same
substance.

otherwise
the
tissue
concentration
used
in
the
calculation
of
the
BCF
should
be'
that
of
the
material
and
its
metabolites
which
are
structurally
similar
and
are
not
much
more
soluble
in
water
than
the
parent
material.

D.
1.
A
BCF
should
be
used
only
if
the
test
was
flow­
through,
the
BCF
was
calculated
based
on
measured
concentrations
of
the
test
material
in
tissue
and
in
the
test
solution,
and
the
exposure
continued
at
least
until
either
apparent
steady­
state
or
2
8
days
was
reached.

Steady­
state
is
reached
when
the
BCF
does
not
change
significantly
over
a
period
of
time,

such
a
2
days
o
r
16
percent
of
the
length
of
the
exposure,
whichever
is
longer.
The
BCF
used
from
a
test
should
be
the
highest
of
(
a)

the
apparent
steady­
state
BCF,
if
apparent
steady­
state
was
reached,
(
b)
the
highest
BCF
obtained,
if
apparent
steady­
state
was
not
reached,
and
(
c)
the
projected
steady­
state
BCF,
if
calculated.

2
.
Whenever
a
BCF
is
determined
for
a
lipophilic
material,
the
percent
lipids
should
also
be
determined
in
the
tissue(
s)
for
which
the
BCF
was
calculated.

3
.
A
BCF
obtained
from
an
exposure
that
adversely
affected
the
test
organisms
may
be
used
only
if
it
is
similar
to
a
BCF
obtained
with
unaffected
organisms
of
the
same
species
at
lower
concentrations
that
did
not
cause
adverse
effects.

4
.
Because
maximum
permissible
tissue
concentrations
are
almost
never
based
on
dry
weights,
a
BCF
calculated
using
dry
tissue
weights
must
be
converted
to
a
wet
tissue
weight
basis.
If
no
conversion
factor
is
reported
with
the
BCF,
multiply
the
dry
weight
BCF
by
0.1
for
plankton
and
by
0
.
2
for
individual
species
of
fishes
and
invertebrates
~
1
7
1
.

5.
If
more
than
one
acceptable
BCF
is
available
for
a
species,
the
geometric
mean
of
the
available
values
should
be
used,
except
that
if
the
BCFs
are
from
different
lengths
of
exposure
and
the
BCF
increases
with
length
of
exposure,
the
BCF
f
o
r
t
h
e
l
o
n
g
e
s
t
exposure
should
be
used.

E.
I
f
enough
p
e
r
t
i
n
e
n
t
data
exist,
several
residue
v
a
l
u
e
s
can
be
c
a
l
c
u
l
a
t
e
d
by
d
i
v
i
d
i
n
g
maximum
p
e
r
m
i
s
s
i
b
l
e
t
i
s
s
u
e
concentrations
by
appropriate
BCFs
:

1.
For
each
a
v
a
i
l
a
b
l
e
maximum
acceptable
d
i
e
t
a
r
y
intake
derived
from
a
chronic
feeding
study
o
r
a
long­
t
e
r
m
f
i
e
l
d
s
t
u
d
y
w
i
t
h
w
i
l
d
l
i
f
e
,

i
n
c
l
u
d
i
n
g
b
i
r
d
s
and
a
q
u
a
t
i
c
organisms,
t
h
e
appropriate
BCF
is
based
on
t
h
e
whole
body
of
aquatic
species
which
constitute'
or
represent
a
major
p
o
r
t
i
o
n
of
t
h
e
d
i
e
t
of
t
h
e
t
e
s
t
e
d
w
i
l
d
l
i
f
e
species.

2
.
For
an
FDA
action
l
e
v
e
l
f
o
r
f
i
s
h
o
r
s
h
e
l
l
f
i
s
h
,

t
h
e
a
p
p
r
o
p
r
i
a
t
e
BCF
is
t
h
e
h
i
g
h
e
s
t
geometric
mean
s
p
e
c
i
e
s
BCF
f
o
r
t
h
e
e
d
i
b
l
e
p
o
r
t
i
o
n
(
muscle
f
o
r
decapods,
muscle
with
o
r
without
skin
f
o
r
f
i
s
h
e
s
,
adductor
muscle
f
o
r
s
c
a
l
l
o
p
s
,

and
t
o
t
a
l
s
o
f
t
t
i
s
s
u
e
f
o
r
o
t
h
e
r
b
i
v
a
l
v
e
molluscs)
of
a
consumed
species.
The
highest
species
BCF
is
used
because
FDA
action
l
e
v
e
l
s
a
r
e
applied
on
a
species­
by­
species
basis.

F.
For
l
i
p
o
p
h
i
l
i
c
materials,
it
might
be
possible
t
o
c
a
l
c
u
l
a
t
e
additional
residue
values.
Because
the
steady­
state
BCF
f
o
r
a
l
i
p
o
p
h
i
l
i
c
material
Seems
t
o
be
p
r
o
p
o
r
t
i
o
n
a
l
t
o
p
e
r
c
e
n
t
l
i
p
i
d
s
from
one
t
i
s
s
u
e
t
o
another
and
from
one
species
t
o
another
[
18­
20],
extrapolations
can
be
made
from
tested
tissues
or
species
to
untested
tissues
or
species
an
the
basis
of
percent
lipids.

1.
For
each
BCF
for
which
the
percent
lipids
is
known
for
the
same
tissue
for
which
the
BCF
was
measured,
normalize
the
BCF
to
a
1
percent
lipid
basis
by
dividing
the
BCF
by
the
percent
lipids.
This
adjustment
to
a
1
percent
lipid
basis
is
intended
to
make
all
the
measured
BCFs
far
a
material
comparable
regardless
of
the
species
or
tissue'with
which
the
BCF
was
measured.

2
.
calculate
the
geometric
mean
normalized
BCF.

Data
for
both
saltwater
and
freshwater
species
should
be
used
to
determine
the
mean
normalized
BCF,
unless
the
data
show
that
the
normalized
BCFs
are
probably
not
similar,

3
.
Calculate
all
possible
residue
values
by
dividing
the
available
maximum
permissible
tissue
concentrations
by
the
mean
normalized
BCF
and
by
the
percent
lipids
values
appropriate
to
the
maximum
permissible
tissue
concentrations,
i.
e.,

(
maximum
permissible
tissue
concentration)
Residue
value
=
(
mean
normalized
BCF)(
appropriate
percent
lipids)

tissue
concentration)
Residue
value
=
(
mean
normalized
BCF)
(
appropriate
percent
lipids)

a.
For
an
FDA
action
level
for
fish
oil,
the
appropriate
percent
lipids
value
is
100.

b.
For
an
FDA
action
level
for
fish,
the
appropriate
percent
lipids
value
is
11
for
freshwater
criteria
and
10
for
saltwater
criteria
because
FDA
action
levels
are
applied
on
a
species­
by­

species
basis
to
commonly
consumed
species.
The
highest
lipid
contents
in
the
edible
portions
of
important
consumed
species
are
about
11
percent
for
both
the
freshwater
chinook
salmon
and
lake
trout
and
about
10
percent
for
the
saltwater
Atlantic
herring
[
21].

c.
For
a
maximum
acceptable
dietary
intake
derived
froma
chronic
feeding
studyora
long­
term
field
study
with
wildlife,
the
appropriate
percent
lipids
is
that
of
an
aquatic
species
o
r
group
of
aquatic
species
which
constitute
a
major
portion
of
the
diet
of
the
wildlife
species.

G.
The
Final
Residue
Value
is
obtained
by
selecting
the
lowest
of
the
available
residue
values.
NOTE:
In
some
cases
the
Final
Residue
Value
will
not
be
l
o
w
enough.
For
example,
a
residue
value
calculated
from
an
FDA
action
level
will
probably
result
in
an
average
concentration
in
the
edible
portion
of
a
fatty
species
that
is
at
the
action
level.
Some
individual
organisms,
and
possibly
some
species,
will
have
residue
concentrations
higher
than
the
mean
value
but
no
mechanism
has
been
devised
to
provide
appropriate
additional
protection.
Also,
some
chronic
feeding
studies
and
long­
term
field
studies
with
wildlife
identify
concentrations
that
cause
adverse
effects
but
do
not
identify
concentrations
which
do
not
cause
adverse
effects;
again,
no
mechanism
has
been
devised
to
provide
appropriate
additional
protection.
These
are
some
of
the
species
and
uses
that
are
not
protected
at
all
times
in
all
places.

­
x.
other
Data
Pertinent
information
that
could
not
be
used
in
earlier
sections
might
be
available
concerning
adverse
effects
on
aquatic
organisms
and
their
uses.
The
most
important
of
these
are
data
on
cumulative
and
delayed
toxicity,
flavor
impairment,
reduction
in
survival,

growth,
or
reproduction,
or
any
other
adverse
effect
that
has
been
shown
to
be
biologically
important.

Especially
important
are
data
for
species
for
which
no
~~
other
data
are
available.
Data
from
behavioral,

biochemical,
physiological,
microcosm,
and
field
studies
might
also
be
available.
Data
might
be
available
from
tests
conducted
in
unusual
dilution
water
(
see
1V.
D
and
VI.
D),
from
chronic
tests
in
which
the
concentrations
were
not
measured
(
see
VI.
B),
from
tests
with
previously
exposed
organisms
(
see
1I.
F)
,

and
from
tests
on
formulated
mixtures
or
emulsifiable
concentrates
(
see
1I.
D).
Such
data
might
affect
a
criterion
if
the
data
were
obtained
with
an
important
species,
the
test
concentrations
were
measured,
and
the
endpoint
was
biologically
important.
'

XI.
Criterion
_.

A.
A
criterion
consists
of
two
concentrations:
the
Criterion
Maximum
Concentration
and
the
Criterion
Continuous
Concentration.

B.
The
Criterion
Maximum
Concentration
(
CMC)
is
equal
to
one­
half
the
Final
Acute
Value.

C.
The
Criterion
Continuous
Concentration
(
CCC)
is
equal
to
the
lowest
of
the
Final
Chronic
Value,

the
Final
Plant
Value,
and
the
Final
Residue
Value,
unless
other
data
(
see
Section
X)
show
that
a
lower
value
should
be
used.
If
toxicity
is
related
to
a
water
quality
characteristic,
the
Criterion
continuous
concentration
is
obtained
from
the
Final
Chronic
Equation,
the
Final
Plant
Value,
and
the
Final
Residue
Value
by
selecting
the
one,
or
the
combination,
that
results
in
the
lowest
concentrations
in
the
usual
range
of
the
water
quality
characteristic,
unless
other
data
(
see
Section
X)
show
that
a
lower
value
should
be
used.

D.
Round
[
14]
both
the
Criterion
Maximum
Concentration
and
the
Criterion
Continuous
Concentration
to
two
significant
digits.

E.
The
criterion
is
stated
as:

The
procedures
described
in
the
Guidelines
for
Deriving
Numerical
National
Water
Quality
Criteria
for
the
Protection
of
Aquatic
Organisms
and
Their
Uses
indicate
that,
except
possibly
where
a
locally
important
species
is
very
sensitive,
(
1)

aquatic
organisms
and
their
uses
should
not
be
0
affected
unacceptably
if
the
4­
day
average
concentration
of
(
2)
does
not
exceed
(
3)
ug/
L
more
than
once
every
3
years
on
the
average
and
if
the
1­
hour
average
concentration
does
not
exceed
(
4)

ug/
L
more
than
once
every
3
years
on
the
average.

where
(
1)
=
insert
nlfreshwaternl
or
nrsaltwaternl
(
2)
=
insert
name
of
material
(
3)
=
insert
the
Criterion
Continuous
Concentration
(
4)
=
insert
the
Criterion
Maximum
Concentration.
Final
R
e
v
i
e
v
A.
T
h
e
d
e
r
i
v
a
t
i
o
n
o
f
t
h
e
c
r
i
t
e
r
i
o
n
s
h
o
u
l
d
be
c
a
r
e
f
u
l
l
y
reviewed
by
rechecking
each
s
t
e
p
of
t
h
e
Guidelines.
I
t
e
m
s
t
h
a
t
s
h
o
u
l
d
be
e
s
p
e
c
i
a
l
l
y
checked
are:

1.
I
f
unpublished
d
a
t
a
a
r
e
used,
are
t
h
e
y
w
e
l
l
documented?

A
r
e
a
l
l
required
data
a
v
a
i
l
a
b
l
e
?
2
.

3.
Is
the
range
of
acute
v
a
l
u
e
s
f
o
r
any
species
greater
than
a
f
a
c
t
o
r
of
l
o
?

4
.
Is
the
range
of
Species
Mean
Acute
Values
for
any
genus
g
r
e
a
t
e
r
than
a
f
a
c
t
o
r
of
lo?

5
.
Is
t
h
e
r
e
more
t
h
a
n
a
f
a
c
t
o
r
of
10
difference
between
t
h
e
'
f
o
u
r
lowest
Genus
Mean
Acute
Values?

6
.
A
r
e
any
of
t
h
e
f
o
u
r
lowest
Genus
Mean
Acute
Valuesquestionable?

7
.
Is
the
F
i
n
a
l
Acute
Value
reasonable
i
n
comparison
with
t
h
e
Species
Mean
Acute
Values
and
Genus
Mean
Acute
Values?

8.
For
any
c
o
m
m
e
r
c
i
a
l
l
y
or
r
e
c
r
e
a
t
i
o
n
a
l
l
y
important
s
p
e
c
i
e
s
,
is
t
h
e
geometric
mean
of
t
h
e
a
c
u
t
e
v
a
l
u
e
s
from
flow­
through
tests
i
n
which
t
h
e
concentrations
of
t
e
s
t
material
were
measured
lower
than
the
F
i
n
a
l
Acute
Value?

9.
Are
any
of
the
chronic
values
questionable?

0
10.

11.

12
13.

14
*

15.

16.
Are
chronic
values
available
for
acutely
sensitive
species?

Is
the
range
of
acute­
chronic
ratios
greater
than
a
factor
of
10?

Is
the
Final
Chronic
Value
reasonable
in
comparison
with
the
available
acute
and
chronic
data?

Is
the
measured
or
predicted
chronic
value
for
any
commercially
or
recreational
ly
important
species
below
the
Final
Chronic
Value?

Are
any
of
the
other
data
important?

Do
any
data
look
like
they
might
be
outliers?

Are
there
any
deviations
from
the
Guidelines?

Are
they
acceptable?

B.
On
the
basis
of
a
l
l
available
pertinent
laboratory
and
field
information,
determine
if
the
criterion
is
consistent
with
sound
scientific
evidence.
If
it
is
not,
another
criterion,
either
higher
or
lower,
should
be
derived
using
appropriate
modifications
of
these
Guidelines.
APPENDIX
B
SUMHARY
OF
THE
1980
AQUATIC
LIFE
GUIDELINES
The
Guidelines
f
o
r
Deriving
Water
Q
u
a
l
i
t
y
C
r
i
t
e
r
i
a
f
o
r
t
h
e
P
r
o
t
e
c
t
i
o
n
of
Aquatic
L
i
f
e
and
i
t
s
U
s
e
s
were
developed
t
o
describe
an
objective,
i
n
t
e
r
n
a
l
l
y
consistent,
and
appropriate
way
of
ensuring
t
h
a
t
water
q
u
a
l
i
t
y
c
r
i
t
e
r
i
a
f
o
r
a
q
u
a
t
i
c
l
i
f
e
would
provide,
on
t
h
e
average,
a
reasonable
amount
of
protection.
The
r
e
s
u
l
t
i
n
g
c
r
i
t
e
r
i
a
are
n
o
t
intended
t
o
provide
100
percent
protection
of
a
l
l
species
and
a
l
l
uses
of
aquatic
l
i
f
e
a
l
l
of
the
t
i
m
e
,
b
u
t
t
h
e
y
a
r
e
i
n
t
e
n
d
e
d
to
p
r
o
t
e
c
t
most
species
i
n
a
balanced,
healthy
aquatic
community.

Minimum
data
requirements
are
i
d
e
n
t
i
f
i
e
d
i
n
f
o
u
r
a
r
e
a
s
;

acute
t
o
x
i
c
i
t
y
to
animals
(
eight
data
points),
chronic
t
o
x
i
c
i
t
y
t
o
animals
(
three
data
points),
t
o
x
i
c
i
t
y
t
o
p
l
a
n
t
s
,
and
residues.

Data
on
a
c
u
t
e
t
o
x
i
c
i
t
y
a
r
e
needed
f
o
r
a
v
a
r
i
e
t
y
of
f
i
s
h
and
invertebrate
species
and
a
r
e
used
t
o
derive
a
Final
Acute
Value.

By
taking
i
n
t
o
account
t
h
e
number
and
r
e
l
a
t
i
v
e
s
e
n
s
i
t
i
v
i
t
i
e
s
of
t
h
e
tested
species,
the
Final
Acute
Value
is
designed
t
o
protect
most,
b
u
t
n
o
t
n
e
c
e
s
s
a
r
i
l
y
a
l
l
,
of
t
h
e
tested
and
u
n
t
e
s
t
e
d
species.

Data
on
chronic
t
o
x
i
c
i
t
y
to
animals
can
be
used
to
d
e
r
i
v
e
a
F
i
n
a
l
Chronic
V
a
l
u
e
by
two
d
i
f
f
e
r
e
n
t
means.
If
c
h
r
o
n
i
c
v
a
l
u
e
s
a
r
e
a
v
a
i
l
a
b
l
e
f
o
r
a
s
p
e
c
i
f
i
e
d
number
and
a
r
r
a
y
of
s
p
e
c
i
e
s
,
a
F
i
n
a
l
Chronic
Value
can
be
c
a
l
c
u
l
a
t
e
d
d
i
r
e
c
t
l
y
.
I
f
n
o
t
,
an
acute­
chronic
ratio
is
derived
and
then
used
with
the
Final
Acute
Value
to
obtain
t
h
e
Final
Chronic
Value.

The
F
i
n
a
l
P
l
a
n
t
Value
is
obtained
by
s
e
l
e
c
t
i
n
g
t
h
e
lowest
p
l
a
n
t
t
o
x
i
c
i
t
y
value
based
on
measured
concentrations.
The
Final
Residue
Value
is
intended
t
o
protect
w
i
l
d
l
i
f
e
which
consume
a
q
u
a
t
i
c
organisms
and
t
h
e
m
a
r
k
e
t
a
b
i
l
i
t
y
of
a
q
u
a
t
i
c
organisms.
Protection
of
the
marketability
of
aquatic
organisms
i
s
,
i
n
a
c
t
u
a
l
i
t
y
,
p
r
o
t
e
c
t
i
o
n
o
f
a
u
s
e
o
f
t
h
a
t
water
body
(
commercial
f
i
s
h
e
r
y
)
.
Two
kinds
of
d
a
t
a
are
necessary
t
o
c
a
l
c
u
l
a
t
e
t
h
e
F
i
n
a
l
Residue
Value:
a
b
i
o
c
o
n
c
e
n
t
r
a
t
i
o
n
f
a
c
t
o
r
(
BCF)
and
a
maximum
permissible
t
i
s
s
u
e
concentration,
which
can
be
an
FDAaction
l
e
v
e
l
or
c
a
n
b
e
t
h
e
r
e
s
u
l
t
o
f
a
c
h
r
o
n
i
c
w
i
l
d
l
i
f
e
feeding
study.
For
l
i
p
i
d
­
s
o
l
u
b
l
e
p
o
l
l
u
t
a
n
t
s
,
t
h
e
BCF
is
normalized
f
o
r
percent
l
i
p
i
d
s
and
then
t
h
e
F
i
n
a
l
Residue
Value
is
c
a
l
c
u
l
a
t
e
d
by
d
i
v
i
d
i
n
g
t
h
e
maximum
permissible
t
i
s
s
u
e
concentration
by
the
normalized
BCF
and
by
an
appropriate
percent
l
i
p
i
d
v
a
l
u
e
.
BCFs
are
normalized
f
o
r
p
e
r
c
e
n
t
l
i
p
i
d
s
s
i
n
c
e
the
BCF
measured
f
o
r
any
i
n
d
i
v
i
d
u
a
l
a
q
u
a
t
i
c
s
p
e
c
i
e
s
is
g
e
n
e
r
a
l
l
y
proportional
t
o
t
h
e
percent
l
i
p
i
d
s
i
n
t
h
a
t
species.
a
If
s
u
f
f
i
c
i
e
n
t
data
are
a
v
a
i
l
a
b
l
e
t
o
demonstrate
t
h
a
t
one
o
r
more
of
t
h
e
f
i
n
a
l
v
a
l
u
e
s
should
be
related
t
o
a
water
q
u
a
l
i
t
y
c
h
a
r
a
c
t
e
r
i
s
t
i
c
,
such
as
s
a
l
i
n
i
t
y
,
hardness,
or
suspended
s
o
l
ids,

t
h
e
f
i
n
a
l
v
a
l
u
e
(
s
)
a
r
e
e
x
p
r
e
s
s
e
d
a
s
a
f
u
n
c
t
i
o
n
o
f
t
h
a
t
c
h
a
r
a
c
t
e
r
i
s
t
i
c
.

A
f
t
e
r
t
h
e
four
f
i
n
a
l
values
(
Final
Acute
Value,
Final
Chronic
Value,
F
i
n
a
l
P
l
a
n
t
Value,
and
F
i
n
a
l
Residue
Value)
have
been
obtained,
the
c
r
i
t
e
r
i
o
n
is
established
w
i
t
h
the
Final
Acute
value
becoming
t
h
e
maximum
v
a
l
u
e
and
t
h
e
lowest
of
t
h
e
o
t
h
e
r
t
h
r
e
e
values
becoming
the
24­
hour
average
value.
A
l
l
of
the
data
used
t
o
c
a
l
c
u
l
a
t
e
the
four
f
i
n
a
l
values
and
any
additional
pertinent
information
are
then
reviewed
to
determine
i
f
t
h
e
c
r
i
t
e
r
i
o
n
i
s
reasonable.
I
f
sound
s
c
i
e
n
t
i
f
i
c
evidence
i
n
d
i
c
a
t
e
s
t
h
a
t
t
h
e
criterion
should
be
raised
or
lowered,
appropriate
changes
are
made
as
necessary.

The
November
28,
1980,
Guidelines
have
been
revised
from
the
earlier
published
versions
(
4
3
FR
21506,
May
18,
1978;
43
FR
29028,
July
5,
1978;
4
4
FR
15926,
March
15,
1979).
Details
have
been
added
in
many
places
and
the
concept
of
a
minimum
data
base
has
been
incorporated.
In
addition,
three
adjustment
factors
and
the
species
sensitivity
factor
have
been
deleted.
These
modifications
were
the
result
of
the
Agency's
analysis
of
public
comments
and
comments
received
from
the
Science
Advisory
Board
on
earlier
versions
of
the
Guidelines.
These
comments
and
the
Resultant
modifications
are
addressed
fully
in
Appendix
D
to
this
notice.

Criteria
for
the
Protection
of
Human
Health
Interpretation
of
the
Human
Health
Criteria
0
The
human
health
criteria
issued
today
are
summarized
in
Appendix
A
of
this
Federal
Register
notice.
Criteria
for
the
protection
of
human
health
are
based
on
their
carcinogenic,

toxic,
or
organoleptic
(
taste
and
odor)
properties.
The
meanings
and
practical
uses
of
the
criteria
values
are
distinctly
different
depending
on
the
properties
on
which
they
are
based.

The
objective
of
the
health
assessment
portions
of
the
criteria
documents
is
to
estimate
ambient
water
concentrations
which,
in
the
case
of
noncarcinogens,
prevent
adverse
health
effects
in
humans,
and
in
the
case
of
suspect
or
proven
carcinogens,
represent
various
levels
of
incremental
cancer
risk.
Health
assessments
typically
contain
discussions
of
four
elements:
exposure,
pharmacokinetics,
toxic
effects,
and
criterion
formulation.

The
exposure
section
summarizes
information
on
exposure
routes:
ingestion
directly
from
water,
indirectly
from
consumption
of
aquatic
organisms
found
in
ambient
water,
other
dietary
sources,
inhalation,
and
dermal
contact.
Exposure
assumptions
are
used
to
derive
human
health
criteria.
Most
criteria
are
based
solely
on
exposure
from
consumption
of
water
containing
a
specified
concentration
of
a
toxic
pollutant
and
through
consumption
of
aquatic
organisms
which
are
assumed
to
have
bioconcentrated
pollutants
from
the
water
in
which
they
live.
Other
multimedia
routes
of
exposure
such
as
air,

nonaquatic
diet,
or
dermal
are
not
factored
into
the
criterion
formulation
for
the
vast
majority
of
pollutants
because
of
lack
of
data.
The
criteria
are
calculated
using
the
combined
aquatic
exposure
pathway
and
also
using
the
aquatic
organism
ingestion
exposure
route
alone.
In
criteria
reflecting
both
the
water
consumption
and
aquatic
organism
ingestion
routes
of
exposure,

the
relative
exposure
contribution
varies
with
the
propensity
of
a
pollutant
to
bioconcentrate,
with
the
consumption
of
aquatic
organisms
becoming
more
important
as
the
bioconcentration
factor
(
BCF)
increases.
As
additional
information
on
total
exposure
is
assembled
for
pollutants
for
which
criteria
reflect
only
the
two
specified
aquatic
exposure
routes,
adjustments
in
water
concentration
values
may
be
made.
The
demonstration
of
significantly
different
exposure
patterns
will
become
an
element
of
a
process
to
adapt/
modify
human
health­
based
criteria
to
local
conditions,
somewhat
analogous
to
the
aquatic
life
criteria
modification
process
discussed
previously.
It
is
anticipated
that
States
at
their
discretion
will
be
able
to
set
appropriate
human
health
criteria
based
on
this
process.
0
Specific
health­
based
criteria
are
developed
only
if
a
weight
of
evidence
supports
the
occurrence
of
the
toxic
effect
and
if
dose/
response
data
exist
from
which
criteria
can
be
estimated.

The
pharmacokinteics
section
reviews
data
on
absorption,

distribution,
metabolism,
and
excretion
to
assess
the
biochemical
fate
of
the
compounds
in
the
human
and
animal
system.
The
toxic
effects
section
reviews
data
on
acute,
subacute,
and
chronic
toxicity,
synergistic
and
antagonistic
effects,
and
specific
information
on
mutagenicity,
teratogenicity,
and
carcinogenicity.

From
this
review,
the
toxic
effect
to
be
protected
against
is
identified
taking
into
account
the
quality,
quantity,
and
weight
of
evidence
characteristic
of
the
data.
The
criterion
formulation
section
reviews
the
highlights
of
the
text
and
specifies
a
rationale
for
criterion
development
and
the
mathematical
derivation
of
the
criterion
number.

Within
the
limitations
of
time
and
resources,
current
published
information
of
significance
was
incorporated
into
the
human
health
assessments.
Review
articles
nad
reports
were
used
for
data
evaluation
and
synthesis.
Scientific
judgment
was
exercised
in
reviewing
and
evaluating
the
data
in
each
criteria
document
and
in
identifying
the
adverse
effects
for
which
protective
criteria
were
published.
0
>.
Criteria
for
suspect
or
proven
carcinogens
are
presented
as
concentrations
in
water
associated
with
a
range
of
incremental
cancer
risks
to
man.
Criteria
for
noncarcinogens
represent
levels
at
which
exposure
to
a
single
chemical
is
not
anticipated
to
produce
adverse
effects
in
man.
In
a
few
cases,
organoleptic
(
taste
and
odor)
data
form
the
basis
for
the
criterion.
While
this
type
of
criterion
does
not
represent
a
value
which
directly
affects
human
health,
it
is
presented
as
an
estimate
of
the
level
of
a
pollutant
that
will
not
produce
unpleasant
taste
or
odor
either
directly
from
water
consumption
or
indirectly
by
consumption
of
aquatic
organisms
found
in
ambient
waters.
A
criterion
developed
in
this
manner
is
judged
to
be
as
useful
as
other
types
of
criteria
in
protecting
designated
water
uses.
In
addition,
where
data
are
available,
toxicity­
based
criteria
are
also
presented
for
pollutants
with
derived
organoleptic
criteria.

The
choice
of
criteria
used
in
water
quality
standards
for
these
pcllutants
will
depend
upon
the
designated
use
to
be
protected.

In
the
case
of
a
multiple
use
water
body,
the
criterion
protecting
the
most
sensitive
use
will
be
applied.
Finally,
for
several
pollutants
no
criteria
are
recommended
because
insufficient
information
is
available
for
quantitative
criterion
formulation.

Risk
Extrapolation
Because
methods
do
not
exist
to
establish
the
presence
of
a
threshold
for
carcinogenic
effects,
EPAIs
policy
is
that
there
is
no
scientific
basis
for
estimating
"
safe"
levels
for
carcinogens.

The
criteria
for
carcinogens,
therefore,
state
that
the
recommended
concentration
for
maximum
protection
of
human
health
is
zero.
In
addition,
the
Agency
has
presented
a
range
of
concentrations
corresponding
to
incremental
cancer
risks
of
to
(
one
additional
case
of
cancer
in
populations
ranging
from
10
million
to
100,000,
respectively).
Other
concentrations
representing
different
risk
levels
may
be
calculated
by
use
of
the
Guidelines.
The
risk
estimate
range
is
presented
for
information
purposes
and
does
not
represent
an
Agency
judgment
on
a
"
acceptable"
risk
level.

Summary
of
the
Human
Health
Guidelines
The
health
assessments
and
corresponding
criteria
were
derived
based
on
Guidelines
and
Methodology
Used
in
the
Preparation
of
Health
Effect
Assessment
Chapters
of
the
Consent
Decree
Water
Criteria
Documents
(
the
Guidelines
)
developed
by
EPA'S
Office
of
Research
and
Development.
The
estimation
of
0
health
risk
associated
with
human
exposure
to
environmental
pollutants
requires
predicting
the
effect
of
low
doses
for
up
to
a
lifetime
in
duration.
A
combination
of
epidemiological
and
animal
dose/
response
data
is
considered
the
preferred
basis
for
quantitative
criterion
derivation.

No­
effect
(
noncarcinogen)
or
specified
risk
(
carcinogen)

concentrations
were
estimated
by
extrapolation
from
animal
toxicity
or
human
epidemiology
studies
using
the
following
basic
exposure
assumptions:
a
70­
kilogram
male
person
(
Report
of
the
Task
Group
on
Reference
Man,
International
Commission
for
Radiation
Protection,
November
23,
1957)
as
the
exposed
individual;
the
average
daily
consumption
of
freshwater
and
0
estuarine
fish
and
shellfish
products
equal
to
6.5
grams/
day;
and
the
average
ingestion
of
2
liters/
day
of
water
(
Drinking
Water
and
Health,
National
Academy
of
Sciences,
National
Research
Council,
1977).
Criteria
based
on
these
assumptions
are
estimated
to
be
protective
of
an
adult
male
who
experiences
average
exposure
conditions.

Two
basic
methods
were
used
to
formulate
health
criteria,

depending
on
whether
the
prominent
adverse
effect
was
cancer
or
other
toxic
manifestations.
The
 01
lowing
sections
detail
these
methods.

Carcinogens
Extrapolation
of
cancer
responses
from
high
to
low
doses
and
subsequent
risk
estimation
from
animal
data
are
performed
using
a
linearized
multi­
stage
model.
This
procedure
is
flexible
enough
to
fit
all
monotonically­
increasing
dose
response
data,
since
it
incorporates
several
adjustable
parameters.
The
multi­
stage
model
is
a
linear
nonthreshold
model
as
was
the
"
one­
hit"
model
original
1
y
used
in
the
proposed
criteria
documents.
The
1
inear
nonthreshold
concept
has
been
endorsed
by
the
four
agencies
in
the
Interagency
Regulatory
Liaison
Group
and
is
less
likely
to
underestimate
risk
at
the
low
doses
typical
of
environmental
exposure
than
other
models
that
could
be
used.
Because
of
the
uncertainties
associated
with
dose
response,
animal­
to­
human
extrapolation,
and
other
unknown
factors;
because
of
the
use
of
average
consumptions;
and
because
of
the
serious
public
health
consequences
that
could
result
if
risks
were
underestimated,
EPA
believes
that
it
is
prudent
to
use
conservative
methods
to
estimate
risk
in
the
water
quality
criteria
program.
The
linearized
multistage
model
is
more
systematic
and
invokes
fewer
arbitrary
assumptions
than
the
"
one­
hit"
procedure
previously
used.
0
It
should
be
noted
that
extrapolation
models
provide
estimates
of
risk
since
a
variety
of
assumptions
are
built
into
any
model.
Models
using
widely
different
assumptions
may
produce
estimates
ranging
over
several
orders
of
magnitude.
Since
there
is
at
present
no
way
to
demonstrate
the
scientific
validity
of
any
model,
the
use
of
risk
extrapolation
models
is
a
subject
of
debate
in
the
scientific
community.
However,
risk
extrapolation
is
generally
recognized
as
the
only
tool
available
at
this
time
 or
estimating
the
magnitude
of
health
hazards
associated
with
nonthreshold
toxicants
and
has
been
endorsed
by
numerous
Federal
agencies
and
scientific
organizations,
including
EPA`
s
Carcinogen
Assessment
Group,
the
National
Academy
of
Sciences,
and
the
Interagency
Regulatory
Liaison
Group,
as
a
useful
means
of
assessing
the
risks
of
exposure
to
various
carcinogenic
pollutants.

Noncarcinogens
Health
criteria
based
on
toxic
effects
of
pollutants
other
than
carcinogenicity
are
estimates
of
concentrations
which
are
not
expected
to
produce
adverse
effects
in
humans.
They
are
based
upon
Acceptable
Daily
Intake
(
ADI)
levels
and
are
generally
derived
using
no­
observed­
adverse­
ef
fect­
level
data
from
animal
studies
although
human
data
are
used
wherever
available.
The
AD1
is
calculated
using
safety
factors
to
account
for
uncertainties
0
inherent
in
extrapolation
from
animal
to
man.
In
accordance
With
the
National
Research
Council
recommendations
(
Drinking
Water
and
Health,
National
Academy
of
Sciences,
National
Research
Council,

1977),
safety
factors
of
10,
100,
or
1,000
are
used,
depending
on
the
quality
and
quantity
of
data.
In
some
instances
extrapolations
are
made
from
inhalation
studies
or
limits
to
approximate
a
human
response
from
ingestion
using
the
Stokinger­

Woodward
model
(
Journal
of
American
Water
Works
Association,

1958).
Calculations
of
criteria
from
ADIS
are
made
using
the
standard
exposure
assumptions
(
2
liters
of
water,
6.5
grams
of
edible
aquatic
products,
and
an
average
body
weight
of
7
0
kg).
APPENDIX
C
.
THE
PHILOSOPHY
OF
TRE
1976
WATER
QUALITY
CRITERIA
­
­­­

Water
quality
criteria
specify
concentrations
of
water
constituents
which,
if
not
exceeded,
are
expected
to
support
an
organic
ecosystem
suitable
for
the
higher
uses
of
water.
Such
criteria
are
derived
from
scientific
facts
obtained
from
experimental
or
L~
J
g&
gg
observations
that
depict
organic
responses
to
a
defined
stimulus
or
material
under
identifiable
or
regulated
environmental
conditions
for
a
specified
time
period.

Water
quality
criteria
are
not
intended
to
offer
the
same
degree
of
strategy
for
survival
and
propagation
at
all
times
to
all
organisms
within
a
given
ecosystem.
They
are
intended
not
only
to
protect
essential
and
significant
life
in
water
and
the
direct
users
o
f
w
a
t
e
r
,
b
u
t
a
l
s
o
t
o
p
r
o
t
e
c
t
life
that
isdependent
on
life
in
water
for
its
existence,
or
that
may
consume
intentionally
or
unintentionally
any
edible
portion
of
such
life.

The
criteria
levels
for
domestic
water
supply
incorporate
available
data
for
human
health
protection.
Such
values
are
different
from
the
criteria
levels
necessary
for
protection
of
aquatic
life.
The
Agency's
interim
primary
drinking
water
regulations
(
40
Federal
Register
59566
December
24,
1975),
as
required
by
the
Safe
Drinking
Water
Act
(
4
2
U.
S.
C.
300f,
et
seq.)
,
incorporate
applicable
domestic
water
supply
criteria.

Where
pollutants
are
identified
in
both
the
quality
criteria
for
domestic
water
supply
and
the
Drinking
Water
Standards,
the
concentration
levels
are
identical.
Water
treatment
may
not
significantly
affect
the
removal
of
certain
pollutants.
What
is
essential
and
significant
life
in
water?
Do
Daphnia
or
stonefly
nymphs
qualify
as
such
life?
Why
does
1/
100th
of
a
concentration
that
is
lethal
to
5
0
percent
of
the
test
organisms
(
LC50)
constitute
a
criterion
in
some
instances,
whereas
1/
2
or
l/
lOth
of
some
effect
levels
constitutes
a
criterion
in
other
instances?
These
are
questions
that
often
are
asked
of
those
who
undertake
the
task
of
criteria
formulation.

The
universe
of
organisms
composing
life
in
water
is
great
in
both
kinds
and
numbers.
As
in
the
human
population,

physiological
variability
exists
among
individuals
of
the
same
species
in
response
to
a
given
stimulus.
A
much
greater
response
variation
exists
among
species
of
aquatic
organisms.
Thus,

aquatic
organisms
do
not
exhibit
the
same
degree
of
harm,

individually
or
by
species,
from
a
given
concentration
of
a
toxicant
or
potential
toxicant
within
the
environment.
In
establishing
a
level
or
concentration
of
a
quality
constituent
as
a
criterion
it
is
necessary
to
ensure
a
reasonable
degree
of
safety
for
those
more
sensitive
species
that
are
important
to
the
functioning
of
the
aquatic
ecosystem
even
though
data
on
the
response
of
such
species
to
the
quality
constituent
under
consideration
may
not
be
available.
The
aquatic
food
web
is
an
intricate
relationship
of
predator
and
prey
organisms.
A
water
constituent
that
may
in
some
way
destroy
or
eliminate
an
important
segment
of
that
food
web
would,
in
all
likelihood,

destroy
or
seriously
impair
other
organisms
associated
with
it.
1
0
Although
experimentation
relating
to
the
effects
of
particular
substances
under
control
led
conditions
began
in
the
early
19OO's,
the
effects
of
any
substance
on
more
than
a
few
of
the
vast
number
of
aquatic
organisms
have
not
been
investigated.

Certain
test
animals
have
been
selected
by
investigators
for
intensive
investigation
because
of
their
importance
to
man,
their
availability
to
the
researcher,
and
their
physiological
responses
to
the
laboratory
environment.
As
general
indicators
of
organism
responses
such
test
organisms
are
representative
of
the
expected
results
for
other
associated
organisms.
In
this
context
Daphnia
or
stoneflies
or
other
associated
organisms
indicate
the
general
levels
of
toxicity
to
be
expected
among
untested
species.
In
addition,
test
organisms
are
themselves
vital
1
inks
within
the
food
web
that
results
in
the
fish
population
in
a
particular
waterway.

The
ideal
data
base
for
criteria
development
would
consist
of
information
on
a
large
percentage
of
aquatic
species
and
would
show
the
community
response
to
a
range
of
concentrations
for
a
tested
constituent
during
a
long
time
period.
This
information
is
not
available
but
investigators
are
beginning
to
derive
such
information
for
a
few
water
constituents.
Where
only
96­
hour
bioassay
data
are
available,
judgmental
prudence
dictates
that
a
substantial
safety
factor
be
employed
to
protect
all
life
stages
of
the
test
organism
in
waters
of
varying
quality,
as
well
as
associated
organisms
within
the
aquatic
environment
that
have
not
been
tested
and
that
may
be
more
sensitive
to
the
test
constituent.
Application
factors
have
been
used
to
provide
the
degree
of
protection
required.
Safe
levels
for
certain
chlorinated
hydrocarbons
and
certain
heavy
metals
were
estimated
by
applying
an
0.01
application
factor
to
the
96­
hour
LC50
value
a
for
sensitive
aquatic
organisms.
Flow­
through
bioassays
have
been
conducted
for
some
test
indicator
organisms
over
a
substantial
period
of
their
life
history.
In
a
few
other
cases,

information
is
available
for
the
organism's
natural
life
or
for
more
than
one
generation
of
the
species.
Such
data
may
indicate
a
minimal
effect
level,
as
well
as
a
no­
effect
level.

The
word
'*
criterion*
I
should
not
be
used
interchangeably
with
or
as
a
synonym
for
the
word
'*
standard.**
The
word
'*
criterion"

represents
a
constituent
concentration
or
level
associated
with
a
degree
of
environmental
effect
upon
which
scientific
judgment
may
be
based.
A
s
it
is
currently
associated
with
the
water
environment
it
has
come
to
mean
a
designated
concentration
of
a
constituent
that,
when
not
exceeded,
will
protect
an
organism,

an
organism
community,
or
a
prescribed
water
use
or
quality
with
an
adequate
degree
of
safety.
A
criterion,
in
some
cases,
may
be
a
narrative
statement
instead
of
a
constituent
concentration.
on
the
other
hand,
a
standard
connotes
a
legal
entity
for
a
particular
reach
of
waterway
or
 or
an
effluent.
A
water
quality
standard
may
use
a
water
quality
criterion
as
a
basis
for
regulation
or
enforcement,
but
the
standard
may
differ
from
a
criterion
because
of
prevailing
local
natural
conditions,
such
as
naturally
occurring
organic
acids,
or
because
of
the
importance
of
a
particular
waterway,
economic
considerations,
or
the
degree
of
safety
to
a
particular
ecosystem
that
may
be
desired.

Toxicity
to
aquatic
life
generally
is
expressed
in
terms
of
acute
(
short
term)
or
chronic
(
long­
term)
effects.
Acute
toxicity
refers
to
effects
occurring
in
a
short
time
period:
often
death
is
the
end
point.
Acute
toxicity
can
be
expressed
as
the
lethal
concentration
for
a
stated
percentage
of
organisms
tested,
or
the
reciprocal,
which
is
the
tolerance
limit
of
a
percentage
of
surviving
organisms,
Acute
toxicity
for
aquatic
organisms
generally
has
been
expressed
for
2
4
to
96­
hOur
exposures.

chronic
toxicity
refers
to
effects
through
an
extended
time
period.
Chronic
toxicity
may
be
expressed
in
terms
of
an
observation
period
equal
to
the
lifetime
of
an
organism
o
r
to
the
time
span
of
more
than
one
generation.
Some
chronic
effects
may
be
reversible,
but
most
are
not.

Chronic
effects
often
occur
in
the
species
population
rather
than
in
the
individual.
o
r
the
sperm
does
not
remain
viable,
the
species
would
be
eliminated
from
an
ecosystem
because
of
reproductive
failure.
Physiological
stress
may
make
a
species
less
competitive
with
others
and
may
result
in
a
gradual
population
decline
o
r
absence
from
an
area.
The
elimination
of
a
microcrustacean
that
serves
as
a
vital
food
during
the
larval
period
of
a
fish's
life
could
result
ultimately
in
the
elimination
of
the
fish
from
an
area.
The
phenomenon
of
bioaccumulation
of
certain
materials
may
result
in
chronic
toxicity
to
the
ultimate
consumer
in
a
rood
chain.
Thus,
fish
may
mobilize
lethal
toxicants
from
their
fatty
tissues
during
periods
of
physiological
stress.
Egg
shells
of
predatory
birds
may
be
weakened
to
a
point
of
destruction
in
the
nest.
Bird
chick
embryos
may
have
increased
mortality
rates.
There
may
be
a
hazard
to
the
health
of
man
if
aquatic
organisms
with
toxic
residues
are
consumed.
If
eggs
fail
to
develop
The
fact
that
living
systems,
i.
e.,
individuals,
populations,

species,
and
ecosystems,
can
take
up,
accumulate,
and
bioconcentrate
manmade
and
natural
toxicants
is
well
documented.

In
aquatic
systems
biota
are
exposed
directly
to
pollutant
toxicants
through
submersion
in
a
relatively
efficient
solvent
(
water)
and
are
exposed
indirectly
through
food
webs
and
other
biological,
chemical,
and
physical
interactions.
Initial
toxicant
levels,
if
not
immediately
toxic
and
damaging,
may
accumulate
in
the
biota
or
sediment
over
time
and
increase
to
levels
that
are
lethal
or
sublethally
damaging
to
aquatic
organisms
or
to
consumers
of
these
organisms.
Water
quality
criteria
reflect
a
knowledge
of
the
capacity
for
environmental
accumulation,
persistence,
and
effects
of
specific
toxicants
in
specific
aquatic
systems.

Ions
of
toxic
materials
frequently
cause
adverse
effects
because
they
pass
through
the
semipermeable
membranes
of
an
organism.
Molecular
diffusion
through
membranes
may
occur
for
some
compounds
such
as
pesticides,
polychlorinated
biphenyls,

and
other
toxicants.
Some
materials
may
not
pass
through
membranes
in
their
natural
or
waste­
discharged
state,
but
in
water
they
may
be
converted
to
states
that
have
increased
ability
to
affect
organisms.
For
example,
certain
microorganisms
can
methylate
mercury,
thus
producing
a
material
that
more
readily
enters
physiological
systems.
Some
materials
may
have
multiple
effects:
for
example,
an
iron
salt
may
not
be
toxic;
an
iron
floc
or
gel
may
be
an
irritant
or
clog
fish
gills
to
effect
asphyxiation;
iron
at
l
o
w
concentrations
can
be
a
trace
nutrient
but
at
high
concentrations
it
can
be
a
toxicant.
Materials
also
can
affect
organisms
if
their
metabolic
byproducts
cannot
be
excreted.
Unless
otherwise
stated,
criteria
are
based
on
the
total
concentration
of
the
substance
because
an
ecosystem
can
produce
chemical,
physical,
and
biological
changes
that
may
be
detrimental
to
organisms
living
in
or
using
the
water.

In
prescribing
water
quality
criteria,
certain
fundamental
principles
dominate
the
reasoning
process.
In
establishing
a
level
or
concentration
as
a
criterion
for
a
given
constituent
it
was
assumed
that
other
factors
within
the
aquatic
environment
are
acceptable
to
maintain
the
integrity
of
the
water.

Interrelationships
and
interactions
among
organisms
and
their
environment,
as
well
as
the
interrelationships
of
sediments
and
the
constituents
they
contain
to
the
water
above,
are
recognized
0
as
fact.

Antagonistic
and
synergistic
reactions
among
many
quality
constituents
in
water
also
are
recognized
as
fact.
The
precise
definition
of
such
reactions
and
their
relative
effects
on
particular
segments
of
aquatic
life
have
not
been
identified
with
scientific
precision.
Historically
much
of
the
data
to
support
criteria
development
was
of
an
ambient
concentration­
organism
response
nature.
Recently,
data
are
becoming
available
on
long­

term
chronic
effects
on
particular
species.
Studies
now
determine
carcinogenic,
teratogenic,
and
other
insidious
effects
of
toxic
materials.

Some
unpolluted
waters
in
the
Nation
may
exceed
designated
criteria
for
particular
constituents.
There
is
variability
in
the
natural
quality
of
water
and
certain
organisms
become
adapted
to
that
quality,
which
may
be
considered
extreme
in
other
areas.

Likewise,
it
is
recognized
that
a
single
criterion
cannot
identify
minimal
quality
for
the
protection
of
the
integrity
of
water
for
every
aquatic
ecosystem
in
the
Nation.
To
provide
an
adequate
degree
of
safety
to
protect
against
long­
term
effects
may
result
in
a
criterion
that
cannot
be
detected
with
present
analytical
tools.
In
some
cases,
a
mass
balance
calculation
can
provide
a
means
of
assurance
that
the
integrity
of
the
waterway
is
not
being
degraded.

Water
quality
criteria
do
not
have
direct
regulatory
impact,

but
they
form
the
basis
for
judgment
in
several
Environmental
Protection
Agency
programs
that
are
derived
from
water
quality
considerations.
For
example,
water
qual
ity
standards
developed
by
the
States
under
section
303
of
the
Act
and
approved
by
EPA
are
to
be
based
on
the
water
quality
criteria,
appropriately
modified
to
take
account
of
local
conditions.
The
local
conditions
to
be
considered
include
actual
and
projected
uses
of
the
water,
natural
background
levels
of
particular
constituents,

the
presence
or
absence
of
sensitive
important
species,

characteristics
of
the
local
biological
community,
temperature
and
weather,
flow
characteristics,
and
synergistic
or
antagonistic
effects
of
combinations
of
pollutants.

Similarly,
by
providing
a
judgment
on
desirable
levels
of
ambient
water
qual
ity,
water
quality
criteria
are
the
starting
point
in
deriving
toxic
pollutant
effluent
standards
pursuant
to
section
307(
a)
of
the
Act.
Other
EPA
programs
that
use
water
qual
ity
criteria
involve
drinking
water
standards,
the
ocean
dumping
program,
designation
of
hazardous
substances,
dredge
spoil
criteria
development,
removal
of
in­
place
toxic
materials,

thermal
pollution,
and
pesticide
registration.

To
provide
the
water
resource
protection
for
which
they
are
designed,
quality
criteria
should
apply
to
virtually
all
of
the
Nation's
navigable
waters
with
modifications
for
local
conditions
as
needed.
To
violate
quality
criteria
for
any
substantial
length
of
time
or
in
any
substantial
portion
of
a
waterway
may
result
in
an
adverse
affect
on
aquatic
life
and
perhaps
a
hazard
to
man
o
r
other
consumers
of
aquatic
life.

Quality
criteria
have
been
designed
­
to
provide
long­
term
protection.
Thus,
they
may
provide
a
basis
for
effluent
standards,
but
it
is
not
intended
that
criteria
values
become
effluent
standards.
It
is
recognized
that
certain
substances
may
a
be
applied
to
the
aquatic
environment
with
the
concurrence
of
a
governmental
agency
for
the
precise
purpose
of
controlling
or
managing
a
portion
of
the
aquatic
ecosystem:
aquatic
herbicides
and
piscicides
are
examples
of
such
substances.
For
such
occurrences,
criteria
obviously
do
not
apply.
It
is
recognized
further
that
pesticides
applied
according
to
official
label
instructions
to
agricultural
and
forest
lands
may
be
washed
to
a
receiving
waterway
by
a
torrential
rainstorm.
Under
such
conditions
it
is
believed
that
such
diffuse
source
inflows
should
receive
consideration
similar
to
that
of
a
discrete
effluent
discharge
and
that
in
such
instances
the
criteria
should
be
applied
to
the
principal
portion
of
the
waterway
rather
than
to
that
peripheral
portion
receiving
the
diffuse
inflow.
e
­.
,
The
format
f
o
r
presenting
water
quality
criteria
includes
a
concise
statement
of
the
dominant
criterion
o
r
criteria
for
a
particular
constituent
followed
by
a
narrative
introduction,
a
rationale
that
includes
justification
for
the
designated
criterion
or
criteria,
and
a
listing
of
the
references
cited
within
the
rationale.
An
effort
has
been
made
to
restrict
supporting
data
to
those
which
either
have
been
published
o
r
are
in
press
awaiting
publication.
A
particular
constituent
may
have
more
than
one
criterion
to
ensure
more
than
one
water
use
or
condition,
i.
e.,
hard
or
soft
water
where
applicable,
suitability
as
a
drinking
water
supply
source,
protection
of
human
health
when
edible
portions
of
selected
biota
are
consumed,
provision
for
recreational
bathing
or
waterskiing,
and
permitting
an
appropriate
factor
of
safety
to
ensure
protection
for
essential
warm­
or
coldwater
associated
biota.

Criteria
are
presented
 or
those
substances
that
may
occur
in
water
where
data
indicate
the
potential
for
harm
to
aquatic
life,

or
to
water
users,
or
to
the
consumers
of
the
water
o
r
aquatic
life.
Presented
criteria
do
not
represent
an
all­
inclusive
list
of
constituent
contaminants.
omissions
from
criteria
should
not
be
construed
to
mean
that
an
omitted
quality
constituent
is
either
unimportant
o
r
non­
hazardous.
BIBLIOGRAPHY
Adelman,
I.
R.
and
L.
L.
Smith,
1970.
Effect
of
hydrogen
sulfide
on
northern
pike
eggs
and
sac
fry.
99:
501.
Trans.
Amer.
Fish.
SOC.,

Agriculture
Handbook
No.
60,
1954.
Diagnosis
and
improvement
of
saline
and
alkali
soils.
L.
A.
Richards,
editor,
U.
S.
Govern­
ment
Printing
Office,
Washington,
D.
C.

Anderson,
B.
G.,
1960.
The
toxicity
of
organic
insecticides
to
Daphnia.
Second
Seminar
on
Biol.
Problems
in
Water
Pollution.
Robert
A.
Taft
Sanitary
Engineering
Center
Technical
Report
W60­
3,
Cincinnati,
Ohio.

Bahner,
L.
H.
and
D.
R.
Nimmo,
1974.
Methods
to
assess
effects
of
combinations
of
toxicants,
salinity,
and
temperature
on
estu­
marine
animals.
Proc.
9th
Am.
Conf.
on
Trace
Substances
in
Env.
Health,
Univ.
Miss.?
olumbia,
Mo.

Ballentine,
R.
X.
and
F.
W.
Kittrell,
1968.
Observations
of
fecal
colifonns
in
several
recent
stream
pollution
studies.
ings
of
the
Symposium
on
Fecal
Coliform
Bacteria
in
Water
and
Wastewater,
May
21­
22,
1968,
Bureau
of
Sanitary
Engineering,
California
State
Department
of
Public
Health.

Bell,
H.
L.,
1971.
Effect
of
low
pH
on
the
survival
and
emergence
of
aquatic
insects.
Water
Res.,
5:
313.

Bellan,
et
al.,
1972.
The
sublethal
effects
of
a
detergent
on
the
reproduction,
development,
and
settlement
in
the
polychae­
tous
annelid
Capitella
capitata.
Marine
Biology,
14:
183.
Proceed­

Bender,
M.
E.,
1969.
The
toxicity
of
the
hydrolysis
and
breakdown
products
of
malathion
to
the
fathead
minnow
(
PimephalE
promelas,
Rafinesque).
Water
Res.,
3:
571.

methyl
parathion,
parathion
and
azinphosmethyl
in
mice
and
fish:
Onset
and
recovery
of
inhibition.
Bull.
Environ.
Contam.
Toxicol.,
12:
117.

Berg,
G.,
1974.
Reassessment
of
the
virus
problem
in
sewage
and
in
surface
and
renovated
waters.
Sixth
International
Water
poll.
Res.
California.
Pergamon
Press.

Bigqar,
J.
W.
and
M.
Fireman,
1960.
Boron
abosrption
and
release
by
soils.

Billard,
R.
and
dekinkelin,
1970.
Sterilization
of
the
testicles
of
guppies
by
means
of
non­
lethal
doses
of
parathion.
D
Hydrobiologie,
9(
1):
91.
Benke,
G.
M.
and
S.
D.
Murphy,
1974.
Anticholinesterase
action
of
Soil
Sci.
SOC.
Amer.
Proc.,
24:
115
Annales
Black,
E.
C.,
1953.
Upper
lethal
temperatures
of
some
British
Columbia
freshwater
fishes.
Jour.
Fish.
Res.
Bd.
Can.,
10:
196
Blumer,
M.,
1970.
Oil
contamination
and
the
living
resources
of
the
sea.
F.
A.
O.
Tech.
Conf.
Rome.
FIR:
MP/
7O/
R­
lIllP.

Bollard,
E.
G.
and
G.
W.
Butler,
1966.
Mineral
nutrition
of
plants.
Ann.
Rev.
Plant
Physiol.,
17:
77.

Bonde,
G.
J.,
1966.
Bacteriological
methods
for
the
estimation
of
water
pollution.
Hlth.
Lab.
Sci.
3:
124.

Bonde,
G.
J.,
1974.
Bacterial
indicators
of
sewage
pollution.
International
Symposium
on
Discharge
of
Sewage
from
Sea
Out­
falls.
Pergamon
Press.

Bookhout,
C.
G.
­
et
­.
I
a1
1973.
Effects
of
mirex
on
the
larval
development
of
two
crabs.
Water,
Air,
and
Soil
Pollution,
1:
165.

channel
catfish.
Trans.
Amer.
Fish.
SOC.,
96:
31.

mirex
in
selected
estuaries
of
South
Carolina,
1969­
71.
Pesti­
cides
Monit.
Jour.,
7:
6.

an
reproduction
of
fish
exposed
to
gas
supersaturated
water.
Unpublished
report.
U.
S.
Environmental
Protection
Agency,
Western
Fish
Toxicology
Station,
Corvallis,
Oregon.

Boyd,
C.
E.
and
D.
E.
Ferguson,
1964.
Susceptibility
and
resistance
of
mosquitofish
to
several
insecticides.
Econ.
Entomo
1.
,
57:
430.

Boyle,
H.
W.,
1967.
Taste/
odor
contamination
of
fish
from
the
Ohio
River.
Fed.
Water
Poll.
Cont.
Admin.,
Cincinnati,
Ohio.

Bradford,
G.
R.,
1966.
Boron
[
toxicity,
indicator
plants]
,
in
diagnostic
criteria
for
plants
and
soils.
H.
D.
Chapman,
Ed.,
University
of
California,
Division
of
Agricultural
Science,
Berkeley,
p.
33.

development
and
weight
of
sockeye
salmon
embryos
and
alevins.
International
Pacific
Salmon
Fisheries
Commission,
Progress
Report
No.
12,
pp.
1­
26,
mimeo.

of
speckled
trout.
Trans.
Amer.
Fish.
SOC.,
70:
397.

of
fishes.
Quarterly
Rev.
Biol.,
31:
75.
Boon,
C.
W.
and
B.
J.
Follis,
1967.
Effects
of
hydrogen
sulfide
on
Borthwick,
P.
W.,
et
al.,
1973.
Accumulation
and
movement
of
Bouck,
G.
R.,
et
al.,
1975.
Mortality,
saltwater
adaption
Brannon,
E.
L.,
1965.
The
influence
of
physical
factors
on
the
Brett,
J.
R.,
1941.
Tempering
versurs
acclimation
in
the
planting
Brett,
J.
R.,
1956.
Some
principals
in
the
thermal
requirements
Brett,
J.
R.,
1960.
Thermal
requirementsof
f
ish­­
three
decades
of
study,
1940­
1970.
In:
Biological
problems
in
water
pollution.
C.
M.
Tarzwell
(
ed.)
Dept.
of
Health,
Education
and
Welfare,
Public
Health
Service.

Brinley,
F.
J.
,
1944.
Biological
Studies.
House
Document
266,
78th
Congress,
1st
session:
Part
11,
Supplement
F.
pp.
1275­
1353.
0
Brungs,
W.
A.,
1971.
Chronic
effects
of
low
dissolved
oxygen
concentrations
on
fathead
minnow
(
Pimephales
promelas).
­
J.
Fish.
Res.
Bd.
Canada.
28:
1119­
1123.

taste
in
domestic
water.
Jour.
Amer.
Water
Works
Assn.,
61:
575.

Bugbee,
S.
L.
and
C.
M.
Walter,
1973.
The
response
of
macro­
invertebrates
to
gasoline
pollution
in
a
mountain
stream.
In:
Prevention
and
control
of
oil
spills,
proceedings
of
symposium
March
13­
17,
Washington,
D.
C.,
p.
725.

insecticides
to
resistant
and
susceptible
mosquitofish
in
static
and
flowing
solutions.
,
Mosquito
News,
29(
1):
96.

Burson,
B.,
1938.
Seasonal
temperature
variations
in
relation
to
water
treatment.
Jour.
Amer.
Water
Works
Assn.,
30:
793.

Butler,
P.
A.,
1963.
Commercial
fisheries
investigations.
In:
pesticide
wildlife
Studies
during
1961
and
1962.
U.
S.
Fish.
Wildl.
Serv.
Circ.
167,
Washington,
D.
C.

organisms.
In:
A
Guide
to
Marine
Pollution,
E.
D.
Goldberg,
ed.
Gordon
and
Beach,
NY.

Calabrese,
A.,
1969.
Individual
and
combined
effects
of
salinity
and
temperature
on
embryos
and
larvae
of
the
coot
clam,
Mulinia
­
lateralis
(
say),
Biol.
Bull.
137,
3:
417.

embryos
of
the
american
oyster,
Crassotrea
virginica.
Marine
Biol.
,
18:
162.

Camp,
T.
R.,
1963.
Water
and
its
impurities.
Reinhold
Publishing
Corp,
New
York,
New
York.

Capurro,
L.
R.
A.,
1970.
Oceanography
for
practicing
engineers.
Barnes
and
Noble
Inc.,
New
York.

Carlson,
A.
R.,
et
al.,
1974.
Effects
of
lowered
dissolved
oxygen
concentrations
on
channel
catfish
(
Ictalurus
embryos
and
larve.
Trans.
Amer.
Fish.
SOC.
103:
623­
626.
Bruvold,
W.
H.,
et
,
al.,
1969.
Consumer
assessment
of
mineral
Burke,
W.
D.
and
D.
E.
Ferguson,
1969.
Toxicities
of
four
0
Butler,
P.,
et
al.,
1972.
Test,
monitoring
and
indicator
Calabrese,
A.,
et
al.,
1973.
The
toxicity
of
heavy
metals
to
Chin,
E.,
1961.
A
trawl
study
of
an
estuarine
nursery
area
in
Galveston
Bay
with
particular
reference
to
penaeid
shrimp.
P~.
D.
Dissertation,
University
of
Washington.

significance
of
lead
isotopes
in
pelagic
sediments.
Cosmochim.
Acta.,
26:
263.

Clay,
A.,
et
a1
.,
1975.
Experimental
induction
of
gas
bubble
disease
in
menhaden.
Presented
a
the
American
Fisheries
Society,
September,
1975,
Las
Vegas,
Nevada.
New
England
Aquarium,
Boston,
Mass.

Coble,
D.
W.,
1961.
Influence
of
water
exchange
and
dissolved
oxygen
in
redds
on
survival
of
steelhead
trout
embryos,
Trans.
Amer.
Fish.
SOC.
90~
469­
474.

Cooper,
A.
C.,
1965.
The
effect
of
transported
stream
sediments
on
the
survival
of
sockeye
and
pink
salmon
eggs
and
alevin.
International
Pacific
Salmon
Fisheries
Commission,
Bulletin
Chow,
T.
J.,
and
C.
C.
Patterson,
1962.
The
occurrences
and
Geochim.

18
:
1­
71.

Coppage,
D.
L.,
1972.
Organophosphate
pesticides:
specific
level
of
brain
ACHE
inhibition
related
to
death
in
sheepshead
minnows.
Trans.
Amer.
Fish.
SOC.,
101:
534.

Coppage,
D.
L.
and
T.
W.
Duke,
1971.
Effects
of
pesticides
in
estuaries
along
the
Gulf
and
Southeast
Atlantic
Coasts.
In:
Proceedings
of
the
2nd
Gulf
Coast
Conference
on
Mosquito
Suppression
and
Wildlife
Management
(
C.
H.
Schmidt,
ed.)
National
Mosquito
Control­
Fish
and
Wildlife
Management
Coordinating
Committee,
Washington,
D.
C.

Coppage,
D.
L.
and
E.
Matthews,
1974.
Short­
term
effects
of
organophosphate
pesticides
on
cholinesterases
of
estuarine
fishes
and
pink
shrimp.
Bull.
Environ.
Contam.
Toxicol.,
11:
483.

inhibition
in
fish
as
a
diagnosis
of
environmental
poisoning
by
malathion,
0,
0­
dimethyl
S­(
1,
l­
dicarbethoxy­
ethyl)
phosphoro­
dithioate.
Pesticide
Biochemistry
and
Physiology
(
in
press).

temperatures
on
larval
development
in
the
mud
crab,
Rhitropanopeus
harrisii.
In:
Fourth
European
Marine
Biology
Symposium.
D.
I.
Crisp
(
ed.)
,
Cambridge
University
Press,
London.

review
of
the
lterature
of
1967.
Jour.
Water
Poll.
Cont.
Fed.,
40:
1047.

Coutant,
C.
C.,
1969.
Thermal
pollution
effects:
A
review
of
the
Coppage,
D.
L.
.
et
al.,
1975.
Brain
acteylcholinesterase
Costlow,
Jr.,
J
.
D
.
and
C.
G.
Bookhout,
1971.
The
effect
of
cyclic
Coutant,
C.
C.,
1968.
Thermal
pollution­­
biologicaly
effects:
A
literature
of
1968.
Jour.
Water
Poll.
Cont.
Fed.,
41:
1036.
Coutant,
C.
C.,
1970.
Thermal
pollution
effects:
A
review
of
the
literature
of
1970.
Thermal
pollution
effects:
A
review
of
the
literature
of
1969.
Jour.
Water
Poll.
Cont.
Fed.,
42:
1025.

Coutant,
C.
C.,
1971.
Thermal
pollution
effects:
A
review
of
the
literature
of
1970.
Jour.
Water
Poll.
Cont.
Fed.,
43:
1292.

Coutant,
C.
C.,
1972.
Biological
aspects
of
thermal
pollution,
11.
Scientific
basis
for
water
temperature
standards
at
power
plants.
CRC
Critical
Rev.
in
Environ.
Cont.,
3:
1.

Coutant,
C.
C.,
1975.
Temperature
selection
by
fish­­
a
factor
in
power
plant
impact
assessments.
In:
Symposium
on
the
physical
and
biological
effects
on
the
environment
of
cooling
systems
and
thermal
discharges
at
nuclear
power
stations.
Int'l
Atomic
Energy
Agency
(
In
press).

Coutant,
C.
C.,
and
C.
P.
Goodyear,
1972.
Water
pollution­
thermal
pollution:
A
review
of
the
literature
of
1971.
Jour.
Water
Poll.
Cont.
Fed.,
45:
1250.

Coutant,
C.
C.,
and
H.
A.
Pfuderer,
1973.
Water
pollution­
thermal
effects:
A
review
of
the
literature
of
1973.
Jour.
Water
Poll.
Cont.
Fed.,
45:
1331.

effects:
A
review
of
the
literature
of
1974.
Jour.
Water
Poll.
Cont.
Fed.,
45:
1476.

marine
animals.
Nature,
179:
1138.
Coutant,
C.
C.,
and
H.
A.
Pfuderer,
1974.
Water
pollution­
thermal
Crisp,
D.
J.,
1957.
Effect
of
low
temperature
on
the
breeding
of
Datta,
N.
and
J.
Olearte,
1974.
R­
Factors
in
strains
of
Salmonella
typhi
and
Shigella
dysenteri
I
isolated
during
efi7emic.
s
in
Mexico:
­
Classification
by
comrJabilitv.
Antimicrobial
Agents
and
Chemotherapy
5:
310.

embryonic
development
of
clams
and
oysters
and
on
survival
and
grouwth
of
the
larvae.
U.
S.
Fish
and
Wildlife
Service,
Fishery
Bulletin,
67:
393.
­
­
­

Davis,
H.
C.
and
H.
Hindu,
1969.
Effects
of
pesticides
on
Dawlev,
E.
M.
and
W.
J.
Ebel.
1975.
Effectw
of
various
~~

concentrations
of
dissolved
atmospheric
gas
on
juvenile
chinook
salmon,
Oncorhynchus
tsawytscha,
and
steelhead
trout.
Fish.
Bull.,
U.
S.
(
In
press)

Dawley,
E.,
et
al.,
1975.
Bioassays
of
total
dissolved
gas
pressure.
(
Unpublished
report.
National
Marine
Fisheries
Services,
Seattle,
Washington.

Delfino,
J.
J.
and
G.
F.
Lee,
1971.
Variation
of
manganese,
dissolved
oxygen
and
related
chemical
parameters
in
the
bottom
waters
of
Lake
Mendota,
Wisconsin.
Water
Res.,
5:
1207.
DeMont,
W.
J.,
et
al.,
1975.
Effect
of
atmospheric
gas
supersaturation
caused
by
dams
on
salmon
and
steelhead
trout
to
the
Snake
and
Columbia
Rivers.
Final
Report.
Northwest
Fisheries
Center,
NMFS,
Seattle,
Washington.

De
Sylva,
D.
P.,
1969.
Theoretical
considerations
of
the
effects
of
heated
effluents
on
marine
fisher.
In:
Biological
aspects
of
thermal
pollution.
P.
A.
Krenkel
and
F.
C.
Parker
(
eds.),
Vanderbilt
University
Press.

photoperiod
on
reproductive
cycling
of
the
estuarine
gobiid
fish,
Gillichthys
mirabilk.
Fishery
Bull.
70,
4:
1137.

deVlaming,
V.
L.,
1971.
The
effects
of
food
deprivation
and
salinity
changes
on
reproductive
function
in
the
estuarine
gobiid
fish,
Gillichtys
mirabilis.
­
Biol.
Bull.,
141:
450.

Diaz­
Piferrer,
1962.
The
effects
on
an
oil
spill
on
the
shore
of
Guanica,
Puerto
Rico
(
abstract)
Ass.
Isl.
Mar.
Labs,
4th
Mtg.
Curaco,
12­
13.

Doudoroff,
P.
and
Katz,
M..
1953.
Critical
review
of
literature
deVIaming,
V.
L.,
1972.
The
effects
of
temperature
and
*
­

on
the
toxicity
of
industrial
wastes
and
their
components
to
fish.
11.
The
metals
and
salts.
Sew.
and
Ind.
Wastes,
25,
p.
002.

Doudoroff,
P.
and
D.
L.
Shumway,
1970.
Dissolved
oxygen
requirements
of
fresh
water
fishes.
FA0
Fish.
Tech.
Paper
No.
06.

Dow,
R.
L.,
1973.
Fluctuations
in
marine
species
abundance
dur
climatic
cycles.
Mar.
Tech.
SOC.
Jour.,
7,4:
30.

Dowden,
B.
F.,
1966.
Effects
of
five
insecticides
on
the
oxygen
consumption
of
the
bluegill
sunfish,
Le
omis
macrochirus.
Ph.
D.
Thesis,
Louisiana
State
Univers
f
i
t
o
n
Rouge,
LA.

Eaton,
J.
G.,
1970.
Chronic
malathion
toxicity
to
the
bluegill
(
Lepomis
macrochirus
Rafinesque)
.
Water
Research,
4:
673.

Ebel,
W.
J.,
et
al.,
1975.
Effect
of
atmospheric
gas
supersaturation
caused
by
dams
on
salmon
and
steelhead
trout
of
the
Snake
and
Columbia
Rivers.
Final
Report.
Northwest
Fisheries
Center,
NMFS,
Seattle,
Washington.

sediment
studied
in­
situ
and
in
the
laboratory.
Water
Research,
7:
1285.
Edberg,
N.
and
Hofstan,
B.
V.,
1973.
Oxygen
uptake
of
bottom
Eddy,
R.
M.,
1972.
The
influence
of
dissolved
oxygen
.
concentration
and
temperature
on
the
survival
and
growth
of
chinook
salmon
embryos
and
fry.
M.
S.
Thesis,
Oregon
State
Univ.,
June,
1972.
Eichelberger,
J.
W.
and
J.
J.
Lichtenberg,
1971.
Persistence
of
pesticides
in
river
water.
Environ.
Sci.
&
Technol.,
5:
541.

Eisler,
R.,
1969.
Acute
toxicities
of
insecticides
to
marine
decapod
crustaceans.
Crustaceana,
16:
302.

Eisler,
`
R.,
1970.
Acute
toxicities
of
organochlorine
and
organophosphorus
insecticides
to
estuarine
fishes.
U.
S.
Bureau
of
Sport
Fisheries
and
Wildlife,
Technical
Paper
46.

Ellis,
M.
M.,
1937.
Detection
and
measurement
of
stream
pollution.
Bull.
US.
Buraue
of
Sport
Fisheries,
and
Wildlife.
48
(
22)
:
365­
437.

Environmental
Protection
Agency,
1973.
The
control
of
pollution
irom
hydrographic
modifications.
EPA
430/
9­
73­
017,.
U.
S.
Government
Printing
Office,
Washington,
D.
C.

Environmental
Protection
Agency,
1985a.
Technical
Support
Document
for
Water
Quality­
Based
Toxics
Control.
Office
of
Water,
Washington,
D.
C.

Everett,
G.
V.,
1973.
Rainbow
trout,
Salmon
gairdneri
(
Rich.)
,
fishery
of
Lake
Titicaca.
J.
Fish.
Biol.,
5:
429­
440.

European
Inland
Fisheries
Advisory
Commission,
1969.
Water
quality
criteria
for
Eurpoean
freshwater
fish­­
extreme
pH
values
and
inland
fisheries,
prepared
by
EIFAC
Working
Party
on
Water
Quality
Criteria
for
European
Freshwater
Fish.
Water
Research,
3:
593.

ERL­
Duluth,
Environmental
Reseach
Laboratory,
1976,
Procedures
for
developing
temperature
criteria
for
freshwater
fish.
Ecological
Research
Series,
Report
in
Preparation.

Fairbridge,
R.
W.
(
ed.)
,
1966.
The
encyclopedia
of
oceanography.
Reinhold,
New
York,
New
York.

Federal
Water
Pollution
Control
Administration,
1967.
Temperature
and
aquatic
life.
Laboratory
Investigations­
No.
6,
Technical
Advisory
and
Investigations
Branch,
Cincinnati,
Ohio.
Fetterolf
,
C.
M.
Jr.,
1973.
Mixing
zone
concepts:
Biological
Methods
for
the
Assesment
of
Water
Quality,
ASTM
STP­
528,
American
Society
for
Testing
and
Materials,
pp.
31­
45.

Food
and
Drug
Administration,
1974.
Poisonous
or
deleterious
substances
in
peanuts,
evaporated
milk,
fish
and
shellfish.
Proposed
Rules.
Federal
Register,
December
6,
1974,
Washington
,
D.
C.

temperature,
In:
Thermobiology.
A.
H.
Rose
(
ed.)
,
Academic
Press,
New
York.
Fry,
F.
E.
J.,
1967.
Responses
of
vertebrate
poikilotherms
to
Furch,
X.,
1972.
Der
Einfluss
einer
Vorbehandlung
mit
konstanten
und
wechselnden
temperaturen
auf
die
hitzerestistenz
von
Gammarus
­­
salinus
und
Idotea
­
balthica.
Mar.
Biol.,
15:
12.

Gallagher,
T.
P.,
et
al.,
1969.
Pollution
affecting
shellfish
harvesting
in
Mobile
Bay,
Alabama,
Tech.
Servs.
F.
W.
P.
C.
A.
Southeast
Water
Lab.,
Athens,
GA.

Gammon,
J.
R.,
1970.
The
effect
of
inorganic
sediment
on
stream
biota.
Environmental
Protection
Agency.
Water
Poll.
Cont.
Res.
Series,
18050
DWC
12/
70,
USGPO,
Washington,
D.
C.

Garside,
E.
T.,
1966.
Effect
of
oxygen
in
relation
to
temperature
on
the
development
of
embryos
of
brook
trout
and
rainbow
trout.
J.
Fish.
Res.
Bd.
Canada.
23:
1121­
1134.

insecticides
to
various
aquatic
invertebrates.
Water
and
Sew,
Works,
112:
276.

Geldreich,
E.
E.
and
Kenner,
1969.
Concepts
of
fecal
streptococci
in
stream
pollution.
Jour.
Water
Poll.
Contr.
Fed.
41:
R336.

Geldreich,
E.
E.,
1972.
Buffalo
Laboratory
recreational
water
quality:
A
study
in
bacteriological
data
interpretation.
Water
Ree.
6:
912.

Geldreich,
E.
E.,
1974.
Microbiological
criteria
concepts
for
coastal
bathing
waters.
Ocean
Mgt.
(
In
press).

Geldreich,
E.
E.
,
1974.
Monitoring
marine
waters
for
microbiological
quality.
WHO
Conference
­
Scientific
Aspect
of
Marine
Pollution.
Geneva
Switzerland.

Gibson,
J.
R.,
­
et
­
­
.
I
a1
1969.
Sources
of
error
in
the
use
of
fish­
brain
acetylcholinesterase
activity
as
a
monitor
for
pollution.
Bull.
Environ.
Contam.
Toxicol.,
4:
17.

Gillette,
L.
A.,
et
al.!
1952.
Appraisal
of
a
chemical
waste
problem
by
fish
toxicity
tests.
Sewage
Ind.
Wastes,
24:
1397.

Glude,
J.
B.,
1954.
The
effects
of
temperature
and
predators
on
the
abundance
of
the
soft­
shell
clam,
arenaria,
in
New
England.
Trans.
Amer.
Fish.
SOC.,
04:
13.

Radioactivity
in
the
marine
enviroment.
Sciences,
Washington,
D.
C.,
p.
137.

Gonzalez
J.
G.,
1972.
Seasonal
variation
in
the
responses
of
estuarine
populations
to
heated
water
in
the
vicinity
of
a
steam
generating
plant.
Ph.
D.
Dissertation,
Uinv.
Rhode
Island,
p.
142.
Gaufin,
A.
R.,
et
al.,
1965.
The
toxicity
of
ten
organic
Goldgerg,
E.
D.,
et
al.,
1971.
Marine
chemistry.
In:
National
Academy
of
Griffin,
A.
E.,
1960.
Significance
and
removal
of
manganese
in
water
supplies.
Jour.
Amer.
Water
Works
Assn.,
52:
1326.

Griffith,
E.
W.,
1963.
Salt
as
a
possible
limiting
factor
to
the
Suisan
Marsh
pheasant
population.
Wildlife
Protection
Study,
Cooperative
Study
of
California.

Guarraia,
L.
J.,
1972.
Brief
literature
review
of
gebsiellg
as
pathogens.
In
seminar
on
the
significance
of
Fecal
coliform
in
Industrial
Waste,
E.
P.
A.,
T.
R.
3,
National
Field
Investigations
Center,
Denver,
CO,
p.
94.

­
lucius
L.)
when
incubated
under
different
oxygen
conditions.
Probs.
of
Ichthyol.,
9:
841­
851.

Haas,
A.
R.
C.,
1932.
Nurtirional
aspects
in
mottleleaf
and
other
physiological
diseases
of
citrus.
Hilgardia,
6:
483.

Hamdorf,
K.,
1961.
Die
Beeinflussung
der
Embryonal
­
und
Larvalentqicklung
der
Regenbogenforelle
(
Salmo
iridues
Gibb.)
durch
die
Umwelgfactoren
02­
Partialdruck
and
Temperatur.
verg
1.
Physiol
.
44
:
4
51­
4
62.

15:
8.

techniques
for
coagulation­
filtration.
jour.
Amer.
Water
Works
Assn.,
59:
1149.

Harmeson,
R.
H.,
et
al.,
1971.
The
nitrate
situation
in
Illinois.
Jour.
Amer.
Water
Works
Assn.,
63:
303.

Heath,
R.
G.,
et
al.,
1972.
Comparative
dietary
toxicities
of
pesticides
to
birds.
Wildlife
Report
No.
152,
U.
S.
Dept.
of
the
Interior,
Washington,
D.
C.,
p.
5
1
.

environment;
A
continuing
controversy.
Intercont.
Med.
Book
Corp.,
N.
Y.,
pp.
421­
435.

concentrations
on
the
growth
of
junenile
coho
salmon.
Trans.
Amer.
Fish.
SOC.
91:
155­
167.

inorganic
pollutants
on
young
salmon
and
trout.
Department
Fish.
Res.,
Bull.
No.
5,
p.
264.
Annual
Report
Belta
Fish
and
0
Gulidov,
M.
V.,
1969.
Embryonic
develoopment
of
the
pike
(=

2.

Hampson,
G.
R.,
and
H.
L.
Sanders,
1969.
Local
oil
spill.
Oceanus,

Hannah,
S.
A.,
J
.
M
.
Cohen,
and
G.
G.
Robeok,
1967.
Control
Bureau
of
Sport
Fisheries
and
Wildlife,

Heath,
R.
G.,
and
J.
W.
Spann,
1973.
Pesticides
in
the
Hermann,
R.
B.,
et
al.,
1962.
Influence
of
dissolved
oxygen
Holland,
G.
A.,
et
al.,
1960.
Toxic
effects
of
organic
and
Washington,

Holland,
HeT.,,
a&.,
1967.
Use
of
fish
brain
.
L
acetylocholinesterase
to
monitor
pollution
by
organophosphorus
­
w
pesticides.
Bull.
Environ.
Contsm.
Toxicol.,
2:?~
56.­

HOSS,
D.
I.,
et
al.,
1975.
Effects
of
temperature,
copper
and
chlorine
on
fish
during
simulated
entrainment
in
power
plant
condenser
cooling
systems.
Symp.
Physical
and
Biological
Effects
on
the
Environment
of
Cooling
Systems
and
Thermal
Discharges
at
Nuclear
Power
Plants.
Intll
Atomic
Energy
Agency
(
In
press).

comparison
of
total
coliform
and
fecal
coliform
values
in
shellfish
growing
areas
and
a
proposed
fecal
coliform
growing
area
standard.
Presented
at
8th
National
Shellfihs
Sanitation
Workshop.
(
FDA,
Washington,
D.
C.)
Hunt,
D.
A.
and
J.
Springer,
1974.
Preliminary
repdrt
on
tlA1t
Hutchinson,
G.
E.,
1957.
A
treatise
on
limnology.
John
Wiley
&

Hyde,
X.
M.,
et
e.,
1974.
The
effect
of
mirex
on
channel
catfish
Idler,
D.
R.,
1969.
Coexistence
of
a
fishery
and
a
major
industry
Illig,
G.
L.
Jr.,
1960.
Significance
and
removal
of
manganese
in
Itazawa,
Y.
1971.
An
estimation
of
the
minimum
level
of
Sons,
New
York.

production,

in
Placentia
Bay.
Chemistry
in
Canada,
21(
11):
16.

water
supplies.
Jour.
Uer.
Water
works
Assn.,
52:
1326.

dissolved
oxygen
in
water
required
for
normal
life
of
fish.
Bull.
Jap.
SOC.
Sci.
Fish.
37:
273­
276.
Trans.
Amer.
Fish
SOC.,
103:
366.

Jackim,
E.,
et
a1
.,
1970.
Effects
of
metal
poisoning
for
five
1
iver
enzymes
in
the
kill
if
ish
(
Fundulus
­
heterocl
itus).
Fish.
Res.
Bd.
Can.,
27:
383.
Jour.

Jacobson,
S.
M.
and
D.
B.
Boylan,
1973.
Effect
of
seawater
soluble
fraction
of
kerosene
on
chemotaxis
in
a
marine
snail,
Nassarius
obsoletus.
Nature,
241:
213.

Jangaard,
P.
M.
1970.
The
role
played
by
the
Fisheries
Research
Board
of
Canada
in
the
Isredll
herring
phosphorus
pollution
crisis
in
Placentia
Bay,
Newfoundland.
Fisheries
Research
Board,
Atlantic
Regional
Office
(
Circular
No.
1)
Halifax,
Nova
Scotia.

Jef
fries,
H.
P.
and
W.
C.
Johnson,
1974.
Seasonal
distributions
of
bottom
fishes
in
the
Narragansett
Bay
area:
Seven­
year
variations
in
the
abundance
of
winter
flounder
(
Pseudopleuronectes
americanus)
.
Jour.
Fish.
Res.
Bd.
Can.,
31:
1057.

Jensen,
L.
D.
and
A.
R.
Gaufin,
1966.
Acute
and
long­
term
effects
of
organic
insecticides
on
two
species
of
stonefly
naiads.
Jour.
Water
Poll.
Cont.
Fed.,
38:
1273.
Johns
Hopkins
University,
1956.
Final
report
to
the
water
quality
subcommittee
of
the
American
Petroleum
Institute,
Project,
PG
49:
41.

Jones,
J.
R.
E.,
1964.
Fish
and
river
pollution.
Butterworth
and
Co.,
Ltd.,
London.

Kaiser,
K.
L.
E.,
1974.
Mirex:
an
unrecognized
contaminant
of
fishes
from
Lake
Ontario.
Science
185:
523.

Karnak,
R.
E.
and
W.
J.
Collins,
1974.
The
susceptibility
to
selected
insecticides
and
acetylchlolinesterase
activity
in
a
0
~

laboratory
colony
of
midge
larvae,
Chironomus
tentans
(
Dintera:
chironomidae)
.
Bull.
Enivorn.
Contam.
Toxicol.,
i2
:­
a.

Katz,
M.,
1961.
Aucte
toxicity
of
some
organic
insecticides
to
three
species
of
salmonids
and
to
the
threespine
strickleback.
Trans.
Amer.
Fish.
SOC.,
90:
264.

Kelly,
C.
B.
and
W.
Arcisy,
1954.
Survival
of
enteric
organisms
in
shellfish.
Pub.
Health
Repts.
69~
1205.

Keup,
L.
E.,
1975.
Factors
in
fish
kill
investigations.
Water
and
Sewage
Works,
121:
48.

Kinne,
O.,
1970.
Temperatur­
animals­
invertebrates,
In:
Marine
ecology.

Klinger,
R.,
1957.
Sodium
nitrite,
a
slow
acting
fish
poison.
Schweiz,
2.
Hydrol.
19(
2):
565.

Knepp,
G.
L.
and
G.
F.
Arkin,
1973.
Ammonia
toxicity
levels
and
nitrate
tolerance
of
channel
catfish.
The
Progressive
Fish­
Culturist,
35:
221.

insecticides
to
striped
bass,
Morone
szxatilk.
Calif.
Fish
and
Game,
60:
128.

Kopfler,
F.
C.,
1974.
The
accumulation
of
organic
and
inorganic
mercury
compounds
by
the
eastern
oyster
(
Crassostrea
virginica).
Bull.
Environ.
Contam.
Toxicol.
11:
275.

United
States.
U.
S.
Dept.
of
the
Interior,
Federal
Water
Pollution
Control
Admin.,
Cincinnati,
Ohio.

Kovacs,
N.,
1959.
Enteric
fever
in
connection
with
pollution
of
seawater.
Western
Australia.
Renort
of
the
Commissioner
Of
0.
Kinne
(
ed.),
John
Wiley
and
Sons,
New
York.
0
Korn,
S.
and
R.
Earnest,
1974.
Acute
toxicity
of
twenty
Kopp,
J.
F.
and
R.
C.
Kroner,
1967.
Trace
metals
in
waters
of
the
k
l
i
c
Health
for
the
year
1958.
­

0
Kraus,
as
cited
by
Mitchess,
J.
W.,
R.
E.
Hogson,
and
C.
R.
Gaetjens,
1946.
Tolerance
of
farm
animals
to
feed
containing
2,
4­
dichlorophenoxyacetic
acid.
Jour.
Animal.
Sci.,
5:
226.

Lange,
N.
A.,
Ed.,
1961.
Handbook
of
chemistry,
10th
ed.
McGraw­
Hill
Book
Co.,
New
York.

Langelier,
W.
F.,
1936.
The
analytical
control
of
anti­
corrision
water
treatment.
Jour.
Amer.
Water
Works
Assn.,
28:
1500.

Le
Clerc,
E.
and
F.
Devlaminck,
1955.
Fish
toxicity
tests
and
water
quality.
Bull.
de
Belge
Condumeiit
Eaux.,
28:
ll.

Le
Clerc,
E.,
1960.
The
self­
purification
of
streams
and
the
relationship
between
chemical
and
biological
tests.
Proc.
Second
symposium
on
Treatment
of
Waste
Waters,
Pergamon
Press,
London,
England,
p.
281.

Leggett,
W.
C.
R.
R.
Whitney,
1972.
Water
temperature
and
the
migrations
of
American
shad,
Fish.
Bull.
70,
3:
659.

Lehman,
A.
J.,
1965.
Summaries
of
pesticide
toxicity.
Association
of
Food
and
Drug
Officials
of
the
U.
S.,
Topeka,
Kansas,
pp.
1­
40.

pressure.
Inst.
Freshwater
Res.
Drottningholm,
Report
#
29.
Fish.
Ed.
of
Sweden
(
Annual
report
for
1948),
49­
50.

water.
Arch.
Hydrobiology,
53
:
589.
Lindroth,
A.,
1949.
Vitality
of
salmon
parr
at
low
oxygen
Lindroth,
A.,
1957.
Abiogenic
gas
supersaturation
of
river
Lockhart,
E.
E.,
­
et
­.
I
a1
1955.
The
effect
of
water
impurities
on
Lowe,
J.
I.,
et
al.,
1971.
Effects
of
mirex
on
selected
estuarine
the
flavor
of
brewed
coffee.
Food
Research,
20:
598.

organisms.
In:
Transactions
of
the
36th
North
American
Wildlife
Resources
Conference.
pp.
171­
186.

Lowman,
F.
G.,,
et
al.,
1971.
Accumulation
and
redistribution
of
radionuclides
by
marine
organisms,
In:
Radioactivity
in
the
Marine
Environment,
National
Academy
of
Sciences,
Washington,
D.
C.,
p.
161.

Ludke,
J.
L.,
et
al.,
1971.
Toxicity
of
mirex
to
crayfish,
Procambarus
blandingi.
­
Bull.
Environ.
Contam.
Toxicol.,
6:
89.

Lurid,
W.
A.
and
G.
C.
Maltezos,
1970.
Movements
and
migrations
of
the
bluefish,
Pomatomus
etarix,
tagged
in
waters
of
New
York
and
southern
New
England.
Trans.
Amer.
Fish.
SOC.,
99:
719.

Macek,
X.
J.
and
McAllister,
1970.
Insecticide
susceptibility
of
some
common
fish
family
representatives.
Trans.
her.
Fish
SOC.,
99:
20.
Macek,
K.
J.,
et
al.,
1974.
Chronic
toxicity
of
lindane
to
selected
aquatic
invertebrates
and
fishes.
Environmental
Protection
Agency
Research
Contract
Report,
Enviornmental
Protection
Agency
Ecol.
Res.
Series,
(
In
preparation).

Mackenthun,
K.
M.,
1973.
Toward
a
cleaner
aquatic
environment.
US.
Government
Printing
Off
ice,
Washington,
D.
C.

Malouf,
R.,
1972.
Occurrence
of
gas
bubble
disease
in
three
species
of
bivalve
mollusks.
Jour.
Fish.
Res.
Bd.
Can.,
0
29:
588.

Marcello,
R.
A.,
et
al.,
1975.
Evaluation
of
alternative
solutions
to
gas
buble
disease
mortality
of
menhaden
at
Pilgrim
Nuclear
Power
Station.
Yankee
Atomic
Electric
Co.,
Westboro,
Mass.
YAEC­
1087.

Massengill,
R.
R.,
1973.
Change
in
feeding
and
body
condition
of
brown
bullheads
overwintering
in
the
heated
effluent
of
a
power
plant.
Ches.
Sci.
14,2:
138.

McCord,
C.
P.,
1951.
Beryllium
as
a
sensitizing
agent.
Ind.
Med.
Surg.,
20:
236.

McCoy,
E.
F.,
1972.
Role
of
bacteria
in
thenitrogen
cycle
in
lakes.
Enviornmental
Protection
Agency,
Water
Pollution
Control
Research
Series,
U.
S.
Govenrment
Printing
O
f
f
ice
(
EP
2.10:
16010
EHR
03/
72).
Washington,
D.
C.

McKim,
J.
M.,
et
al.,
1976.
Long­
term
effects
of
methyl­
mercuric
chloride
on
three
generations
of
brook
trout
(
Salvelinus
__
fontinalis)
:
Toxicity,
accumulation,
distribution,
and
elsination.
Jour.
Fish.
Res.
Board
Can.
33:
2726.

McKee,
J.
E.
and
H.
W.
Wolf,
1963.
Water
quality
criteria.
State
Water
Quality
Control
Board,
Sacramento,
CAI
pub.
3­
A.

Merna,
J.
W.
and
P.
J.
Eisele,
1973.
The
effects
of
methoxychlor
on
aquatic
biota.
U.
S.
EPA
Ecological
Res.
Series,
No.
EPA­
R3­
73­
046.
U.
S.
Government
Printing
Office,
Washington,
D.
C.

Bull.
World
Health
Organization,
44:
363.

shellfish
in
esturary
waters.
Jour.
San.
Eng.
Div.
Proc.
Amer.
SOC.
Civ.
Eng.,
94:
595.

selective
fish
eradicator.
Trans.
Amer.
Fish.
SOC.,
94:
203.
Metcalf,
R.
L.,
&
a&.,
1971.
Biodegradable
analogs
of
DDT.

Metcalf,
T.
G.
and
W.
C.
Stiles,
1968.
Viral
pollution
of
Meyer,
F.
P.,
1965.
The
experimental
use
of
guthion
as
a
Minchew,
G.
D.
and
D.
E.
Ferguson,
1970.
Toxicities
of
six
insecticides
to
resistant
and
susceptible
green
sunfish
and
golden
shiners
in
static
bioassays.
Jour.
Miss.
Acad.
Sci.,
15:
29.

Mironov,
O.
G.,
1967.
Effects
of
low
concentrations
of
petroleum
and
its
products
on
thedevelopment
of
roeoftheBlackSea
flatfish.
Vop
Ikhtiol.,
7:
557.

Mironov,
O.
G.,
1970.
The
effect
of
oil
pollution
on
flora
and
fauna
of
the
Black
Sea.
In
Poceedings:
FA0
Conference
on
Marine
Pollution
and
its
effects
on
living
resources
and
fish.
Rome,
Dec.,
1970.,
E­
92.
Food
and
Agriculture
Organization
of
the
United
Nations.

Moore,
E.
W.,
1952.
Phisiological
effects
of
the
consumption
of
saline
drinking
water.
National
Res.
Council,
Div.
of
Medical
Sciences,
Bull.
San.
Engr.,
and
Environment,
Appendix
E.

Moore,
S.
F.,
et
al.,
1973.
A
preliminary
assessment
of
the
environmental
vulnerability
of
Machias
Bay,
Maine
to
oil
super­
tankers.

Mount,
D.
I.,
1960.
Effects
of
various
dissolved
oxygen
levels
on
fish
activity.
Ohio
State
Univ.
Natur.
Resources
Inst.,
Ann.
Fisheries
Res.
Rept.
pp.
13­
33.

freshwater
fishes.
In:
Biological
aspects
of
thermal
pollution.
P.
A.
Krenkel
and
F.
L.
Parkers,
(
eds.)
,
Vanderbilt
University
Press.

survival,
growth
and
reproduction.
Water
Res.,
7:
987.

acceptable
toxicant
limits
for
fish­
malathion
and
the
butoxyethanol
ester
of
2,4­
D.
Trans.
Amer.
Fish.
SOC.,
21:
185.

Mulla,
M.
S.
and
A.
M.
Khasawina,
1969.
Laboratory
and
field
evaluations
of
larvicides
against
chironomid
midges.
Jour.
Econ.
Entomol.,
62:
37.

National
Academy
of
Sciences
­
Conmittee
on
Biologic
Effects
of
Atmospheric
Pollutants,
1972.
Lead:
Airborne
lead
in
perspective.
The
National
Academy
of
Sciences,
Washington,
D.
C.

National
Acadmey
of
Sciences,
National
Academy
of
Engineering,
Mount,
D.
I.,
1969.
Developing
thermal
requirements
for
Mount,
D.
I.,
1973.
Chronic
effect
of
low
pH
on
fathead
minnow
Mount,
D.
T.
and
C.
E.
Stephan,
1967.
A
method
for
establishing
1974.
Water
quality
criteria,
1972,
U.
S.
Government
Printing
Off
ice,
Washington,
D.
C.
National
Institute
for
Occupational
Safety
and
Health,
1972.
Occupational
exposure
to
beryllium.
U.
S.
Department
of
Health,
Education
and
Welfare,
Health
Services
and
Mental
Health
Admin.,
National
Institute
for
Occupational
Safety
and
Health.
U.
S.
Government
Printing
Off
ice,
Washington,
D.
C.

National
Research
Council,
1954.
Sodium
restricted
diets.
Publication
325,
Food
and
Nutrition
Board,
Washington,
D.
C.

National
Technical
Advisory
Committee
to
the
Secretary
of
the
Interior,
1968.
Water
quality
criteria,
U.
S.
Government
Printing
Office,
Washington,
D.
C.

Nagvi,
S.
M.
and
D.
E.
Ferguson,
1969.
Pesticide
tolerances
of
selected
freshwater
invertebrates.
Jour.
Miss.
Acad.
Sci.
,
14
:
121.

Nebeker,
A.
V.
and
A.
R.
Gaufin,
1964.
Bioassays
to
determine
pesticide
toxicity
to
the
amphipod
crustacean,
Gammarus
lactuis.
Proc.
Utah
Acad.
Sci.
Arts
and
Letters,
41:
64.

Nebeker,
A.
V.,
et
al.,
1976a.
Nitrogen,
oxygen,
and
carbon
dioxide
as
factors
affecting
fish
survival
in
gas
supersaturated
water.
Trans.
Amer.
Fish.
SOC.
(
In
Press).

salmon
molts
in
seawater
after
exposure
to
gas
supersaturated
water.
Trans.
Amer.
Fish.
SOC.
(
In
Press).

water
on
freshwater
invertebrates.
Proc.
Gas
Bubble
Disease
Workshop.
Battelle
Northwest,
ERDA
Special
Report
(
In
Press).

Neill,
W.
H.
and
J.
J.
Magnuson,
1974.
Distributional
ecology
and
behavioral
thermoregulation
of
fishes
in
relation
to
heated
effluent
from
a
power
plant
at
Lake
Monona,
Wisconsin.
Trans.
Amer.
Fish.
SOC.
103,
4:
663.

animals,
p.
273­
380.
In:
P.
Hepple
(
ed.),
Water
Pollution
by
oil.
I.
P.
London.

Copiea
No.,
55:
37.

chloramphenical,
amphicil
lin
and
other
anrimicrobial
agents:
Strains
isolated
during
an
extensive
typhoid
fever
epidemic
in
Mexico.
Antimicrobial
Agents
and
Chmeothearpy,
4:
597.
0
Nebeker,
A.
V.,
et
al.,
197633.
Survival
of
coho
and
sockeye
Nebeker,
A.
V.,
et
al.,
1975.
Effects
of
gas
supersaturated
Nelson­
Smith,
A.,
1971.
Effects
of
oil
on
marine
plants
and
Nichols,
J.
T.,
1918.
An
abnormal
winter
flounder
and
others.

Olearte,
J.
and
E.
Galindo,
1973.
­
Salmonella
typhi
resistance
to
Olla,
B.
L.
and
A.
L.
Studholme,
1971.
The
effect
of
temperature
on
the
activity
of
bluefish,
Pomatomus
saltatrix
(
1)
Biol.
Bull.,
141:
337.
0
Olson,
G.
F.,
­
_
et
al.,
1975.
Mercury
residues
in
fathead
minnows,
Pimephales
promelas
Rafinesque,
chronically
exposed
to
methyl­
mercury
in
water.
Bull.
Environ.
Contam.
Toxcol.,
14:
129.

hydrogen
sulfide
to
Gammarus
psedol
­
imnaeus.
Trans.
her.
Fish.
sac.,
103
(
In
press).
Oseid,
D.
M.
and
L.
L.
Smith,
Jr.,
197a.
Chronic
toxicity
of
Oseid,
D.
M.
and
L.
L.
Smith,
197413.
Long­
term
effects
of
hydrogen
sulfide
on
Hexaqenia
limbata
(
Ephemeroptera)
.
Environmental
Ecology
(
In
press).

oseid,
D.
M.
and
L.
L.
Smith,
1974c.
Factors
influencing
acute
toxicity
estimates
of
hydrogen
sulfide
to
freshwater
invertebrates.
Water
Research
8
(
In
press).

Palmer,
J.
S.
and
R.
D.
Radeleff,
1964.
The
toxicologic
effects
of
certain
fungicides
and
herbicides
on
sheep
and
cattle.
Ann.
N.
Y.
Acad.
Sci.,
111:
729.

Paris,
D.
F.,
et
all
1975.
Rates
of
degradation
of
malathion
by
bacteria
isolated
from
aquatic
systems.
Environ.
Sci.
&
Technol.,
9:
135.

Parker,
F.
L.
and
P.
A.
Krenkel,
1969.
Thermal
pollution:
Status
of
the
art.
Report
No.
3,
Vanderbilt
University,
School
of
Engineering,
Nashville,
Tennessee.

Patrick
R.,
1969a.
Some
effects
of
temperature
on
freshwater
algae.
In
Biological
aspects
of
thermal
pollution.
P.
A.
Krenkel
and
F.
L.
Parker,
(
eds.)
,
Vanderbilt
University
Press.

Patrick,
R.,
1969b.
Discussion
of
engineering
aspects,
soucres
and
magnitude
of
thermal
pollution.
By
P.
A.
Krenkel
and
F.
L.
Parker.
In:
Biological
aspects
of
thermal
pollution.
P.
A.
Krenkel
and
F.
L.
Parker,
(
eds.)
,
Vanderbilt
University
Press.

mineralized
water
on
houshold
plumbing
and
appliances.
Jour.
Amer.
Water
Works
Assn.,
60:
1060.

Pearce,
J.
B.,
1969.
Thermal
addition
and
the
benthos,
Capd
Cod
Canal.
Ches.
Sc.,
10:
227.

Pearse,
J.
S.,
1970.
Reproductive
periodicities
of
Indo­
Pacif
ic
invertebrates
in
the
6ulf
of
Suez.
111.
The
echinoid
Diadema
setosum
(
Leske).
Bull.
Mar.
Sci.,
20,
3:
697.

organisms
after
contamination
of
the
bottom
of
Long
Harbour,
Placentia
Bay,
Newfoundland
with
elemental
phosphorus.
In:
Effects
of
Elemental
Phosphorus
on
Marine
Life,
Fish.
Res.
Bd4
of
Canada,
Circular
2,
pp.
181­
186.
Patterson,
W.
L.
and
R.
F.
Banker,
1968.
Effects
of
highly
Peer,
D.
L.,
1972.
Observations
on
mortalities
of
e\
benthis
Phelps,
E.
B.,
1944.
Stream
sanitation.
John
Wiley
and
sons,

Phillips,
A.
M.,
Jr.,
D.
L.
Livingston
and
R.
F.
Dumas,
1960.
Inc.
,
New
York.

Effects
of
starvation
on
the
chemical
composition
of
brook
trout.
Prog.
Fish.
Cult.,
22:
147.

phosphorus
inseFticides
to
different
species
of
warmwater
fishes.
Trans.
Amer.
Fish.
SOC.,
91:
175.

Plotkin,
S.
A.
and
M.
Katz,
1967.
Minimal
infective
doses
of
viruses
for
man
by
the
oral
route.
In:
Transmission
of
Viruses
by
the
Water
Route,
G.
Berg,
Ed.,
John
Wiley
Interscience,
N.
Y.,
p.
155.

using
standard
methods.
In:
Marine
pollution
and
sea
life
(
Ruivo,
Ed.).
Fishing
News
(
Books)
Ltd,
London,
England,
pp.
212­
217.
Pickering,
Q.
H.,
et
a&.,
1962.
The
toxicity
of
organic
Portman,
J.,
1972.
Results
of
acute
tests
with
marine
organisms,

Post,
G.
and
T.
R.
Schroeder,
1971.
The
toxicity
of
four
insecticides
to
four
salmonid
species.
Bull.
Environ.
Contam.
Toxicol.,
6:
144.

Post,
G.
and
R.
A.
Leasure,
1974.
Sublethal
effect
of
malathion
to
three
salmonid
species.
Bull.
Environ.
Contam.
Toxicol.,
12:
312.

Presnell,
N.,
1974.
Discussion
of
fecal
coliforms
for
shellfish
growing
waters.
Proc.
7th
National
Shellfish
Sanitation
Workshop,
Oct.
21­
22,
1971.
Ratcliff
and
Wilt,
Eds.,
FDA.,
Washington,
D.
C.

Raney,
E.
C.,
1969.
Discussion
of
effects
of
heated
discharges
on
freshwater
fish
in
Britian.
By
J.
S.
Alabaster.
In:
Biological
aspects
of
thermal
pollution.
P.
A.
KrenkeL
and
F.
L.
Parker
(
eds.)
,
Vanderbilt
University
Press.

Rawson,
D.
S.
and
J.
E.
Moore,
1944.
The
saline
lakes
of
Saskatchewan.
Canadian
Jour.
of
Res.
,
22
:
141.

Reeves,
A.
L.,
1965.
Absorption
of
beryllium
from
the
gastrointestinal
tract.
A.
M.
A.
Arch.
Envir.
Health,
11:
209.

Reid,
L.
C.
and
D.
A.
Carlson,
1974.
Chlorine
disinfection
of
low
temperature
waters.
Jour.
Environ.
Eng.
Div.,
ASCE,
Vol.
100,
No.
EE2:
339.

Reitler,
R.
and
R.
Seligaman,
1957.
Pseudomonas
aeruginosa
in
drinking
water.
Jour.
A
p
p
l
.
Bact.,
20:
145.

Renfro,
W.
C.,
1963.
Gas
bubble
mortality
of
fishes
in
Galveston
Bay,
Texas.
Trans.
Amer.
Fish.
SOC.,
92:
320.
Richards,
L.
A.
(
Ed.)
,
1954.
Diagnosis
and
improvement
of
saline
.,

and
alkali
soils.
Agriculture
Handbook.
No.
60,
U.
S.
Government
Printing
Office,
Washington,
D.
C.

humans.
Am.
J.
Physiol.,
126:
l.

supplies.
Jour.
Amer.
Water
Works
Assn.,
50:
680.

elements
in
species
of
phytoplankton
grown
in
culture.
Jour.
Mar.
Biol.
Assn.
U.
X.,
51:
63.

Ringem,
L.
M.
and
C.
H.
Drake,
1952.
A
study
of
the
incidence
of
Ricter,
C.
O.,
and
A.
MacLean,
1939.
Salt
taste
threshold
of
Riddick,
J.
M.,
et
al.,
1958.
Iron
and
Manganese
in
water
Riley,
J.
P.
and
I.
Roth,
1971.
The
distribution
of
trace
Pkeudomonas
aeruginesa
from
various
national
sources.
Jour.
Bact.
,
64
:
841.

Romney,
E.
M.,
et
al.,
1962.
Beryllium
and
the
growth
of
bush
Romney,
E.
M.
and
J.
D.
Childress,
1965.
Effects
of
beryllium
in
beans.
Science,
135:
786.

'
plants
and
soil.
Soil
Sci.,
100:
210.

Rounsefell,
G.
A.
and
W.
H.
Everhart,
1953.
Fishery
science,
its
methods
and
applications.
John
Wiley
and
Sons,
Inc.,
New
York.

Rucker,
R.
R.,
1974.
Gas
bubble
disease:
Mortalities
of
coho
salmon,
Oncorhynchus
kisutch,
in
water
with
constant
total
gas
pressure
and
different
oxygen­
nitrogen
ratios.
National
Oceanic
and
Atmas.
Admain.,
Natl.
Mar.
Fish.
Serv.,
Northwest
Fish
Center,
Seattle,
Washington,
unpublished
manuscript.
0
Rulifson,
R.
L.
and
G.
­
el,
1971.
Nitrogen
suspesaturation
in
the
Columbia
and
Snake
Rivers.
Tech
Rept.
TS­
09­
70­
208­
016.2,
Environmental
Protection
Agency,
Region
X,
Seattle,
Washington.

trout
(
Salmon
­­­­
gaikdneri).
­­­
Russo,
R.
C.,
et
al.,
1974.
Acute
toxicity
of
nitrite
to
rainbow
Jour.
Fish.
Res,
Bd.
Can.,
31:
1653.

Russo,
R.
C.
and
R.
V.
Thurfiton,
1975,
Acute
toxicity
of
nitrite
to
cutthroat
trout
(
Salmon
clarki).
Fisheries
Bioassay
Laboratory
Tech.
Report
No.
75­
3,
Montana
State
University.

Saeki,
A.,
1965.
Stu6ie.
s
on
fish
culture
in
filtered
closed­
circulating
aquaria.
11.
on
the
carp
culture
experiments
in
the
systems.
Bull.
Jap.
Soa.
Sci.
Fish.,
31:
916.

Sanders,
H.
O.,
1969.
Toxicity
of
pesticides
to
the
crustacean,
Gammarus
lacustris.
US.
Department
of
the
Interior,
Washington,
D.
C.,
Bureau
of
Sport
Fisheries
and
Wildlife
Technical
Paper
­
No.
25.
­
Sanders,
H.
O.,
1972.
Toxicity
of
some
insecticides
to
four
­
species
of
malacostracan
crustaceans.
U.
S.
Department
of
the
Interion,
Washington,
D.
C.
Bureau
of
Sport
Fisheries
and
Wild­
life
Technical
Paper
No.
66.

pesticides
to
two
species
of
cladocerans.
Trans.
Amer.
Fish.
Sanders,
H.
O.
and
O.
B.
Cope,
1966.
Toxicities
of
several
'

SOC.,
95:
165.
.?

Sanders,
H.
O.
and
O.
B.
Cope,
1968.
The
relative
toxicities
of
several
pesticides
to
naiads
of
three
species
of
stonefly.
Limnol.
&
Oceanong.,
13:
112.

Sastry,
A.
N.,
1975.
Physiology
and
ecology
of
reproduction
in
marine
invertebrates.
In:
F.
J.
and
W.
B.
Vernberg
(
eds.)
,
Physiological
ecology
of
estuarine
organisms.
Belle
W.
Baruch
Library
of
Marine
Species,
Univ.
South
Carolina
Press,
Columbia,
South
Carolina.

Sattlemacher,
P.
G.,
1962.
Methemoblobinemia
from
nitrates
in
drinking
water.
SchrReihe.
Ver.
Wasser­,
Boden­
u.
Luthyg.
No.
21,
Fishcer,
Stuttgart.

Hi1
1
,
New
York.

as
indicators
of
sewage
pollution.
from
Sear
Outfalls.
Pergamon
Press.

Seabury,
J.
H.,
1963.
Toxicity
of
2,
4­
dichlorophenoxyacatic
acid
for
man
and
dog.
Arch.
Envir.
Health,
7:
202.

Shumway,
D.
L.
and
J.
R.
Palensky,
1973.
Impairment
of
the
flavor
of
fish
by
water
pollutants.
U.
S.
Environmental
Protection
Agency,
EPA­
R3­
73­
OlOI
U.
S.
Government
Printing
Office,
Washington,
D.
C.

Shuval,
H.
J.,
et
aJ.,
1971.
Natural
inactivation
processes
of
viruses
=
seawater.
Proc.
Natural
Specialty
Conf.
on
Disinfection.
Amer.
SOC.
of
Civil
Engineers,
N.
Y.
Saywer,
C.
N.,
1960.
Chemistry
for
sanitary
engineers.
McGraw­

Scarpino,
D.,
1974.
Human
enteric
virsuses
and
bacteriolphages
International
Symposium
0
Siefert,
R.
E.,
et
al.,
1973.
Effects
of
reduced
oxygen
concentrations
on
northern
pike
(
Esox
­
­
lucius)
embryos
and
larvae.
J.
Fish.
Res.
Bd.
Canada.
30:
849­
852.

concentrations
on
the
early
life
stages
of
mountain
whitefish,
smallmouth
bass,
and
white
bass.
Accepted
for
publication
in
the
Prog.
Fish.
Cult.
Siefert,
R.
E.,
et
al.,
1975.
Effects
of
reduced
oxygen
S
i
e
f
e
r
t
,
.
R
E
.
and
W.
A.
Spoor.,
1974.
E
f
f
e
c
t
s
o
f
reduced
oxygeriie
concantrations
on
embryos
and
larvae
of
white.
sucker,
coho
salmon,
brook
t
r
o
u
t
,
and
walleye.
Procaedings'..
of:
an
'

International
Symposium
on
t
h
e
E
a
r
l
y
Life
H
i
s
t
o
r
y
of
Fish.

Springer­
Verlag
B
e
r
l
i
n
Heidelberg
New
York.

diseases
sf
herring.
from
t
h
e
western
North
Atlantic..
Spec.:
Publ.
Corn.
Horthw.
A
t
l
a
n
t
i
c
Fish.
No.
6:
603­
610.

Slanetz,
L.
W.,
e
t
a1
1965.
Correlation
of
coliform
and
f
e
c
a
l
streptococci
indices
.
with
'
t
h
e
presence
of
Salmonella
­­
and,
,
'

enteric
viruses
i
n
sea
water
and
she2lfish.
.
Adv....
in
W
a
t
e
r
Pollution
R
e
s
.
2nd
International
Confr.,
Tokyo,
3:
17.

Slonim,
A.
R:,
1973.
Acute
t
o
x
i
c
i
t
y
of
beryllium
s
u
l
f
a
t
e
t
o
t
h
e
common
guppy.
Jour.
Water
P
o
l
l
.
Cont.
Fed.
45:
2110.

Slonim,
A.
R.
and
E.
E.
Ray,
1975.
Acute.
toxic
y
of
beryllium
s
u
l
f
a
t
e
t
o
salamander
larvae
(
Ambystoma
q~).
'
'
B
u
l
l
.
Envir.
Contam.
Toxicol.
,
1
3
:
307.

Slonim,
C.
B.
and
A.
R..;
sibnim,
1973..
Effecf':
oi
5
hardness
,
on
t
h
e
tolerance
of
t
h
e
guppy
t
o
berylliumzsul~
f
B
u
l
l
.
Envir.
Contam.
Toxicol.,
10:
295.

Smith,
R.
S.,
e
t
a
l
.
,
1951.
Bathing
water.
q
u
a
l
i
t
y
and
he.
alth.
I.
Great
Lakes
(
U.
S.
Public
Health
Service,
Cincinnati,
ohio).

Smith,
L.
L.,
1971.
Influence
of
hydrdgen.
s
u
l
f
i
d
e
.
bn'.
f
i
s
h
and
­'

arthropods.
Environinental
Protection
Agency,
:,
Project
18050
PCG,
Washington,
D.
C.

Smith,
L.
L.
and'
D.
K.
O
s
e
i
d
,
1972.
Effects
of
.
hydrogen
s
u
l
f
i
d
e
­
&
f
i
s
h
eggs
and
fry.
Water
Research,
6:
7P1.

Smith,
R.
S.
and
T.
B.
Woolsey,
,1932.
''"'
Bathing
wator
q
u
a
l
i
t
y
and.'.
health.
11.
I
n
l
a
n
d
river
and
pool.
'(
U.
S..
,&
bl'ic
Health
I
Smith,
K.
S.
and
T.
D.
Woolsey,
1961.
Bathing
'
w
a
t
e
r
q
u
a
l
i
t
y
and:"
'
Obarr,
Scotland,
May
17­
23,
1973,
Edited
by
J.
H.
S.
.,
Blaxter,
'*
<
.

pp.­
487­
495,
.
.
,

Sinderman,
C.
J.,
1965.
Effects
of
environment
on
several
:
..
.
.

­
­,.?
­
­
.
I
.
.
.
,
.
...
.
!
>
"

r;,

,
,

.
.

.
.

.
,

,
.
.
.
Service,
Cincinnati,
Ohio).
~
.
.,

Public
Health.
III.
Coastal
Waters
(
U.
5;
'
Public
Health
Snow,
J.
R.,
1958.
A
preliminary
report
on
the
comparative
.­,
.
.
..
..
..
..
.
,
.
.
.
1
.

.
.
,
..
.
Serviee,
Cincinnati,
Ohio).
.
.

t
e
s
t
i
n
g
of.
some
of
t
h
e
newer
herbicides.
Proc.
11th
Ann.
.'
Conf.,
Southeast
ASSOC.
G
a
m
e
F
i
s
h
COI~
IIB.,.
pp.
,125­
132.

,
Report
of
t
h
e
Commissioner
'
uf
h
b
l
i
c
H
e
a
l
t
i
....
..

Snow,
D.
J.
R.,
1959.
Typhoid
and
C
i
t
y
Beach.
Western
Australia,
or
t
h
e
year
1958;
p.
52.

P
h
i
l
adel
phia,
Pennsylvania.
Sollman,
T.
H..,
1957.
A
manual
of
pharmacology.
W.
B.
Saunders,
South,
G.
E.
and
R.
D.
H
i
l
l
,
1970.
Stud$
e,
s.
on
marine
a.
lgae.
of..
­:
Newfoundland.
I.
Occurrence
and
­
distr.
ibution
of­,
free
....
.
.
i
i
G
.
h
g
..,
.'

Ascophyl
__
lum
nOdOSum
i
n
Newfoundian&"
.
.
.
:
Can.
Jour.
Bot.,
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
/
46:
1697.

Sprague,
J.
B.,
1963.
Resistance
.'&
f6ur:.
fresh,
wat
to
l
e
t
h
a
l
,
high
temperature
and,
low
.
.
.
oxygen.
J..
Canada.
,
2
0
:
387.

1971.
13th
Edition,
APHA,
Amer.
.­
.
.
.
;

Cont.
Fed.
.
'
~
.
.
.
.

.
.

Standard.
Methods
for
t
h
e
Examinatio
­

Stev~
enson,
A.
H.,
­
1.953.
~

Studies
..
of
bathing
water.
q
u
a
l
i
t
y
.
.
.
and
'
'
'

Stewart,
B.
A.,
e
t
a
l
.
,
i96
N'itrate
and
other
p
utants
under
f
i
e
l
d
s
and
feedl.
ots..
.~
.
.
E
m
i
r
.
.
_
.
Sci.
.
.
.
.
Tech.,:,
.
.
.
.1:
73.6..
.
.
.

Stewart,
N.
E.,
et'
a
l
.
,
1967.
Influence
o
f
oxygen
concentration
on
t
h
e
growth
of
juvenile.,.
largemouth
bass,,
J.
Fish.,
Re's.
Bd.
Canada.,
24~
475­
494.
.
.
health..
Amer,
Sour.
'
h
b
l
i
c
.
­
.
.
.
.
Health,
43~
529;
.
.
.
.
.
.
.
.
.
.
.
.
.

­
.
.
,
,
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.

..
.
.
.
.
.
.
.
.
.
.

~
..
.
,
.
,

S
t
i
c
k
e
l
,
L.
F.,
1973.
Pesticide
residues
i
n
­
bi'rds.
and
mammals.

~

In:
Environmental
.,
po&
lution
ides.,
Jc.
p.
Edwards,
Ed..)

S
t
i
f
f
,
M.
J.,
1971.
Copper/
bicarbonate
e
q
u
i
l
i
b
r
i
a
i
n
solutions
of
Plenum
Pxess,
New
York.
pp.,
.....
:.
.
.
.
.
.
.
.

bicarbonate
ion
a
t
concentrations
simi.
lar
t,
o
tho.
se
found
i
n
natural
water.
.
Water
...
Research,
...
.
.
.
.
.
.
.
.
5:
171.:
.
.
.
.
.
.

experimental
beryl
1i.
m
poisoning.
Jour.:
.
.
.
.
.
.
Lab.
C
1
in.
Med.,
38:
173.

and
sons,
Inc.,
New
...
york',
,.
Chapter'
4.
.
.

I
i
i
i
t
i
n
g
factor.;
r
.
s
t
r
i
p
e
d
.
..,
.
.
.
.
.
bass.
.
'
Tran
'
h
e
r
.
Fish.
soc.,
Special
Pub.,
3
:
3
i
:
.
.
.
.
.
.
.
.
.
.
Stokinger,
H.
E.
and
C.
A.:.
Stroud,
1951,
Anemia
i
n
a
c
u
t
e
.
.
.

.
S
t
i
i
m
,
W.
and
J.
J.
Morgan,
.
L970.
Aquat.
ic­.
.
.
.
.
.
.
Chemistry..
John
Wiley
Talbot,
G.
B.,
,1966
,..,.
stuarine
environmental
requirements
and
.
­.
.
.
.

.
I
..

Tarzwell,
C.
M.
.
and
A.
R..
Gaufin,
.
,1953..
some^
important
biolbgica3.
..­

e
f
f
e
c
t
s
of
p
o
l
l
u
t
i
o
n
.
otte
proceedings
of
the
8th
PU
Reprinted
i
n
Biology
of
Water
Pollution,
i967.
D
e
p
C
Of
I
n
t
e
r
i
o
r
,
Washington,
­
D.
c.
.
.
.
,
..
.
.
sregarded
i
n
stream
surveys,
.,
%
.
:

i
n
d
i
s
t
r
i
a
i
.
.
'
waste
confeyence.,,
..
~

.
.­
,
.
.
.
.
J.
'
2
Tarzwell,
C.
M.
and
C.
Henderson,.
1960.

$
aylor,
E.
W.,
1968.
$
oFdrty­
secop
f
.
Director
of
Water:.
.­'
Toxicity
.
.
of
less'.
co
.
.
~
­
.
.
,
metals
to
fishes.
'
I
n
d
u
s
t
r
i
a
l
.
Wastes,
5i12.

Examination,
19
65­
66
,
Metropo
er
Board,
Londori,'~$:~
117.

..
.......
.
.
.
.
.
.
.
.
.
.
.
..

..
.
­
.
_.
^,
*.>

.
.
.
.
.
.
.
.
..
I.
.
.
.
.~
?
.
~
,
.
:
­
.
.
.
.
.
>
.
.
=.
Vc
.:
­
.
,­
..
.
.
_
.
U.
S.
Bureau
of
Sp.
o.
rt
Fisheries
and
Wildlife,
1970.
Resource
wa&
hinGtok,
&
Q~;.
.?!<.
I
.
i:
.
C
i
.
.
.
.
.
.
.
....
ii_
_:_..
:.­
i:
2
::
i
?;,,­.:,:,~.
v:%.:;
2
,
­
U.
S.
Environmental
Protection
Agency.
1985s.
Techiii&
alF*
Sup@
oft
Document
for
Water
Qualitv­
Based
Toxics
Control.
Office
of
.
.
...
i
,
......
:__­
I
5
,,'.
2<,
L..
:
\
_
"
>
Van
Donsel,
.
D:.
J..­
:
and.
.
E.
E.
.
Geld
Relations.
bips
o
f
c,
dciT;

sediments,:
..
­
W$*
k­
.
BgS:'­
,
to
gecal.
c
o
l
i
f
o
.
.
.
&',.
­
1
v;
.
­
,.
.
...
.~.
2
.
.
,
'
.
,
'
~
I
Z
si;
.
.
.
­
~
Van
Horn,
W.
M.,
1958.

Van
Ords+%
nd,
'
H.
The
effect
of
pul~
p
and
paper
m
i
l
l
wastes"
on
aquatic.­
1.
iif.
e
.
,.
 '
roc.
..
Ontari
nf.,,
5~
50.
,.
.
s..;:<
ie.
ri:
..
:.­
..

Amer.
Med.
Assn.,
1=:
1084.

van
&?
a
1,
in
,..
C,
zF.
w.
ate.
r
.
f
iqhes.

.

a0:
1119.

nput_
qutput,
mqdeJ!
s.
r
Schweiz
2.

waters.
Sewage
Ind.
Wastes,

:
.
r+~.
at+
ng
..,
to
infant
aminated
water.
Amer.""'
..'
i.'(
,
~
,
.
,
.
.
.~
.
~
*
.
L
l
i
::.
2
9.:
4
9
5
.
..
,..

.
.

,
~
..
.
.,
.
.
_._

­
I..
.,
:",
L,­.,
.,.*
~

­
­
r
.
.
"
.
.
.
­
~.
..
.
i
,

*,
uJ.
QOVERNMENT
PRlNTlNQ
OFflCE]
9
8
6
­
1
5
9­
3
0
0
5
0
4
7
2
QUALITY
CRITERIA
FOR
WATER
1986
UPDATE
#
I
*
ALDRIN­
DIELDRIN
CRITERIA:

Aquatic
Life
Dieldrin
For
dieldrin
the
criterion
to
protect
freshwater
aquatic
life
as
derived
using
the
Guidelines
is
0.0019
ug/
L
as
a
24­
hour
average,
and
the
concentration
should
not
exceed
2.5
ug/
L
at
any
time.

For
dieldrin
the
criterion
to
protect
saltwater
aquatic
life
as
derived
using
the
Guidelines
is
0.0019
ug/
L
as
a
24­
hour
average,
and
the
concentration
should
not
exceed
0.71
ug/
L
at
any
time.

Aldrin
For
freshwater
aquatic
life
the
conckntration
of
aldrin
should
not
exceed
3.0
ug/
L
at
any
time.
No
data
are
available
concerning
the
chronic
toxicity
of
aldrin
to
sensitive
freshwater
aquatic
life.

For
saltwater
aquatic
life
the
concentration
of
aldrin
should
not
exceed
1.3
ug/
L
at
any
time.
No
data
are
available
concerning
the
chronic
toxicity
of
aldrin
to
sensitive
saltwater
aquatic
life.

Human
Health
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
aldrin
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
*
Indicates
suspended,
canceled
or
restricted
by
U.
S.
EPA
office
of
Pesticides
and
Toxic
Substances
ambient
water
concentration
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,

the
levels
which
may
result
in
incremental
increase'of
cancer
risk
over
the
lifetime
are
estimated
at
and
The
corresponding
recommended
criteria
are
0.74
ng/
L,
0.074
ng/
L,

and
0.0074
ng/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
0.79
ng/
L,
0.079
ng/
L,
and
0.0079
ng/
L,
respectively.
0
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
dieldrin
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
conceqtration
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
loe6
and
The
corresponding
recommended
criteria
are
0.71
ng/
L,
0.071
ng/
L,
and
0.0071
nq/
L,
respectively.
If
these
above
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
0.76
ng/
L,
0.076
ng/
L,
and
0.0076
ng/
L,

respectively.

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
AMMONIA
sUMlrmy:

~
1
1
concentrations
used
herein
are
expressed
as
un­
ionized
ammonia
(
NH3),
because
NH3,
not
the
ammonium
ion
(
NH4')
has
been
demonstrated
to
be
the
principal
toxic
form
of
ammonia.
The
data
used
in
deriving
criteria
are
predominantly
from
flow
through
tests
in
which
ammonia
concentrations
were
measured.

Ammonia
was
reported
to
be
acutely
toxic
to
freshwater
organisms
at
concentrations
(
uncorrected
for
pH)
ranging
from
0.53
to
22.8
mg/
L
NHj
for
19
invertebrate
species
representing
14
families
and
16
genera
and
from
0.083
to
4.60
mg/
L
NH3
for
29
fish
species
from
9
families
and
18
genera.
Among
fish
species,
reported
96­

hour
LC50
ranged
from
0.083
to
1.09
mg/
L
for
salmonids
and
from
0
0.14
to
4.60
mg/
L
NH3
for
nonsalmonids.
Reported
data
from
chronic
tests
on
ammonia
with
two
freshwater
invertebrate
species,
both
daphnids,
showed
effects
at
concentrations
(
uncorrected
for
pH)
ranging
from
0.304
to
1.2
mg/
L
NH3,
and
with
nine
freshwater
fish
species,
from
five
families
and
seven
genera,
ranging
from
0.0017
to
0.612
mg/
L
NH3.

Concentrations
of
ammonia
acutely
toxic
to
fishes
may
cause
loss
of
equilibrium,
hyperexcitability,
increased
breathing,

cardiac
output
and
oxygen
uptake,
and,
in
extreme
cases,

convulsions,
coma,
and
death.
At
lower
concentrations
ammonia
has
many
effects
on
fishes,
including
a
reduction
in
hatching
success,
reduction
in
growth
rate
and
morphological
development,

and
pathologic
changes
in
tissues
of
gills,
livers,
and
kidneys.
Several
factors
have
been
shown
to
modify
acute
NH3
toxicity
in
fresh
water.
Some
factors
alter
the
concentration
of
un­

ionized
ammonia
in
the
water
by
affecting
the
aqueous
ammonia
equilibrium,
and
some
factors
affect
the
toxicity
Of
un­
ionized
ammonia
itself,
either
ameliorating
or
exacerbating
the
effects
of
ammonia.
Factors
that
have
been
shown
to
affect
ammonia
toxicity
include
dissolved
oxygen
concentration,
temperature,

p
~
,
previous
acclimation
to
ammonia,
fluctuating
or
intermittent
exposures,
carbon
dioxide
concentration,
salinity,
and
the
presence
of
other
toxicants.

The
most
well­
studied
of
these
is
pH;
the
acute
toxicity
of
NH3
has
been
shown
to
increase
as
pH
decreases.
Sufficient
data
exist
from
toxicity
tests
conducted
at
different
pH
values
to
formulate
a
mathematical
expression
to
describe
pH­
dependent
acute
NH3
toxicity.
The
very
limited
amount
of
data
regarding
effects
of
pH
on
chronic
NH3
toxicity
also
indicates
increasing
NH3
toxicity
with
decreasing
pH,
but
the
data
are
insufficient
to
derive
a
broadly
applicable
toxicity/
pH
relationship.
Data
on
temperature
effects
on
acute
NH3
toxicity
are
limited
and
somewhat
variable,
but
indications
are
that
NH3
toxicity
to
fish
is
greater
as
temperature
decreases.
There
is
no
information
available
regarding
temperature
effects
on
chronic
NH3
toxicity.

Examination
of
pH
and
temperature­
corrected
acute
NH3
toxicity
values
among
species
and
genera
of
freshwater
organisms
showed
that
invertebrates
are
generally
more
tolerant
than
fishes,
a
notable
exception
being
the
fingernail
clam.
There
is
no
clear
trend
among
groups
of
fish;
the
several
most
sensitive
tested
species
and
genera
include
representatives
from
diverse
faailies
(
Salmonidae,
Cyprinidae,
Percidae,
and
Centrarchidae).

Available
chronic
toxicity
data
for
freshwater
organisms
also
indicate
invertebrates
(
cladocerans,
one
insect
species)
to
be
more
tolerant
than
fishes,
again
with
the
exception
of
the
fingernail
clam.
When
corrected
for
the
presumed
effects
of
temperature
and
pH,
there
is
alsonoclear
trend
among
groups
of
fish
for
chronic
toxicity
values,
the
most
sensitive
species
including
representatives
from
five
families
(
Salmonidae,

Cyprinidae,
Ictaluridae,
Centrarchidae,
and
Catostomidae)
and
having
chronic
values
ranging
by
not
much
more
than
a
factor
or
two.
The
range
of
acute­
chronic
ratios
for
10
species
from
6
families
was
3
to
4
3
,
and
acute­
chronic
ratios
were
higher
for
the
species
having
chronic
tolerance
below
the
median.

Available
data
indicate
that
differences
in
sensitivities
between
warm
and
coldwater
families
of
aquatic
organisms
are
inadequate
to
warrant
discrimination
in
the
national
ammonia
criterion
between
bodies
of
water
with
"
warm"
and
"
coldwater"
fishes:

rather,
effects
of
organism
sensitivities
on
the
criterion
are
most
appropriately
handled
by
site­
specific
criteria
derivation
procedures.

Data
for
concentrations
of
NH3
toxic
to
freshwater
phytoplankton
and
vascular
plants,
although
limited,
indicate
that
freshwater
plant
species
are
appreciably
more
tolerant
to
NH3
than
are
invertebrates
or
fishes.
The
ammonia
criterion
appropriate
for
the
protection
of
aquatic
animals
will
therefore
in
all
likelihood
be
sufficiently
protective
of
plant
life.
Available
acute
and
chronic
data
for
ammonia
with
saltwater
organisms
are
very
limited,
and
insufficient
to
derive
a
saltwater
criterion.
A
few
saltwater
invertebrate
species
have
been
tested,
and
the
prawn
Macrobrachiurn
__­­
rosenbergiA
was
the
most
sensitive.
The
few
saltwater
fishes
tested
suggest
greater
sensitivity
than
freshwater
fishes.
Acute
toxicity
of
NH3
appears
to
be
greater
at
low
pH
values,
similar
to
findings
in
freshwater.
Data
for
saltwater
plant
species
are
limited
to
diatoms,
which
appear
to
be
more
sensitive
than
the
saltwater
invertebrates
for
which
data
are
available.

More
quantitative
information
needs
to
be
published
on
the
toxicity
of
ammonia
to
aquatic
life.
Several
key
research
needs
must
be
addressed
to
provide
a
more
complete
assessment
of
ammonia
toxicity.
These
are:
(
1)
acute
tests
with
additional
saltwater
fish
species
and
saltwater
invertebrate
species;
(
2
)

life­
cycle
and
early
life­
stage
tests
with
representative
freshwater
and
saltwater
organisms
from
different
families,
with
particular
attention
to
trends
of
acute­
chronic
ratios;
(
3
)

fluctuating
and
intermittent
exposure
tests
with
a
variety
of
species
and
exposure
patterns;
(
4
)
more
complete
tests
of
the
individual
and
combined
effects
of
pH
and
temperature,
especially
for
chronic
toxicity;
(
5)
more
histopathological
an6
histochenical
research
with
fishes,
which
would
provide
a
rapid
means
of
identifying
and
quantifying
sublethal
ammonia
effects;

and
(
6)
studies
on
effects
of
dissolved
and
suspended
solids
on
acute
and
chronic
toxicity.
NATIONAL
CRITERIA:

T
h
e
p
r
o
c
e
d
u
r
e
s
d
e
s
c
r
i
b
e
d
i
n
t
h
e
G
u
i
d
e
l
i
n
e
s
f
o
r
D
e
r
i
v
i
n
g
Numerical
National
Water
Q
u
a
l
i
t
y
C
r
i
t
e
r
i
a
f
o
r
t
h
e
Protection
of
Aquatic
Organisms
and
T
h
e
i
r
U
s
e
s
i
n
d
i
c
a
t
e
t
h
a
t
,
except
p
o
s
s
i
b
l
y
where
a
l
o
c
a
l
l
y
important
s
p
e
c
i
e
s
is
v
e
r
y
s
e
n
s
i
t
i
v
e
,
freshwater
a
q
u
a
t
i
c
o
r
g
a
n
i
s
m
s
a
n
d
t
h
e
i
r
u
s
e
s
s
h
o
u
l
d
n
o
t
be
a
f
f
e
c
t
e
d
unacceptably
i
f
:
8
(
1)
t
h
e
1­
hour*
average
concentration
o
f
un­
ionized
ammonia
(
i
n
mg/
L
NH3)
does
n
o
t
exceed,
more
o
f
t
e
n
t
h
a
n
once
every
3
years
on
t
h
e
average,
t
h
e
numerical
v
a
l
u
e
g
i
v
e
n
by
0.52/
FT/
FPH/
2,

where:

FT
­
­
10
°
.
03
20­
TCAP);
TCAP
<
T
5
3
0
­

0.03
20­
T);
0
<
T
<
TCAP
10
­
­

FPH
=
1
;
8
<
p
H
<
9
1+
107.4­
PH
1.25
;
6.5<
pH
­
­
<
7
.
7
TCAP
=
2
0
C
;
Salmonids
o
r
o
t
h
e
r
s
e
n
s
i
t
i
v
e
coldwater
species
p
r
e
s
e
n
t
coldwater
s
p
e
c
i
e
s
absent
=
25
C;
Salmonids
and
o
t
h
e
r
s
e
n
s
i
t
i
v
e
(*
An
a
v
e
r
a
g
i
n
g
p
e
r
i
o
d
of
1
h
o
u
r
may
n
o
t
be
a
p
p
r
o
p
r
i
a
t
e
i
f
e
x
c
u
r
s
i
o
n
s
o
f
c
o
n
c
e
n
t
r
a
t
i
o
n
s
t
o
g
r
e
a
t
e
r
t
h
a
n
1.5
times
t
h
e
average
occur
during
t
h
e
hour;
i
n
such
c
a
s
e
s
,
a
shorter
averaging
period
may
be
needed.)

(
2)
t
h
e
4­
day
a
v
e
r
a
g
e
c
o
n
c
e
n
t
r
a
t
i
o
n
of
un­
ionized
ammonia
(
i
n
mg/
L
NH3)
does
n
o
t
exceed,
more
o
f
t
e
n
t
h
a
n
once
every
3
years
on
t
h
e
a
v
e
r
a
g
e
,
t
h
e
a
v
e
r
a
g
e
*
n
u
m
e
r
i
c
a
l
v
a
l
u
e
g
i
v
e
n
by
0.80/
FT/
FPH/
RATIO,
where
FT
and
FPH
a
r
e
as
above
and:
RATIO
=
16
:
7.7
­
<
pH
­
<
9
TCAP
=
15
C;
S
a
l
m
o
n
i
d
s
o
r
o
t
h
e
r
s
e
n
s
i
t
i
v
e
coldwater
s
p
e
c
i
e
s
p
r
e
s
e
n
t
coldwater
species
absent
=
2
0
c
;
S
a
l
m
o
n
i
d
s
a
n
d
o
t
h
e
r
s
e
n
s
i
t
i
v
e
(*
Because
t
h
e
s
e
formulas
a
r
e
n
o
n
l
i
n
e
a
r
i
n
pH
and
temperature,
t
h
e
c
r
i
t
e
r
i
o
n
s
h
o
u
l
d
be
t
h
e
a
v
e
r
a
g
e
of
s
e
p
a
r
a
t
e
e
v
a
l
u
a
t
i
o
n
s
o
f
t
h
e
f
o
r
m
u
l
a
s
r
e
f
l
e
c
t
i
v
e
o
f
t
h
e
f
l
u
c
t
u
a
t
i
o
n
s
of
f
l
o
w
,
pH,
and
temperature
w
i
t
h
i
n
t
h
e
averaging
p
e
r
i
o
d
;
it
is
n
o
t
a
p
p
r
o
p
r
i
a
t
e
i
n
g
e
n
e
r
a
l
t
o
simply
a
p
p
l
y
t
h
e
formula
t
o
average
p
H
,
temperature,

and
flow.)

The
extremes
f
o
r
temperature
(
0,
30)
and
pH
(
6.5,
9)
given
i
n
t
h
e
a
b
o
v
e
f
o
r
m
u
l
a
s
a
r
e
a
b
s
o
l
u
t
e
.
I
t
is
n
o
t
permissible
w
i
t
h
c
u
r
r
e
n
t
d
a
t
a
t
o
conduct
any
e
x
t
r
a
p
o
l
a
t
i
o
n
s
beyond
these
l
i
m
i
t
s
.

I
n
p
a
r
t
i
c
u
l
a
r
,
t
h
e
r
e
i
s
r
e
a
s
o
n
t
o
b
e
l
i
e
v
e
t
h
a
t
a
p
p
r
o
p
r
i
a
t
e
c
r
i
t
e
r
i
a
a
t
p
H
>
9
w
i
l
l
be
l
o
w
e
r
t
h
a
n
t
h
e
p
l
a
t
e
a
u
between
pH
6
and
9
g
i
v
e
n
above.

C
r
i
t
e
r
i
a
c
o
n
c
e
n
t
r
a
t
i
o
n
s
f
o
r
t
h
e
p
H
r
a
n
g
e
6.5
t
o
9.0
and
t
h
e
t
e
m
p
e
r
a
t
u
r
e
r
a
n
g
e
0
C
t
o
30
C
a
r
e
p
r
o
v
i
d
e
d
i
n
t
h
e
f
o
l
l
o
w
i
n
g
t
a
b
l
e
s
.
T
o
t
a
l
ammonia
c
o
n
c
e
n
t
r
a
t
i
o
n
s
e
q
u
i
v
a
l
e
n
t
t
o
e
a
c
h
un­

ionized
ammonia
concentration
are
a
l
s
o
provided
i
n
these
t
a
b
l
e
s
.

There
a
r
e
l
i
m
i
t
e
d
d
a
t
a
on
t
h
e
e
f
f
e
c
t
of
t
e
m
p
e
r
a
t
u
r
e
on
c
h
r
o
n
i
c
t
o
x
i
c
i
t
y
.
EPA
w
i
l
l
be
c
o
n
d
u
c
t
i
n
g
a
d
d
i
t
i
o
n
a
l
research
on
t
h
e
e
f
f
e
c
t
s
o
f
t
e
m
p
e
r
a
t
u
r
e
on
ammonia
t
o
x
i
c
i
t
y
i
n
o
r
d
e
r
t
o
f
i
l
l
p
e
r
c
e
i
v
e
d
d
a
t
a
g
a
p
s
.
Because
of
t
h
i
s
u
n
c
e
r
t
a
i
n
t
y
,
a
d
d
i
t
i
o
n
a
l
s
i
t
e­
s
p
e
c
i
f
i
c
i
n
f
o
r
m
a
t
i
o
n
s
h
o
u
l
d
be
d
e
v
e
l
o
p
e
d
b
e
f
o
r
e
t
h
e
s
e
0
criteria
are
used
in
wasteload
allocation
modeling.
For
example,

the
chronic
criteria
tabulated
for
sites
lacking
salmonids
are
less
certain
at
temperatures
much
below
20
C
than
those
tabulated
at
temperatures
near
20
C.
Where
the
treatment
levels
needed
to
meet
these
criteria
below
20
C
may
be
substantial,
use
of
site­

specific
criteria
is
strongly
suggested.
Development
of
such
criteria
should
be
based
upon
site­
specific
toxicity
tests.
I)

Data
available
for
saltwater
species
are
insufficient
to
derive
a
criterion
for
saltwater.

The
recommended
exceedence
frequency
of
3
years
is
the
Agency's
best
scientific
judgment
of
the
average
amount
of
time
it
will
take
an
unstressed
system
to
recover
from
a
pollution
event
in
which
exposure
to
ammonia
exceeds
the
criterion.
A
stressed
system,
for
example,
one
in
which
several
outfalls
occur
in
a
limited
area,
would
be
expected
to
require
more
time
for
recovery.
The
resilience
of
ecosystems
and
their
ability
to
recover
differ
greatly,
however,
and
site­
specific
criteria
may
be
established
if
adequate
justification
is
provided.

The
use
of
criteria
in
designing
waste
treatment
facilities
requires
the
selection
of
an
appropriate
wasteload
a1
location
model.
Dynamic
models
are
preferred
for
the
application
of
these
criteria.
Limited
data
or
other
factors
may
make
their
use
impractical,
in
which
case
one
should
rely
on
a
steady­
state
model.
The
Agency
recommends
the
interim
use
of
1Q5
or
lQlO
for
Criterion
Maximum
Concentration
design
flow
and
745
or
7410
for
the
Criterion
Continuous
Concentration
design
flow
in
steady­

state
models
for
unstressed
and
stressed
systems
respectively.
12)
4­
day
average
concentrations
for
m
m
n
l
a
.
.
\

PU
o
c
5
c
I0
c
15
C
20
c
25
c
W
C
L
.
S
a
l
m
I
6
a
or
Other
Smsltlv.
Cold.
ater
S
D
~
C
I
~
S
Present
Un­
1onlz.
d
Amonla
l
m
g
/
l
l
t
r
NH,)

6.50
6.75
7
.
w
7.25
7.50
7.73
8.00
8.25
8.50
8.75
9
.
oo
6.50
6.75
7
.
oo
7.25
7.40
7.75
8.00
8.25
8.50
8.75
9
.
oo
0.0007
0.0012
0.0021
0.0037
0.0066
0.0105
0.0126
0.0126
O
.
O
l
2
t
0.0126
0.0126
2.5
2.5
2.5
2.5
2.5
2.3
1.53
0.87
0.49
0.28
0.16
0
.
oow
0.0017
0.0029
0.0052
0.0093
0.01
53
0.0177
0.0177
0.0177
0.0177
0.0177
2.4
2.4
2.4
2.4
2.4
2.2
I
.44
0.82
0.47
0.27
0.16
0.0013
0.0023
0.0042
0.0074
0.0132
0.022
0.025
0.025
0.025
0.025
0.025
0.0019
0.0033
0
m
5
9
0.0105
0.0186
0.03
1
0.035
0.035
0.035
0.035
0.035
0.0019
0.0033
0.0059
0.0105
0.0186
0.03
I
0.035
0.035
0.035
0.035
0.035
2.2
2.2
2.2
2.2
2.2
2.1
1.37
0.70
0.45
0.26
0.16
2.2
2.2
2
3
2.2
2.2
2
.
o
I
.33
0.76
0.44
0.27
0.16
1
A9
I
,49
1.49
I
.50
I
.%
I
.40
0.93
0.54
0.32
0.19
0.13
0.0019
0.0033
0.00%
0.0105
0.0186
0.031
0
.
a35
0.035
0.035
0.035
0.035
I
.
M
1.04
1
.04
I
.
M
1.05
0.99
0.66
0.39
0.23
..
0.15
0.10
0.0019
0.0033
0.0059
0.0105
0.0186
0.031
­
0.035
0.035
0.035
0.035
0.035
0.73
0.73
Q.
74
0.74
0.74
0.71
0.47
0.28
0.17
0.11
0
.
oc
8
;
oo
0;
0126
8.25
0.0126
8.50
0.0126
8.75
0.0120
9.00
0.0126
8.50
o;
r9
8.75
0.28
9
.
oo
0.16
0.0009
0.0017
0.0029
0.0052
0.0093
0.0153
0.0177
0.0177
0.0177
0.0177
0.0177
0.0013
0.0023
0.0042
0.0074
0.0132
0.022
0.025
0.025
0.025
0.025
0.025
0.0019
0.0033
0.0059
0.0105
0.0186
0.03
1
0.035
0.035
0.035
0.035
0.035
0.0026
0.0047
0
.
Og83
0.0148
0.026
0.043
0.0%
0.050
0.0%
0.050
0.050
Total
Amonla
(
mg/
lltor
NH,)

2.1
2.
I
2.1
2.
I
2.1
I
.98
1.51
0.76
c.
45
0.27
0.17
0
.0026
0.0047
0.0083
0.0148
0.026
0.043
0.0%
0.050
0.0%
0.0%
o.
om
I
.
a
I
A
7
I
A
7
I
.48
1
A9
1.39
0.93
0.54
0.33
0.21
0.14
0.
W26
0.0047
0.0083
0.0148
0.026
0.043
0.050
0.050
0.050
0.050
0.050
I
.03
1­
04
1.04
1.05
I
.06
I
.
oo
0.67
0.40
0
2
5
0.16
0.11
To
convert
meso
v
a
l
w
s
to
np/
ll?
er
N,
m
u
l
t
l
p
l
y
by
0.822.

t
S
l
t
c
~
p
~
~
l
t
I
c
crltrla
dorolopment
Is
rtronplv
ruggost.
d
a
t
TmDOratyr"
abwm
20
C
buause
of
the
Iln1t.
d
data
avallable
to
generato
tho
c
r
l
t
u
l
a
rrm.
nd.
tlon,
and
at
t
m
p
e
r
a
t
u
r
*
s
blow
20
C
b
W
u
s
e
of
T
b
l
i
m
1
t
.
d
data
and
b
o
a
u
Y
mall
hmgor
In
the
c
r
l
t
o
r
l
a
may
have
rlgnlflcant
Impact
on
the
Inol
Of
trmtl(
lt
r.
qu1r.
l
In
Irotlng
tho
r.
6umNnd.
d
crltorla.
o
c
5
c
10
c
I5
c
M
C
25
C
K
J
C
vH
A.
Salmcnldr
or
Omer
Sensltlve
CoId.
atu
5
p
u
l
.
r
Present
6.50
35
6.75
32
i
;
oo
28
7
­
25
23
7.50
17.4
7.75
12.2
8
.
oo
8
.
o
8.25
4.5
8.30
.
2.6
8.75
I
.47
9
.
oo
0.86
Vn­
lonlz.
d
Amanla
0.0129
0.0182
0.02
1
0.030
0.033
0
.
w
0.048
0.
W8
0
.
ow
0.091
o.
oa0
0,113
0.092
0.133
0.092
0.130
0.092
0.130
0.092
0.130
0.092
0.1x
(
mq/
llter
0.026
0.042
0
.
W6
0.095
0.128
0.159
0.1W
0.184
0.
lW
0.184
0.184
NH3)

0.036
0.059
0.093
0.135
0.181
0.22
0
2
6
0.26
0
2
6
0.26
0
3
6
33
33
26
22
I63
11.4
7.5
4.2
I
.40
0.83
2.4
31
28
25
20
15.5
10.9
7
.
I
4.1
2.3
I
.37
0
.83
30
27
24
19.7
14.9
10.5
6.9
2.3
I
.)
8
0
3
6
4.0
29
27
23
19.2
14.6
10.3
6
.8
3.9
2.3
1.42
0.91
0.0s
0.059
0.093
0.135
0.101
0.22
0.26
0.26
0.26
0.26
0.26
20
18.6
16.4
13.4
10.2
,
7
.2
4.8
2.8
1.71
1.07
0.72
0.036
0.059
0.093
0.135
0.181
0.22
0
3
6
0.26
0.26
0.26
0
2
6
14.3'
13.2
11.6
9.5
7.3
5.2
3.5
2.1
1.28
0.83
0
.58
6.50
6.75
7
.
w
7
;
25
7.
w
7.75
8
.
oo
8.25
8
:
59
8.75
9
.
oo
6.50
6.75
1.00
7.25
7.50
7
.7)
8
.
oo
0.25
8.50
8.75
9
.
w
0.009
I
0.0149
0.023
0.034
0.045
0.056
0.065
0.065
0.065
0.065
0.065
35
32
28
2)
17
A
12.2
8
.
o
4.5
2.6
I
.41
0.86
0.0129
0.02
I
0.033
0.046
0
;
OM
0.080
0
.
ox
0.092
0.092
0.092
0
.
o
n
0.0182
0.030
0.046
0.068
0.091
0.113
0.130
0.130
0.130
0.150
0.130
3s
30
26
22
16.3
4.2
2.4
I
.40
0.83
0.026
0.042
0.066
0.095
0.128
0.159
0.184
0.184
0.184
0.184
0.184
0.036
0.059
0.093
~.~
~

0.135
0.181
0.22
0.26
0.26
0;
26
0.26
0.26
31
28
25
20
15.5
10.9
7
.
I
4;
l
2.3
1.31
0.83
30
27
24
19.7
14.9
10.5
6.9
4.0
2
i3
1.33
0.86
29
27
23
19.2
14.6
10.3
6.8
3.9
2.3
1.42
0.91
0.051
0.084
0.131
0.190
0
2
6
0.32
0.37
0.31
0.37
0.37
0
3
7
29
26
23
19.0
14.5
10.2
6.8
4
.
O
2
;
4
1.52
1.01
0.051
0.131
0.190
0.26
0.32
0.37
0.37
0.37
0.37
0.37
0.084
20
18.6
16.4
13.5
10.3
7.3
4.9
2;
9
1
.
el
1.18
0.82
T
h
e
Agency
a
c
k
n
o
w
l
e
d
g
e
s
t
h
a
t
t
h
e
C
r
i
t
e
r
i
o
n
C
o
n
t
i
n
u
o
u
s
Concentration
stream
flow
averaging
period
used
f
o
r
steady­
s
t
a
t
e
w
a
s
t
e
l
o
a
d
a
l
l
o
c
a
t
i
o
n
modeling
may
be
a
s
l
o
n
g
a
s
30
d
a
y
s
i
n
s
i
t
u
a
t
i
o
n
s
i
n
v
o
l
v
i
n
g
POTWs
d
e
s
i
g
n
e
d
t
o
remove
ammonia
where
l
i
m
i
t
e
d
v
a
r
i
a
b
i
l
i
t
y
o
f
e
f
f
l
u
e
n
t
p
o
l
l
u
t
a
n
t
c
o
n
c
e
n
t
r
a
t
i
o
n
and
r
e
s
u
l
t
a
n
t
concentrations
i
n
r
e
c
e
i
v
i
n
g
waters
can
be
demonstrated.

I
n
c
a
s
e
s
where
low
v
a
r
i
a
b
i
l
i
t
y
c
a
n
be
d
e
m
o
n
s
t
r
a
t
e
d
,
l
o
n
g
e
r
a
v
e
r
a
g
i
n
g
p
e
r
i
o
d
s
f
o
r
t
h
e
ammonia
C
r
i
t
e
r
i
o
n
C
o
n
t
i
n
u
o
u
s
C
o
n
c
e
n
t
r
a
t
i
o
n
(
e
.
g
.
,
30­
day
a
v
e
r
a
g
i
n
g
p
e
r
i
o
d
s
)
would
b
e
a
c
c
e
p
t
a
b
l
e
because
t
h
e
magnitude
and
d
u
r
a
t
i
o
n
of
exceedences
a
b
o
v
e
t
h
e
C
r
i
t
e
r
i
o
n
C
o
n
t
i
n
u
o
u
s
C
o
n
c
e
n
t
r
a
t
i
o
n
w
o
u
l
d
be
s
u
f
f
i
c
i
e
n
t
l
y
l
i
m
i
t
e
d
.
These
m
a
t
t
e
r
s
are
discussed
i
n
more
d
e
t
a
i
l
i
n
t
h
e
Technical
Support
Document
f
o
r
Water
Quality­
Based
Toxics
C
o
n
t
r
o
l
(
U.
S.
EPA,
1985a).

(
50
F.
R.
30784,
J
u
l
y
29,
1985)
SEE
APPENDIX
A
FOR
METHODOLOGY
BERYLLIUM
CRITERIA:

8
Aquatic
Life
The
available
data
for
beryllium
indicate
that
acute
and
chronic
toxicity
to
freshwater
aquatic
life
occur
at
concentrations
as
low
as
130
and
5.3
ug/
L,
respectively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
Hardness
has
a
substantial
effect
on
acute
toxicity.

The
limited
saltwater
data
base
available
for
beryllium
does
not
permit
any
statement
concerning
acute
or
chronic
toxicity.

Human
Health
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
beryllium
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentration
should
be
zero,
based
on
the
non
threshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
and
The
corresponding
recommended
criteria
are
68
ng/
L,
6.8
ng/
L,
and
0.68
ng/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
1170
ng/
L,
117.0
nq/
L,
and
11.71
ng/
L,

respectively.

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
a
BORON
CRITERION:

750
ug/
L
f
o
r
long­
term
i
r
r
i
g
a
t
i
o
n
on
s
e
n
s
i
t
i
v
e
crops.

INTRODUCTION:

Boron
i
s
n
o
t
found
i
n
i
t
s
e
l
e
m
e
n
t
a
l
form
i
n
n
a
t
u
r
e
:
it
i
s
u
s
u
a
l
l
y
found
a
s
a
sodium
o
r
c
a
l
c
i
u
m
b
o
r
a
t
e
s
a
l
t
.
Boron
s
a
l
t
s
a
r
e
u
s
e
d
i
n
f
i
r
e
r
e
t
a
r
d
a
n
t
s
,
t
h
e
p
r
o
d
u
c
t
i
o
n
of
g
l
a
s
s
,
l
e
a
t
h
e
r
t
a
n
n
i
n
g
and
f
i
n
i
s
h
i
n
g
i
n
d
u
s
t
r
i
e
s
,
c
o
s
m
e
t
i
c
s
,
photographic
m
a
t
e
r
i
a
l
s
,
m
e
t
a
l
l
u
r
g
y
a
n
d
f
o
r
h
i
g
h
e
n
e
r
g
y
r
o
c
k
e
t
f
u
e
l
s
.

Elemental
boron
a
l
s
o
can
be
used
i
n
n
u
c
l
e
a
r
r
e
a
c
t
o
r
s
f
o
r
neutron
absorption.
Borates
a
r
e
used
a
s
"
burnable"
poisons.

RATIONALE:

Boron
is
an
e
s
s
e
n
t
i
a
l
element
f
o
r
growth
of
p
l
a
n
t
s
b
u
t
t
h
e
r
e
i
s
no
e
v
i
d
e
n
c
e
t
h
a
t
it
is
r
e
q
u
i
r
e
d
by
a
n
i
m
a
l
s
.
The
maximum
c
o
n
c
e
n
t
r
a
t
i
o
n
found
i
n
1
,
5
4
6
s
a
m
p
l
e
s
o
f
r
i
v
e
r
and
l
a
k
e
w
a
t
e
r
s
from
v
a
r
i
o
u
s
p
a
r
t
s
o
f
t
h
e
United
S
t
a
t
e
s
was
5.0
mg/
L:
t
h
e
mean
v
a
l
u
e
was
0.1
mg/
L
(
Kopp
and
Kroner,
1
9
6
7
)
.
Ground
waters
could
c
o
n
t
a
i
n
s
u
b
s
t
a
n
t
i
a
l
l
y
h
i
g
h
e
r
c
o
n
c
e
n
t
r
a
t
i
o
n
s
a
t
c
e
r
t
a
i
n
places.

The
concentration
i
n
seawater
i
s
reported
a
s
4.5
mg/
L
i
n
t
h
e
forn
o
f
b
o
r
a
t
e
(
N
A
S
,
1
9
7
4
)
.
N
a
t
u
r
a
l
l
y
o
c
c
u
r
r
i
n
g
c
o
n
c
e
n
t
r
a
t
i
o
n
s
of
boron
should
have
no
effects
on
a
q
u
a
t
i
c
l
i
f
e
.

The
minimum
l
e
t
h
a
l
dose
f
o
r
minnows
exposed
t
o
b
o
r
i
c
acid
a
t
20
OC
f
o
r
6
h
o
u
r
s
was
r
e
p
o
r
t
e
d
t
o
be
1
8
,
0
0
0
t
o
19,000
mg/
L
i
n
d
i
s
t
i
l
l
e
d
water
and
19,000
t
o
19,500
mg/
L
i
n
hard
water
(
Le
Clerc
and
Devlaminck,
1955:
Le
C
l
e
r
c
,
1960).

I
n
t
h
e
d
a
i
r
y
cow,
1
6
t
o
2
0
g/
day
of
b
o
r
i
c
a
c
i
d
f
o
r
4
0
days
produced
no
ill
e
f
f
e
c
t
s
(
McKee
and
Wolf,
1963).

S
e
n
s
i
t
i
v
e
c
r
o
p
s
h
a
v
e
shown
t
o
x
i
c
e
f
f
e
c
t
s
a
t
1
0
0
0
ug/
L
o
r
less
o
f
b
o
r
o
n
(
R
i
c
h
a
r
d
s
,
1954).
Bradford
(
1966),
i
n
a
review
o
f
b
o
r
o
n
d
e
f
i
c
i
e
n
c
i
e
s
and
t
o
x
i
c
i
t
i
e
s
,
s
t
a
t
e
d
t
h
a
t
when
t
h
e
boron
c
o
n
c
e
n
t
r
a
t
i
o
n
i
n
i
r
r
i
g
a
t
i
o
n
w
a
t
e
r
s
was
g
r
e
a
t
e
r
t
h
a
n
0.75
ug/
L,

some
s
e
n
s
i
t
i
v
e
p
l
a
n
t
s
s
u
c
h
a
s
c
i
t
r
u
s
began
t
o
show
i
n
j
u
r
y
.

Biggar
and
Fireman
(
1960)
showed
t
h
a
t
with
n
e
u
t
r
a
l
and
a
l
k
a
l
i
n
e
s
o
i
l
s
o
f
h
i
g
h
a
b
s
o
r
p
t
i
o
n
c
a
p
a
c
i
t
i
e
s
,
water
c
o
n
t
a
i
n
i
n
g
2
ug/
L
boron
m
i
g
h
t
b
e
u
s
e
d
f
o
r
some
t
i
m
e
w
i
t
h
o
u
t
i
n
j
u
r
y
t
o
s
e
n
s
i
t
i
v
e
p
l
a
n
t
s
.
The
c
r
i
t
e
r
i
o
n
o
f
750
ug/
L
i
s
t
h
o
u
g
h
t
t
o
p
r
o
t
e
c
t
s
e
n
s
i
t
i
v
e
c
r
o
p
s
during
long­
term
i
r
r
i
g
a
t
i
o
n
.

(
QUALITY
CRITERIA
 OR
WATER,
JULY
1976)
PB­
263943
SEE
APPENDIX
C
 OR
METHODOLOGY
CHLORINATED
BENZENES
CRITERIA
e
Aquatic
Life
The
available
data
for
chlorinated
benzenes
indicate
that
acute
toxicity
to
fresh
water
aquatic
life
at
concentrations
as
low
as
250
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
the
more
toxic
of
the
chlorinated
benzenes
to
sensitive
freshwater
aquatic
life
but
toxicity
occurs
at
concentrations
as
low
as
50
ug/
L
for
a
fish
species
exposed
for
7.5
days.

The
available
data
for
chlorinated
benzenes
indicate
that
acute
and
chronic
toxicity
to
saltwater
aquatic
life
occur
at
concentrations
as
low
as
160
and
129
ug/
L,
respectively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.

Human
Health
For
comparison
purposes,
two
approaches
were
used
to
derive
criterion
levels
for
monochlorobenzene.
Based
on
available
toxicity
data,
for
the
protection
of
public
health,
the
derived
level
is
4
8
8
ug/
L.
Using
available
organoleptic
data,
for
controlling
undesirable
taste
and
odor
quality
of
ambient
water,

the
estimated
level
is
20
ug/
L.
It
should
be
recognized
that
organoleptic
data
as
a
basis
for
establishing
a
water
quality
criteria
have
limitations
and
have
no
demonstrated
relationship
to
potential
adverse
human
health
effects.
Trichlorobenzenes
Due
to
the
insufficiency
in
the
available
information
for
the
trichlorobenzenes,
a
criterion
cannot
be
derived
at
this
time
using
the
present
guidelines.

1,2,4,5­
Tetrachlorobenzene
For
the
protection
of
human
health
from
the
toxic
properties
of
1,2,4,5­
tetrachlorobenzene
ingested
through
water
and
contaminated
aquatic
organisms,
the
ambient
water
criterion
is
determined
to
be
38
ug/
L.

For
the
protection
of
human
health
from
the
toxic
properties
of
1,2,4,5­
tetrachlorobenzene
ingested
through
contaminated
aquatic
organisms
alone,
the
ambient
water
criterion
is
determined
to
be
4
8
ug/
L.

Pentachlorobenzene
For
the
protection
of
human
health
from
the
toxic
properties
of
pentachlorobenzene
ingested
through
water
and
contaminated
aquatic
organisms,
the
ambient
water
criterion
is
determined
to
be
74
ug/
L.

For
the
protection
of
human
health
from
the
toxic
properties
of
pentachlorobenzene
ingested
through
contaminated
aquatic
organisms
alone,
the
ambient
water
criterion
is
determined
to
be
8
5
ug/
L.

Hexachlorobenzene
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
due
to
exposure
of
hexachlorobenzene
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentration
should
be
zero
based
on
the
non­
threshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefor,
the
levels
which
may
result
in
incramental
increase
of
cancer
risk
0
over
the
lifetime
are
estimated
at
and
The
corresponding
recommended
criteria
are
7.2
ng/
L,
0.72
ng/
L,
and
0.072
ng/
L,
respectively.
If
the
above
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
7.4
ng/
L,
0
.
7
4
ng/
L
and
0
.
0
7
4
.
ng/
L
respectively.

(
4
5
F.
R.
79316,
November
28,
1980)
SEE
APPERDIX
B
FOR
METHODOLOGY
DICHMROPROPANES/
DICHMROPROPENES
CRITERIA:

Aquatic
Life
The
available
data
for
dichloropropanes
indicate
that
acute
and
chronic
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
23,000
and
5,700
ug/
L,
respectively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.

The
available
data
for
dichloropropene
indicate
that
acute
and
chronic
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
6,060
and
244
ug/
L,
respectively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.

The
available
data
for
dichloropropane
indicate
that
acute
and
chronic
toxicity
to
saltwater
aquatic
life
occur
at
concentrations
as
low
as
10,300
and
3,040
ug/
L,
respectively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.

The
available
data
for
dichloropropene
indicate
that
acute
toxicity
to
saltwater
aquatic
1
ife
occurs
at
concentrations
as
low
as
790
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
dichloropropene
to
sensitive
saltwater
aquatic
life.
Human
Health
Using
the
present
guidelines,
a
satisfactory
criterion
cannot
be
derived
at
this
time
because
of
insufficient
avaiiable
data
for
dichloropropanes.

For
the
protection
of
human
health
from
the
toxic
properties
of
dichloropropenes
ingested
through
water
and
contaminated
aquatic
organisms,
the
ambient
water
criterion
is
determined
to
be
87
ug/
L.

For
the
protection
of
human
health
from
the
toxic
properties
of
dichloropropenes
ingested
through
contaminated
aquatic
organisms
alone,
the
ambient
water
criterion
is
determined
to
be
14.1
mg/
L.

(
45
F
.
R
.
79318,
November
2
8
,
1980)
SEE
APPENDIX
B
 OR
METHODOLOGY
CRITERIA:

­
For
endrin
the
crit
*
ENDRIN
Aquatic
Life
rion
to
protect
fr
shwater
aquatic
li
e
as
derived
using
the
Guidelines
is
0.0023
ug/
L
as
a
24­
hour
average,
and
the
concentration
should
not
exceed
0.18
ug/
L
at
any
time.

For
endrin
the
criterion
to
protect
saltwater
aquatic
life
as
derived
using
the
Guidelines
is
0.0023
ug/
L
as
a
24­
hour
average,

and
the
concentration
should
not
exceed
0.037
ug/
L
at
any
time.

Human
Health
The
ambient
water
quality
criterion
for
endrin
is
recommended
to
be
identical
to
the
existing
water
standard
which
is
1.0
ug/
L.

Analysis
of
the
toxic
effects
data
resulted
in
a
calculated
level
which
is
protective
of
human
health
against
the
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms.
The
calculated
value
is
comparable
to
the
present
standard.
For
this
reason
a
selective
criterion
based
on
exposure
solely
from
consumption
of
6.5
g
of
aquatic
organisms
was
not
derived.

*
Indicates
suspended,
canceled
or
restricted
by
U.
S.
EPA
Office
of
Pesticides
and
Toxic
Substances
(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
HEPTACHMR
CRITERIA:

Aquatic
Life
For
heptachlor
the
criterion
to
protect
freshwater
aquatic
0
life
as
derived
using
the
Guidelines
is
0.0038
ug/
L
as
a
24­
hour
average,
and
the
concentration
should
not
exceed
0.52
ug/
L
at
any
time.

For
heptachlor
the
criterion
to
protect
saltwater
aquatic
life
as
derived
using
the
Guidelines
is
0.0036
ug/
L
as
a
24­
hour
average,
and
the
concentration
should
not
exceed
0.053
ug/
L
at
any
time.

Human
Health
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
heptachlor
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentration
should
be
zero,
based
on
the
non
threshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
and
The
corresponding
recommended
criteria
are
2.78
ng/
L,
0.28
ng/
L,
and
0.026
ng/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
2.85
ng/
L,
0.29
ng/
L,
and
0.029
ng/
L,

respectively.

(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
CRITERIA
:
HE?
ACHLOROCYCU3HEXANE
Aquatic
Life
 or
lindane
the
criterion
to
protect
freshwater
aquatic
life
as
derived
using
the
Guidelines
is
0.080
ug/
L
as
a
24­
hour
average
and
the
concentration
should
not
exceed
2
.
0
ug/
L
at
any
time.

For
saltwater
aquatic
life
the
concentration
of
lindane
should
not
exceed
0.16
ug/
L
at
any
time.
No
data
are
available
concerning
the
chronic
toxicity
of
lindane
to
sensitive
saltwater
aquatic
life.

BHC
The
available
data
for
a
mixture
of
isomers
of
BHC
indicate
that
acute
toxicity
to
freshwater
aquatic
life
occurs
at
concentrations
as
low
as
100
ug/
L
and
would
occur
at
lower
concentrations
anong
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
a
mixture
of
isomers
of
BHC
to
sensitive
freshwater
aquatic
life.

The
available
data
for
a
mixture
of
isomers
of
BHC
indicate
that
acute
toxicity
to
saltwater
aquatic
life
occurs
at
concentrations
as
l
o
w
as
0.34
ug/
L
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.
No
data
are
available
concerning
the
chronic
toxicity
of
a
mixture
of
isomers
of
BHC
to
sensitive
saltwater
aquatic
life.
Human
Health
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
alpha­

hexachlorocyclohexane
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentrations
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
10­
6
and
The
corresponding
recommended
criteria
are
9
2
ng/
L,
9.2
ng/
L,
and
.92
ng/
L,

respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
310
ng/
L,
31.0
ng/
L,
and
3.10
ng/
L,
respectively.

For
the
maximum
protection
ofrhuman
health
from
the
potential
carcinogenic
effects
of
exposure
to
beta­
hexachlorocyclohexane
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentrations
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
thc
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
lo+,
and
The
corresponding
recommended
criteria
are
163
ng/
L,
16.3
ng/
L,

and
1.63
ng/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
547
ng/
L,
54.7
ng/
L,
and
5.47
ng/
L,

respectively.
For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
due
to
exposure
of
gama­

hexachlorocyclohexane
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentrations
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chenical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
lom5,
and
The
corresponding
recommended
criteria
are
186
ng/
L,
18.6
ng/
L,
and
1.86
ng/
L,

respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
625
ng/
L,
62.5
ng/
L,
and
6.25
ng/
L,
respectively.

For
the
maximum
protection
of
human
health
from
the
potential
carcinogenic
effects
of
exposure
to
technical­

hexachlorocyclohexane
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentrations
should
be
zero,
based
on
the
nonthreshold
assumption
for
this
chemical.
However,
zero
level
may
not
be
attainable
at
the
present
time.
Therefore,
the
levels
which
may
result
in
incremental
increase
of
cancer
risk
over
the
lifetime
are
estimated
at
and
The
corresponding
recommended
criteria
are
123
ng/
L,
12.3
ng/
L,
and
1.23
ng/
L,
respectively.
If
these
estimates
are
made
for
consumption
of
aquatic
organisms
only,
excluding
consumption
of
water,
the
levels
are
414
ng/
L,
41.4
ng/
L,
and
4.14
ng/
L,

respectively.
0
Using
the
present
guidelines,
satisfactory
criteria
cannot
be
derived
at
this
tine
for
delta
and
epsilon
hexachlorocyclohexane
because
of
insufficient
available
data.
a
(
45
F.
R.
79318,
November
28,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
MIREX
CRITERION:

0.001
ug/
L
f
o
r
freshwater
and
marine
a
q
u
a
t
i
c
l
i
f
e
.

RATIONALE:
0
Mirex
is
used
t
o
c
o
n
t
r
o
l
t
h
e
imported
f
i
r
e
a
n
t
S
o
l
e
n
s
s
i
s
saevissima
r
i
c
h
t
e
r
i
i
n
t
h
e
southeastern
United
S
t
a
t
e
s
.
Its
use
is
e
s
s
e
n
t
i
a
l
l
y
l
i
m
i
t
e
d
t
o
t
h
e
c
o
n
t
r
o
l
of
t
h
i
s
i
n
s
e
c
t
and
it
i
s
a
l
w
a
y
s
p
r
e
s
e
n
t
e
d
i
n
b
a
i
t
.
I
n
t
h
e
most
common
f
o
r
m
u
l
a
t
i
o
n
,

t
e
c
h
n
i
c
a
l
grade
mirex
i
s
d
i
s
s
o
l
v
e
d
i
n
soybean
o
i
l
and
sprayed
on
corncob
g
r
i
t
s
.
The
b
a
i
t
produced
i
n
t
h
i
s
manner
c
o
n
s
i
s
t
s
of
0.3
__

p
e
r
c
e
n
t
mirex,
14.7
p
e
r
c
e
n
t
soybean
o
i
l
and
85
percent
corncob
g
r
i
t
s
.
b
a
i
t
o
f
t
e
n
i
s
a
p
p
l
i
e
d
a
t
a
r
a
t
e
o
f
1
.
4
kg/
ha,

e
q
u
i
v
a
l
e
n
t
t
o
4.2
grams
of
t
o
x
i
c
a
n
t
per
hectare.
The
mirex
R
e
l
a
t
i
v
e
l
y
f
e
w
s
t
u
d
i
e
s
have
been
made
of
t
h
e
effects
of
mirex
on
freshwater
i
n
v
e
r
t
e
b
r
a
t
e
s
o
f
'
these,
o
n
l
y
Ludke
e
t
a
l
.
(
1
9
7
1
)

r
e
p
o
r
t
chemical
a
n
a
l
y
s
e
s
of
mirex
i
n
t
h
e
w
a
t
e
r
.
T
h
e
i
r
s
t
u
d
y
r
e
p
o
r
t
e
d
e
f
f
e
c
t
s
on
two
c
r
a
y
f
i
s
h
s
p
e
c
i
e
s
exposed
t
o
mirex
by
three
techniques.
F
i
r
s
t
,
f
i
e
l
d­
c
o
l
l
e
c
t
e
d
c
r
a
y
f
i
s
h
were
exposed
t
o
s
e
v
e
r
a
l
s
u
b
l
e
t
h
a
l
c
o
n
c
e
n
t
r
a
t
i
o
n
s
o
f
t
e
c
h
n
i
c
a
l
g
r
a
d
e
mirex
s
o
l
u
t
i
o
n
s
f
o
r
v
a
r
i
o
u
s
p
e
r
i
o
d
s
o
f
t
i
m
e
:
second,
c
r
a
y
f
i
s
h
were
e
x
p
o
s
e
d
t
o
m
i
r
e
x
l
e
a
c
h
e
d
f
r
o
m
b
a
i
t
(
0.3
p
e
r
c
e
n
t
a
c
t
i
v
e
i
n
g
r
e
d
i
e
n
t
)
;
and
t
h
i
r
d
,
t
h
e
c
r
a
y
f
i
s
h
were
fed
mirex
b
a
i
t
.

Procambarus
­
blandingi
j
u
v
e
n
i
l
e
s
were
exposed
t
o
1
o
r
5
ug/
L
f
o
r
6
t
o
1
4
4
h
o
u
r
s
,
t
r
a
n
s
f
e
r
r
e
d
t
o
c
l
e
a
n
w
a
t
e
r
and
observed
f
o
r
1
0
days.
A
f
t
e
r
5
days
i
n
c
l
e
a
n
water,
95
percent
o
f
the
animals
exposed
t
o
1
ug/
L
f
o
r
1
4
4
h
o
u
r
s
were
dead.
Exposure
t
o
5
ug/
L
f
o
r
6
,
2
4
,
and
58
h
o
u
r
s
r
e
s
u
l
t
e
d
i
n
2
6
,
5
0
,
and
9
8
p
e
r
c
e
n
t
m
o
r
t
a
l
i
t
y
10
d
a
y
s
a
f
t
e
r
t
r
a
n
s
f
e
r
t
o
c
l
e
a
n
water.
C
r
a
y
f
i
s
h
,
Procambarus
w,
were
exposed
to
0.1
and
0.5
ug/
L
for
48
hours.

Four
days
after
transfer
to
clean
water,
6
5
percent
of
the
animals
exposed
to
0.1
ug/
L
were
dead.
At
the
0.5
ug/
L
concentration,
71
percent
of
the
animals
were
dead
after
4
days
in
clean
water.
Tissue
residue
accumulations
(
wet
weight
basis)

ranged
from
940­
to
27,210­
fold
above
water
concentrations.
In
leached
bait
experiments,
10
bait
particles
were
placed
in
2
liters
of
water
but
isolated
from
20
juvenile
crayfish.
Thirty
percent
of
the
crayfish
were
dead
in
4
days
and
9
5
percent
were
dead
in
7
days.
Water
analysis
indicated
mirex
concentrations
of
0.86
ug/
L.
In
feeding
experiments,
108
crayfish
each
were
fed
one
bait
particle.
Mortality
was
noticed
on
the
first
day
after
feeding,
and
by
the
sixth
day
77
percent
were
dead.
In
another
­

*

experiment,
all
crayfish
were
dead
4
days
after
having
been
fed
2
bait
particles
each.
From
this
report
it
is
obvious
that
mirex
is
extremely
toxic
to
these
species
of
crayfish.
Mortality
and
accumulation
increase
with
time
of
exposure
to
the
insecticide.

Concentrations
as
l
o
w
as
0.1
ug/
L
or
the
ingestion
of
one
particle
resulted
in
death.

Research
to
determine
effects
of
mirex
on
fish
has
been
concentrated
on
species
which
have
economic
and
sport
fishery
importance.
Hyde
et
al.
(
1974)
applied
mirex
bait
(
0.3
percent
mirex)
at
the
standard
rate
(
1.4
kg/
ha)
in
four
ponds
containing
channel
catfish,
­
Ictalurus
punctatus.
Three
applications
were
made
over
an
8­
month
period
with
the
first
application
8
days
after
fingerling
(
average
weight
18.4
g
)
catfish
were
placed
in
the
ponds.
Fish
were
collected
at
each
subsequent
application
­

*
(
approximately
4­
month
intervals).
Two
and
one
half
months
after
the
final
application,
the
ponds
were
drained,
all
fish
were
measured
and
weighed,
and
the
percent
survival
was
calculated.

Mirex
residues
in
the
fish
at
termination
of
the
experiment
ranged
from
0.015
ug/
g
(
ppm)
in
the
fillet
to
0.255
ug/
g
in
the
fat.
0
In
another
study,
Van
Valin
et
al.
(
1968)
exposed
bluegills,

­
LeEmis
­_
macrochirus,
­­
and
the
goldfish,
Carassius
auratus,
to
mirex
by
feeding
a
mirex­
treated
diet
(
1,
3,
and
5
mg
mirex
per
kg
body
weight)
or
by
treating
holding
ponds
with
mirex
bait
(
1.3,
100,
and
1000
ug/
L
computed
water
concentration).
They
reported
no
mortality
or
tissue
pathology
for
the
bluegills:

however,
after
56
days
of
exposure,
gill
breakdown
in
goldfish
was
found
in
the
100
and
1000
ug/
L
contact
exposure
ponds,
and
kidney
breakdown
was
occurring
in
the
1000
ug/
L
ponds.
Mortality
in
the
feeding
experiments
was
not
related
to
the
level
of
exposure,
although
growth
of
the
bluegills
fed
5
ug/
L
mirex
was
reduced.

In
laboratory
and
field
test
systems,
reported
concentrations
of
mirex
usually
are
between
0.5
and
1.0
ug/
L
(
Van
Valin
et
al.

1968:
Ludke
et
al.
1971).
Although
mirex
seldom
is
found
above
1
ug/
L
in
the
aquatic
environment,
several
field
studies
have
shown
that
the
insecticide
is
accumulated
through
the
food
chain.

Borthwick
et
al.
(
1973)
reported
the
accumulation
of
mirex
in
South
Carolina
estuaries.
Their
data
revealed
that
mirex
was
transported
from
treated
land
and
marsh
to
the
estuary
animals
and
that
accumulation,
especially
in
predators,
occurred.
In
the
test
area,
water
samples
consistently
were
less
than
0.01
ug/
L.
R
e
s
i
d
u
e
s
i
n
f
i
s
h
v
a
r
i
e
d
from
non­
d
e
t
e
c
t
a
b
l
e
t
o
0.8
ug/
g
w
i
t
h
1
5
percent
of
t
h
e
samples
containing
residues.
The
amount
of
mirex
and
t
h
e
percent
of
samples
containing
mirex
increased
a
t
higher
t
r
o
p
h
i
c
l
e
v
e
l
s
.
F
i
f
t
y­
f
o
u
r
p
e
r
c
e
n
t
of
t
h
e
raccoons
sampled
c
o
n
t
a
i
n
e
d
mirex
r
e
s
i
d
u
e
s
up
t
o
4
.
4
ug/
g
and
7
8
p
e
r
c
e
n
t
of
t
h
e
b
i
r
d
s
c
o
n
t
a
i
n
e
d
r
e
s
i
d
u
e
s
up
t
o
17
ug/
g.
Nagvi
and
d
e
l
a
Cruz
(
1
9
7
3
)
r
e
p
o
r
t
e
d
a
v
e
r
a
g
e
r
e
s
i
d
u
e
s
f
o
r
m
o
l
l
u
s
c
s
(
0.15
ug/
g),
f
i
s
h
(
0
.
2
6
u
g
/
g
)
,
i
n
s
e
c
t
s
(
0.29
u
g
/
g
)
,
c
r
u
s
t
a
c
e
a
n
s
(
0.44
ug/
g)
and
a
n
n
e
l
i
d
s
(
0
.
6
3
u
g
/
g
)
.
They
also
r
e
p
o
r
t
e
d
t
h
a
t
mirex
was
found
i
n
a
r
e
a
s
n
o
t
t
r
e
a
t
e
d
w
i
t
h
mirex
which
s
u
g
g
e
s
t
s
movement
of
t
h
e
p
e
s
t
i
c
i
d
e
i
n
t
h
e
environment.
Wolfe
and
Norment
(
1973)
sampled
a
n
area
f
o
r
one
y
e
a
r
f
o
l
l
o
w
i
n
g
a
n
a
e
r
i
a
l
a
p
p
l
i
c
a
t
i
o
n
of
mirex
b
a
i
t
(
2
.
1
g
mirex/
ha).
C
r
a
y
f
i
s
h
r
e
s
i
d
u
e
s
ranged
from
0.04
t
o
0.16
ug/
g.
F
i
s
h
r
e
s
i
d
u
e
s
were
about
2
t
o
2
0
t
i
m
e
s
g
r
e
a
t
e
r
t
h
a
n
t
h
e
c
o
n
t
r
o
l
s
and
averaged
from
0.01
t
o
0.76
ug/
g.
K
a
i
s
e
r
(
1
9
7
4
)
,

r
e
p
o
r
t
e
d
t
h
e
p
r
e
s
e
n
c
e
of
mirex
i
n
f
i
s
h
from
t
h
e
Bay
of
Q
u
i
n
t
e
,

Lake
O
n
t
a
r
i
o
,
Canada.
C
o
n
c
e
n
t
r
a
t
i
o
n
s
range
from
0.02
ug/
g
i
n
t
h
e
gonads
of
t
h
e
northern
long
nose
gar,
Lepistosteus
osseus,
t
o
0.05
ug/
g
i
n
t
h
e
post­
anal
f
i
n
of
t
h
e
northern
pike,
Esox
l
u
c
i
u
s
.

Mirex
has
never
been
r
e
g
i
s
t
e
r
e
d
f
o
r
use
i
n
Canada.

Mirex
does
n
o
t
appear
t
o
be
g
r
e
a
t
l
y
t
o
x
i
c
t
o
b
i
r
d
s
,
w
i
t
h
LC5O's
f
o
r
the
young
of
f
o
u
r
s
p
e
c
i
e
s
ranging
from
547
t
o
g
r
e
a
t
e
r
t
h
a
n
1667
ug/
g
(
Heath
e
t
al.
1
9
7
2
)
.
Long­
term
d
i
e
t
a
r
y
dosages
caused
no
adverseeffect
a
t
3
u
g
/
g
w
i
t
h
m
a
l
l
a
r
d
s
and
13
u
g
/
g
w
i
t
h
pheasants
(
Heath
and
Spann,
1973).
However,
it
has
been
reported
(
S
t
i
c
k
e
l
e
t
a
l
.
1
9
7
3
)
t
h
a
t
t
h
e
p
e
r
s
i
s
t
e
n
c
e
of
mirex
i
n
b
i
r
d
t
i
s
s
u
e
exceeds
t
h
a
t
of
a
l
l
organochlorine
compounds
tested
except
for
DDE.
Delayed
mortality
occurred
among
birds
subjected
to
doses
above
expected
environmental
concentration.

A
summary
examination
of
the
data
available
a*
this
time
shows
a
mosaic
of
effects.
Crayfish
and
channel
catfish
survival
.
is
affected
by
mirex
in
the
water
or
by
ingestion
of
the
bait
particles.
­
Bioaccumulation
is
well
established
for
a
wide
variety
of
organisms
but
the
effect
of
this
bioaccumulation
on
the
aquatic
ecosystem
is
unknown,
There
is
evidence
that
mirex
is
very
persistent
in
bird
tissue.
Considering
the
extreme
toxicity
and
potential
for
bioaccumulation,
every
effort
should
be
made
to
keep
mirex
bait
particles
out
of
water
containing
aquatic
organisms
and
water
concentrations
should
not
exceed
0.001
uq/
L
mirex.
This
value
is
based
upon
an
application
factor
of.
0.01
applied
to
the
lowest
levels
at
which
effects
on
crayfish
have
been
observed.

Data
upon
which
to
base
a
marine
criterion
involve
several
estuarine
and
marine
crustaceans.
A
concentration
of
0.1
ug/
L
technical
grade
mirfx
in
flowing
seawater
was
lethal
to
juvenile
pink
shrinp,
Penaeus
durorarum,
in
a
3­
week
exposure
(
Lowe
et
al.

1971).
In
static
tests
with
larval
stages
(
megalopal)
of
the
mud
crab,
Rhithropanopeus
harrisii,
reduced
survival
was
observed
in
B.
l­
ug/
L
mirex
(
Bookhout
et
al.
1972).
In
three
of
four
28­
day
seasonal
flow­
through
experiments,
Tagatz
et
al.
(
1975)
found
Educed
survival
of
Call
inectes
sapidus,
Penaeus
durorarum,
and
grass
shrimp,
Palaemonetes
pugio,
at
levels
of
0.12
ug/
L
in
summer,
0.06
ug/
L
in
fall
and
0.09
ug/
L
in
winter.

Since
two
reports,
Lowe
et
al.
(
1971)
and
Bookhout
et
al.

(
1972),
stated
that
effects
of
mirex
on
estuarine
and
marine
crustaceans
were
observed
only
after
considerable
time
had
elapsed,
it
seems
reasonable
that
length
of
exposure
is
an
important
consideration
for
this
chemical.
This
may
not
be
the
case
in
fresh
water
since
the
crayfish
were
affected
within
48
hours.
Therefore,
a
3­
to
4­
week
exposure
night
be
considered
tlacutetB
and
by
applying
an
application
factor
of
0.01
to
a
reasonable
average
of
toxic­
effect
levels
as
summarized
above,
a
recommended
marine
criterion
of
0.001
ug/
L
results.

(
QUALITY
CRITERIA
FOR
WATER,
JULY
1976)
PB­
263943
SEE
APPENDIX
C
FOR
METHODOMGY
NICKEL
CRITERIA:

Aquatic
Life
For
t
o
t
a
l
r
e
c
o
v
e
r
a
b
l
e
n
i
c
k
e
l
t
h
e
c
r
i
t
e
r
i
o
n
(
i
n
ug/
L)
t
o
0
p
r
o
t
e
c
t
freshwater
a
q
u
a
t
i
c
l
i
f
e
a
s
derived
using
t
h
e
Guidelines
i
s
t
h
e
numerical
v
a
l
u
e
g
i
v
e
n
by
e(
0.76[
ln(
hardness)]+
l.
06)
a
s
a
24­
hour
a
v
e
r
a
g
e
,
and
t
h
e
c
o
n
c
e
n
t
r
a
t
i
o
n
(
i
n
ug/
L)
s
h
o
u
l
d
n
o
t
exceed
t
h
e
numerical
v
a
l
u
e
given
by
.(
0.76[
I
n
(
hardness)
]+
4.02)

a
t
any
t
i
m
e
.
For
example,
a
t
hardnesses
of
50,
1
0
0
,
and
200
mg/
L
a
s
CaC03
t
h
e
c
r
i
t
e
r
i
a
a
r
e
56,
96,
a
n
d
160
ug/
L,
r
e
s
p
e
c
t
i
v
e
l
y
,
a
s
24­
hour
averages,
and
t
h
e
concentrations
should
n
o
t
exceed
1
,
1
0
0
,

1
,
8
0
0
,
and
3,100
ug/
L,
r
e
s
p
e
c
t
i
v
e
l
y
,
a
t
any
t
i
m
e
.

For
t
o
t
a
l
r
e
c
o
v
e
r
a
b
l
e
n
i
c
k
e
l
t
h
e
c
r
i
t
e
r
i
o
n
t
o
p
r
o
t
e
c
t
s
a
l
t
w
a
t
e
r
a
q
u
a
t
i
c
l
i
f
e
a
s
d
e
r
i
v
e
d
u
s
i
n
g
t
h
e
G
u
i
d
e
l
i
n
e
s
i
s
7.1
ug/
L
a
s
a
24­
hour
a
v
e
r
a
g
e
,
a
n
d
'
t
h
e
c
o
n
c
e
n
t
r
a
t
i
o
n
s
h
o
u
l
d
n
o
t
0
exceed
1
4
0
ug/
L
a
t
any
t
i
m
e
.

Human
Health
For
t
h
e
p
r
o
t
e
c
t
i
o
n
of
human
h
e
a
l
t
h
from
t
h
e
t
o
x
i
c
p
r
o
p
e
r
t
i
e
s
of
n
i
c
k
e
l
i
n
g
e
s
t
e
d
t
h
r
o
u
g
h
w
a
t
e
r
and
contaminated
a
q
u
a
t
i
c
organisms,
t
h
e
ambient
w
a
t
e
r
c
r
i
t
e
r
i
o
n
i
s
determined
t
o
be
1
3
.
4
W
/
L
.

For
t
h
e
p
r
o
t
e
c
t
i
o
n
of
human
h
e
a
l
t
h
from
t
h
e
t
o
x
i
c
p
r
o
p
e
r
t
i
e
s
of
n
i
c
k
e
l
i
n
g
e
s
t
e
d
t
h
r
o
u
g
h
contaminated
a
q
u
a
t
i
c
organisms
a
l
o
n
e
,
t
h
e
ambient
water
c
r
i
t
e
r
i
o
n
i
s
determined
t
o
be
1
0
0
.

Ug/
L*

(
45
F.
R.
79318,
November
2
8
,
1980)
SEE
APPENDIX
B
FOR
METHODOLOGY
2,3,7,8­
TETRACHLORODIBENZO­
P­
DIOXI l
CRITERIA:

Aquatic
Life
Not
enough
data
are
available
concerning
the
effects
of
a
2,3,7,8­
TCDD
on
aquatic
life
and
its
uses
to
allow
derivation
of
national
criteria.
The
available
information
indicates
that
acute
values
for
some
freshwater
animal
species
are
>
1.0
ug/
L;

some
chronic
values
are
<
0.01
ug/
L;
and
the
chronic
value
for
rainbow
trout
is
<
0.001
ug/
L.
Because
exposures
of
some
species
of
fishes
to
0.01
ug/
L
for
<
6
days
resulted
in
substantial
mortality
several
weeks
later,
derivation
of
aquatic
life
criteria
for
2,3,7,8­
TCDD
may
require
special
consideration.
Predicted
bioconcentration
factors
(
BCFs)
for
2,3,7,8­
TCDD
range
from
3,000
to
900,000,
but
the
available
measured
BCFs
range
from
390
to
13,000.
If
the
BCF
is
5,000,

concentrations
>
0.00001
ug/
L
should
result
in
concentrations
in
edible
freshwater
and
saltwater
fish
and
shellfish
that
exceed
levels
identified
in
a
U
S
.
FDA
health
advisory.
If
the
BCF
is
>
5,000
or
if
uptake
in
a
field
situation
is
greater
than
that
in
laboratory
tests,
the
value
of
0.00001
ug/
L
will
be
too
high.
0
Human
Health
 or
the
mxirnux
protection
of
human
health
from
the
potential
carcinogenic
effects
of
2,3,7,8­
TCDD
exposure
through
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms,
the
ambient
water
concentration
should
be
zero.
This
criterion
is
based
on
t
h
e
nonthreshold
assumption
f
o
r
2,3,7,8­
TCDD.
However,

z
e
r
o
m
a
y
n
o
t
b
e
a
n
a
t
t
a
i
n
a
b
l
e
l
e
v
e
l
a
t
t
h
i
s
time.
T
h
e
r
e
f
o
r
e
,
t
h
e
l
e
v
e
l
s
t
h
a
t
may
r
e
s
u
l
t
i
n
an
i
n
c
r
e
a
s
e
of
cancer
r
i
s
k
over
t
h
e
l
i
f
e
t
i
m
e
a
r
e
e
s
t
i
m
a
t
e
d
a
t
a
n
d
The
c
o
r
r
e
s
p
o
n
d
i
n
g
recommended
c
r
i
t
e
r
i
a
a
r
e
1
.
3
~
1
0
­
~
,
1
.
3
~
1
0
­
~
and
1.3x10­'
ug/
L,
r
e
s
p
e
c
t
i
v
e
l
y
.
If
t
h
e
above
estimates
are
made
f
o
r
consumption
of
a
q
u
a
t
i
c
organisms
o
n
l
y
,
excluding
consumption
of
w
a
t
e
r
,
t
h
e
l
e
v
e
l
s
a
r
e
1
.
4
x
3
0
­
~
,
1.4x10­*
and
1.4x10­'
ug/
L,

r
e
s
p
e
c
t
i
v
e
l
y
.
I
f
t
h
e
s
e
e
s
t
i
m
a
t
e
s
a
r
e
made
f
o
r
comsumption
of
water
o
n
l
y
,
t
h
e
l
e
v
e
l
s
are
2
.
2
~
1
0
­
~
,
2.2xlO­'
and
2
.
2
~
1
0
­
~
ug/
L,

r
e
s
p
e
c
t
i
v
e
l
y
.

(
4
9
F.
R.
5831,
February
15,
1984)
SEE
APPENDIX
B
 OR
METHODOLOGY
ALKALINITY
BARIUM
REFERENCES
CITED
Browning,
E
1961.
Toxicity
of
industrial
metals.
Butterworth,
London.
b
t
r
.
M.,
et
d.
1970.
Effeeta
of
pollution
on
fmh
Me,
heavy
m
e
a
Annd
Literature
review.
J
w
.
Water
Poll.
Cont
Fed.
42W.
L.
ngr.
N
A
1961.
Handbook
of
cbemistry.
lorh
ed
Yd;
fiw­
Hill,
Book
Co..
New
Ywk.
little,
A.
D.
1971.
Iwrg.
oic
cbemical
pollution
of
fnah
water.
Water
q
d
t
y
&
la
lmk
Val.
2.
US.
Envimnmental
Proteetion
Agency,
la010
DPV.
pp.
2e26.
McKee,
J.
E.
ud
W.
W.
Wolf.
1963.
Water
quality
criteria
Calif&
Skte
Watm
M
u
m
a
Contrd
Board,
Pub.
No.
3­
A
National
Aademy
of
Seieooes.
National
Academy
of
Engioeering.
191+
Water
qwlity
criteria,
1972
U
S
Government
Printing
Off=,
W.
shin@
n,
D.
C.
Patty,
FAl962lndustrialbygieneudtoximlogy,
VoL1I.
JohnA
W
i
l
e
y
,
N
e
r
Y
a
t
p
p
%
loo2
Cited
in
US.
Department
of
Health,
Eduatian
ud
Welfare.
1M.
Public
Health
SeMce.
1WB.
Drinking
water
quality
of
rkasd
intast.
te
eUria
water
supplies.
U.
S.
Depar(
ment
of
Health
Eduatian
ud
Welfy
W..
hin%
on,
D.
C.
Sollmann.
T.
H.
1957.
A
manud
of
pharmamlogy
and
its
applition
to
ud
toljcolcgy.
8th
ed.
W.
B.
Saunden
Co.
Philadelphia
Stokinger,
H.
E.,
radRLWmdwaardl968.
Toljcdogiemetbodrfaat.
blirhin~
water
standards.
Jour.
Ame?.
Water
Worlrs
h.
50.615.
U.
S
Deprtment
of
Health,
Education
and
Welfue.
1968.
Preliminvy
air
pdlutioa
r
w
e
y
of
buiurn
and
its
compouinja,
a
literature
review.
N
a
W
Air
P
o
U
d
n
CODtml
Administntion
Publiation
No.
AIT
D
69­
23,
Raleigh,
N.
C.

BORON
REFERENCES
CITED
Bw.
J.
W..
and
M.
Fireman
1960.
Boron
*
tion
and
mlean~
by
m
i
k
Soil
sd.
SOe
Amer.
Ra2­
l:
115.
B
d
w
d
.
G.
R
1966.
h
n
[
toxicity,
indicator
plants],
in
di.
gwstic
criteria
for
plants
and
.
oik
H.
D.
Chapman.
ed
Univwsity
of
California,
Diviaion
of
Agridtural
Science,
Berkeley,
p.
53.
Le
Clem
E
1960.
The
self­
purifkation
of
strams
ud
the
relatia~
hip
between
chemical
md
biologid
test&
Ra
Seeond
Symposium
on
Treatment
of
Want?
Watem.
Pergamon
h.
London,
p.
231.
Le
Clem.
E,
and
F.
D
e
v
h
i
n
d
L
1456.
Fiah
toxicity
tats
and
water
q
d
t
y
.
Bull.
de
Beige
Condument
E.
u.
28:
11.
Kopp
J.
F..
ud
RC.
Kmner.
1961.
Trace
met&
in
waten
of
the
UnitEd
Slate..
F
e
d
4
Water
Pollution
Conbol
Administration,
US.
Department
of
Interior,
Cincinnati.
Ohio.
YcKee,
J.
E,
ud
A.
W.
Wolf.
1963.
Water
quality
criteria
State
Water
Qrulity
Conmi
Bcwd.
haamento,
Calif.
Pub.
3­
A.
National
Ardemy
of
Seieocpa,
National
Academy
of
Engineering.
1974
Water
p
d
i
t
y
mi­
1972
US.
Government
Printing
Offii.
W.
shingon,
D.
C.
R
i
W
.
LA.
ed.
1954.
Diagnosis
and
impmvement
of
saline
md
I
w
i
mi.
Agrieult­
Hmdbmk
No.
 U.
US.
Government
F'rinting
Offioe,
Wpshington,
D.
C.
CHMROPHENOXY
HERBICIDES
REFERENCES
CITED
Knua.
u
dted
by
Mitcheaa.
J.
W.,
RE
Hogmn,
and
C.
R.
Gaetjem.
1946.
Tol­
of
fum
.
ni&
to
feed
mnUining
2,4diehlomphenoxyacetie
d
b
Jour.
A
n
i
d
.
sd.
5:
226.
Lthrmn.
AJ.
1986.
Sum­
of
p
t
i
c
i
d
e
toxidty.
Assodrtbn
of
Food
md
Dnrg
Oificialn
of
the
US..
T
o
p
h
KAM.
pp.
1­
40,
Palmer.
J.
S..
and
RD.
Radeleff.
1964.
The
toximlcgk
elf&
of
certain
fungiddea
md
herbicida
on
sheep
and
cattle.
Ann.
N.
Y.
A
d
.
Sei.
111:
729.
Wury,
J.
H.
1963
Toxicity
of
24dichloruphenoxyacetic
acid
for
man
and
dog.
Arch.
Envir
Health.
7:
282
COLOR
REFERENCES
CITED
hceriom
Water
Works
Assodrtion.
1971.
Water
qudity
and
treatment
Brd
ed
Mdmw­

Birge.
E
X
.
md
C.
Juday.
1990.
A
amnd
report
on
.
oh
radiation
and
inland
Up..

B
U
A.
P.,
and
RF.
chrisrmaa
l96sachvrrtMstw
.
of
colored
surface
waters
Jau.
Hill
Book
CO.,
New
York
Tnns.
W
k
.
A
d
.
Science,
AM.
Let.
a:
285.
'

hmer.
Water
works
kdn.
55:
7s.
El&.
AP..
and
RF.
chrisrmaa
1953b.
Cbemi~
I
rby.
ctaufKI
'
of
fdvk
dda
J
a
u
.
'

~~

Akner.
w.
7eTworksAasa,
55:
w.
N
a
t
i
o
~
l
A
d
e
m
y
of
Mencm
National
Academy
of
Engirraing.
1974
W
s
t
a
q
d
t
y
N
~
t
i
o
~
l
Technical
Adviawy
Committee
to
tbe
Secreq
of
tbe
IntpiOr.
1988.
Wata
Public
iiealth
service.
1952
Drinkira
water
st.
nduda,
1962
PAS
F'uM
No.
SS6.
US.
criterir
1972
US.
Government
Rinting
Office.
Wnhingmn,
D.
C.

quality
criteria
US.
Government
Printing
Office,
Wuhingron,
D.
C.

Government
Printing
o
i
f
i
i
,
Washir;
gOn.
D.
C.
Sawyer.
C.
N.
1960.
Chemistry
for
..
nitary
engineera.
Mdmw­
Hill
Bodr
Co.,
New
YOrL.
Shn~
w.
J.
lSi
Effect
of
yellow
o
d
e
d
d
s
on
Lon
md
other
meub
in
water.
Jour.
A'­:
Water
works
h
t
i
56:
1062'
Stllldvd
methods
for
the
examination
of
water
and
wastewater.
l8tb
ed.
1971.
Edited
by
Michael
C.
Rand.
et
d.
M
o
a
n
Public
Health
Aran.
American
Wster
Works
kdn.
Water
Pollution
Conml
Federation.
Wrahington.
D.
C.
Welch,
P.
S.
1952
Imnology.
M&
nw­
Hill
Book
CO.,
New
York.

DEMETON
REFERENCES
CITED
Butler,
PA.
1964.
COmwrd.
l
fiihery
isvestigatbm.
P.
ges
61B
in
Pedicide
vildliie
atudiea.
1969.
U.
S.
b
h
md
Wildlife
W
e
e
Cinuh
149.
WMhingtOn,
D.
C.

fuh.
Tnnh
Amer.
Fish
Sa
67:
39.
Ludenunn,
D.,
rad
H.
Neurmnn
ube
die
wirh~
ng
der
d
t
l
i
c
h
e
n
konwdnsedd.
tidde
auf
die
tiere
de
auasr.
mers.'
higu
fur
schdlingskunde
und
PILuneMut~.
s:
5.
M
e
k
.
 3..
md
W.
A.
MeAllkter.
1970.
lnseetidde
ausoeptibility
of
*
we
mmmon
fuh
funily
repesentativea.
Tnna
Ama.
h
h
.
Sa
992l.

toxic
levels
of
patieides
Tnna
Amer.
Fiah
k
101:
817.

­
of
oumw~
terfiishesTnns.
Amer.
Fiah.~
81:
175.
Henderson.
C..
md
Q.
H.
Mering.
EX8.
Toxicity
of
organic
pimphaw
i­

to
YcCmn,
JA.,
md
RL
Jq.
WR
Vcrrebnl
d.
nuge
to
Mueeilb
expxed
to
rutely
Pickering,
Q.
H.,
et
d.
196L
Tbe
toxieity
of
ap.
nie
phoaphow
inaect*
idpa
to
different
Sanders,
H.
O.
1912
Toxicity
of
come
iMeetiddes
to
four
SpeCKa
of
malamstracan
w
t
x
m
n
s
.
Burnu
of
Sport
Fisheries
and
Wildlife,
Tech
P
l
p
u
No.
66
US.
Department
of
tbe
I
n
k
r
i
a
,
Washington.
D.
C.
pp.
S­
19.
We&+
C.
M.
1968
Thl
Qtamirution
of
cholinedteme
in
the
h
a
i
n
timueof
thR0
lpgies
of
fnshwstu
Tub
md
ib
iructividon
in
vim.
Emlogy,
89:
194
We&.
C.
M.
1968.
R
e
a
p
o
~
of
fish
to
sublethal
expos­
of
­*
phmphom
iweeticidea.
Sew.
and
Ind.
Wastes.
31:
580.
W
e
b
.
C.
M.
1961.
PhVsologieal
effect
of
organic
phosphorus
inseetieides
on
several
speciesol
fish.
Trans.
Amer.
M.
Soc.
90:
143
W
e
b
.
C.
M..
and
J.
H.
Gakstatter.
1964s.
Detection
of
­
ticides
in
water
by
biwhemical
aaaay.
Jour.
Water
Poll.
Cant.
Fed.
36:
240.
Weisa.
C.
M..
and
J.
H.
Gakatatter.
1
W
b
.
The
decay
of
anticholineste­
activity
of
organic
phmphorua
inseeticidea
on
storage
in
waters
of
different
pH.
Advances
in
Water
Poll.
Research,
I:&%.

GASSES,
TOTAL
DISSOLVED
REFERENCES
CITED
b
u
c
k
.
G.
R,
et
al.
1975.
Monality.
saltwater
adaptation
and
reproduction
of
fMh
expeed
to
gas
supmaturated
water.
US.
Environmental
Roteaion
Agency,
Western
Fish
Toxicology
Station.
Cawallis.
OR.
Unpublished
reprt.
Clay,
A..
et
al.
1975.
Erperimental
induction
of
gas
bubble
W
in
menhaden.
PRsented
at
the
American
Fisheries
Society,
September
1975,
Ira
Vegu,
Nev.
New
England
Aquarium,
Boston.
Dewley.
EM.,
and
WJ.
&
I.
1975.
Effect,
of
various
mnoenhLions
of
diaaolved
atmmphenc
g
e
on
juvenile
chinook
salmon.
Onmrhynchw
tdmqfda,
and
ateelhesd
trout.
Fish.
Bull.
Dawley,
E,
et
al.
1975.
BioaaMp
of
total
dissolved
gas
p
m
u
~
National
Mviw
Fisheries
Sen&,
Seattle,
Waab
Unpublhhed
~
prt
DeMont,
J.
D.,
and
RB'.
Miller.
1972
first
reported
incidewe
of
p
bubble
d
k
w
in
the
heated
effluent
of
a
steam
electric
generating
station.
Roe.
Bth
l
a
n
d
meeting,
Southerat
Assoe.
of
Game
and
Fiah
Commi\
aown
Ebel,
WJ.,
et
al.
1975.
Effect
of
atmospheric
gas
superssturation
carved
by
dams
on
salmon
and
steelhead
trout
of
the
Snake
and
Columbia
Riven.
Final
Report.
N
h
w
e
s
t
Fisheries
Center.
National
Mpriw
Fisheries
Service.
Seattle.
Wash.
Keup.
L.
E
1975.
How
toreadafishkill.
WaterandSew%
ge
Work8,121:
18.
Lindmth,
A
1957.
Abiogenic
gas
iupnstmtion
of
river
water.
Arch.
Iiydrobiology,
59:
589.
Mrlouf,
R
1972
Oenurenoe
of
g.
s
bubble
disease
in
W
species
of
bivalve
mollusks
Jour.
Fish.
Res.
Bd.
Can.
B:
SBB.
Mlsoello,
RA,
et
al.
1975.
Evaluation
of
alternative
m
l
u
t
i
o
~
to
gas
bubble
disease
monality
of
menhaden
at
Pilgrim
Nudear
Power
Station.
Yankee
Atomic
Electric
Ca..
Weatborn.
M­
YAEX­
IoB7.
~~~
~~
,~~~
~
~~~

Kaoonal
Academy
of
Saenapa.
National
Academy
of
hgimennp.
1974
Water
quality
cntena.
1972
V.
S.
Government
h
u
n
g
Office.
Washington.
D.
C.
Nebeker.
AV.,
et
al.
1975.
Effects
of
gas
supmaturate­
water
on
freshwater
inverteb
ntes.
Roe.
Gas
Bubble
M
Worlrshop.
Battelle
Northwest,
Energy
Reeeych
l
a
d
Development
Agency
Spgial
Report
Nebeker.
AV.,
et
al.
1976~.
Nitrogen,
oxygen,
and
carton
dioxide
as
factom
d
f
d
n
g
fMh
B
h
V
d
h
8y
Npenatmted
water.
h.
h
e
r
.
Fish.
SOe
Nebeker.
AV.,
et
al.
197%.
SvVival
of
mho
and
scckeye
salmon
smolt,
in
seawater
d
t
e
r
exposure
to
gas
8upenaUMt.
d
water.
TraM.
Amer.
Fish
Soe.
Renfm,
W.
C.
1W.
Gas
bubble
morlality
of
flahes
in
Gilveston
Bay,
Tex.
h.
Amer.
Fish.
5%
92320.
Ruder.
RR
1974.
Gas
bubble
disease:
Monalities
of
mho
salmon,
Chmrh&
ud
&
A,
in
water
with
mustant
total
gas
presure
and
different
oxygen­
nitrogen
ntim.
National
Oceanic
and
Atmm.
Admix,
Natl.
bhr.
Fish.
SW.,
Northwest
Fish
Center,
Seattle.
WMh
Unpublished
m
u
s
c
r
i
p
t
Rulifmn,
RL
and
G:
Abel.
197L
Nivogen
supenaturation
in
the
Columbia
and
Snake
Riven.
T
e
d
Rep;
TS­
Cr%'/
O­
Z%
0162
U.
S.
EnvLonmental
Rotenion
Agency,
w
o
n
X,
Seattle,
Wash.
Van
Slyke.
D.
D.,
et
al.
I
W
.
Studies
of
gaa
md
electrolyte
equilibri.
in
Moo&
XVIII.
Solubility
and
phyxieal
state
of
a
t
m
p
h
e
r
i
c
ni­
n
in
blood
fell8
and
plrama
Jour.
Biol.
Chem.
1&:
571.
GUTHION
REFERENCES
CITED
Adelmn.
I.
R,
md
LL
Smith.
Unpublished
data
Deprtment
of
htomlogy,
Fisberiea
and
W'ildlife,
University
of
Minnesota,
St
Paul.
Benlre,
G.
M..
md
S.
D.
Murphy.
1974.
Anticbollnesterase
setion
of
w
t
h
y
l
purtbjon,
ppnrhion
and
ezinphoamethyl
in
mi@
and
fmh:
Onset
ud
remveryof
inhibition.
Bull.
Envimn.
Contam.
Toxiwl.
12:
117.
Butler,
P
A
1963.
CommerCial
fuheries
investiptiona.
I
n
Peaticirk­
wildlife
rtudim
during
1961
and
19EZ
US.
Fish
and
Wildlife
Service
Circ
167.
W.
abin%
on,
D.
C.
Caner,
F.
L
1971.
In
tIw
studies
of
brain
acetylcholines­
inhibition
by
wkwpbosp
bate
and
earbarnate
inseetieides
in
fish
PhD.
Dhenation,
LouLiuu
State
Univerrity.
Baton
Rouge
Coppage.
D.
L
1972
Orgmophosphate
perticidea:
Speific
level
of
bnin
ACHE
inhibition
relawd
to
denth
in
sheephead
minnows.
Trans.
Amer.
Fish
Soe.
101:
534
Coppcge,
D.
L.
and
T.
W.
Duke.
1971.
Effeets
of
penicidfa
in
eatuvica
dong
the
gulf
and
mutheart
Atlantic
mas­.
In
C.
H.
Schmidt,
ed.
F?
a+
nga
of
the
2nd
gulf
mt
mnference
on
m
q
u
i
t
a
suppmion
and
wildlife
management
National
Mosquito
Control­
Fish
and
Widlife
Management
Coordinating
Committee,
Wuhingtan.
D.
C.
Coppage,
D.
L
and
E
Mrtthews.
lPi4.
Sbon
term
effectsof
orgawpholph.
te
p
t
i
c
i
d
c
a
on
eholinfaterace
of
estuarine
fuhea
and
pink
ahrimp.
Bull.
Envimn.
Contam
Toxiool.
11:
483.
Coppage.
D.
L
e
t
al.
In
preaa.
Brain
seetylcholinesterase
inhibition
in
fmh
M
a
diagnmk
of
environmental
poisoning
by
malathion.
0
,
M
m
e
t
h
y
l
S(
l.
24imrbethoxyethyl)
phoaphorodithioate.
Pesticide
Biochemistry
and
Physiology.
Davis,
H.
C..
and
H.
Hidu.
1969.
Effects
of
pesticides
onembryonic
development
of
d
m
and
optera
and
on
survival
and
growth
of
the
larvae.
U.
S.
Fish
and
Wildlife
Service,
Fishery
Bulletin,
67:
993.
a
t
o
n
.
J.
G.
1970.
Chmnic
malathion
toxicity
to
the
blvegiU
(
kpomis
&
inrs
ILfinesque).
Water
Researrh,
4:
6iS.
Flinl,
D.
R.
et
d.
1970.
Soil
runoff,
leaching.
md
.
dsorption
.
nd
water
stability
studies
with
guthion.
C
h
e
w
Rept
No.
m.
Gaufn,
A.
R.
et
d.
1965.
The
toxicity
of
10
organic
insecticides
t
o
,
v
v
i
o
w
4
u
s
t
i
c
invertebratea.
Water
and
Sew.
Worlw.
112276.
Hollnnd.
H.
T.,
et
al.
1967.
Use
of
fmh
brain
acetylehalireterme
to
monitor
pollution
by
organopbmphoruc
pesticides.
Bull.
Envimn.
Contan
Toximl.
2:
156.
Jensen.
LD.,
and
AR
Caufn.
1966.
Acute
and
lpng
term
effeets
of
organic
inmaicides
on
two
Species
of
stowfly
nriadr.
Jour.
Water
PolL
Cont
Fed.
98:
1zTJ.
b
t
r
.
M.
1961.
Acute
tadcity
of
mme
organic
inseeriddes
to
thrse
@
of
vlrnonida
and
to
the
thresapine
sdddeback.
T
m
.
Amer.
Fish.
Soc
90:!
264.
W
v
.
Y.,
and
S.
Sai+
g.
1989.
Se~
itivity
of
pond
fuh
to
mtnion
(
uinphmethyl)
ud
panthion.
Bamidgeh.
21:
67.
Macek,
KJ..
et
a].
1969.
The
effects
of
temperature
on
thesuseoptibilityofthe
bluegills
nnd
rainbow
mut
to
selected
pesticide%
Bull.
h
v
i
m
n
.
Conram
Toxiad.
3174.
Macel,
KJ.,
and
W.
A.
MeAllister.
1970.
Insecticide
susoeptibility
of
some
mmmon
fmh
family
repmntativea.
TMa
h
e
r
.
Fish.
sOe.
99:
P.
Meyer.
F.
P.
1965.
The
experimental
u8e
of
guthion
as
a
selective
fvh
eradicator.
Trms
A
m
.
Fiah
Sor
44:'
ZX.
Nebeker,
A.
V..
and
A.
R
G
n
u
h
1964.
Bioaarays
to
determine
psticide
todcity
to
ilw
amphipal
m
w
n
.
Gummanu
hcwtris.
Roe.
Utah
A
d
.
sd.
M
a
and
Letters.
41:
64.
Piekering,
Q.
H.,
et
al.
1962
The
toxicity
of
organic
phaphonu
insecticidca
to
different
species
of
w
m
w
a
t
e
r
fmhea
Trana.
d.
Fish.
k
91:
1'
75.
Ponman.
J.
1912
Resulfs
of
acute
testa
with
marine
organisms.
usingitandvd
methods.
P.
ges
212­
217
in
Ruivo.
ed.
M
e
n
e
pollution
and
s
e
~
life.
Fishing
N
e
m
(
Books)
Ltd..
London.
PmL
G..
and
RA.
Lersure.
1974.
Sublethal
effect
of
malathion
to
three
d
m
n
i
d
spied.
Bull.
hvimn.
Conulm.
Toximl.
12:
912
sudem.
H.
O.
1969.
Toxicity
of
pesticides
to
the
matacean.
Gammanu
hcwtri~
BWU
of
Sport
Fisheries
and
Wildlife
Twb.
Paper
No.
25,
US.
Depanment
of
the
Interior,
WMbinRWn.
D.
C.
W
e
n
.
KO.
197!
2
Toxicity
of
wome
i&
dea
to
four
apeEiea
of
­­
c
n
~
­
US.
Deprtment
of
the
Interior.
Bureru
of
Sport
Fisberiea
and
Wildlife
Tecb
Parer
No.
66.
Wnhington,
D.
C.
W
e
n
,
E.
O.,
ud
O.
B.
Cape.
19BB.
Tne
relative
toxicities
of
wed
ptiddea
to
nrids
Tucker.
RK.
md
D.
G.
Crabtree.
1970.
Handbook
of
toxicity
of
peatiades
to
wildlife.
Bvreru
of
Spwr
FLberica
ud
Wildlife
Raavor
F'ubl.
No.
sl
U
S
Ikprrmantof
tbe
Interior,
W.
rhin%
on,
D.
C.
We%
C.
Y
1868
Tbe
determimtion
of
eholinertanse
in
the
b
n
i
n
tissue
of
throe
spedea
of
freshwater
fuh
ud
itr
rrivation
is
viw.
Emloay.
S:
1%
Weim,
C.
Y.
1868.
Rasponr
of
r
i
to
mbleth.
l
expama
of
0rp.
niC
i
d
&
Sew.
and
Ind
W.
stea,
31:
580.
W&.
C.
M.
1961.
Physiological
effects
of
w
p
.
n
i
c
pbaspkru
ineriddps
on
severr.!

W
e
b
,
CX,
lad
J.
H.
c.
Lsuttu.
1864
Detection
of
m
d
e
a
in
Gtu
by
bkkw
o
f
L
h
r
e
e
~
o
f
a
t
o
n
e
i
l
i
e
a
~
o
l
.
~
o
c
e
a
n
o
g
.
l
3
:
1
l
z
.
Fe&
of
T
i
TMa
Amu.
Fhh
k
s
o
:
143.

UUY.
Jw.
Water
Poll.
Cont
Fed.
86:
240.
a
HARDNESS
a
REFERENCES
CITED
Doudomff,
P.
and
M.
b
t
z
.
1953.
Critical
review
cf
li+
rstweon
the
toxicity
of
induarrial
wastes
and
their
mmponenh
to
fish.
11:
The
metabasa.
d@.
Sew.
andind.
Wasted,
25:
Bm
National
Academy
of
Science.
National
Academy
of
Engineering.
1974.
Water
quality
criteria.
1912.
U
S
Government
t
i
n
t
i
n
g
Office.
Washingon,
D.
C.
National
Technical
Ad­
r).
Committee
to
the
Secretary
of
the
Interior.
1963.
Water
quality
enterin.
US.
Government
t
i
n
t
i
n
g
Office.
Washington,
D.
C.
Sawyer.
C.
H.
1960.
Chemisq
for
sanitary
engineen.
McGraw­
Hill
Book
Co..
New
York.
Stiff,
MJ.
1971.
Copper/
biearbonatP
equilibria
in
m
l
u
t
i
o
~
of
bicarbonate
ion
at
mncentrations
similar
to
t
h
m
found
in
n
a
t
d
water.
Water
Research.
5:
171.

MALATHION
REFERENCES
CITED
Andemon.
B.
G.
1W.
The
toxicity
of
organic
insecticides
to
hphnia
Semad
Seminar
on
Biol.
Pmblems
in
Water
Pollution.
Robert
A.
Taft
Sanitaryhgineering
Center
Tgh.
Report
W603,
Cincinnati,
Ohio.
Bender,
M.
E
19%.
The
toxicity
of
the
hydrolysis
and
breakdown
p
d
u
a
a
of
maLthion
to
the
fathead
minnow
(
Anuphaler
pmmcloa
Rainesque).
Water
&
a
8:
571.
B
w
n
u
of
Sport
Fisheries
and
Wildlife.
1970.
Resourn
Publication
106.
W.
abiagton,
D.
C.
Butler,
PA.
1963.
Commercial
fisherie
investigations.
In
Pd&
wildlife
studies
during
1961
and
1962
U.
S.
Fish
and
Wildlife
senice
cirr
167.
Wdiagton,
D.
C.
Coppage,
D.
L,
and
T.
W.
Duke.
1
9
i
l
.
~
f
e
c
t
3
c
f
~
t
i
c
i
d
e
s
i
n
e
s
t
u
~
e
s
~
o
n
g
t
b
e
~
~
southeast
Atlantic
masts.
In
C.
H.
Schmidt,
ed.
Roceeding~
of
the
2nd
gulf
m
t
mnference
on
mosquito
suppression
and
wildlife
m
a
m
r
n
e
n
t
.
N
a
t
i
o
d
Mosquito
Control­
Fish
and
Wildlife
Manapmeat
Ccardimting
Committee,
W
d
n
g
t
o
n
.
D.
C.
Coppage.
D.
L.
and
E.
Matthema.
1974.
Short
term
effects
of
organophosphate
peatieides
on
c
h
o
l
i
n
e
s
t
e
m
of
estuarine
fiahes
and
pink
shrimp.
Bull.
Envirun
Coatam
Toxiad.
11:
w.
Coppa@?,
D.
L.
e
l
al.
1975.
Brain
acetylcholinesterase
inhibition
in
fuh
M
adiagnosis
of
envimnmental
poisoning
by
malathion,
0,
Odirnethyl
s(
l.
Zdiearbetboxysthy1)
phmphorcdithioste.
Peaticide
Bioehemistrj
and
Physiology.
Cope.
O.
B.
1965.
Sport
fuher).
investigations.
In
The
effect3
of
peaticides
on
fiah
and
wiidlife.
Fiah
and
Wildlife
Service
Cidar
266,
US.
Depanment
of
the
Interior,
Washinkon.
D.
C.
Da\%
H.
C..
and
H.
Hidu.
1989.
Effeets
of
pesticides
on
embryonic
development
of
dam­
and
optera
and
on
sUrnval
and
gmalh
of
the
larvae.
U.
S.
Fish
e,
nd
Wildlife
Service,
Fisbery
Bulletin.
6i:
399.
Eaton.
J.
G.
1970.
Chronic
malathion
toxicity
to
the
bluegill
(
kpomis
moFTochltlyl
Rafinesque).
Water
Res.
4:
6?
3.
Eiehelberger,
J.
W..
and
JJ.
Licbtenterg,
1971.
Persistence
of
peaticidea
in
river
water.
hvimn.
sd.
Technol.
5541.
Eider,
R
1988.
Acute
toxiatiea
of
i
d
a
d
e
s
to
marine
d
a
p
d
Crust.~
1­
aa
1
6
:
m
Eisler.
R
1970.
Acute
toxicities
of
organochlorine
and
organophosphorus
i~
eetiddea
to
atuvine
fiahes.
U.
S.
Burerru
of
Sport
Fisheries
and
Wildlife,
Tech.
Paper
46,
W
a
s
h
i
w
n
.
D.
C.
Henderson.
C..
and
Q.
H.
Pickering.
1958.
Toxicity
of
organic
phosphorus
insecticides
to
k
h
.
T
m
.
Amer.
Fish
Sor
S:
39.
M,
RE.
and
WJ.
Collins.
1974.
The
susceptibility
to
sekcted
insectidea
md
umtylcholiwsterase
activity
in
a
laboratory
mlony
of
midge
larvae,
Chimnmnus
trmonr
(
Dipters:
chimnomidae).
Bull.
Envimn.
Contarn.
Toximl.
E62
K
a
q
M.
1961.
Acute
toxicity
of
some
organic
insecticides
to
three
species
of
Sllmonids
and
to
tbe
threespine
stickleback.
Trana.
Amer.
Fish
Soe.
W.%.
M
a
d
.
KJ..
and
W.
A.
McAllister.
1970.
Inseeticides
suaeeptibility
of
~
omemmmon
fmh
family
repRsentatives.
Trans.
h
e
r
.
Fish.
Soe.
99:
ZJ.
Mount,
D.
T..
and
C.
E.
Stephan.
196i.
A
method
for
establishing
sxeptahle
toxiant
limits
for
fish­
dathion
and
the
butoxyethanol
ester
of
44­
D.
Tram
Arner.
Fish
Sa
21:
185.
Mulla.
Y.
S.,
and
A.
M.
Khssauinah.
1969.
Laboratory
and
field
evaluation8
of
I
d
d
e
a
.
gainst
ehimnomid
m
i
d
p
.
Jour.
Ecan.
Entomol.
6237.
Puis.
D.
F..
et
al.
1975.
Rates
of
depadation
of
malathion
by
M
a
klafed
fmm
aquatic
system.
Envimn.
Sci.
Technol.
9:
la.
Piekering.
Q.
H..
et
d.
1962
The
toxicity
of
orpanic
phmphom
insectickb
to
different
s
p
i
e
s
of
w
m
w
a
t
e
r
fishes.
Trans.
Amer.
Fish.
SOC.
91:
175.
pmt,
G.,
and
T.
R
Sehroder.
IKI.
The
toxicity
of
four
iruecticides
to
four
dmonid
species
Bull.
Envimn.
Contam.
Toximl.
6:
144.
Sanden,
H.
O.
1W.
Toxicity
of
Wticides
to
the
crustacean.
Cammnu
W
n
S
.
B
m
U
of
Sport
Fisher%
and
Wildlie
Tech.
Paper
No.
25.
U.
S.
Department
of
the
Interior.
WashingWn.
D.
C.
Sanden.
H.
O.
I972
Toxicity
of
some
insecticides
to
four
v
i
e
s
of
m
d
r
m
s
W
n
cmsu~
oedns.
Bureau
of
Sport
Fisherier
and
Wildlife
Tech.
Paper
No.
66,
US.
Deportment
of
the
Interior.
Washiwpn,
D.
C.
Sandera.
H.
O.,
and
O.
B.
Cope.
1966.
Toxicities
of
several
peatieides
to
two
Of
cladmeram
Tram.
Amer.
Fish.
Soc.
%:
I&.
Sanders,
H.
O..
and
O.
B.
Gap.
196%
The
relativetoxicitiesof
several
peatiadesto
lyiads
of
rhree
species
of
stoneflier.
Limnol.
Oeeanog.
13:
112
Weiss,
C.
M..
and
J.
H.
Gakntatter.
1964.
The
decay
of
anticholinesteRas
Vtinty
of
orpanic
nhaaohom
insecticida
on
storag~
in
waten
of
diiferent
pH.
A
d
v
m
in
Water
k'oll;
tion
Fk%
e.
rch,
1:
83
YIGANESE
REFERENCES
CITED
F.
irbndge.
RW..
ed
1966.
The
encyclopedia
of
axanogmpby.
Reinhold.
New
York.
Griffin.
A.
E
1960.
Si&
icanoe
ud
~
m
o
v
a
l
of
rmnganeee
in
water
aupdica.
J
w
.
Amer.
..
Water
Worka
Asan.
52:
1926.

Amer.
Water
WorLs
h
n
.
52:
W7.
Illig.
G.
L
Jr.
1960.
Use
of
d
u
r
n
heurnetaphosphate
in
man~.
neest.
biliutioa
J
w
,

McKee.
J.
E..
and
H.
W.
Wolf.
1963.
Water
quality
eriterir
State
Water
Quality
Control
Natioml
Academy
of
Sciences,
National
Aeademy
of
Engineering.
1974.
Water
g
d
i
t
y
Riddirk,
J.
M..
et
d.
1958.
Iron
and
manganese
in
water
suppliea
Jour.
Amer.
Water
Board.
Sacramento.
calif.
Pub.
3­
A.

criteria,
1972.
US.
Government
Rinting
Office.
Washinkon,
D.
C.

Works
Asan.
54:
688.
sollman.
T.
H.
1457.
A
manual
of
pharmacology.
W.
B.
Saunden,
Philadelphia
METHOXYCHMR
REFERENCES
CITED
Bdmer.
LE..
and
D.
R
Nimmo.
1974.
Methods
to
ae.
s&
as
effeets
of
combinations
of
t
o
h
t
r
,
salinity,
ud
temperature
on
wtuuine
minula
Roe.
9th
A
n
n
d
Conference
on
Trace
Submama
in
hvironmenul
Health,
University
of
Yiamuri.
Columbia
Butler.
P
A
19n.
Influence
of
petiadea
on
mMne
emsyatems.
Roe.
Royal
Soi
b
h
n
,

Eisele.
PJ.
1974.
The
effeets
of
mqbxyehlor
on
q
u
a
t
i
c
invertohate
papuktiona
md
communities.
PkD.
%
is,
University
of
Michigan,
Ann
hrbor.
Heath.
RG.,
et
d.
1972
Comparative
diewry
toxicities
of
peStieidea
to
birds.
Burerruof
S
p
n
FLberiea
4
Wildlife.
Widlife
Report
No.
1%
US.
Depvrmeat
of
tbe
Intcrior,
Washinrton
D.
C.
in:
sn.

Hodge,
Hk.,
et
d.
1954.
Short
term
oral
toxiaty
testa
of
methoxyehlor
Q
ntrd
+.
Jour.
Pbumaml.
Exo.
Theraav
99.
la
,
­,.­­
Kern.
S..
and
R
Earnest
1974.
Acute
toxicity
of
a,
innecrjddm
ta
d
i
p
a
d
hr
M­
amtilu.
Womb
Fish
ud
Game,
60:
B
Lehnun.
AJ.
Summvlesof
peaticidetodaty.&
socofFoodudhvgOffdrbof
the
US..
Topeka
h.
pp.
1
4
.
Y
m
,
J.
W..
ud
P
J
.
b
l
e
.
lgls.
The
effem
of
methoxyeblor
on
4
u
c
i
e
biota
U.
S.
hvimnmenul
Rotenion
Agency,
Ecologid
Re.
Senea,
So.
EPA­
WXun&
U.
S.
Covemment
Printing
Office.
Wpahiunpn,
D.
C.
Lehnun.
A
J
.
1W.
Summvles
of
pesticidetodaty.&
socofFoodudDrupOffdrbof
.­­­
­...
r­_._
r
r
~

­
J
lgls
effem
4
u
t
l
e
us
hvvonmenul
EC0logx.
i
So
U
S
hung
Office.
W
p
a
h
n
p
n
.
D
C
McM,
RL.
et
d.
1911.
Biodegradable
analog3
of
DM.
Bull.
World
Health
Omanh­

smden.
H.
O.
1972
Toxicity
oS
sue
iwetieidea
to
four
sped=
of
daooanCan
BURIU
of
spwl
Fiahenes
and
Wildlife
Tech.
Paper
No.
66,
US.

Stkkd,
LF.
1919.
Pesticide
residues
in
bin$
and
mammals.
Page4
254­
312
in
C.
A.
tiW.
44m.

Governaclrt
Rinting
Off=.
Washinkon.
D.
C.

Ed&
ed.
Environmental
pollution
by
peaticides.
Plenum
Reas.
New
York.

HIREX
REFERENCES
CITED
Borthwick,
P.
W.,
et
al.
1973.
Accumulation
and
movement
of
mirex
inselwted
esturka
of
South
Carolina,
196s­
71.
Pmtiaded
Monit
Jour.
1:
6.
Bookhout.
C.
G..
et
d.
1972
Effm
of
mirex
on
the
l
w
a
l
development
of
two
enbs.
Water,
Air.
and
Sail
Pollution,
1:
Ia.
Heath,
RG.,
and
J.
W.
Spann.
1973.
Pesticidfs
in
the
environment:
A
mntinuing
mntmversy.
lntemont
Med.
Bmk
C~
Q.,
New
Yark,
pp.
01­
4s
Heath.
RG..
et
al.
1972
Comparative
dietary
toxicities
of
pesticides
to
bi&.
B
m
u
of
Spn
Fisherie
and
Wildlife,
WildliIe
Repan
KO.
152,
US.
Department
of
Inrerior.
Wahington.
D.
C.
Hydf.
K.
M.,
et
al.
1974.
The
effect
of
mirex
onchannel
caUishpduction.
Trana.
Amer.
Fish.
Soe.
1(
n:
366.
Kaiser,
K.
L.
E.
1974.
Mirex:
An
unremp&
conraminant
of
fihes
from
M
e
Ontario.
Seience
185:
523
Lowe.
J.
I..
et
81.
1971.
Effwts
of
mirex
on
selected
estuarine
organisms.
P
a
p
171­
186
in
Ludke.
J.
L.,
et
a].
1971.
Toxicity
of
mirex
to
crayfish.
I
f
o
m
d
m
blandinqi
Bull.

Naq\
i.
S.
M.,
and
A.
de
Ih
C
w
.
1973
Mirex
inmrporation
in
theenvlonment:
Residue
in
Stickel.
WH.,
et
al.
1973.
Pesticides
and
the
envimnment:
A
mntinuing
mntmvmy.

Tapau.
M.
E.,
et
al.
(
19%).
Seasonal
effecfs
of
leached
&
x
on
selected
esuurine
Van
Valin,
C..
et
al.
1
s
.
Some
effects
of
mirex
on
two
w
m
water
fihes.
Tram.
Amer.

Wolfe.
J.
L.,
and
B.
R.
Nonnent
I973
Ammulation
of
mirex
midues
in
selected
organkms
after
an
aerial
treatment.
Mississippi,
1971­
72
Pesticides
Monit.
Jour.
7:
lrz
Transactions
of
the
36th
h'onh
American
Wildlife
W
u
m
Conference.

Environ.
Conram.
Toxicol.
6:
89.

non­
tar@
organisms,
1972
Pesticides
MoNt
Joui.
7:
104.

lntermnt.
Med.
Book
Corp..
New
York,
pp.
43747.

animals.
hrch.
Envimn.
Contam.
Toxicol.
3:
371.

Fish.
Sa.
9i:
1%.

NITRATES,
NITRITES
REFERENCES
CITED
Gillette,
U,
et
al.
1952
Apprakl
of
a
chemical
w
e
problem
by
fiah
toxicity
tests.

Humemn.
RH..
et
d.
1971.
The
nitrate
mtuation
in
Illinoia
Jour.
Amer.
Water
Works
Sewage
Ind
Wastes.
24:
I
S
.

b
n
.
b3:
m.
Klinger,
K
1957.
Sodium
nitrite.
a
slow
acting
f
i
h
poison.
Schweu
2
Bydml.
19(
2):
565.
Kwpp.
G.
L,
and
G.
F.
A
r
b
.
1973.
Ammonia
toxicity
levels
.
ad
nitrate
tolennce
of
channeles~~
hTheProgRsJiveFLahCulturist,
35:
221.
U&
y,
EF.
1972
Role
of
bacteria
in
the
nitrogen
cycle
in
Inkes.
Water
Pollution
Conml
Reaeuch
Series,
(
EP
210:
16010
EHR
W
E
)
,
US.
hvimnmenul
P
m
W
o
n
Agency,
US.
Government
Rinting
Office,
Washington.
D.
C.
National
A
d
e
m
y
of
Sciences.
1972
Aocumulation
of
nitrate.
National
Academy
of
Sdenw,
Wlshington.
D.
C.
Narionsl
A
d
e
m
y
of
Sciences.
National
Academy
of
Engiarering.
W97C
Water
quality
criteria,
1972
U
S
Government
Rinting
Offioe,
Wrshingmn,
D.
C.
Public
H
d
t
h
Sernce.
1961.
Groundwater
mnumination:
prmedinga
of
1961
symposi­
um
Tph.
Rpt
W61­
5,
RA
Taft
Sanitary
Engineering
Center,
U.
S.
Public
Health
Senice,
Department
of
Health,
Education
and
Welfare,
Cincinnati.
Ohio,
Ruso,
RC..
et
d.
1974.
Acute
toxicity
of
nitrite
to
rainbow
h
u
t
(
Sdmo
pidmi).
Jour.
Fish.
Res.
Bd.
Can.
31:
1663.
Rueo,
RC.;
and
RV.
Thurston.
1975.
Acute
toxicity
of
nitrite
tocutthroat
tmut
(
Sdmo
elark').
Fkherka
Bicamy
Lsboretory
Tech.
Repn
No.
754.
Yontms
State
Uwvvai­
tr.
W.
A.
1965.
Studia
on
f
s
h
culture
in
filtered
doeed­
drculating
quaria
11.
On
the
a
r
p
culture
exprimenu
in
the
syayatems.
Bull.
Jap.
Scc.
sd.
hab.
31:
916.
httelmarher,
P.
G.
1962
Methemoglobinemia
fmm
nitrate.
in
drinking
ntu.
SehrReihe.
Ver.
W.
aser­.
Boden­
u.
Lufthyg.
Ko.
21.
*
her,
Stuttgan
Smith,
C.
E.
and
W.
G.
Williams.
1974.
Erpenmntal
nitrite
toxicity
in
rainbow
h
u
t
and
chinmk
d
m
o
n
.
Trans.
Amer.
Fish.
See.
laS:
389.
Stewut.
B.
A.
et
al.
1967.
Nitrate
and
other
pollutants
under
fields
and
feedlots.
Envir
Sn.
Tech.
I:%.
.~~~~
~
.~
Trarm.
F.
B.
19%.
The
.
cute
toxicity
of
m
m
mmmon
salts
of
mdium.
pomsium.
and
d
d
u
m
10
the
mmmn
bluegill
(
Lrpnla
marrochinu
Railwsque).
Ra
A
d
N
a
t
Sci..
Philsdelphie
106:
185.
Vipil,
J..
et
d.
1965.
Nitrates
in
municipal
water
supplies
muse
metbemoglobinemia
in
infanta.
Public
Health
Rept
80:
1119.
Wdlen.
I.
E.
e
t
d.
1957.
Toxicity
to
Gambuaio
affinisof
certain
purechemiakin
turbid
watem.
Seam
Ind.
Wmm.
29:
695
Wdton.
G.
161.
Survey
of
literature
relating
to
infant
methemogiobinemi.
due
to
WVestin.
D.
T.
1974.
Nitrate
and
nitrite
toxicity
to
salmonid
fuhes.
The
w
v
e
Fish­
nitraw­
cvntaminated
water.
Amer.
Jou.
Public
Health.
41:
M.

C"
lt"
rist.
%!@&
WoUf.
L
A
,
and
Wakrman.
1972
Nitrates,
nitrite.,
and
niuoaamine.
Science,
177:
15.

OIL
AND
GREASE
REFERENCES
CITED
Bellan.
e
t
.
I.
1972
The
sublethal
effeeta
of
a
detergent
on
the
reprodudion.
development,
and
settlement
in
tbe
polyeh.
etous
annelid
Gpitclb
mpilda.
Yyim
BWlo([
y.
14:
uE3.
Blumer,
M.
1970.
Oil
matahnation
and
the
living
reanvces
of
the
.
e
~
Fmd
ud
Ag~
iculture
Orgsnizdtjon
Tech
Cod.
Rome.
FIR:
MPrlO/
Rl,
llP.
Bugbee,
S.
L.
a
d
C.
Y.
Walter.
1819.
The
~
spow
of
maaminvertebratea
to
penline
pollution
in
a
mountain
stream.
Page
7%
in
Revention
unl
mntml
of
oil
spilla,
proeeedinga
of
symposium
March
l
b
1
7
.
W.
ahiMton,
D.
C.
.
.
­
Diaz­
Pilerrer.
1962
The
effects
of
an
oil
spill
on
the
ahore
of
Cuania,
Puerb
Rim
Hampaon.
G.
R,.
and
H.
L
Sanden.
1989.
Local
oil
spill.
Oceanus.
15:
8.
I
m
b
m
n
.
S.
M..
and
D.
B.
Boylan.
19TJ.
Mfect
of
seawater
mluble
fraction
of
kemeene
on
chernotaxis
in
a
marine
snail,
Naamrius
ddclua.
Nature,
241:
213.
Johns
Hookins
UniveniQ.
1956.
Final
report
to
the
water
quality
submmmittee
of
.
the
(
ahtract).
Ass.
Isl.
Mar.
Late,
4th
Mtg.
C
u
m
,
12­
13.

Am&&
Petmleum
I&
titute,
Project~
49.41.
MeKee
and
Wolf.
19'
3.
Water
quality
m'teria
Slate
Water
Resource
Control
Board.
Saeramento.
Calif.
Pub.
No.
%­
A.
Yimnov,
O.
C.
1961.
Mfeets
of
low
mnaentrations
of
ptroleum
and
its
pmducts
on
the
development
of
me
of
the
Black
Sea
flatfish.
Vop
Ikhtiol.
7:
557.
Yimnov.
0.
G.
1970.
The
e
f
f
m
of
oil
pollution
on
flora
and
fauna
of
the
Black
Sea
In
Rmedirq:
FA0
mnferenoe
on
marine
pollution
and
its
effects
on
living
resouroes
and
fish.
Rome.
December
1970,
692
Food
and
Agriculture
Organization
of
the
United
Natiom.
Moore.
S.
F..
et
d.
1973
A
pmliminwy
assessment
of
the
environmental
vulnerability
of
Machi=
Bay.
Maine
tooil
superlankem.
Report
No.
MITSG
734.
Nelsan­
Smith.
A.
1971.
Effects
of
oil.
on
marine
0lmt8
and
a
n
i
d
.
P
a
a
e
~
m
i
n
P.
Hepple.
ed.
Water
pollutlon
by
oil.
London.
.
.

Nelmn­
Smith,
A.
1973
Oil
pollution
and
marine
emlogy
Plenum
PreSB.
New
Yo&.
PARATHION
REFERENCES
CITED
Bill&,
R.
and
deKinkelin.
1970.
Sterilization
of
the
Mticles
of
guppies
by
me8116
of
non­
lethal
doeeaof
parathion.
Annales
D
Hydrobiologie,
1(
1):
91.,
Burke.
W.
D.,
and
D.
E
Ferguson.
1989.
Toxicities
of
four
imticides
to
&
stant
and
susoeptible
mosquitofuh
in
static
and
flowingmlutions.
M
q
u
i
t
o
News.
291):
96.
Carter,
F.
L
1971.
In
mw
studies
of
brain
aoetylcholine~
terme
inhibition
by
organophmp
bate
and
earbarnate
iaseeticides
in
fish.
Ph.
D.&
nation,
&
MUM
State
University.
Baton
Rouge.
Coppage,
D.
L
1972
Orghnophcephate
peaticid­:
spedfic
level
of
brain
AChE
inhibition
related
to
death
in
~
heepshead
minnows.
TRM.
Amer.
Fish.
h.'
l01:
594.
~
oppage,
D.
L.
and
1.
W.
Duke.
1971.
Effecta
of
pesticides
in
estuaries
along
the
gulf
and
mutheas
Atlantic
m
t
a
.
In
C.
H.
Sehmidt,
ed.
Roeeglinga
of
the
2nd
Gull
Caa
Conferem
on
Mcaquito
Suppression
and
Wildlife
Management
National
Mmquitn
Conuol
­
Pish
and
Wildlife
Management
Cmdinating
Committee.
R'sahingvn,
D.
C.
Coppape.
D.
L.
and
E.
Matthews.
1974.
Short
term
effead
of
organophaephate
pesticides
on
cholinestenues
of
estuarine
fuhes
and
pink
shrimp.
Bull.
Environ.
Contam
Toximl.
11:
w.
Coppage.
D.
L,
et
al.
1975.
Brain
scetylcholinestenue
inhibition
in
fuh
a8
a
di.
gnoaie
of
environmental
poisoning
by
malathion,
0.
O­
dimethyl
S(
1,
ldicarbethoxysthyI)
phob
phorodithioate.
Pesticide
biochemisuy
and
phpiology.
Dowden,
B.
F.
19%.
Effecta
of
five
insecticides
on
the
oxygen
mmumption
of
the
bluegill
sunfish.
LqmnW
mnnahinu.
Ph.
D.
Thais.
Louisiana
State
University.
Baton
m.
Eaton,
J
.
G.
1970.
Chronic
malathion
toxicity
to
the
bluegill,
(
Lcpmna
mannhirw
Winesque).
Water
Reseueh.
4:
673.
Eichelberger.
J.
W.,
and
J.
J.
Lichtenberg.
1971.
Penistenoe
of
pesticided
in
river
water.
Environ.
Sci.
and
Technol.
5541.
Enrimnmental
h
t
e
c
t
i
o
n
Agency,
Office
of
Pesticides
Program.
1975.
Initial
sientific
and
miniemnomic
review
of
parathion.
Repon.
No.
EF'A­
540/
1­
75ml.
US.
Environ­
mental
h
t
g
t
i
o
n
Agency.
Natiod
Technical
Information
service,
Springfield,
Va
Gaufin.
A.
R.
et
al.
1465.
Toxicity
of
10
organic
insectjades
to
various
aquatic
inuertebnrter.
Water
and
Sew.
Works,
112276.
Gibmn.
J
.
R.
et
al.
1969.
Sources
of
ermr
in
the
use
of
fish­
brain
aoetylcho1ioeste­
k
activity
as
a
monitor
for
pollution.
Bull.
Envimn.
Contam.
Toximl.
437.
Jeaaen,
L
D.,
and
A.
R
Gaufin.
1964.
Lanp
term
effects
of
organic
i
d
c
i
d
e
s
on
two
species
of
stonefly
naiads.
Trans.
Amer.
Fish.
Soe.
93:
357.
Korn.
S..
and
R.
Earneat
1974.
Acute
toxicity
of
20
insecticides
to
striped
bass.
M
m
&
ilk.
Califomin
Fish
and
Game,
60:
128.
Lahay,
M..
and
S.
Sarig.
1989.
Sensitivity
of
pond
fuh
to
mtnion
(
azinphwmethyl)
and
parathion.
Bamidgeh.
Bull.
Fbh.
Cult
Israel.
2
6
7
.
Leland.
H.
V.
!
I.
1968.
Bimhemid
facton
affecting
toxicity
of
parathion
and
selected
lrnalop
to
fuhes.
Univemity
of
Michigan,
Ann
Arbor.
Lowe.
J.
I.,
et
al.
1970.
Laboratory
bioausys.
In
Fhgrss
report
for
f
4
y
y
1969.
Pesticide
Field
Station,
Gulf
Breeze,
Fla
U.
S.
Fish
and
Wildlife
Service
Cue
F.
Ludke.
J.
L
1970.
Mechanism
of
resistance
to
parathion
in
mmquitofuh.
GambvM
afini8.
Ph.
D.
Thesis,
Mississippi
State
Univenity.
Univemity.
Mount,
D.
1..
and
H.
W.
Boyle.
1969.
Parathion
­
y9e
of
blmd
mnoentmtion
to
diagnose
mortality
of
fmh.
Environ.
Sci.
and
Technol.
3:
1163.
Mulla,
M.
S..
and
A
M.
Khasa&
nah.
1969.
Lsborntory
and
field
evaluation
of
larvicides
againat
Chimnomic
mid­.
Jour.
Emn.
Entomol.
62:
s
Murphy.
S.
D..
et
al.
1963
Comparative
anticholineaterase
action
of
organophosphorus
inseeticides
in
invertebra­.
Toximl.
Appl.
Phannaeol.
12:
P
Pcst.
G..
and
R
A
Leasure.
1974.
Sublethal
effect
of
malathion
to
thRe
dmonid
sped­.
Bull.
Envimn
Cordam.
Toximl.
12:
312
Sanden.
H.
0.
1972
Toxicity
of
some
insecticides
to
four
species
of
malamstracan
mwtaeeans.
Bureau
of
Sport
fisheriea
and
Wildlife,
Tech.
Paper
No.
66.
U.
S.
Depanment
of
the
Interior,
Wsahingvn.
D.
C.
W
e
n
.
H.
0..
and
0.
B.
Cop.
19%.
Toxicities
of
several
pesticides
to
two
species
of
CIadooenuls.
TRlu.
h
e
r
.
Fish.
Sa
95:
165.
Spacie,
A.
1975.
Acute
and
chronic
parathion
toxicity
to
fuh
and
invertebrates.
U.
S.
Environmental
Pmteetion
Agency
Emlogical
Reeeareh
Series.
W
e
b
,
C.
M.
195%.
The
determination
of
cholinestersse
in
the
brain
tissue
of
three
a+
es
of
freshwaterfhhandiuinaeti~.
atiDninvivo.
Eeology,
39:
194.
We&,
C.
M.
1959.
hponse
of
fish
!
A
sublethal
expmures
of
organic
phoaphom
insefticides.
Sew.
and
Ind.
Was@,
91:
580.
Web.
C.
M.
1961.
Physiological
effect
of
organic
phmphorus
ilueeticide
on
severd
species
of
fuh.
Trans.
Amer.
Fish.
Soe.
90:
143.
We*,
C.
M.,
and
J.
H.
Gakstatter.
1964.
The
decsy
of
anticholinestersse
activity
of
organic
phcephom
insecticides
on
storage
in
waten
of
different
pH.
Adan­
in
Water
Poll.
Res.
I:@.
0
.

REFERENCES
CITED
Bell,
H.
L
1971.
Effect
of
low
pH
on
the
survival
and
emergem
of
equatic
insetu.
Water
ILea.
5:
313
~~~~
~
~

Butterfield.
C.
T.
1
M
.
Bacterial
propetties
of
free
and
mmbined
available
chlorine.
Jw.
A
m
.
Water
Works
Assn.
40;
1%.
Gpuro.
LRA.
1970.
Oeeawgrnphy
for
practicing
engineers.
Barnes
md
Noble,
Inr,
New
York.
DeUino.
JJ..
and
G.
F.
Lee.
1971.
Variation
of
manganese.
diamlved
oxygen
and
relaled
chemical
parameters
in
the
bottom
waters
of
M
e
Yendota,
WM.
Water
b.
5
lXY/.
Ewpem
Inland
Fiaheriea
Advimry
Cornmisaion.
1969.
Water
q
d
i
t
y
aiteria
for
European
freshwater
fish­
extreme
pH
values
and
inland
fuheries.
,
Repred
by
EIFAC
Working
Pnrty
on
Water
Q
d
i
t
y
Criteria
for
Eumpean
h
b
w
b
k
r
h
h
.
Water
Reseuch.
3:
593.
Joned.
J.
RE.
1964.
Fish
and
river
pollution.
Butterworth,
London.
Lngelier.
W.
F.
1936.
The
rJlalytical
m
n
m
l
of
a
n
t
i
a
m
i
o
n
water
t
R
l
t
m
e
n
t
J&.
Water
Worka
Asan.
28:
1500.
Mnunc
D.
I.
1m.
~
hmnic
effect
of
low
PH
on
fa­
minnow
mm+
v.
l,
rod
and
~..
~
.~
~

reproduction.
Water
Rea.
7
:
W
uiterin.
1972
US.
Government
Printing
Office.
Washin@
n.
D.
C.

Enviro.
Engineering
Div.,
ASCE
100:
No
EEL
Ror
Papers
1W43.539.
National
Aodemy
of
Soenaes,
National
Academy
of
Engineering,
1974.
Water
q
d
i
t
y
Reid,
L.
C.,
and
D.
A.
Cprlson.
1`
374.
Chlorine
disinfection
of
low
temperature
watem.
Jour.

Sawyz,
C.
N.
1960.
Chemistry
for
saniuus.
engineers.
MeGraw­
Hill
Book
Co..
New
York.
Stumm.
W.,
and
JJ.
Morgan.
1970.
Aquatic
chemiaty.
John
WIley
and
Sons,
I=.
New
Zirino.
A.,
and
S.
Yamamoto.
1572
A
pHdependent
model
for
the
chemical
sppi.
6~
0
of
.
.

York.

mpper,
nnc.
cadmium
and
lead
in
seawater.
Limn.
and
oeploog.
11:
661.

PHOSPHORUS
REFERENCES
CITED
Fletcher.
G.
L.
md
RJ.
Hoyle.
1972
Acute
toxicity
of
yellow
phmphom
14
Allantic
ad
(
Wud
mdua)
and
Atlantic
salmon
(
Sdmo
sdar)
smolts.
Jour.
Fish.
Rea
Bd.
of
­
29:
l.
aS.
Hutchinmn,
G.
E
1957.
A
treatise
on
limnology.
John
Wiley
and
Sons,
New
York.
Idler.
D.
R
1963.
Coexistence
of
a
fuhery
and
a
major
industy
in
Plaoentia
Bay.
Chembtry
in
Camda,
21(
11):
16.
Imm
B.
G.
IW.
Toxicity
of
elementpry
phosphorus.
Jour.
Water
Poll.
Convol
Fed.
32:
1312
Jurgaard,
P.
Y.
1970.
Ihe
mle
played
by
the
Fiaheriea
Research
Board
of
Canada
in
the
"
d
haring
pbosphwus
pollution
crioia
in
Pkcentia
Bay.
Newfoundland.
Circular
No.
1.
FLheries
Reseuch
Bard,
Atlantic
Regional
Mfioe,
Hrli!
u.
Nova
Smtk
Yrkenthun,
K.
M.
197S.
Toaud
a
derner
quatic
environment.
US.
Environmental
Roteaion
Agency,
W.
shington.
D.
C.
P
e
,
D.
L
1972
Obervations
onmorulitiesof
tenthicwgraismsiltercuntaminationof
tbe
bottom
of
&
n!
g
Hybour,
Aaoentia
Bay,
Newfoundl.
od
with
ekmental
p
h
p
b
Rpea
181­
186
in
Eff­
.
of
elemental
phosphorus
on
mvine
life.
Fishries
r!
efsmb
Board
of
csnadq
c
i
h
r
2
Volknweider.
RA
1973.
Input
output
nudela
ScbwaebZ
Hydml.
SOLIDS
(
DISSOLVED)
AND
SALINITY
REFERENCES
CITED
Apiculture
Handbook
h'o.
60.19%.
Diagnosis
and
improvement
of
saline
and
alkali
soils.
L.
A.
Richards,
ed.
U.
S.
Government
RintingOffice,
Washingl0n.
D.
C.
Bruvold,
W.
H.,
et
al.
1969.
Consumer
­
merit
of
mineral
tante
in
domestic
water.
Jour.
Amer.
Water
W
o
r
k
s
h
n
.
61:
515.
Capurm.
LRA.
1970.
Oceanopaphy
for'practicing
engineem.
Barnes
and
Noble
Inc.,
h'ea
York.
Griffith,
W.
H.,
Jr.
1963.
Salt
as
a
possible
limiting
factor
to
the
Suisun
Manh
phearant
population.
Annual
repon
Delta
Rsh
and
Wildlife
Protection
Study,
Cooperative
Study
of
California.
hhman,
J.
A.
1964
Control
of
mmaion
in
water
systems.
Jour.
Amer.
Water
Works
h
n
.
56:
1009.
Lodihart,
EE,
e
t
al.
1955.
The
effect
of
water
impurities
on
the
flavor
of
brewed
mffee.
Fmd
Research,
20:
5%.
Moore,
EW.
1952
Physiological
effeets
of
the
mnsumption
of
saline
drinking
water.
National
Res.
Counal,
Div.
of
M
e
d
i
d
Seienw,
Bull.
San.
Engr.,
and
Environment.
Appendix
E.
National
Academy
of
Sciences,
National
Academy
of
Gngineering.
1974.
Water
quality
criteria,
1972.
US.
Government
Printing
Office,
Washington,
D.
C.
National
Research
Council.
1954.
Sodium
resvicted
diets.
Publication
925,
Food
and
Nutrition
Board,
Washington,
D.
C.
National
Technid
Advisory
Committee
to
the
Secretary
of
the
Interior.
1968.
Water
quality
criteria.
US.
Government
Rinting
Office,
W
a
s
h
i
a
n
,
D.
C.
Patterson,
W.
L.,
and
RF.
Banker.
1963.
Effecta
of
higbly
mineralized
water
on
housebold
plumbing
and
appliances.
Jour.
Am­.
Water
Worka
h
n
.
60:
1060.
Public
Health
Service.
1962
Drinking
water
atandarda,
1962
US.
Government
Printing
Office,
W
a
a
h
i
a
n
,
D.
C.
Ra­
n.
D.
S..
and
J.
E.
Moore.
lSi.
4.
The
d
i
n
e
lakes
of
Saskatchewan.
crnadian
Jour.
of
Res.
22:
141.
Ricter.
C.
O.,
and
A.
MaeLean.
1939.
Salt
tante
threshold
of
humans.
Am.
J.
Phyiol.
126:
l.
Rouwfell,
G.
A.,
and
W.
H.
Everhart
1953.
Fisheryseience,
its
metbods
and
applications.
'
.

John
Wile?.
and
Sons.
he.,
New
York.
Sawyer.
C.
N.
1960.
Chembtrj
for
3aniUry
epgineem.
Mffiraw­
Hill
Book
Co.,
New
York.
Standard
methods
for
the
examination
of
waier
and
wastewater,
18thed..
1971.
Edited
by
Michael
C.
Rand.
et
a].
American
Public
Health
h
n
.
,
American
Water
Works
Asan.,
Water
Pollution
Control
Federation,
Washington,
D.
C.

SOLIDS
(
SUSPENDED,
SETTLEABLE)
AND
TURBIDITY
REFERENCES
CITED
Edberg.
N.
and
Hofsten,
B.
V.
1973.
Oxygen
uptake
of
bottom
sediment
studied
in­
situ
and
in
the
Inboratory.
Water
Resemth,
I
:
1285.
European.
Inland
fisheries
Advisary
Commiasion.
1965.
Water
quality
criteria
for
European
freshwater
fiih,
report
on
finely
divided
solids
and
inland
fisheries.
InL
Jour.
Air
Water
Poll.
9:
151.
Gammon.
J.
R
1970.
The
effect
of
inorganic
sediment
on
stmam
biota
Water
Poll.
Cont.
Rea.
Serieb,
18060
DWC
W
O
.
U.
S.
Environmental
Proteetion
Agency.
US.
Govern­
ment
Rinting
Office.
Washington,
D.
C.
Mackenthun.
K.
M.
1973.
Toward
a
cleaner
aquatic
envimnment
US.
Government
Rinting
Office,
Washington.
D.
C.
National
Academy
of
Sciences,
National
Academy
of
Engineering.
1974.
Water
quality
criteria,
1972
US.
Government
Printing
Office,
Washington,
D.
C.
Standard
methala
for
the
eramination
of
water
and
wastewater,
1971.13th
ed.
Edited
by
Mictwl
C.
Rand,
et
al.
American
Public
Health
h
n
.
,
American
Water
Works
h
n
.
,
Water
Pollution
Control
Federation.
T
m
e
l
l
.
C.
M.,
and
A.
R
Gnufin.
1959.
Some
important
hiologieal
effecta
of
pollution
often
disregarded
in
StRam
survey.
Roceedings
of
the
0th
RYdue
industrial
waste
mnferrnce.
Reprinted
in
Biology
of
water
pollution,
1367.
US.
Depanmenr
of
Interior,
Washingl0n.
D.
C.
Tebo,
L.
B.,
Jr.
1955.
Effects
of
siltation,
resulting
from
improper
logging,
on
the
bottom
fauna
of
a
smll
trout
stream
in
the
southern
Appalachians.
The
RogRaaive
fish
Culturist.
11:
 4.
SULFIDE
­
HYDROGEN
SULFIDE
REFERENCES
CITED
Adelman.
1.
R.
and
L.
L.
Smith.
1970.
Effect
of
hy­
3
sulfide
on
nwthern
pike
egga
and
Rmn.
C.
W..
and
B
J
.
Follis.
1967.
Effects
of
hydmren
sulfide
onchanneleaffuh
Tram.
aae
fry.
Trans.
Amer.
Fish.
Soe
99:
501.
,
Amer
Fiih.
Scc
96
31
d
m
o
n
and
mut
Wash.
Ikpt.
Fish.
Res.
Bull.
No.
5.
Jom.
3
.
R
1364.
Fish
and
river
pollution.
Butteworth,
London.
National
W
e
m
y
of
Sciences.
National
Academy
of
Engineering.
1974.
Water
qUditY
criteria
1972
U.
S.
Government
Printing
Offioe.
WashingWn.
D.
C.
b
i
d
.
D.
M.,
and
L.
L
Smith,
1974a
Chronic
toxicity
of
hydrogen
sulide
to
Gammnur
pruddimnneus.
Trans.
Amer.
Fish.
Soe.
Im.
h
i
d
.
D.
M..
and
L.
L
Smith.
1974b.
Long
termeffectsof
hydmgensulfideonHc2ogen~
limbat0
(
Ephemeroptera).
Environmental
Ecology.
h
i
d
,
D.
M.,
and
LL.
Smith.
1974~.
Facton
influencing
acute
coxicity
estimates
of
hydrogen
sulfide
to
freshwater
invertebrates
Water
Research
8.
Smith,
LL
1971.
Influence
of
hydrogen
sulfide
on
fish
and
arthmpods.
R
o
j
d
18050
PCG,
U.
S.
Environmental
R
o
w
t
i
o
n
Agency.
WaahingWn,
D.
C.
Smith,
L.
L..
and
D.
M.
h
i
d
,
1972
Effects
of
hydrogen
sulfide
on
fish
eggs
and
fry.
Water
Reeesrch,
G:
711.
Smith,
LL,
Jr.,
and
D.
M.
&
id.
1974.
Effwt
of
hy­
n
sulfide
on
development
and
sunrival
of
eight
freshwater
fish
speck.
Pagw
415­
470
in
J.
H.
S.
Blaner,
ed.
Tbe
early
life
history
of
fish.
Sprinpr­
Verlag.
New
York.
Theede.
H.,
et
81.
1969.
Studies
on
the
resistance
of
marine
bottom
invertebrates
to
oxygen
deficiencies
and
hydrogen
sulfrde.
Mar.
Biol.
2:
325.
Van
Horn,
W.
M.
1958.
The
effect
of
pulp
and
paper
mill
wastes
on
aquatic
l
i
k
Roe.
Ontario
Indust.
Wmte
Cod.
5
60.

TAINTING
SUBSTANCES
REFERENCES
CITED
Boyie.
H.
W.
1967.
TaJte/
dor
mntamination
of
fish
from
theOhio
River.
Federal
Water
Pollution
Conml
Administration,
Cincinnati.
Ohio.
National
Academy
of
Sciences,
National
Academy
of
Engineering.
1974.
Water
quality
criteria,
1972.
U.
S.
Government
Printing
Office.
Washington,
D.
C.
Shumwsy,
D.
L..
and
J.
R.
Palensky.
1973.
Impairment
of
the
flavor
of
fish
by
water
pollutants.
U.
S.
Environmental
PmLection
Agency,
EPA­
R.
3­
7M10,
US.
Government
Rinting
Office,
Waahingwn,
D.
C.
Thomas.
N.
A.
1973
Aasesament
of
fish
flesh
t
i
n
t
i
n
g
substances.
In
J.
Cairns
and
K.
L.
Dicksan,
eds.
Biological
methods
for
the
assessment
of
water
quality.
Amer.
Soeiety
for
Testing
and
Materials.
Tech.
Publ.
528.
Philadelphia
TEMPERATURE
REFERENCES
CITED
Ballentine.
RK.,
and
F.
W.
Kitmll.
1%.
Obeervations
of
fed
mlifom
in
several
m
n
t
stream
pollution
studies.
RoeeedingJ
of
the
Symposium
on
Fecal
Coliform
Bacterh
in
Water
and
Watewater,
May
21­
2,196s.
Bureau
of
Sanituy
Eqmeering,
California
State
Department
of
Public
Health.
BlncL,
E.
C.
1953.
Upper
lethal
temperatm
of
Mme
British
Columbia
frsahwater
fiihe.
Jour.
Fish.
Res
Bd.
Can.
10
196.
BretL
J
.
R
W
I
.
Tempering
v
e
n
u
acclimation
in
the
planting
of
spgWedbvut
Trura.
Amer.
Fish.
Sx.
70:
391.
BretL
J.
R
1%.
Some
principles
in
the
thermal
requirements
of
rubes.
M
y
Rev.
Biol.
31:
75.
Brett
J.
R
1W.
Thermal
requirements
of
fwh­
threo
d
s
d
e
s
of
study,
194&
1970.
In
C.
M.
Tarzwell,
ed.
Biol*
cal
pmblem
in
water
pollution.
Public
H
d
t
h
Senriee.
U.
S.
Dept
of
Health.
Education
and
Welfare.
Burnson.
B.
19S.
Seamnal
temperarm
variations
in
relation
to
anter
mtment.
J
w
.
h
e
r
.
Water
Warlo
&
an.
30:
793.
Cairns.
J..
Jr.
1456.
Eifects
of
in&
temperatures
on
qutk
organkm
I
d
4
W
a
r
n
.
1:
50.
CalabRse.
A
1968.
Individual
and
mmbined
e
f
f
e
of
lulinity
d
temperature
on
e
d
r
y
o
e
and
lmae
of
the
mot
clam,
Mulink
Mia
(
Spy).
Biol.
Bull.
lS7,
S:
417.
Camp,
T.
R
1963.
Water
and
iw
impurities.
Reinhold
Publishing
%.,
New
Y&
Chin,
E.
1961.
A
trawl
study
of
an
estuarine
nwsery
wea
in
Gdveaton
Bay
ritb
panicular
referenoe
la
penaeid
rhrimp.
Ph.
D.
Dissenation,
Univvsityof
Wuhin%
on
Gmtloa,
J.
D..
Jr..
d
C.
G.
BmWlout
1971.
The
effect
of
eydie
t
e
m
p
l
r
t
u
r
a
on
h
d
development
in
the
mud
aab,
JZhitAmpowp~
d
haorririi.
In
D.
I.
Crisp,
ed.
F
d
European
marine
biolcgy
sympium
Cambridge
Univemity
Rsa
h
d
o
n
Coutant.
C.
C.
IW.
Thermal
pollution­
biological
effects:
A
review
of
the
Literature
of
1967.
Jour.
Water
Poll.
Cont
Fed.
40:
1M7.
Coutant,
C.
C.
1968.
Thermal
pollution­
biological
effest.
5:
A
review
of
the
literature
of
1968.
Jour.
Water
Poll.
Cont
Fed.
41:
1836.
h
u
n
t
,
C.
C.
1970,
Thermal
pllution­
biological
effects:
A
review
of
the
Literature
of
1989.
Jour.
Water
PoU.
Cont
Fed.
4:
1025.
Couunt,
C.
C.
1971.
Thermal
pollution­
biological
effects:
A
review
of
the
literature
of
1970.
Jour.
Water
PoU.
Cont
Fed.
43:
1292
Coumt,
C.
C.
1972
Biological
~
p
e
e
t
s
of
thermal
pollution,
11.
SdentifK
hsis
for
r
a
t
e
temperature
a
t
~
n
d
u
d
s
nt
power
phnnta.
CRC
Critical
Rev.
in
hviron
Cant
8:
L
Coutant.
C.
C.
1975.
Temperature
selection
hy
&
h­
a
f
m
r
in
pow­
plant
impact
smeaumenfll
In
Symposium
on
the
physical
d
biological
e
f
f
e
on
the
environment
of
m
l
i
n
g
srstema
d
thermal
discharp
at
nuclear
p
w
e
r
ations
Intemstjorul
Atomic
hergy
Agency.
Coumt,
C.
C..
m
d
C.
P.
coodyear.
1972
Water
pollution­
tkd
pdlution:
Areview
of
the
literature
of
1971
J
w
.
Water
Poll.
Cont
Fed.
44:
1250.
CoutlnL
C.
C.,
and
H.
A
Pfuderer.
1978.
Tbermal
effe­
litenture
r&
ew
iaw
Jour.
Water
Poll.
Cont
Fed.
45:
EX.
Couunt,
C.
C..
d
H
A
Pfuderer.
1974.
Tbernul
effe­
titmature
review
iaw
Jour.
Water
Poll.
Cont
Fed
16:
1476.
Crisp,
D
J
.
1957.
Wfea
of
low
temperature
on
the
k
d
m
g
of
myiDe
dnvk
Nature,
179:
1x78.
De
Sylvr,
D.
P.
1983.
l
'
h
r
e
t
k
l
mnsidemtions
of
tbe
e
f
f
d
of
heated
effluents
on
mvine
f'sbes
I
n
PA
KreJel
d
F.
C.
M
u
,
eds.
BioloBiaJ
.
spgu
of
thermal
pollution
V.
aderbilt
Uuivemity
Freat,
NMhville.
Tern
dtvluuiug,
V.
L
1971.
Tbe
effecb
of
food
deprivation
d
vlinity
ehnges
on
repdue­
tive
fm&
m
in
the
rstuuine
pbiid
fmh,
C;
Uiehuy#
m'mbilia.
Biol.
Bull
141:
4.58.
devluning,
V.
L
l977.
The
e
f
f
e
of
temperature
ud
pbotoperiod
on
reproductive
cycling
of
tbe
mtuuine
pbiid
f
M
h
,
Gilliehuya
nirrdilu.
fLhry
Bull.
70,4:
1151.
Dow.
RL
W73.
Fluctutioua
in
mvine
apgies
.
buwhce
during
clinutic
c
y
c
k
Mu,
T
e
h
Sx.
Jour.
7
k
5%.
Environmental
Resevch
Lboramy.
1976.
prmdwra
for
developing
tempeatwe
miteria
for
frahwp.
ter
f&
Emlogid
Reaeych
series.
repfi
in
­
tion
ERL'
Duluth.
Mjm
PLdenl
Wrtpr
Pollution
Contml
Administnrion
l967.
Tempenture
ud
q
u
t
i
c
life.
L.
bontmy
laveatigatiorm
No.
6.
Technical
Ad.
risory
lad
Invmtigatiom
Branch,
C
i
d
t
i
.
Ohio.
Fry,
F.
EJ.
1967.
Rmpomz
of
vertebrate
poiLilotbenns
to
tewersture
I
n
AH.
Rose.
ed.

­­

h=

xa
eiwr
Vorbehandlung
nit
koast.
nten
urd
wghaelnden
tempentauen
uti
die
hiWemistenr
von
Gommonu
sdiw
uod
I&
ta
kJuim
&.
Biol.
15:
12
Clude,
J.
B.
ISM
The
e
f
f
d
of
temperature
d
psd.
tora
on
the
.
bundance
of
the
mfubell
drm,
dlyoamk,
in
New
Fh&
d.
TMb
Amer.
Fish
Soc
&
19.
I
'
Rcm,
NewYok
Glynn.
P.
W.
1963.
b
m
o
d
i
t
i
e
s
of
eehiwids
andocberrpefllrtog.
nismsmi~
nt
6th
midday.
low
water
erposursa
in
Puerto
Rieo.
Mar.
Biol.
3
226.
Gouda.
J.
G.
1~
iZ
h
M
[
vanition
in
tbe
reapow
of
&
ne
popuhtioy
to
k
t
e
d
warn
in
tbe
vidnity
of
m
gewratang
phnt
Ph.
D.
DLssenrtaon,
Umversty
of
RbodekLnd.
H
d
.
S
A
et
J.
1967.
Contml
techniques
for
mrgulrtiobfdbation
Jwr.
Amer.
Water
I
T
I
.
I'

I
E
!
!
!

,
.
.
.
­­­
...
~
..,

J
~
V
O
L
.,
I..

7
.
,
I
I
!
UNITED
STATES
ENVIRONMENTAL
PROTECTION
AGENCY
OFFICE
OF
WATER
REGULATIONS
AND
STANDARDS
CRITERIA
AND
STANDARDS
DIVISION
UPDATE
t2
to
"
QUALITY
CRITERIA
FOR
WATER
1986"

May
1,
1987
This
is
the
second
update
to
the
EPA
document
"
Quality
Criteria
for
Water
1986".
Included
in
this
package
are
criteria
summaries
for
contaminants
that
were
recently
revised
as
well
as
a
criteria
summary
for
a
new
contaminant.
Several
hand
corrections
are
also
included.

@
Revised
New
Hand
corrections
INDEX
CHLORPYRIFOS
AMMONIA
SUMMARY
CHART
CYANJDE
NICKEL
CHLORINATED
ETHANES
P
ARATH
I
ON
PENTACHLOROPHENOL
TOXAPHENE
ZINC
Directions:
a.
Replace
sections
that
have
been
revised
with
b.
Insert
new
section
alphabetically.
c.
Make
the
identified
hand
corrections.
new
sections.

For
additional
information
contact:
EPA's
Criteria
and
Standards
Division
at
(
202)
475­
7315.
I
HAND
CORRECTIONS
.
.

.
.
.
.
oaia
­
page
5,
center
page,
third
line
of
equation
should
be
changed
from
.'
.
"
FPH
=
1
:
8
<
pH
<
9"

to
.
.
,

;
8
5
pH
<
9"
"
FPH
=
1
­
'
'.
.
.

.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.

Cyanide
­
last
page,
include
the
following
after
the
last
line
.
.
.
.
"(
45
F
­
R
­
79318,
NOV.
2
8
,
1980)
(
50
F
.
R
.
30784,
July
2
9
.
1985)"
."
SEE
APPENDIX
A
FOR
METHODOLOGY"

.,
.
.
.
.
.
,
.

­
.
':.'
Chlorinated
Ethanes
­
last
page,
change
.
.

1
"
1.03
ug/
l"

to
"
1.03
g/
L"
,
INDEX
INTRODUCTION
SUMMARY
CHART
NTIS
No.

PB
81­
117269
PB
81­
117277
PB
81­
117285
PB­
263943
PB­
263943
PB
81­
117301
PB
85­
227114
PB
81­
117319
PB
85­
227445
PB
81­
117335
PB­
263943
PB
81­
117293
PB
81­
117343
PB
81­
117350
PB­
263943
PB
85­
227031
PB
81­
117376
PB
81­
1'
17384
PB
81­
117392
PB
81­
117400
PB
81­
117426
PB
85­
227429
PB
81­
117434
PB
81­
117418
PB
81­
117442
PB­
263943
PB
87­
105359
PB
85­
227478
PB
81­
117459
PB­
263943
PB
85­
227023
PB
85­
227460
PB
81­
117491
PB­
263943
PB
81­
117509
PB
81­
117517
PB
81­
117525
PB
81­
117533
PB
86­
158­
045
 i
PB­
263943
Acenaphthene
Acrolein
Acrylonitrile
Aesthetics
Alkalinity
AldrinDieldrin
Ammonia
Antimony
Arsenic
Asbestos
Bacteria
Barium
Benzene
Benzidine
Beryllium
Boron
Cadmium
Carbon
Tetrachloride
Chlordane
Chlorinated
Benzenes
Chlorinated
Ethanes
Chlorinated
Naphthalenes
Chlorine
Chlorinated
Phenols
Chloroalkyl
Ethers
Chloroform
Chlorophenoxy
Herbicides
Chlorpyrifos
Chromium
2­
Chlorophenol
Color
Copper
Cyanide
DDT
and
Metabolites
Demeton
Dichlorobenzenes
Dichlorobenzidine
Dichloroethylenes
2.4.
­
DichloroDhenol
Dichloropropanes/
Dichloropropenes
2.4.
­
Dimethylphenol
Dinitrotoluene
Diphenylhydrazine
Endosulfan
Endrin
Ethylbenzene
Fluoranthene
Gasses,
Total
Dissolved
PB
81­
117541
PB
81­
117558
PB
81­
117566
PB
81­
117731
PB
81­
117574
PB
81­
117582
PB
81­
117590
PB­
263943
PB
8i­
ii76oa
Guthion
Haloethers
Halomethanes
Hardness
Heptachlor
Hexachlorobutadiene
Hexachlorocyclohexane
Hexachlorocyclopentadiene
Iron
Isophorone
Lead
Ma
la
thion
Manganese
Mercury
Methoxychlor
Mirex
Naphthalene
Nickel
Nitrates.
Nitrites
Nitrobenzene
Nitrophenols
Nitrosamines
Oil
and
Grease
Oxygen,
Dissolved
Parathion
Pentachlorophenol
Ph
Phenol
Phosphorus
Phthalate
Esters
Polychlorinated
Biphenyls
Polynuclear
Aromatic
Hydrocarbons
Selenium
Silver
Solids
(
Dissolved)
h
Salinity
Solids
(
Suspended)
h
Turbidity
Sulfides,
Hydrogen
Sulfide
Tainting
Substances
Temperature
2.3.7.8­
Tetrachlorodibenzo­
p­
dioxin
Tetrachloroethylene
Thallium
Toluene
Toxaphene
Trichlorocthylene
Vinyl
Chloride
Zinc
PB­
263943
PB
81­
117616
PB
81­
117624
PB­
263943
PB
81­
117632
PB
81­
117640
PB
81­
117657
PB
81­
117665
PB­
263943
PB
81­
117673
PB
85­
227437
PB­
263943
PB­
263943
PB
85­
227452
PB­
263943
PB­
263943
PB
81­
117707
PB
87­
105359
PB­
263943
PB
81­
117723
PB
81­
117749
PB
81­
117756
PB­
263943
PB
86­
208253
PB
87­
105383
PB
87­
105391
PB­
263943
PB
81­
117772
PB­
263943
PB
81­
117780
PB
81­
117798
PB
81­
117806
PB
81­
117814
PB
81­
117822
PB­
263943
1
PB­
263943
PB­
263943
PB­
263943
PB­
263943
EPA
#
440/
5­
84­
007
PB
81­
117830
PB
81­
117848
PB
81­
117855
PB
87­
105375
PB
81­
117871
PB
81­
117889
PB
87­
153581
APPENDIX
A
Methodology
for
Developing
Criteria
APPENDIX
B
Methodology
for
Developing
Criteria
APPENDIX
C
Methodology
 or
Developing
Criteria
BIBLIOGRAPHY
a
CHLORPYRIFOS
Summary
The
acute
values
for
eighteen
freshwater
species
in
fifteen
genera
range
from
0.11
ug/
L
for
an
amphipod
to
greater
than
806
ug/
L
for
two
fishes
and
a
snail.
The
bluegill
is
the
most
acutely
sensitive
fish
species
with
an
acute
value
of
10
ug/
L,

but
seven
intervetebrate
genera
are
more
sensitive.
Smaller
organisms
seem
to
be
more
acutely
sensitive
than
larger
ones.

Chronic
toxicity
data
are
available
for
one
freshwater
species,
the
fathead
minnow.
Unacceptable
effects
occurred
in
second
generation
larvae
at
0.12
ug/
L,
which
was
the
lowest
concentration
tested.
The
resulting
acute­
chronic
ratio
was
greater
than
1,417.

Little
information
is
available
on
the
toxicity
of
chlorpyrifos
to
freshwater
plants,
although
algal
blooms
frequently
follow
field
applications
of
chlorpyrifos.
The
only
available
bioconcentration
test
on
chlorpyrifos
with
a
freshwater
species
was
with
the
fathead
minnow
and
resulted
in
a
bioconcen­

tration
factor
of
1,673.

The
acute
toxicity
of
chlorpyrifos
has
been
determined
for
15
species
of
saltwater
animals
in
12
genera
with
the
acute
values
ranging
from
0.01
ug/
L
for
the
Korean
shrimp,
Palaemon
macrodactvlus,
to
1.911
ug/
L
for
larvae
of
the
eastern
oyster,

Crassostrea
virginica.
Arthropods
are
particularly
sensitive
to
chlorpyrifos.
Among
the
10
species
of
fish
tested,
the
96­
hr
LC50s
range
from
0
.
5
8
ug/
L
for
striped
bass
to
520
ug/
L
for
gulf
toadfish.
Fish
larvae
are
more
sensitive
than
other
life
stages.

Growth
of
the
mysid,
Mysidolssis
bahia,
was
reduced
at
0.004
ug/
L
in
a
life­
cycle
test.
In
early
life­
stage
tests,
the
California
grunion,
Leuresthes
tenuis,
was
the
most
sensitive
of
the
six
a.

fishes,
with
growth
being
reduced
at
0.30
ug/
L.
Of
the
seven
acute­
chronic
ratios
that
have
been
determined
with
saltwater
species,
the
five
lowest
range
from
1.374
to
12.50,
whereas
the
highest
is
228.5.

Concentrations
of
chlorpyrifos
affecting
six
species
of
saltwater
phytoplankton
range
from
138
to
10,000
ug/
L.
BCFs
ranged
from
100
to
5,100
when
the
gulf
toadfish
was
exposed
to
.
oncentrations
increasing
from
1.4
to
150
ug/
L.
Steady­
state
BCFs
averaged
from
100
to
757
for
five
fishes
exposed
in
early
life­
stage
tests.

National
Criteria
The
procedures
described
in
the
"
Guidelines
for
Deriving
Numerical
National
Water
Quality
Criteria
for
the
Protection
of
Aquatic
Organisms
and
Their
Uses"
indicate,
that
except
possibly
where
a
locally
important
species
is
very
sensitive,
freshwater
aquatic
organisms
and
their
uses
should
not
be
affected
unacceptably
if
the
four­
day
average
concentration
of
chlorpyrifos
does
not
exceed
0.041
ug/
L
more
than
once
every
three
years
on
the
average
and
if
the
one­
hour
average
concentration
does
not
exceed
0.083
ug/
L
more
than
once
every
three
years
on
the
average.

The
procedures
described
in
the
"
Guidelines
for
Deriving
Numerical
National
Water
Quality
Criteria
for
the
Protection
of
Aquatic
Organisms
and
Their
Uses"
indicate,
that
except
possibly
a
where
a
locally
important
species
is
very
sensitive,
saltwater
aquatic
organisms
and
their
uses
should
not
be
affected
unacceptably
if
the
four­
day
average
concentration
of
chlorpyrifos
does
not
exceed
0.0056
ug/
L
more
than
once
every
three
years
on
the
average
and
if
the
one­
hour
average
concentration
does
not
exceed
0.011
ug/
L
more
than
once
every
three
years
on
the
average.

Three
yeaks
is
the
Agency's
best
scientific
judgment
of
the
average
amount
of
time
aquatic
ecosystems
should
be
provided.

between
excursions.
The
resiliences
of
ecosystems
and
their
abilities
to
recover
differ
greatly,
however,
and
site­
specific
allowed
excursion
frequencies
may
be
established
if
adequate
justification
is
provided.

a
Use
of
criteria
for
developing
water
quality­
based
permit
limits
and
for
designing
waste
treatment
facilities
requires
selection
of
an
appropriate
wasteload
allocation
model.
Dynamic
models
are
preferred
for
the
application
of
these
criteria.

Limited
data
or
other
consideretions
might
make
their
use
impractical,
in
which
case
one
must
rely
on
a
steady­
state
model.

(
51
F.
R.
43665,
December
3
,
1986)
SEE
APPENDIX
A
FOR
METHODOLOGY
a
Summary
Acute
values
with
twency­
one
freshwater
species
in
18
genera
range
from
1,101
ug/
L
for
a
cladoceran
to
43,240
ug/
L
for
a
fish.

Fishes
and
invertebrates
are
both
spread
throughout
the
range
of
sensitivity.
Acute
values
with
four
species
are
significantly
correlated
with
hardness.
Data
are
available
concerning
the
chronic
toxicity
of
nickel
to
two
invertebrates
and
two
fishes
in
freshwater.
Data
available
for
two
species
indicate
that
chronic
toxicity
decreases
as
hardness
increases.
The
measured
chronic
values
ranged
from
14.77
ug/
L
with
DaDhnia
maqna
in
soft
water
to
526.7
ug/
L
with
the
fathead
minnow
in
hard
water.
Five
acute­

chronic
ratios
are
available
for
two
species
in
soft
and
hard
water
and
range
from
14
to
122.

Nickel
appears
to
be
quite
toxic
to
freshwater
algae,
with
a
concentrations
as
low
as
50
ug/
L
producing
significant
inhibition.
Bioconcentration
factors
for
nickel
range
from
0.8
for
fish
muscle
to
193
for
a
cladoceran.
I
Acute
values
for
twenty­
three
saltwater
species
in
twenty
genera
range
from
151.7
ug/
L
with
juveniles
of
a
mysid
of
to
1,100,000
ug/
L
with
juveniles
and
adults
of
a
clam.
The
acute
values
for
the
four
species
of
fish
range
from
7,598
to
350.000
ug/
L.
The
acute
toxicity
of
nickel
appears
to
be
related
to
salinity,
but
the
form
of
the
relationship
appears
to
be
species­

dependent.

Mysidopsis
bahia
is
the
only
saltwater
species
with
which
an
acceptable
chronic
test
has
been
conducted
on
nickel.
Chronic
0
exposure
to
141
ug/
L
and
greater
resulted
in
reduced
survival
and
reproduction.
The
measured
acute­
chronic
ratio
was
5.478.
01
Bioconcentration
factors
in
saltwater
range
from
261.8
with
a
oyster
to
675
with
a
brown
alga.

National
Criteria
The
procedures
described
in
the
"
Guidelines
for
Deriving
Numerical
National
Water
Quality
Criteria
for
the
Protection
of
Aquatic
Organisms
and
Their
Uses"
indicate,
that
except
possibly
where
a
locally
important
species
is
very
sensitive,
freshwater
aquatic
organisms
and
their
uses
should
not
be
affected
unacceptably
if
the
four­
day
average
concentration
of
nickel
(
in
(
0.8460[
In
ug/
L)
does
not
exceed
the
numerical
value
given
by
e
(
hardness)
1+
1.1645)
more
than
once
every
three.
years
on
the
average
and
if
the
one­
hour
average
concentration
(
in
ug/
L)
does
not
exceed
the
numerical
value
given
by
e
(
0.846011n
0)

more
than
once
every
three
years
on
the
(
hardness)
1+
3.3612)

average.
For
example,
at
hardnesses
of
50.
100,
and
200
mg/
L
as
CaC03
the
four­
day
average
concentrations
of
nickel
are
88,
160.

and
280
ug/
L,
respectively,
and
the
one­
hour
average
concentrations
are
790,
1400.
and
2500
ug/
L.

The
procedures
described
in
the
"
Guidelines
for
Deriving
Numerical
National
Water
Quality
Criteria
for
the
Protection
of
Aquatic
Organisms
and
Their
Uses"
indicate,
that
except
possibly
where
a
locally
important
species
is
very
sensitive,
saltwater
aquatic
organisms
and
their
uses
should
not
be
affected
unacceptably
if
the
four­
day
average
concentration
of
nickel
does
not
exceed
8.3
ug/
L
more
than
once
every
three
years
on
the
average
and
if
the
one­
hour
average
concentration
does
not
exceed
0
)
a
75
ug/
L
more
than
once
every
three
years
on
the
average.

"
Acid­
soluble''
is
probably
the
best
measurement
at
present
for
expressing
criteria
for
metals
and
the
criteria
for
nickel
were
developed
on
this
basis.
However,
at
this
time,
no
EPA
approved
method
for
such
a
measurement
is
available
to
implement
criteria
for
metals
through
the
regulatory
programs
of
the
Agency
and
the
States.
The
Agency
is
considering
development
and
approval
of
a
method
for
a
measurement
such
as
"
acid­
soluble."

Until
one
is
approved,
however,
EPA
recommends
applying
criteria
for
metals
using
the
total
recoverable
method.
This
has
two
impacts:
(
1)
certain
species
of
some
metals
cannot
be
measured
because
the
total
recoverable
method
cannot
distinguish
between
individual
oxidation
States,
and
(
2
)
in
some
cases
these
criteria
might
be
overly
protective
when
based
on
the
total
recoverable
method
.

Three
years
is
the
Agency's
best
scientific
judgment
of
the
average
amount
of
time
aquatic
ecosystems
should
be
provided
between
excursions.
The
resiliences
of
ecosystems
and
their
abilities
to
recover
differ
greatly,
however,
and
site­
specific
allowed
excursion
frequencies
may
be
established
if
adequate
justification
is
provided.

Use
of
criteria
for
developing
water
quality­
based
permit
limits
and
for
designing
waste
treatment
facilities
requires
selection
of
an
appropriate
wasteload
allocation
model.
Dynamic
models
are
preferred
for
the
application
of
these
criteria.

Limited
data
or
other
considerations
might
make
their
use
impractical,
in
which
case
one
must
rely
on
a
steady­
state
model.
(
51
F
.
R
.
43665,
December
3,
1986)
SEE
APPENDIX
A
FOR
METHODOLOGY
PARATHION
Summary
The
acute
values
for
thirty­
seven
freshwater
species
in
thirty­
one
genera
range
from
0.04
ug/
L
for
an
early
instar
of
a
crayfish,
Orconectes
nais,
to
5,230
ug/
L
for
two
species
of
tubificid
worms.
For
Daphnia
mama,
the
chronic
value
and
acute­

chronic
ratio
are
0.0990
ug/
L
and
10.10
respectively.
Chronic
toxicity
values
are
available
for
two
freshwater
fish
species,

the
bluegill
and
the
fathead
minnow,
with
chronic
values
of
0.24
ug/
L
and
6.3
ug/
L,
and
acute­
chronic
ratios
of
2.121
and
79.45.

respectively.
Two
freshwater
algae
were
affected
by
toxaphene
concentrations
of
30
and
390
ug/
L,
respectively.

Bioconcentration
factors
determined
with
three
fish
species
ranged
from
27
to
573.

The
acute
values
that
are
available
for
saltwater
species
are
11.5
and
17.8
ug/
L
for
the
Korean
shrimp,
Palaemon
macrodactvlus.

and
17.8
ug/
L
for
the
striped
bass,
Morone
saxatilis.
No
data
are
available
concerning
the
chronic
toxicity
of
parathion
to
saltwater
species,
toxicity
to
saltwater
plants,
or
bioaccumulation
by
saltwater
species.
Some
data
indicate
that
parathion
is
acutely
lethal
to
commercially
important
saltwater
shrimp
at
concentrations
as
low
as
0.24
ug/
L.
Measurement
of
acetylcholinesterase
(
AChE)
in
fish
tissue
might
be
useful
for
diagnosing
fish
kills
caused
by
parathion.

National
Criteria
The
procedures
described
in
the
"
Guidelines
for
Deriving
Numerical
National
Water
Quality
Criteria
for
the
Protection
of
Aquatic
Organisms
and
Their
Uses"
indicate,
that
except
possibly
where
a
locally
important
species
is
very
sensitive,
freshwater
a
aquatic
organisms
and
their
uses
should
not
be
affected
unacceptably
if
the
four­
day
average
concentration
of
parathion
does
not
exceed
0.013
ug/
L
more
than
once
every
three
years
on
the
average
and
if
the
one­
hour
average
concentration
does
not
exceed
0.065
ug/
L
more
than
once
every
three
years
on
the
average.

The
procedures
described
in
the
"
Guidelines
for
Deriving
Numerical
National
Water
Quality
Criteria
for
the
Protection
of
Aquatic
Organisms
and
Their
Uses"
require
the
availability
of
specified
data
for
the
derivation
of
a
criterion.
A
saltwater
criterion
for
parathion
cannot
be
derived
because
very
few
of
the
required
data
are
available.

Three
years
is
the
Agency's
best
scientific
judgment
of
the
.

I
average
amount
of
time
aquatic
ecosystems
should
be
provided
between
excursions.
The
resiliences
of
ecosystems
and
`
their
abilities
to
recover
differ
greatly,
however,
and
site­
specific
allowed
excursion
frequencies
may
be
established
if
adequate
justification
is
provided.

Use
of
criteria
for
developing
water
quality­
based
permit
limits
and
for
designing
waste
treatment
facilities
requires
selection
of
an
appropriate
wasteload
allocation
model.
Dynamic
models
are
preferred
for
the
application
of
these
criteria.

Limited
data
or
other
considerations
might
make
their
use
impractical,
in
which
case
one
must
rely
on
a
steady­
state
model.

(
51
F.
R.
43665,
December
3,
1986)
SEE
APPENDIX
A
FOR
METHODOLOGY
PENTACHLOROPHENOL
(
PCP)

Summary
The
acute
and
chronic
toxicity
of
PCP
to
freshwater
animals
increased
as
pH
and
dissolved
oxygen
concentration
of
the
water
decreased.
Generally,
toxicity
also
increased
with
increased
temperature.
The
estimated
acute
sensitivities
of
36
species
at
pH
=
6.5
ranged
from
4.355
ug/
L
for
larval
common
carp
to
>
43,920
ug/
L
for
a
crayfish.
At
pH
=
6.5,
the
lowest
and
highest
estimated
chronic
values
of
11.835
and
79.06
ug/
L,
respectively,

were
obtained
with
different
cladoceran
species.
Chronic
toxicity
to
fish
was
affected
by
the
presence
of
impurities,
with
industrial­
grade
PCP
being
more
toxic
than
purified
samples.

Mean
acute­
chronic
ratios
for
six
freshwater
species
ranged
from
0.8945
to
>
15.79,
but
the
mean
ratios
for
the
four
most
acutely
sensitive
species
only
range
0.8945
to
5.035.
Freshwater
algae
were
affected
by
concentrations
as
low
as
7.5
ug/
L,
whereas
vascular
plants
were
affected
at
189
ug/
L
and
above.

Bioconcentration
factors
ranqed
from
7.3
to
1,066
for
three
species
of
fish.

Acute
toxicity
values
from
tests
with
18
species
of
saltwater
animals,
representing
17
genera,
range
from
22:
63
ug/
L
for
late
yolk­
sac
larvae
of
the
Pacific
herring,
Clupea
narenqus
pallasi,

to
18,000
ug/
L
for
adult
blue
mussels,
Mvtllus
edulis.
The
embryo
and
larval
stages
of
invertebrates
and
the
late
larval
premetamorphosis
stage
of
a
s
h
appear
to
be
the
most
sensitive
0
life
stages
to
PCF.
With
few
exceptions,
M
s
h
are
more
sensitive
.
than
invertebrates
to
PCP.
Salinity,
temperature,
and
pR
have
a
slight
effect
on
the
toxicity
of
PCP
to
some
saltwater
animals.

Life­
cycle
toxicity
tests
have
been
conducted
with
the
sheepshead
minnow,
Cmrinodon
variegatus.
and
the
polychaete
worm,
Ophrvotrocha
diadema.
The
chronic
value
for
the
minnow
is
64.31
ug/
L
and
the
acute­
chronic
ratio
is
6.873.
Unfortunately,

no
statistical
analysis
of
the
data
from
the
test
with
the
worm
is
available.

The
EC506
for
saltwater
plants
range
from
17.40
ug/
L
for
the
diatom,
Skeletonema
costatum,
to
3.600
ug/
L
for
the
green
alga,

Dunaliella
tertiolecta.
Apparent
steady­
state
BCFs
are
available
for
the
eastern
oyster.
Crassostrea
virginica,
and
two
saltwater
fishes
and
range
from
10
to
8
2
.

National
Criteria
The
procedures
described
in
F
e
"
Guidelines
for
Deriving
Numerical
National
Water
Quality
Criteria
for
the
Protection
of
Aquatic
Organisms
and
Their
Uses"
indicate,
that
except
possibly
where
a
locally
important
species
is
very
sensitive,
freshwater
aquatic
organisms
and
their
uses
should
not
be
affected
unacceptably
if
the
four­
day
average
concentration
(
in
ug/
L)
of
pentachlorophenol
does
not
exceed
the
numerical
value
given
by
e
11.005(
pH)­
5­
2901
more
than
once
every
three
years
on
the
average
and
if
the
one­
hour
average
concentration
(
in
ug/
L)
does
not
exceed
the
numerical
value
given
by
e
[
l.
OOS(
pH)­
4.8301
than
once
every
three
years
on
the
average.
For
example,
at
pH
=

6.5.
7.8,
and
9.0
the
four­
day
average
concentrations
pentachlorophenol
are
3.5,
13,
and
43
ug/
L,
respectively,
and
the
one­
hour
average
concentrations
are
5.5,
20,
and
68
ug/
L.
At
pH
=
6.8,
a
pentachlorophenol
concentration
of
1.74
ug/
L
caused
a
50%
reduction
in
the
growth
of
yearling
sockeye
salmon
in
a
56­

day
test.
a
The
procedures
described
in
the
"
Guidelines
for
Deriving
Numerical
National
Water
Quality
Criteria
for
the
Protection
of
Aquatic
Organisms
and
Their
Uses"
indicate,
that
except
possibly
where
a
locally
important
species
is
very
sensitive,
saltwater
aquatic
organisms
and
their
uses
should
not
be
affected
unacceptably
if
the
four­
day
average
concentration
of
pentachlorophenol
does
not
exceed
7.9
ug/
L
more
than
once
every
three
years
on
the
average
and
if
the
one­
hour
average
concentration
does
not
exceed
13
ug/
L
more
than
once
every
three
years
on
the
average.

Three
years
is
the
Agency's
best
scientific
judgment
of
the
average
amount
of
time
aquatic
ecosystems
should
be
provided
between
excursions.
The
resiliences
of
ecosystems
and
their
abilities
to
recover
differ
greatly,
however,
and
site­
specific
allosred
excursion
frequencies
may
be
established
if
adequate
justification
is
provided.

Use
of
criteria
for
developing
water
quality­
based
permit
limits
and
for
designing
waste
treatment
facilities
requires
selection
of
an
appropriate
wasteload
allocation
model.
Dynamic
models
are
preferred
for
the
application
of
these
criteria.

Limited
data
or
other
considerations
might
make
their
w
e
impractical,
in
which
case
one
must
rely
on
a
steady­
state
model.

(
51
F
.
R
.
43665,
December
3,
1986)
SEE
APPENDIX
A
FOR
METHODOLOGY
TOXAPHENE
Summary
The
acute
sensitivities
of
36
freshwater
species
in
28
genera
ranged
from
0.8
ug/
L
to
500
ug/
L.
Such
important
fish
species
as
the
channel
catfish,
largemouth
bass,
chinook
and
coho
salmon,

brook,
brown
and
rainbow
trout,
striped
bass,
and
bluegill
had
acute
senitivities
ranging
from
0.8
ug/
L
to
10.8
ug/
L.
Chronic
values
for
four
freshwater
species
range
from
less
than
0.039
ug/
L
for
the
brook
trout
to
0.1964
ug/
L
for
the
channel
catfish.

The
growth
of
algae
was
affected
at
100
to
1,000
ug/
L,
and
bioconcentration
factors
from
laboratory
tests
ranged
from
3.100
to
90,000.
Concentrations
in
lake
trout
in
the
Great
Lakes
have
frequently
exceeded
the
U.
S.
FDAJaction
level
of
5
mg/
kg,
even
though
the
concentrations
in
the
water
seem
to
be
only
1
to
4
ng/
L.
These
concentrations
in
lake
water
are
thought
to
have
resulted
from
toxaphene
being
transported
to
the
Great
Lakes
from
remote
sites,
the
locations
of
which
are
not
well
known.
a
The
acute
toxicity
of
toxaphene
to
15
species
of
saltwater
animals
ranges
from
0.53
for
pinfish,
Laqodon
rhomoides.
to
460.000
ug/
L
for
the
adults
of
the
clam,
Ranqia
cuneata.
Except
for
resistant
species
tested
at
concentrations
greater
than
toxaphene's
water
solubility,
acute
values
for
most
species
were
within
a
factor
of
10.
The
toxicity
of
toxaphene
was
found
to
decrease
slightly
with
increasing
salinity
for
adult
blue
crabs.

Callinectes
sapidus,
whereas
no
relationship
between
toxicity
and
salinity
was
observed
with
the
three
spine
stickleback,
Gasterosteus
aculeatus.
Temperature
significantly
affected
the
toxicity
of
toxaphene
to
blue
crabs.

Early
life­
stage
toxicity
tests
have
been
conducted
with
the
sheepshead
minnow,
Cmrinodon
varieqatus,
and
the
longnose
killifish,
Fundulus
similis,
whereas
life­
cycle
tests
have
been
conducted
with
the
sheepshead
minnow
and
a
mysid.
For
the
sheepshead
minnow,
chronic
values
of
1.658
ug/
L
from
the
early
life­
stage
test
and
0.7141
ug/
L
from
the
life­
cycle
toxicity
test
are
similar
to
the
96­
hr
LC50
of
1.1
ug/
L.
Killifish
are
more
chronically
sensitive
with
effects
noted
at
0.3
ug/
L.
In
the
life­
cycle
test
with
the
mysid,
no
adverse
effects
were
observed
at
the
highest
concentration
tested,
which
was
only
slightly
below
the
96­
hr
LC50,
resulting
in
an
acute­
chronic
ratio
of
1.132.

Toxaphene
is
bioconcentrated
by
an
oyster,
Crassostrea
01
virsinica,
and
two
fishes,
C.
variegatus
and
z­
similis,
to
concentrations
that
range
from
9,380
to
70.140
times
that
in
the
test
solution.

National
Criteria
The
procedures
described
in
the
"
Guidelines
for
Deriving
Numerical
National
Water
Quality
Criteria
for
the
Protection
of
Aquatic
Organisms
and
Their
Uses"
indicate,
that
except
possibly
where
a
locally
important
species
is
very
sensitive,
freshwater
aquatic
organisms
and
their
uses
should
not
be
affected
unacceptably
if
the
four­
day
average
concentration
of
toxaphene
does
not
exceed
0.0002
ug/
L
more
than
once
every
three
years
on
the
average
and
if
the
one­
hour
average
concentration
does
not
exceed
0.73
ug/
L
more
than
once
every
three
years
on
the
average.

0
If
the
concentration
of
toxaphene
does
exceed
0.0002
ug/
L.
the
edible
portions
of
consumed
species
should
be
analyzed
to
determine
whether
the
concentration
of
toxaphene
exceeds
the
FDA
action
level
of
5
mg/
kg.
If
the
channel
catfish
is
as
acutely
sensitive
as
some
data
indicate
it
might
be,
it
will
not
be
protected
by
this
criterion.

The
procedures
described
in
the
"
Guidelines
for
Deriving
Numerical
National
Water
Quality
Criteria
for
the
Protection
of
Aquatic
Organisms
and
Their
Uses"
indicate,
that
except
possibly
where
a
locally
important
species
is
very
sensitive,
saltwater
aquatic
organisms
and
their
uses
should
not
be
affected
unacceptably
if
the
four­
day
average
concentration
of
toxaphene
does
not
exceed
0.0002
ug/
L
more
than
once
every
three
years
on
the
average
and
if
the
one­
hour
average
concentration
does
not
exceed
0.21
ug/
L
more
than
once
every
three
years
on
the
average.

If
the
concentration
of
toxaphene
does
exceed
0.0002
ug/
L,
the
edible
portions
of
consumed
species
should
be
analyzed
to
determine
whether
the
concentration
of
toxaphene
exceeds
the
FDA
action
level
of
5
mg/
kg.
e
I
Three
years
is
the
Agency's
best
scientific
judgment
of
the
average
amount
of
time
aquatic
ecosystems
should
be
provided
between
excursions.
The
resiliences
of
ecosystems
and
their
abilities
to
recover
differ
greatly,
however,
and
site­
specific
allowed
excursion
frequencies
may
be
established
if
adequate
justification
is
provided.

Use
of
criteria
for
developing
water
quality­
based
permit
0
limits
and
for
designing
waste
treatment
facilities
requires
selection
of
an
appropriate
wasteload
allocation
model.
Dynamic
models
are
preferred
for
the
application
of
these
criteria.

Limited
data
or
other
considerations
might
make
their
use
impractical,
in
which
case
one
must
rely
on
a
steady­
state
model.

(
51
F.
R.
43665,
December
3,
1986)
SEE
APPENDIX
A
FOR
METHODOLOGY
0)
Summary
Acute
toxicity
values
are
available
for
43
species
of
fresl­
rater
animals
and
data
for
eight
species
indicate
that
acute
toxicity
decreases
as
hardness
increases.
When
adjusted
to
a
hardness
of
50
mg/
L,
sensitivities
range
from
50.70
ug/
L
for
CeriodaDhnia
reticulata
to
88,960
ug/
L
for
a
damselfly.

Additional
data
indicate
that
toxicity
increases
as
temperGture
increases.
Chronic
toxicity
data
are
available
for
nine
freshwater
species.
Chronic
values
for
two
invertebrates
ranged
from
46.73
ug/
L
for
DaDhnia
mama
to
>
5,243
ug/
L
for
the
caddisfly,
Clistoronla
magziifica.
Chronic
values
for
seven
fish
species
ranged
from
36.41
ug/
L
for
the
flagfish,
Jordanella
9
floridae,
to
854.7
ug/
L
for
the
brook
trout,
Salvelinus
fontinalis.
Acute­
chronic
ratios
ranged
from
0.2614
to
41.20.

but
the
ratios
for
the
sensitive
species
were
all
less
than
7.3.

The
sensitivity
range
of
freshwater
plants
to
zinc
is
greater
than
that
for
animals.
Growth
of
the
alga,
Selenastrum
capriocornutum,
was
inhibited
by
30
ug/
L.
On
the
other
hand,

with
several
other
species
of
green
algae,
4­
day
EC5Os
exceeded
200,000
ug/
L.
Zinc
was
found
to
bioaccumulate
in
freshwater
animal
tissues
from
51
to
1,130
times
the
concentration
present
in
the
water.

Acceptable
acute
toxicity
values
for
zinc
are
available
for
33
species
of
saltwater
animals
including
26
invertebrates
and
7
fish.
LCSOs
range
from
191.5
ug/
L
for
cabezon,
ScorDanichthvs
a
marmoratus
to
320.400
ug/
L
for
adults
of
another
clam,
Macoma
balthica.
Early
life
stages
of
saltwater
invertebrates
and
01
fishes
are
generally
more
sensitive
to
zinc
than
juveniles
and
adults.
Temperature
has
variable
and
inconsistent
effects
on
the
sensitivity
of
saltwater
invertebrates
to
zinc.
The
sensitivity
of
saltwater
vertebrate
animals
to
zinc
deyreases
with
increasing
salinity,
but
the
magnitudG
of
the
effect
is
species­
specific.

A
life­
cycle
test
with
the
mysid,
Mvsidopsia
bahia,
found
unacceptable
effects
at
120
ug/
L,
but
not
at
231
ug/
L,
and
the
acute­
chronic
ratio
was
2.997.
Seven
species
of
saltwater
plants
were
affected
at
concentrations
ranging
from
19
to
10,100
ug/
L.

Bioaccumulation
data
for
zinc
are
available
for
seven
species
of
saltwater
algae
and
five
species
of
saltwater
animals.
Steady­

state
zinc
bioconcentration
factors
for
the
twelve
species
range
from
3.692
to
23.820.

National
Criteria
The
procedures
described
in
the
"
Guidelines
for
Deriving
Numerical
National
Water
Quality
Criteria
for
the
Protection
of
Aquatic
Organisms
and
Their
Uses"
indicate,
that
except
possibly
where
a
locally
important
species
is
very
sensitive,
freshwater
aquatic
organisms
and
their
uses
should
not
be
affected
unacceptably
if
the
four­
day
average
concentration
of
zinc
(
in
(
0.847311n
ug/
L)
does
not
exceed
the
numerical
value
given
by
e
more
than
once
every
three
years
on
the
(
hardness)
l+
0.7614)

average
and
if
the
one­
hour
average
concentration
(
in
ug/
L)
does
(
0.8473
[
In
not
exceed
the
numerical
value
given
by
e
more
than
once
every
three
years
on
the
(
hardness)]+
0.8604)
0
average.
For
example,
at
hardnesses
of
50,
100,
and
200
mg/
L
as
CaC03.
the
four­
day
average
concentrations
of
zinc
are
5
9
,
110
and
190
ug/
L,
respectively,
and
the
one­
hour
average
concentrations
are
65,
120,
and
210
ug/
L.
If
the
striped
bass
is
as
sensitive
as
some
data
indicate,
it
will
not
be
protected
by
this
criterion.

The
procedures
described
in
the
"
Guidelines
for
Deriving
Numerical
National
Water
Quality
Criteria
for
the
Protection
of
Aquatic
Organisms
and
Their
U
s
e
s
"
indicate,
that
except
possibly
where
a
locally
important
species
is
very
sensitive,
saltwater
aquatic
organisms
and
their
uses
should
not
be
affected
unacceptably
if
the
four­
day
average
concentration
of
zinc
does
not
exceed
86
ug/
L
more
than
once.
every
three
years
on
the
average
and
if
the
one­
hour
average
concentration
does
not
exceed
95
ug/
L
more
than
once
every
three
years
on
the
average.
a
"
Acid­
soluble''
is
probably
the
best
measurement
at
present
for
expressing
criteria
for
metals
and
the
criteria
for
zinc
were
developed
on
this
basis.
However,
at
this
time
no
FPA
approved
method
for
such
a
measurement
is
available
to
implement
criteria
for
metals
through
the
regulatory
programs
of
the
Agency
and
the
States.
The
Agency
is
considering
development
and
approval
of
a
method
for
a
measurement
such
as
"
acid­
soluble.''
Until
one
is
approved,
however,
EPA
recommends
applying
criteria
for
metals
using
the
total
recoverable
method.
This
has
two
impacts:
(
1)

certain
species
of
Some
metals
cannot
be
measured
because
the
total
recoverable
method
cannot
distinguish
between
individual
oxidation
States,
and
(
2
)
in
some
cases
these
criteria
might
be
overly
protective
when
based
on
the
total
recoverable
method.
Three
years
is
the
Agency's
best
scientific
judgment
of
the
average
amount
of
time
aquatic
ecosystems
should
be
provided
a
between
excursions.
The
resiliences
of
ecosystems
and
their
abilities
to
recover
differ
greatly,
however,
and
site­
specific
allowed
excursion
frequencies
may
be
established
if
adequate
justification
is
provided.

Use
of
criteria
for
developing
water
quality­
based
permit
limits
and
for
designing
waste
treatment
facilities
requires
selection
of
an
appropriate
wasteload
allocation
model.
Dynamic
models
are
preferred
for
the
application
of
these
criteria.

Limited
data
or
other
considerations
might
make
their
use
impractical,
in
which
case
one
must
rely
on
a
steady­
state
model.

(
5
2
F
.
R
.
6213,
March
2
.
1987)
SEE
APPENDIX
A
FOR
METHODOLOGY