Document ID: EPA-HQ-OW-2002-0033-0339
Agency: epa
Document Type: Supporting & Related Material
Title: 
Posted Date: 2003-04-14T04:00Z

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UNITED
STATES
ENVIRONMENTAL
PROTECTION
AGENCY
WASHINGTON,
D.
C.
20460
FEB
22
1994
OFFICE
OF
WATER
EPA­
823­
B­
94­
001
MEMORANDUM
SUBJECT:
Use
of
the
Water­
Effect
Ratio
Standards
FROM
:
Tudor
T.
Davies,
Director
Office
of
Science
and
TO:
Water
Management
Division
Directors,
Regions
I
­
X
State
Water
Quality
Standards
Program
Directors
PURPOSE
There
are
two
purposes
for
this
memorandum.

The
first
is
to
transmit
the
Determination
and
Use
of
Water­
Effect
ratios
for
Metals.
EPA
committed
to
developing
this
guidance
to
support
implementation
of
federal
standards
for
those
States
included
in
the
National
Toxics
Rule.

The
second
is
to
provide
policy
guidance
on
whether
a
State's
application
of
a
water­
effect
ratio
is
a
site­
specific
criterion
adjustment
subject
to
EPA
review
and
approval/
disapproval.

BACKGROUND
In
the
early
1980'
s,
members
of
the
regulated
community
expressed
concern
that
EPA's
laboratory­
derived
water
quality
criteria
might
not
accurately
reflect
site­
specific
conditions
because
of
the
effects
of
water
chemistry
and
the
ability
of
species
to
adapt
over
time.
In
response
to
these
concerns,
EPA
created
three
procedures
to
derive
site­
specific
criteria.
These
procedures
were
published
in
the
Water
Quality
Standards
Handbook,
1983.
2
Site­
specific
criteria
are
allowed
by
regulation
and
are
subject
to
EPA
review
and
approval.
The
Federal
water
quality
standards
regulation
at
section
131.11(
b)(
l)
provides
States
with
the
opportunity
to
adopt
water
quality
criteria
that
are
"...
modified
to
reflect
site­
specific
conditions."
Under
section
131.5(
a)(
2),
EPA
reviews
standards
to
determine
"
whether
a
State.
ham
adopted
criteria
to
protect
the
designated
water
uses."

On
December
22,
1992,
EPA
promulgated
the
National
Toxics
Rule
which
established
Federal
water
quality
standards
for
14
States
which
had
not
met
the
requirements
of
Clean
Water
Act
Section
303(
c)(
2)(
8).
Am
part
of
that
rule,
EPA
gave
the
States
discretion
to
adjust
the
aquatic
life
criteria
for
metals
to
reflect
site­
specific
conditions
through
use
of
a
water­
effect
ratio.
A
water­
effect
ratio
is
a
means
to
account
for
a
difference
between
the
toxicity
of
the
metal
in
laboratory
dilution
water
and
its
toxicity
in
the
water
at
the
site.

In
promulgating
the
National
Toxics
Rule,
EPA
committed
to
issuing
updated
guidance
on
the
derivation
of
water­
effect
ratios.
The
guidance
reflects
new
information
since
the
previous
guidance
and
is
more
comprehensive
in
order
to
provide
greater
clarity
and
increased
understanding.
This
new
guidance
should
help
standardize
procedures
for
deriving
water­
effect
ratios
and
make
results
more
comparable
and
defensible.

Recently,
an
issue
arose
concerning
the
most
appropriate
form
of
metals
upon
which
to
base
water
quality
standards.
On
October
1,
1993,
EPA
issued
guidance
on
this
issue
which
indicated
that
measuring
the
dissolved
form
of
metal
is
the
recommended
approach.
This
new
policy
however,
is
prospective
and
does
not
affect
the
criteria
in
the
National
Toxics
Rule.
Dissolved
metals
criteria
are
not
generally
numerically
equal
to
total
recoverable
criteria
and
the
October
1,
1993
guidance
contains
recommendations
for
correction
factors
for
fresh
water
criteria.
The
determination
of
mite­
specific
criteria
is
applicable
to
criteria
expressed
as
either
total
recoverable
metal
or
am
dissolved
metal.

DISCUSSION
Existing
guidance
and
practice
are
that
EPA
will
approve
site­
specific
criteria
developed
using
appropriate
procedures.
That
policy
continues
for
the
options
met
forth
in
the
interim
guidance
transmitted
today,
regardless
of
whether
the
resulting
criterion
is
equal
to
or
more
or
less
stringent
than
the
EPA
national
304(
a)
guidance.
This
interim
guidance
supersedes
all
guidance
concerning
water­
effect
ratios
previously
issued
by
the
Agency.
3
Each
of
the
three
options
for
deriving
a
final
water­
effect
ratio
presented
in
this
interim
guidance
meets
the
l
ciedific
and
technical
acceptability
test
for
deriving
mite­
specific
criteria.

Option
3
is
the
simplest,
least
restrictive
and
generally
thm
least
expensive
approach
for
situations
where
l
imulatmd
downstream
water
appropriately
represents
a
"
site."
It
is
a
fully
acceptable
approach
for
deriving
the
water­°
ffect
ratio
although
it
will
generally
provide
a
lower
water­
effect
ratio
than
the
other
2
optionm.
The
other
2
option8
may
be
more
costly
and
time
consuming
if
more
than
3
sample
period8
and
water­
effect
ratio
measurements
are
made,
but
are
more
accurate,
and
ray
yield
a
larger,
but
more
scientifically
defensible
site
l
pmcific
criterion.

Site­
specific
criteria,
propmrly
determined,
will
fully
protect
existing
uses.
The
waterbody
or
segment
thereof
to
which
the
site­
specific
criteria
apply
must
be
clearly
defined.
A
mite
can
be
defined
by
the
State
and
can
be
any
mite,
nail
or
large,
including
a
watershod
or
basin.
However,
the
site­
8pocific
criteria
must
protect
the
mite
am
a
whole.
It
is
likely
to
be
more
cost­
effective
to
derive
any
mite­
specific
criteria
for
am
large
an
area
am
pommible
or
appropriate.
Xt
is
whamited
thmt
mite­
specific
criteria
are
ambient
water
quality
criteria
applicable
to
a
site.
They
mre
not
intended
to
be
direct
modifications
to
National
Pollutant
Discharge
Elimination
Systera
(
NPDES)
penit
liritm.
In
romt
cm888
the
naitea
will
be
synonymous
with
a
State's
standards.
"
megm8nt"
in
its
watu
quality
By
defining
l
itmm
on
a
larger
l
cmle,
titiple
dischargers
can
collaborate
on
water­
effect
ratio
tuting
and
attain
appropriate
mite­
specific
crituia
at
a
rmduced
cost.

More
attention
ham
beon
given
to
water­
effect
ratios
recently
because
of
the
numerous
discussions
and
mmetingm
on
the
entire
question
of
metals
policy
and
because
WERm
were
specifically
applied
in
the
National
Toxic8
Rule.
In
comments
on
the
proposed
National
Toxic8
Rule,
the
public
gum8tfonmd
whether
the
EPA
promulgation
l
hould
be
based
solely
on
thm
total
recoverable
form
of
a
metal.
For
the
reasons
set
forth
in
the
final
preamble,
EPA
chose
to
promulgate
the
crit8ria
bm8ed
on
the
total
recwerabl8
form
with
a
provi8ion
for
the
application
of
a
water­
effect
ratio.
In
addition,
this
approach
wmm
chomen
because
of
the
unigue
difficulties
of
attempting
to
authoritm
site­
specific
criteria
modifications
for
nationally
prorulgatmd
criteria.

EPA
now
recommends
the
use
of
dissolved
metal8
for
Statms
revising
their
water
quality
standards.
Dimsolvmd
crit8ria
may
also
be
modified
by
a
site­
specific
adjustment.
4
While
the
regulatory
application
of
the
water­
effect
ratio
applied
oily
to
the
10
juri8dictio~
included
in
the
final
National
Toxicm
Rule:`
for
aquatic
life
retmls
criteria,
we
underlltood
thmt
othu.
Statu
would
F
interested
in
applying
m
to
their
l
dopted
water
quality
stmndard8.
The
guidance
upon
which
to
bm8e
the
judg#
nt
of
the
acceptability
of'the
water­
effect
ratio
applied
by
the
State
i8
contained
in
the
attached
provide8
l
dditionh
information
on
the
recmlculation
for
site­
8pecific
crituia
radificaeioru.
procmdurm
A
central
qu88tion
concerning
UERm
i8
whether
their
use
by
a
Stat8
reeult8
in
8
mite­
8pecific
criterion
8ubject
to
EPA
review
and
approval
under
Section
303(
c)
of
the
Clemn
Water
Act?

Derivation
of
8
vmter­
effect
ratio
by
a
State
is
a
8ite­
specific
criterion
adjwtmnt
l
ubjwt
to
EPA
review
and
approval/
dbapproval
under
Section
303(
c).
There
are
two
optionm
by
Which
thi8
review
Can
be
l
Ccarpli8hed.

Option
1:
A
State
my
derive
and
8­
t
each
individual
watu­
effect
ratio
dewination
to
EPA
for
review
and
8pprovml.
Thi8
would
be
accoBpli8hed
through
the
normal
review
and
revision
process
umed
by
a
State.

Option
2:
A
State
can
Bnd
its
water
quality
standards'to
provide
a
formal
prowdura
which
include8
derivation
of
water­
effect
ratios,
appropriate
definition
of
situ,
and
enforceable
monitoring
provisions
to
amsure
that
demignatti
usei
l
ra
protected.
Both
thi8
ProCedure
and
aa
reSUltin
criteria
would
be
8ubject
to
full
public
participation
requir­
nts.
Public
review
of
a
mite­
8pecific
criterion
could
be
accarplimhed
in
conjunction
with
the
public
review
required
for
permit
i8suance.
EPA
vould
review
and
approve/
disapprove
this
protocol
am
a
rmvimed
8tandard
once.
For
public
information,
we
recommend
thmt
once
a
year
the
State
publi8h
a
list
of
sit­
specific
criteria.

An
exception
to
thi8
policy
appliem
to
the
waterm
of
the
jurisdiction8
included
in
the
National
Toxic8
Rule.
The
EPA
m
is
not
required
for
the
jurisdiction8
included
in
the
National
Toxic8
Rule
where
EPA
established
the
procedure
for
the
State
for
applicmtion
to
the.
criteria
promulgated.
The
National
Toxic%
Rule
wmm
m
forpal
rulamking
process
with
notice
and
cment
by
which
EPA
pre­
authorized
the
use
of
a
correctly
applied
water­
effect
ratio.
That
8­
m
process
has
not
yet
taken
place
in
States
not
included
in
the
National
Toxic8
Rule.
5
However,
the
National
Toxics
Rule
does
not
affect
State
authority
to
establish
scientifically
defensible
procedures
to
determine
Federally
authorized
WERm,
to
certify
those
WERm
in
NPDES
permit
proceedings,
or
to
deny
their
application
based
on
the
State's
risk
management
analysis.

As.
demcribed
in
Section
131.36(
b)(
iii)
of
the
water
quality
standards
regulation
(
the
official
regulatory
reference
to
the
National
Toxics
Rule),
the
water­
effect
ratio
is
a
site­
specific
calculation.
As
indicated
on
page
60866
of
the
preamble
to
the
National
Toxic8
Rule,
the
rule
was
constructed
a8
a
raWtable
presumption.
The
water­
affect
ratio
is
assigned
a
value
of
1.0
until
a
different
water­
effect
ratio
is
derived
from
l
uitable
tests
representative
of
conditions
in
the
affected
waterbody.
It
is
the
responsibility
of
the
State
to
determine
whether
to
rebut
the
assumed
value
of
1.0
in
the
National
Toxic%
Rule
and
apply
another
value
of
the
water­
effect
ratio
in
order
to
establish
a
site­
8pecific
criterion.
The
site­
specific
criterion
is
then
used
to
develop
appropriate
NPDES
permit
liaitm.
The
rule
thus
provides
a
State
with
the
flexibility
to
derive
an
appropriate
site­
specific
criterion
for
specific
waterbodiu.

As
a
point
of
emphamis,
although
a
water­
effect
ratio
affects
permit
ltiitm
for
individual
dischargers,
it
is
the
State
in
all
cases
that
determines
if
derivation
of
a
site­
specific
criterion
based
on
the
water­
effect
ratio
is
allowed
and
it
is
the
State
that
ensures
that
the
calculations
and
data
analysis
are
done
completely
and
correctly.

This
interim
guidance
explains
and
clarifies
the
use
of
site­
specific
criteria.
It
is
issued
as
interim
guidance
because
it
will
be
included
am
part
of
the
process
unduwaf
for
review
and
pOSSibl8
revision
of
the
national
aquatic
life
criteria
development
methodology
guidelines.
Am
part
of
that
review,
this
interim
guidance
is
subject
to
amendment
based
on
comments,
especially
those
from
the
users
of
the
guidance.
At
the
end
of
the
guidelines
revision
process
the
guidance
will
be
issued
am
Wfinal.
W
EPA
is
interested
in
and
encourages
the
submittal
of
high
quality
datamets
that
can
be
used
to
provide
inmights
into
the
use
of
these
guidelines
and
procedures.
Such
data
and
technical
comments
should
be
submitted
to
Charles
E.
Stephan
at
EPA's
Environmental
Research
Laboratory
at
Duluth,
MN.
A
complete
address,
telephone
number
and
fax
number
for
Mr.
Stephan
are
included
in
the
guidance
itself.
Other
questions
or
comments
should
be
directed
to
the
Standards
and
Applied
Science
Division
(
mail
code
4305,
telephone
202­
260­
1315).
6
mara
is
attached
to
this
memorandum
a
simplified
fAow
diagram
and
an
inplementation
procedure.
Thaw
arm
intmded
to
aid
a
user
by
placing
the
water­
affect
ratio
procadura
in
tha
contmct
of
procmding
from
at
8it8­
8pocific
criterion
to
a
permit
limit.
Following
those
attachment8
is
tha
guidance
itself.

cc:
Robert
Fuciawp8,
Ow
l4aehaG.
Prothro,
o(
I
William
Diamond,
SASD
Hargarat
Stasikowski,
HECD
Hike
Cook,
OWEC
Cynthia
Dougherty,
OWE
Lw
Schroer,
OCC
Swan
kpow,
OGC
Courtney
Riordan,
ORD
ORD
(
Duluth
and
Narragansett
Laboratories)
ESD
Diroctorm,
Region8
I
­
VIII,
X
ESDBranch,
Region
IX
Wat8r
Quality
Standards
Coordinator8,
­
ion8
I
­
x
WER
Implementation
Site
Definition
Study
Plan
Development
Sampling
Design
Effluent
Considerations
Receiving
Water
Considerations
Lab
Procedures
Testing
Organisms
Site
Specific
Criteria
Permit
Limits
Monitoring
Requirements
WATER­
EFFECT
RATIO
IMPLEMENTATION
PRELIMINARY
ANALYSIS
41
PLAN
FORMULATION
­
Site
definition
l
How
my
discharges
must
be
accounted
for?
Tributuies?

sea
page
17.

l
What
is
tha
watarbody
typa?
(
i.
e.,
stream,
tidal
rivar,
bay,
etc.).
Sea
page
44
and
Appendix
A.

l
How
can
tha8e
considerations
best
ba
combined
to
defina
the
ralavant
geographic
%
ite"?
Sea
Appendix
A
e
page
02.

­
Plan
Davalopmant
for
Regulatory
Agency
Review
l
Is
wm
wthod
1
or
2
appropriate?
(
e.
g.,
Is
daoign
flaw
a
meaningful
concept
or
ara
other
considerations
paramount?).
Saa
page
6.

l
Define
the
effluent
&
racaiving
water
666pla
location8
l
Dascribe
tha
temporal
8ampla
collection
protocol8
propmad.
See
page
48.

l
Cm
simulated
8ita
watar
procadura
ba
done,
or
is
down6treamr
mampling
required?
See
Appendix
A.

l
De6criba
the
testing
protocol8
­
test
opecias,
te8t
type,
ta8t
length,
etc.
Sea
page
45,
50;
Appendix
I.

l
Describa
the
chemical
testing
proposed.
See
Appendix
C.

l
Describa
other
details
of
8tudy
­
flow
measuramant,
QA/
QC,
numbar
of
8ampling
pariods
proposed,
to
whom
the
results
are
expactad
to
apply,
schedule,
etc.

SANPXZNG
DESIGN
PORSTREAMS
­
Discuss
the
quantification
of
the
design
streamflow
(
a­
g.,
7410)
­
USGS
gage
directly,
by
extrapolation
from
USGS
gage,
or
3
­
Effluents
l
measure
flows
to
datarmina
average
for
sampling
day
l
collect
24
hour
composite
using
%
leanW
equipment
and
appropriata
procadura6;
avoid
the
use
of
the
plant's
daily
composite
sample
as
a
ohortcut.

­
Streams
l
measure
flow
(
use
currant
mater
or
read
from
gage
if
available)
to
determine
dilution
with
affluent:
and
to
check
if
within
accaptabla
range
for
use
of
the
data
(
i.
e.,
design
flow
to
10
times
the
design
flow).

l
collect
24
hour
composite
of
upstream
water.
LABORATORY
PROCEDURES
(
NOTE:
mesa
are
described
in
detail
in
interim
guidance).

­
Select
appropriate
primmy
C
secondary
tests
­
Determine
appropriate
cmcWER
and/
or
cccWER
­
tifOXBChUiS~
U8ingCl.­
p~­,
WithMthod8
that
have
adaquata
SaMitivity
to
maasura
low
concbntration8,
and
u8a
appropriate
QA/
QC
­
Calculate
final
water­
affect
ratio
(
FWER)
for
sita.
Sea
page
36.

IMPLEHENTATION
­
Assign
PWER8
and
the
site
8pecific
criteria
for
oath
metal
to
each
di8ChUgU
(
if
more
than
0~).

­
parform
a
waste
load
allocation
and
total
maximm
daily
load
(
if
appropriate)
80
that
math
discharger
is
provided
l
pumitlimit.

­
l
8tabli8h
monitoring
condition
for
periodic
evaluation
of
irmtroam
biology
(
raconandad)

­
l
8tabli8h
a
permit
condition
for
pariodic
tasting
of
WER
to
verify
site­
spacific
criterion
(
NTR
recommendation)

2
United
States
Office
of
Water
February
1994
Environmental
Protection
Office
of
Science
&
Technology
EPA­
823­
B­
94­
001
Agency
(
Mail
Code
4305)

EPA
Interim
Guidance
on
Determination
and
Use
of
Water­
Effect
Ratios
for
Metals
Interim
Guidance
on
Determination
and
Use
of
Water­
Effect
Ratios
for
Metals
February
1994
U.
S.
Environmental
Protection
Agency
Office
of
Water
Office
of
Science
and
Technology
Washington,
D.
C.

Office
of
Research
and
Development
Environmental
Research
Laboratories
Duluth,
Minnesota
Narragansett,
Rhode
Island
NOTICES
This
document
has
been
reviewed
by
the
Environmental
Research
Laboratories,
Duluth,
MN
and
Narragansett,
RX
(
Office
of
Research
and
Development)
and
the
Office
of
Science
and
Technology
(
Office
of
Water),
U.
S.
Environmental
Protection
Agency,
and
approved
for
publication.

Mention
of
trade
names
or
commercial
products
does
not
constitute
endorsement
or
recommendation
for
use.

ii
FOREWORD
This
document
provides
interim
guidance
concerning
the
experimental
determination
of
water­
effect
ratios
(
WERS)
for
metals;
some
aspects
of
the
use
of
WERs
are
also
addressed.
It
is
issued
in
support
of
EPA
regulations
and
policy
initiatives
involving
the
application
of
water
quality
criteria
and
standards
for
metals.
This
document
is
agency
guidance
only.
It
does
not
establish
or
affect
legal
rights
or
obligations.
It
does
not
establish
a
binding
norm
or
prohibit
alternatives
not
included
in
the
document.
It
is
not
finally
determinative
of
the
issues
addressed.
Agency
decisions
in
any
particular
case
will
be
made
by
applying
the
law
and
regulations
on
the
basis
of
specific
facts
when
regulations
are
promulgated
or
permits
are
issued.

This
do
comment
is
expected
to
be
revised
periodically
to
reflect
advances
in
this
rapidly
evolving
area.
Comments,
especially
those
accompanied
by
supporting
data,
are
welcomed
and
should
be
sent
to:
Charles
E.
Stephan,
U.
S.
EPA,
6201
Congdon
boulevard,
Duluth
MN
55804
(
TEL:
218­
720­
5510;
FAX:
218­
720­
5539).

iii
UNITED
STATES
ENVIRONMENTAL
PROTECTION
AGENCY
WASHINGTON,
D.
C.
20460
FEB
22
1994
OFFICE
OF
WATER
OFFICE
OF
SCIENCE
AND
TECHNOLOGY
POSITION
STATEMENT
Section
131.11(
b)(
ii)
of
the
water
quality
standards
regulation
(
40
CFR
Part
131)
provides
the
regulatory
mechanism
for
a
State
to
develop
site­
specific
criteria
for
use
in
water
quality
standards
.
Adopting
site­
specific
criteria
in
water
quality
standards
is
a
State
option­
not
a
requirement.
The
Environmental
Protection
Agency
(
EPA)
in
1983
provided
guidance
on
scientifically
acceptable
method
by
which
site­
specific
criteria
could
be
developed.

The
interim
guidance
provided
in
this
document
supersedes
all
guidance
concerning
water­
effect
ratios
and
the
Indicator
Species
Procedure
given
in
Chapter
4
of
the
Water
Quality
Standards
Handbook
issued
by
EPA
in
1983
and
in
Guidelines
for
Deriving
Numerical
Aquatic
Site­
Specific
Water
Quality
Criteria
by
Modifying
National
Criteria,
1984.
Appendix
B
also
supersedes
the
guidance
in
these
earlier
documents
for
the
Recalculation
Procedure
for
performing
site­
specific
criteria
modifications.

This
interim
guidance
fulfills
a
commitment
made
in
the
final
rule
to
establish
numeric
criteria
for
priority
toxic
pollutants
(
57
FR
60848,
December
22,
1992,
also
known
as
the
"
National
Toxics
Rule").
This
guidance
also
is
applicable
to
pollutants
other
than
metals
with
appropriate
modifications,
principally
to
chemical
analyses.

Except
for
the
jurisdictions
subject
to
the
aquatic
life
criteria
in
the
national
toxics
rule,
water­
affect
ratios
are
site­
specific
criteria
subject
to
review
and
approval
by
the
appropriate
EPA
Regional
Administrator.
Site­
specific
criteria
are
new
or
revised
criteria
subject
to
the
normal
EPA
review
requirements
established
in
Clean
Water
Act
§
303(
c).
For
the
States
in
the
National
Toxics
Rule,
EPA
has
established
that
site­
specific
water­
affect
ratios
may
be
applied
to
the
criteria
promulgated
in
the
rule
to
establish
Site­
specific
criteria.
The
water­
affect
ratio
portion
of
these
criteria
would
still
be
subject
to
State
review
before
the
development
of
total
maximum
daily
loads,
waste
load
allocations
or
translation
into
NPDES
permit
limits.
EPA
would
only
review
these
water­
affect
ratios
during
its
oversight
review
of
these
State
programs
or
review
of
State­
issued
permits.

IV
Each
of
the
three
options
for
deriving
a
final
water­
effect
ratio
presented
on
page
36
of
this
interim
guidance
meets
the
scientific
and
technical
acceptability
test
for
deriving
site­
specific
criteria
specified
in
the
water
quality
standards
regulation
(
40
CFR
131.11(
a)).
Option
3
is
the
simplest,
least
restrictive
and
generally
the
least
expensive
approach
for
situations
where
simulated
downstream
water
appropriately
represents
a
"
site."
Option
3
requires
experimental
determination
of
three
water­
effect
ratios
with
the
primary
test
species
that
are
determined
during
any
season
(
as
long
as
the
downstream
flow
is
between
2
and
10
times
design
flow
conditions.)
The
final
WER
is
generally
(
but
not
always)
the
lowest
experimentally
determined
WER.
Deriving
a
final
water­
effect
ratio
wing
option
3
with
the
use
of
simulated
downstream
water
for
a
situation
where
this
simulation
appropriately
represents
a
"
site",
is
a
fully
acceptable
approach
for
deriving
a
water­
effect
ratio
for
use
in
determining
a
site­
specific
criterion,
although
it
will
generally
provide
a
lower
water­
effect
ratio
than
the
other
2
options.

As
indicated
in
the
introduction
to
this
guidance,
the
determination
of
a
water­
effect
ratio
may
require
substantial
resources.
A
discharger
should
consider
cost­
effective,
preliminary
measures
described
in
this
guidance
(
e.
g.,
use
of
"
clean"
sampling
and
chemical
analytical
techniques
or
in
non­
NTR
States,
a
recalculated
criterion)
to
determine
if
an
indicator
species
site­
specific
criterion
is
really
needed.
It
may
be
that
an
appropriate
site­
specific
criterion
is
actually
being
attained.
In
many
instances,
use
of
these
other
measures
may
eliminate
the
need
for
deriving
final
water­
effect
ratios.
The
methods
described
in
this
interim
guidance
should
be
sufficient
to
develop
site­
specific
criteria
that
resolve
concerns
of
dischargers
when
there
appears
to
be
no
instream
toxicity
from
a
metal
but,
where
(
a)
a
discharge
appears
to
exceed
existing
or
proposed
water
quality­
based
permit
limits,
or
(
b)
an
instream
concentration
appears
to
exceed
an
existing
or
proposed
water
quality
criterion.

This
guidance
describes
different
methods
for
determining
water­
effect
ratios.
Method
1
has
3
options
each
of
which
may
only
require
3
sampling
periods.
However
options
1
and
2
may
be
expanded
and
require
a
much
greater
effort.
While
this
position
statement
has
discussed
the
simplest,
least
expensive
option
for
method
1
(
the
single
discharge
to
a
stream)
to
illustrate
that
site
specific
criteria
are
feasible
even
when
only
small
dischargers
are
affected,
water­
effect
ratios
may
be
calculated
using
any
of
the
other
options
described
in
the
guidance
if
the
State/
discharger
believe
that
there
is
reason
to
expect
that
a
more
accurate
site­
specific
criterion
will
result
from
the
increased
cost
and
complexity
inherent
in
conducting
the
V
additional
to&
s
and
analyzing
the
results.
Situation8
where
this
could
be
the
ca80
include,
for
rxample,
whore
8ea8onal
l
ffect8
in
receiving
water
quality
or
in
di8Chargm
quality
need
t0
be
a88e88ed.

In
addition,
EPA
will
cormider
Other
8cientifiWlly
dUfUn8ibla
approach88
in
developing
final
Water­°
ffUt
ratio8
a8
authorized
in
40
CFR
131.11.
Howevu,
EPA
8trongly
recomund8
that
beforo
a
State/
di8Charger
impleDent8
aXby
approach
othu
than
one
described
in
thi8
interim
guidance,
di8cu88ion8
be
held
with
appropriate
EPA
regional
offices
and
Office
of
Rm8earch
and
Development'
8
8cienti8t8
befor
actual
testing
begins.
The8e
di8cu88ion8
would
be
to
enare
that
time
and
ruource8
ar8
not
wa8ted
on
8cientifiWlly
and
technically
unaocoptable
approachu.
It
remain8
tpA's
re8ponsibility
to
Bake
final
deCi8ion8
on
tha
8cientific
and
technical
validity
of
altenativo
approaches
to
dovaloping
8itW8pOCifiC
WatU
@
ldllity
UitUia.

EPA
i8
fully
cognizant
of
the
continuing
d&
ate
bew88n
vhat
coxmtitutu
guidance
and
what
i8
a
regulatory
reguirnmt.
Developing
8itw8puific
critrria
i8
a
Stat8
regulatory
option.
U8ing
the
methodology
correctly
a8
de8cribed
in
thi8
guidance
a88Urm8
the
State
that
EPA
will
accept
the
re8ult.
Othu
approaches
are
possible
and
logically
8hould
be
discussed
with
EPA
prior
to
implmaentation.

The
Offiw
of
Science
and
Technology
beliovu
that
thi8
interim
guidance
advance8
the
8cionw
of
determining
8ite­
8pMifiC
Criteria
and
prOVidm8
policy
gui&
nW
that
Stat88
and
EPA
Wn
u8e
in
thi8
complex
aru.
It
refloct8
the
8cientific
adVanCa8
in
the
past
10
yoar8
and
the
mcperienc8
gained
from
dealing
with
these
i88ue8
in
rul
world
situatiom.
Thi8
guidance
will
help
improve
implementation
of
water
quality
8tandard8
and
be
the
basis
for
future
progreW.

Office
of
Science
And
Technology
Office
of
Water
Vi
CONTENTS
Notices......................

Foreword
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.

Office
of
Science
and
Technology
Position
Statement
Appendices
.
.
.
.

Figures
.
.
.
.
.
.

Acknowledgments
.
.

Executive
Summary
.

Abbreviations
.
.
.

Glossary
.
.
.
.
.

Preface
.
.
.
.
.
.

Introduction
.
.
.

Method
1
.
.
.
.
.
A.
B.
C.
D.
E.
F.
G.
H.
I.
J.
Experimental
Design
.
Background
Information
and
Initial
Decisions
Selecting
Primary
and
Secondary
Tests
.
.
.
.
Acquiring
and
Acclimating
Test
Organisms
.
.
Collecting
and
Handling
Upstream
Water
and
Effluent
.
Laboratory
Dilution
Water­.
.
.
.
.
.
.
.
Conducting
Tests
Chemical
and
Other
Measurements
.
.
.
.
Calculating
and
Interpreting
the
Results
Reporting
the
Results
.
.
.
.
.
.
.
.
.
.
.
.

.
.

.
.

.
.

.
.

.
.

.
.

.
.

.
.
.
.

.
.

.
.

.
.

.
.

.
.

.
.

.
.

.
.
.
.

.
.

.
.

.
.

.
.

.
.

.
.

.
.

.
.
.
.
.
.

.
.

.
.

.
.

.
.

.
.

.
.

.
.

.
.
.
.
.
.

.
.

.
.

.
.

.
.

.
.

.
.

.
.

.
.
.
.
......

......

......

......

......

......

......

.
.
.
.
.
.

......
......
.
.

.
.

.
.

.
.

.
.

.
.

.
.

.
.

.
.

.
.

.
.

.
.
.
.
.
.
.
.
.
.
.
.

.
.

.
.

.
.

.
.

.
.

.
.

.
.

.
.

.
.

.
.

.
.
.
.
.
.
.
.
.
.

Method
2
...................

References
..................
.
.
.
.
.
.
.
.
.
.

.
.

.
.
.
.
.
.
.
.
.
.
.
.

.
.

.
.
.
.
.
.
.
.
.
.
.
.

.
.

.
.
Page
.
ii
.
iii
.
iv
viii
.
ix
.
.
x
.
xi
xiii
.

.

.

.
.
.
.
.
.
.
.
.
.
.

.

.
xiv
xvi
.
1
17
17
44
45
47
48
49
50
55
57
62
65
76
vii
APPENDICES
Page
A.

B.

C.

D.

E.

F.

G.

H.

I.

J.
Comparison
of
WERs
Determined
Using
Upstream
and
Downstream
Water
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
79
The
Recalculation
Procedure
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
90
Guidance
Concerning
the
Use
of
"
Clean
Techniques"
and
QA/
QC
when
Measuring
Trace
Metals
.
.
.
.
.
.
.
.
.
.
.
.
98
Relationships
between
WERs
and
the
Chemistry
and
Toxicology
of
Metals
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
109
U.
S.
EPA
Aquatic
Life
Criteria
Documents
for
Metals
.
.
.
134
Considerations
Concerning
Multiple­
Metal,
Multiple­
Discharge,
and
Special
Flowing­
Water
Situations
.
.
.
.
.
135
Additivity
and
the
Two
Components
of
a
WER
Determined
Using
Downstream
Water
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
139
Special
Considerations
Concerning
the
Determination
of
WERs
with
Saltwater
Species
.
.
.
.
.
.
.
.
.
.
.
.
.
145
Suggested
Toxicity
Tests
for
Determining
WERs
for
Metals
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
147
Recommended
Salts
of
Metals
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
153
viii
FIGURES
Page
1.

2.

3.

4.

5.

6.

B1.

D1.

D2.

D3.

D4.

D5.

D6.
Four
Ways
to
Derive
a
Permit
Limit
.
.
.
.
.
.
.
.
.
.
.
16
Calculating
an
Adjusted
Geometric
Mean
.
.
.
.
.
.
.
.
.
71
An
Example
Derivation
of
a
FWER
.
.
.
.
.
.
.
.
.
.
.
.
.
72
Reducing
the
Impact
of
Experimental
Variation
.
.
.
.
.
.
73
Calculating
an
LC50
(
or
EC50)
by
Interpolation
.
.
.
.
.
74
Calculating
a
Time­
Weighted
Average
.
.
.
.
.
.
.
.
.
.
.
75
An
Example
of
the
Deletion
Process
Using
Three
Phyla
.
.
97
A
Scheme
for
Classifying
Forms
of
Metal
in
Water
.
.
.
.
111
An
Example
of
the
Empirical
Extrapolation
Process
.
.
.
.
125
The
Internal
Consistency
of
the
Two
Approaches
.
.
.
.
.
126
The
Application
of
the
Two
Approaches
.
.
.
.
.
.
.
.
.
.
128
A
Generalized
Complexation
Curve
.
.
.
.
.
.
.
.
.
.
.
.
131
A
Generalized
Precipitation
Curve
.
.
.
.
.
.
.
.
.
.
.
.
132
ix
ACKNOWLEDGEMENTS
This
document
was
written
by:

Charles
E.
Stephan,
U.
S.
EPA,
ORD,
Environmental
Research
Laboratory,
Duluth,
MN.

William
H.
Peltier,
U.
S.
EPA,
Region
IV,
Environmental
Services
Division,
Athens,
GA.

David
J.
Hansen,
U.
S.
EPA,
ORD,
Environmental
Research
Laboratory,
Narragansett,
RI.

Charles
G.
Delos,
U.
S.
EPA,
Office
of
Water,
Health
and
Ecological
Criteria
Division,
Washington,
DC.

Guy
A.
Chapman,
U.
S.
EPA,
ORD,
Environmental
Research
Laboratory
(
Narragansett)
,
Pacific
Ecosystems
Branch,
Newport,
OR.

The
authors
thank
all
the
people
who
participated
in
the
open
discussion
of
the
experimental
determination
of
water­
effect
ratios
on
Tuesday
evening,
January
26,
1993
in
Annapolis,
MD.
Special
thanks
go
to
Herb
Allen,
Bill
Beckwith,
Ken
Bruland,
Lee
Dunbar,
Russ
Erickson,
and
Carlton
Hunt
for
their
technical
input
on
this
project,
although
none
of
them
necessarily
agree
with
everything
in
this
document.
Comments
by
Kent
Ballantine,
Karen
Gourdine,
Mark
Hicks,
Suzanne
Lussier,
Nelson
Thomas,
Bob
Spehar,
Fritz
Wagener,
Robb
Wood,
and
Phil
Woods
on
various
drafts,
or
portions
of
drafts,
were
also
very
helpful,
as
were
discussions
with
several
other
individuals.

x
EXECUTIVE
SUMMARY
A
variety
of
physical
and
chemical
characteristics
of
both
the
water
and
the
metal
can
influence
the
toxicity
of
a
metal
to
aquatic
organisms
in
a
surface
water.
When
a
site­
specific
aquatic
life
criterion
is
derived
for
a
metal,
an
adjustment
procedure
based
on
the
toxicological
determination
of
a
water­
effect
ratio
(
WER)
may
be
used
to
account
for
a
difference
between
the
toxicity
of
the
metal
in
laboratory
dilution
water
and
its
toxicity
in
the
water
at
the
site.
If
there
is
a
difference
in
toxicity
and
it
is
not
taken
into
account,
the
aquatic
life
criterion
for
the
body
of
water
will
be
more
or
less
protective
than
intended
by
EPA's
Guidelines
for
Deriving
Numerical
National
Water
Quality
Criteria
for
the
Protection
of
Aquatic
Organisms
and
Their
Uses.
After
a
WER
is
determined
for
a
site,
a
site­
specific
aquatic
life
criterion
can
be
calculated
by
multiplying
an
appropriate
national,
state,
or
recalculated
criterion
by
the
WER.
Most
WERs
are
expected
to
be
equal
to
or
greater
than
1.0,
but
some
might
be
less
than
1.0.
because
most
aquatic
life
criteria
consist
of
two
numbers,
i.
e.,
a
Criterion
maximum
Concentration
(
CMC)
and
a
Criterion
Continuous
Concentration
(
CCC),
either
a
cmcWER
or
a
cccWER
or
both
might
be
needed
for
a
site.
The
cmcWER
and
the
cccWER
cannot
be
assumed
to
be
equal,
but
it
is
not
always
necessary
to
determine
both.

In
order
to
determine
a
WER,
side­
by­
side
toxicity
tests
are
performed
to
measure
the
toxicity
of
the
metal
in
two
dilution
waters.
One
of
the
waters
has
to
be
a
water
that
would
be
acceptable
for
use
in
laboratory
toxicity
tests
conducted
for
the
derivation
of
national
water
quality
criteria
for
aquatic
life.
In
most
situations,
the
second
dilution
water
will
be
a
simulated
downstream
water
that
is
prepared
by
mixing
upstream
water
and
effluent
in
an
appropriate
ratio;
in
other
situations,
the
second
dilution
water
will
be
a
sample
of
the
actual
site
water
to
which
the
site­
specific
criterion
is
to
apply.
The
WER
is
calculated
by
dividing
the
endpoint
obtained
in
the
site
water
by
the
endpoint
obtained
in
the
laboratory
dilution
water.
A
WER
should
be
determined
using
a
toxicity
test
whose
endpoint
is
close
to,
but
not
lower
than,
the
CMC
and/
or
CCC
that
is
to
be
adjusted.

A
total
recoverable
WER
can
be
determined
if
the
metal
in
both
of
the
side­
by­
side
toxicity
tests
is
analyzed
using
the
total
recoverable
measurement,
and
a
dissolved
WER
can
be
determined
if
the
metal
is
analyzed
in
both
tests
using
the
dissolved
measurement.
Thus
four
WERs
can
be
determined:
Total
recoverable
cmcWER.
Total
recoverable
cccWER.
Dissolved
cmcWER.
Dissolved
cccWER.
A
total
recoverable
WER
is
used
to
calculate
a
total
recoverable
site­
specific
criterion
from
a
total
recoverable
national,
state,

xi
or
recalculated
aquatic
life
criterion,
whereas
a
dissolved
WER
is
used
to
calculate
a
dissolved
site­
specific
criterion
from
a
dissolved
criterion.
WERs
are
determined
individually
for
each
metal
at
each
site;
WERs
cannot
be
extrapolated
from
one
metal
to
another,
one
effluent
to
another,
or
one
site
water
to
another.

Because
determining
a
WER
requires
substantial
resources,
the
desirability
of
obtaining
a
WER
should
be
carefully
evaluated:
1.
Determine
whether
use
of
"
clean
techniques"
for
collecting,
handling,
storing,
preparing,
and
analyzing
samples
will
eliminate
the
reason
for
considering
determination
of
a
WER,
because
existing
data
concerning
concentrations
of
metals
in
effluents
and
surface
waters
might
be
erroneously
high.
2.
Evaluate
the
potential
for
reducing
the
discharge
of
the
metal.
3.
Investigate
possible
constraints
on
the
permit
limits,
such
as
antibacksliding
and
antidegradation
requirements
and
human
health
and
wildlife
criteria.
4.
Consider
use
of
the
Recalculation
Procedure.
5.
Evaluate
the
cost­
effectiveness
of
determining
a
WER.
If
the
determination
of
a
WER
is
desirable,
a
detailed
workplan
for
should
be
submitted
to
the
appropriate
regulatory
authority
(
and
possibly
to
the
Water
Management
Division
of
the
EPA
Regional
Office)
for
comment.
After
the
workplan
is
completed,
the
initial
phase
should
be
implemented,
the
data
should
be
evaluated,
and
the
workplan
should
be
revised
if
appropriate.

Two
methods
are
wed
to
determine
WERs.
Method
1,
which
is
used
to
determine
cccWERs
that
apply
near
plumes
and
to
determine
all
cmcWRRs,
uses
data
concerning
three
or
shore
distinctly
separate
sampling
events.
It
is
best
if
the
sampling
events
occur
during
both
low­
flow
and
higher­
flow
periods.
When
sampling
does
not
occur
during
both
low
and
higher
flows,
the
site­
specific
criterion
is
derived
in
a
more
conservative
manner
due
to
greater
uncertainty.
For
each
sampling
event,
a
WER
is
determined
using
a
selected
toxicity
test;
for
at
least
one
of
the
sampling
events,
a
confirmatory
WER
is
determined
using
a
different
test.

Method
2,
which
is
used
to
determine
a
cccWER
for
a
large
body
of
water
outside
the
vicinities
of
plumes,
requires
substantial
site­
specific
planning
and
more
resources
than
Method
1.
WERs
are
determined
using
samples
of
actual
site
water
obtained
at
various
times,
locations,
and
depths
to
identify
the
range
of
WERs
in
the
body
of
water.
The
WERs
are
used
to
determine
how
many
site­
specific
CCCs
should
be
derived
for
the
body
of
water
and
what
the
one
or
more
CCCs
should
be.

The
guidance
contained
herein
replaces
previous
agency
guidance
concerning
(
a)
the
determination
of
WERs
for
use
in
the
derivation
of
site­
specific
aquatic
life
criteria
for
metals
and
(
b)
the
Recalculation
Procedure.
This
guidance
is
designed
to
apply
to
metals,
but
the
principles
apply
to
most
pollutants.

xii
ABBREVIATIONS
ACR:
Acute­
Chronic
Ratio
CCC
:
Criterion
Continuous
Concentration
CMC:
Criterion
Maximum
Concentration
CRM:
Certified
Reference
Material
FAV
:
Final
Acute
Value
FCV
:
Final
Chronic
Value
FW:
Freshwater
FWER:
Final
Water­
Effect
Ratio
GMAV:
Genus
Mean
Acute
Value
HCME:
Highest
Concentration
of
the
Metal
in
the
Effluent
MDR:
Minimum
Data
Requirement
NTR:
National
Toxics
Rule
QA/
QC:
Quality
Assurance/
Quality
Control
SMAV:
Species
Mean
Acute
Value
SW:
Saltwater
TDS:
Total
Dissolved
Solids
TIE:
Toxicity
Identification
Evaluation
TMDL:
Total
Maximum
Daily
Load
TOC:
Total
Organic
Carbon
TRE:
Toxicity
Reduction
Evaluation
TSD:
Technical
Support
Document
TSS:
Total
Suspended
Solids
WER:
Water­
Effect
Ratio
WET:
Whole
Effluent
Toxicity
WLA:
Wasteload
Allocation
xiii
GLOSSARY
Acute­
chronic
ratio
­
an
appropriate
measure
of
the
acute
toxicity
of
a
material
divided
by
an
appropriate
measure
of
the
chronic
toxicity
of
the
same
material
under
the
same
conditions.

Appropriate
regulatory
authority
­
Usually
the
State
water
pollution
control
agency,
even
for
States
under
the
National
Toxics
Rule;
if,
however,
a
State
were
to
waive
its
section
401
authority,
the
Water
Management
Division
of
the
EPA
Regional
Office
would
become
the
appropriate
regulatory
authority.

Clean
techniques
­
a
set
of
procedures
designed
to
prevent
contamination
of
samples
so
that
concentrations
of
trace
metals
can
be
measured
accurately
and
precisely.

Critical
species
­
a
species
that
is
commercially
or
recreationally
important
at
the
site,
a
species
that
exists
at
the
site
and
is
listed
as
threatened
or
endangered
under
section
4
of
the
Endangered
Species
Act,
or
a
species
for
which
there
is
evidence
that
the
loss
of
the
species
from
the
site
is
likely
to
cause
an
unacceptable
impact
on
a
commercially
or
recreationally
important
species,
a
threatened
or
endangered
species,
the
abundances
of
a
variety
of
other
species,
or
the
structure
or
function
of
the
community.

Design
flow
­
the
flow
used
for
steady­
state
wasteload
allocation
modeling.

Dissolved
metal
­
defined
here
as
"
metal
that
passes
through
either
a
0.45­
µ
m
or
a
0.40­
µ
m
membrane
filter'.

Endpoint
­
the
concentration
of
test
material
that
is
expected
to
cause
a
specified
amount
of
adverse
effect.

Final
Water­
Effect
Ratio
­
the
WER
that
is
used
in
the
calculation
of
a
site­
specific
aquatic
life
criterion.

Flow­
through
test
­
a
test
in
which
test
solutions
flow
into
the
test
chambers
either
intermittently
(
every
few
minutes)
or
continuously
and
the
excess
flows
out.

Labile
metal
­
metal
that
is
in
water
and
will
readily
convert
from
one
form
to
another
when
in
a
nonequilibrium
condition.

Particulate
metal
­
metal
that
is
measured
by
the
total
recoverable
method
but
not
by
the
dissolved
method.

xiv
Primary
test
­
the
toxicity
test
used
in
the
determination
of
a
Final
Water­
Effect
Ratio
(
FWER);
the
specification
of
the
test
includes
the
test
species,
the
life
stage
of
the
species,
the
duration
of
the
test,
and
the
adverse
effect
on
which
the
endpoint
is
based.

Refractory
metal
­
metal
that
is
in
water
and
will
not
readily
convert
from
one
form
to
another
when
in
a
nonequilibrium
condition,
i.
e.,
metal
that
is
in
water
and
is
not
labile.

Renewal
test
­
a
test
in
which
either
the
test
solution
in
a
test
chamber
is
renewed
at
least
once
during
the
test
or
the
test
organisms
are
transferred
into
a
new
test
solution
of
the
same
composition
at
least
once
during
the
test.

Secondary
test
­
a
toxicity
test
that
is
usually
conducted
along
with
the
primary
test
only
once
to
test
the
assumptions
that,
within
experimental
variation,
(
a)
similar
WERs
will
be
obtained
using
tests
that
have
similar
sensitivities
to
the
test
material,
and
(
b)
teats
that
are
less
sensitive
to
the
test
material
will
usually
give
WERs
that
are
closer
to
1.

Simulated
downstream
water
­
a
site
water
prepared
by
mixing
effluent
and
upstream
water
in
a
known
ratio.

Site­
specific
aquatic
life
criterion
­
a
water
quality
criterion
for
aquatic
life
that
has
been
derived
to
be
specifically
appropriate
to
the
water
quality
characteristics
and/
or
species
composition
at
a
particular
location.

Site
water
­
upstream
water,
actual
downstream
water,
or
simulated
downstream
water
in
which
a
toxicity
test
is
conducted
side­
by­
side
with
the
same
toxicity
test
in
a
laboratory
dilution
water
to
determine
a
WER.

Static
test
­
a
test
in
which
the
solution
and
organisms
that
are
in
a
test
chamber
at
the
beginning
of
the
test
remain
in
the
chamber
until
the
end
of
the
test.

Total
recoverable
metal
­
metal
that
is
in
aqueous
solution
after
the
sample
is
appropriately
acidified
and
digested
and
insoluble
material
is
separated.

Water­
effect
ratio
­
an
appropriate
measure
of
the
toxicity
of
a
material
obtained
in
a
site
water
divided
by
the
same
measure
of
the
toxicity
of
the
same
material
obtained
simultaneously
in
a
laboratory
dilution
water.

xv
PREFACE
Several
issues
need
consideration
when
guidance
such
as
this
is
written:

1.
Degrees
of
importance:
Procedures
and
methods
are
series
of
instructions,
than
others.
but
some
of
the
instructions
are
more
important
Some
instructions
are
so
important
that,
if
they
are
not
followed,
the
results
will
be
questionable
or
unacceptable;
other
instructions
are
less
important,
but
definitely
desirable.
Possibly
the
best
way
to
express
various
degrees
of
importance
is
the
approach
described
in
several
ASTM
Standards,
such
as
in
section
3.6
of
Standard
E729
(
ASTM
1993a),
which
is
modified
here
to
apply
to
WERs:
The
words
"
must",
"
should",
"
may",
"
can",
and
"
might"
have
specific
meanings
in
this
document.
"
Must"
is
used
to
express
an
instruction
that
is
to
be
followed,
unless
a
site­
specific
consideration
requires
a
deviation,
and
is
used
only
in
connection
with
instructions
that
directly
relate
to
the
validity
Of
toxicity
tests,
WERs,
FWERS,
and
the
Recalculation
Procedure.
"
Should"
is
used
to
state
instructions
that
are
recommended
and
are
to
be
followed
if
reasonably
possible.
Deviation
from
one
"
should"
will
not
invalidate
a
WER,
but
deviation
from
several
probably
will.
Terms
such
as
"
is
desirable",
"
is
often
desirable",
and
"
might
be
desirable"
are
used
in
connection
with
less
important
instructions.
"
May"
is
used
to
mean
"
is
(
are)
allowed
to",
"
can"
is
used
to
mean
"
is
(
are)
able
to",
and
"
might"
is
used
to
mean
"
could
possibly".
Thus
the
classic
distinction
between
"
may"
and
"
can"
is
preserved,
and
"
might"
is
not
used
as
a
synonym
for
either
"
may"
or
"
can"
This
does
not
eliminate
all
problems
concerning
the
degree
of
importance,
however.
For
example,
a
small
deviation
from
a
"
must"
might
not
invalidate
a
WER,
whereas
a
large
deviation
would.
(
Each
"
must"
and
"
must
not"
is
in
bold
print
for
convenience,
not
for
emphasis,
in
this
document.)

2.
Educational
and
explanatory
material:
Many
people
have
asked
for
much
detail
in
this
document
to
ensure
that
as
many
WERs
as
possible
are
determined
in
an
acceptable
manner.
In
addition,
some
people
want
justifications
for
each
detail.
Much
of
the
detail
that
is
desired
by
sow
people
is
based
on
"
best
professional
judgment",
which
is
rarely
considered
an
acceptable
justification
by
people
who
disagree
with
a
specified
detail.
Even
if
details
are
taken
from
an
EPA
method
or
an
ASTM
standard,
they
were
often
included
in
those
documents
on
the
basis
of
best
professional
judgment.
In
contrast,
some
people
want
detailed
methodology
presented
without
explanatory
material.
It
was
decided
to
include
as
much
detail
as
is
feasible,
and
to
provide
rationale
and
explanation
for
major
items.

xvi
3.

4.
Separation
of
"
science",
best
professional
judgment"
and
"
regulatory
decisions":
These
can
never
be
completely"

5.

6.
Alternatives:
When
more
than
one
alternative
is
both
scientifically
sound
and
appropriately
protective,
it
seems
reasonable
to
present
the
alternatives
rather
than
presenting
the
one
that
is
considered
best.
The
reader
CM
then
select
one
based
on
cost­
effectiveness,
personal
preference,
details
of
the
particular
situation,
and
perceived
advantages
and
disadvantages.

separated
in
this
kind
of
document;
for
example,
if
data
are
analyzed
for
a
statistically
significant
difference,
the
selection
of
alpha
is
an
important
decision,
but
a
rationale
for
its
selection
is
rarely
presented,
probably
because
the
selection
is
not
a
scientific
decision.
In
this
document,
an
attempt
has
been
made
to
focus
on
good
science,
best
professional
judgment,
and
presentation
of
the
rationale;
when
possible,
these
are
separated
from
"
regulatory
decisions"
concerning
margin
of
safety,
level
of
protection,
beneficial
use,
regulatory
convenience,
and
the
goal
of
zero
discharge.
Some
"
regulatory
decisions"
relating
to
implementation,
however,
should
be
integrated
with,
not
separated
from,
"
science"
because
the
two
ought
to
be
carefully
considered
together
wherever
science
has
implications
for
implementation.

Best
professional
judgement
:
Much
of
the
guidance
contained
herein
is
qualitative
rather
than
quantitative,
and
much
judgment
will
usually
be
required
to
derive
a
site­
specific
water
quality
criterion
for
aquatic
life.
In
addition,
although
this
version
of
the
guidance
for
determining
and
using
WERs
attempts
to
cover
all
major
questions
that
have
arisen
during
use
of
the
previous
version
and
during
preparation
of
this
version,
it
undoubtedly
does
not
cover
all
situations,
questions,
and
extenuating
circumstances
that
might
arise
in
the
future.
All
necessary
decisions
should
be
based
on
both
a
thorough
knowledge
of
aquatic
toxicology
and
an
understanding
of
this
guidance;
each
decision
should
be
consistent
with
the
spirit
of
this
guidance,
which
is
to
make
best
use
of
"
good
science"
to
derive
the
most
appropriate
site­
specific
criteria.
This
guidance
should
be
modified
whenever
sound
scientific
evidence
indicates
that
a
site­
specific
criterion
produced
using
this
guidance
will
probably
substantially
underprotect
or
overprotect
the
aquatic
life
at
the
site
of
concern.
Derivation
of
site­
specific
criteria
for
aquatic
life
is
a
complex
process
and
requires
knowledge
in
many
areas
of
aquatic
toxicology;
any
deviation
from
this
guidance
should
be
carefully
considered
to
ensure
that
it
is
consistent
with
other
parts
of
this
guidance
and
with
"
good
science".

Personal
bias
Bias
can
never
be
eliminated,
and
some
decisions
are
at
the
fine
line
between
"
bias"
and
"
best
xvii
7.
professional
judgment".
The
possibility
of
bias
CM
be
eliminated
only
by
adoption
of
an
extreme
position
such
as
"
no
regulation"
or
"
no
discharge".
One
way
to
deal
with
bias
is
to
have
decisions
made
by
a
team
of
knowledgeable
people.

Teamwork:
The
determination
of
a
WER
should
be
a
cooperative
team
effort
beginning
with
the
completion
of
the
initial
workplan,
interpretation
of
initial
data,
revision
of
the
workplan,
etc.
The
interaction
of
a
variety
of
knowledgeable,
reasonable
people
will
help
obtain
the
best
results
for
the
expenditure
of
the
fewest
resources.
Members
of
the
team
should
acknowledge
their
biases
so
that
the
team
can
make
best
use
of
the
available
information,
taking
into
account
its
relevancy
to
the
immediate
situation
and
its
quality.

xviii
INTRODUCTION
National
aquatic
life
criteria
for
metals
are
intended
to
protect
the
aquatic
life
in
almost
all
surface
waters
of
the
United
States
(
U.
S.
EPA
1985).
This
level
of
protection
is
accomplished
in
two
ways.
First,
the
national
dataset
is
required
to
contain
aquatic
species
that
have
been
found
to
be
sensitive
to
a
variety
Of
pollutants
Second,
the
dilution
water
and
the
metal
salt
used
in
the
toxicity
tests
are
required
to
have
physical
and
chemical
characteristics
that
ensure
that
the
metal
is
at
least
as
toxic
in
the
tests
as
it
is
in
nearly
all
surface
waters.
For
example,
the
dilution
water
is
to
be
low
in
suspended
solids
and
in
organic
carbon,
and
some
form
of
metal
(
e.
g.,
insoluble
metal
and
metal
bound
by
organic
complexing
agents)
cannot
be
used
as
the
test
material.
(
The
term
"
metal"
is
used
herein
to
include
both
"
metals"
and
metalloids".)

Alternatively,
a
national
aquatic
life
Criterion
might
not
adequately
protect
the
aquatic
life
at
some
sites.
An
untested
species
that
is
important
at
a
site
might
be
more
sensitive
than
any
of
the
tested
species.
Also,
the
metal
might
be
more
toxic
in
site
water
than
in
laboratory
diluting
water
because,
for
example,
the
site
water
has
a
lower
pH
and/
or
hardness
than
most
laboratory
waters.
Thus
although
a
national
aquatic
life
criterion
is
intended
to
be
lower
than
necessary
for
most
sites,
a
national
criterion
might
not
adequately
protect
the
aquatic
life
at
some
sites.

Because
a
national
aquatic
life
criterion
might
be
more
or
less
protective
than
intended
for
the
aquatic
life
in
most
bodies
of
water,
the
U.
S.
EPA
provided
guidance
(
U.
S.
EPA
1983a,
1984)
concerning
three
procedures
that
may
be
used
to
derive
a
site­
specific
criterion:
1.
The
Recalculation
Procedure
is
intended
to
take
into
account
relevant
differences
between
the
sensitivities
of
the
aquatic
organisms
in
the
national
dataset
and
the
sensitivities
of
organisms
that
occur
at
the
site.
2.
The
Indicator
Species
Procedure
provides
for
the
use
of
a
water­
effect
ratio
(
WER)
that
is
intended
to
take
into
account
relevant
differences
between
the
toxicity
of
the
metal
in
laboratory
dilution
water
and
in
site
water.
3.
The
Resident
Species
Procedure
is
intended
to
take
into
account
both
kinds
of
differences
simultaneously.
A
site­
specific
criterion
is
intended
to
come
closer
than
the
national
criterion
to
providing
the
intended
level
of
protection
to
the
aquatic
life
at
the
site,
usually
by
taking
into
account
the
biological
and/
or
chemical
conditions
(
i.
e.,
the
species
composition
and/
or
water
quality
characteristics)
at
the
site.
The
fact
that
the
U.
S.
EPA
has
made
these
procedures
available
should
not
be
interpreted
as
implying
that
the
agency
advocates
that
states
derive
site­
specific
criteria
before
setting
state
1
standards.
Also,
derivation
of
a
site­
specific
criterion
does
not
change
the
intended
level
of
protection
of
the
aquatic
life
at
the
site.
Because
a
WER
is
expected
to
appropriately
take
into
account
(
a)
the
site­
specific
toxicity
of
the
metal,
and
(
b)
synergism,
antagonism,
and
additivity
with
other
constituents
of
the
site
water,
using
a
WER
is
more
likely
to
provide
the
intended
level
of
protection
than
not
using
a
WER.

Although
guidance
concerning
site­
specific
criteria
has
been
available
since
1983
(
U.
S.
EPA
1983a,
1984),
interest
has
increased
in
recent
years
as
states
have
devoted
more
attention
to
chemical­
specific
water
quality
criteria
for
aquatic
life.
In
addition,
interest
in
water­
effect
ratios
(
WERs)
increased
when
the
"
Interim
Guidance"
concerning
metals
(
U.
S.
EPA
1992)
made
a
fundamental
change
in
the
way
that
WERs
are
experimentally
determined
(
see
Appendix
A),
because
the
change
is
expected
to
substantially
increase
the
magnitude
of
many
WERs.
Interest
was
further
focused
on
WERs
when
they
were
integrated
into
some
of
the
aquatic
life
criteria
for
metals
that
were
promulgated
by
the
National
Toxics
Rule
(
57
FR
60848,
December
22,
1992).
The
newest
guidance
issued
by
the
U.
S.
EPA
(
Prothro
1993)
concerning
aquatic
life
criteria
for
metals
affected
the
determination
and
use
of
WERs
only
insofar
as
it
affected
the
use
of
total
recoverable
and
dissolved
criteria.

The
early
guidance
concerning
WERs
(
U.
S.
EPA
1983a,
1984)
contained
few
details
and
needs
revision,
especially
to
take
into
account
newer
guidance
concerning
metals
(
U.
S.
EPA
1992;
Prothro
1993).
The
guidance
presented
herein
supersedes
all
guidance
concerning
WERs
and
the
Indicator
Species
Procedure
given
in
Chapter
4
of
the
Water
Quality
Standards
Handbook
(
U.
S.
EPA
1983a)
and
in
U.
S.
EPA
(
1984).
All
guidance
presented
in
U.
S.
EPA
(
1992)
is
superseded
by
that
presented
by
Prothro
(
1993)
and
by
this
document.
Metals
are
specifically
addressed
herein
because
of
the
National
Toxics
Rule
NTR)
and
because
of
current
interest
in
aquatic
life
criteria
for
metals;
although
most
of
this
guidance
also
applies
to
other
pollutants,
same
obviously
applies
only
to
metals.

Even
though
this
document
was
prepared
mainly
because
of
the
NTR,
the
guidance
contained
herein
concerning
WERs
is
likely
to
have
impact
beyond
its
use
with
the
NTR.
Therefore,
it
is
appropriate
to
also
present
new
guidance
concerning
the
Recalculation
Procedure
(
see
Appendix
B)
because
the
previous
guidance
(
U.
S.
EPA
1983a,
1984)
concerning
this
procedure
also
contained
few
details
and
needs
revision.
The
NTR
does
not
allow
use
of
the
Recalculation
Procedure
in
jurisdictions
subject
to
the
NTR.

The
previous
guidance
concerning
site­
specific
procedures
did
not
allow
the
Recalculation
Procedure
and
the
WER
procedure
to
be
used
together
in
the
derivation
of
a
site­
specific
aquatic
life
criterion;
the
only
way
to
take
into
account
both
species
2
composition
and
water
quality
characteristics
in
the
Louether.
the
Additional
reasons
for
addressing
both
Recalculation
Procedure
and
the
WER
Procedure
in
this
document
are
that
both
procedures
are
based
directly
on
the
guidelines
for
deriving
national
aquatic
life
criteria
(
U.
S.
EPA
1985)
and,
when
the
two
are
used
together,
use
of
the
Recalculation
Procedure
has
specific
implications
concerning
the
determination
of
the
WER.

This
guidance
is
intended
to
produce
WERs
that
may
be
used
to
derive
site­
specific
aquatic
life
criteria
for
metals
from
most
national
and
state
aquatic
life
Criteria
that
were
derived
from
laboratory
toxicity
data.
Except
in
jurisdictions
that
are
subject
to
the
NTR,
the
WRRs
may
also
be
used
with
site­
specific
aquatic
life
criteria
that
are
derived
for
metals
using
the
Recalculation
Procedure
described
in
Appendix
B.
m
obtained
For
exatrlple,
because
they
are
designed
to
be
applied
to
criteria
derived
on
the
basis
of
laboratory
toxicity
tests,
WERs
determined
using
the
methods
described
herein
cannot
be
used
to
adjust
the
residue­
bassd
mercury
Criterion
Continuous
Concentration
(
CCC)
or
the
field­
based
selenium
freshwater
criterion.
For
the
purposes
of
the
NTR,
WERs
may
be
used
with
the
aquatic
life
criteria
for
arsenic,
cadmium,
chromiuxn(~~~),
chromiumWI),
copper,
lead,
nickel,
silver,
and
zinc
and
with
the
Criterion
Maximum
Concsntration
(
CMC)
for
mercury.
WERs
may
also
be
used
with
saltwater
criteria
for
selenium.

The
concept
of
a
WER
is
rather
simple:
YBvo
side­
by­
side
toxicity
tests
are
conducted
­
one
test
using
laboratory
dilution
water
and
the
other
using
site
water.
The
endpoint
obtained
using
site
water
is
divided
by
the
endpoint
obtained
using
laboratory
dilution
water.
The
quotient
is
the
WER,
which
is
multiplied
times
the
national,
state,
or
recalculated
aquatic
life
criterion
to
calculate
the
site­
specific
criterion.
Although
the
concept
is
simple,
the
determination
and
use
of
WERs
involves
many
considerations.

The
primary
purposes
of
this
document
are
to:
1.
Identify
steps
that
should
be
taken
before
the
determination
of
a
WER
is
begun.
2.
Describe
the
methods
ret
omnended
by
the
U.
S.
EPA
for
the
determination
of
WERs.
3.
Address
some
issues
concerning
the
use
of
WERs.
4.
Present
new
guidance
concerning
the
Recalculation
Procedure.

3
Because
a
national
Criterion
is
intended
toqrotect
aquatic
life
in
alsmst
all
bodies
of
water
and
because
a
WRR
is
intended
to
account
for
a
difference
between
the
toxicity
of
a
metal
in
a
laboratory
dilution
water
Md
its
toxicity
in
a
site
water,
dischargers
who
want
higher
permit
limits
than
those
derived
on
the
basis
of
M
existing
aquatic
life
criterion
will
probably
consider
determining
a
WRR.
Use
of
a
WER
should
be
considered
only
as
a
last
resort
for
at
least
three
reasons:
a.
Even
though
some
WERs
will
be
substantially
greater
than
1.0,
sane
will
be
about
1.0
Md
some
will
be
less
than
1.0.
b.
The
detemaination
of
a
WER
requires
substantial
resources.
c.
There
are
other
things
that
a
discharger
CM
do
that
might
be
mOre
cost­
effective
than
detenaining
a
WER.

The
two
situations
in
which
the
determination
of
a
WER
might
appear
attractive
to
dischargers
are
when
(
a)
a
discharge
appears
to
exceed
existing
or
proposed
water
quality­
based
permit
limits,
and
(
b)
M
instresm
concentration
appears
to
axceed
M
existing
or
proposed
aquatic
life
criterion.
Such
situations
result
from
xneasurernent
of
the
concentration
of
a
metal
in
M
effluent
or
a
surface
water.
It
would
therefore
seem
reasonable
to
ensue
that
such
measurements
were
not
subject
to
contamination.
Usually
it
is
much
easier
to
verify
chemical
measurements
by
using
l
clean
techniques'
for
collecting,
handling,
storing,
preparing,
and
analyzing
soles,
than
to
determine
a
WER.
Clean
techniques
and
some
related
QA/
QC
considerations
are
dis~
ssed
in
Appendix
C.

In
addition
to
investigating
the
use
of
.
clean
techniques',
other
steps
that
a
discharger
should
take
prior
to
beginning
the
experimental
determination
of
a
WER
include:
1.
Evaluate
the
potential
for
reducing
the
discharge
of
the
metal.
2.
Investigate
such
possible
constraints
on
permit
limits
as
Mtibacksliding
and
Mtidegradation
requirements
Md
human
health
and
wildlife
criteria.
3.
Obtain
assistance
from
M
aquatic
toxicologist
who
understands
the
basics
of
WERs
(
see
Appendix
D),
the
U.
S.
EPA's
national
aquatic
life
guidelines
(
U.
S.
EPA
19851,
the
guidance
presented
by
Prothro
(
19931,
the
national
criteria
document
for
the
metal(
s)
of
concern
(
see
Appendix
E),
the
procedures
described
by
the
U.
S.
EPA
(
1993a,
b,
c)
for
acute
and
chronic
toxicity
tests
on
effluents
Md
surface
waters,
and
the
procedures
described
by
ASTM
(
1993a,
b,
c,
d,
e)
for
acute
and
chronic
toxicity
tests
in
laboratory
dilution
water.
4.
Develop
M
initial
definition
of
the
site
to
which
the
site­
specific
criterion
is
to
apply.
5.
Consider
use
of
the
Recalculation
Procedure
(
see
Appendix
B).
6.
Evaluate
the
cost­
effectiveness
of
the
detemination
of
a
WER.
Comparative
toxicity
tests
provide
the
most
useful
data,
but
chemical
analysis
of
the
downstream
water
might
be
helpful
4
because
the
following
are
often
true
for
some
metals:
a.
The
lower
the
percent
of
the
total
recoverable
metal
in
the
downstream
water
that
is
dissolved,
the
higher
the
WER.
b.
The
higher
the
concentration
of
total
organic
carbon
(
Tot)
and/
Or
total
suspended
solids
(
TSS),
the
higher
the
WER.
It
is
also
true
that
the
higher
the
concentration
of
nontoxic
dissolved
metal,
the
higher
the
WER.
Although
some
chemical
analyses
might
provide
useful
information
concerning
the
toxicities
of
soms
metals
in
water,
at
the
present
only
toxicity
tests
CM
accurately
reflect
the
toxicities
of
different
forms
of
a
metal
(
see
Appendix
D).
7.
Submit
a
workplan
for
the
experimental
determination
of
the
WER
to
the
appropriate
regulatory
authority
(
and
possibly
to
the
Water
Managemsnt
Division
of
the
EPA
Regional
Office)
for
coscnent.
The
workplan
should
include
detailed
descriptions
of
the
site;
existing
criterion
and
standard;
design
flows;
site
water;
effluent;
sampling
plan;
procedures
that
will
be
used
for
collecting,
handling,
and
analyzing
samples
of
site
water
and
effluent;
primary
and
secondary
toxicity
tests;
quality
assurance/
quality
control
(
QA/
QC)
procedures;
Standard
Operating
Procedures
(
SOPS);
and
data
interpretation.
After
the
workplan
is
completed,
the
initial
phase
should
be
wlexnented;
then
the
data
obtained
should
bs
evaluated,
and
the
workplan
should
be
revised
if
appropriate.
Developing
and
modifying
the
workplan
and
analyzing
and
interpreting
the
data
should
be
a
cooperative
effort
by
a
team
of
knowledgeable
people.

Most
aquatic
life
criteria
contain
both
a
a4C
and
a
CCC,
and
it
is
usually
possible
to
determine
both
a
cmcWRR
anda
cccWER.
The
two
WERs
cannot
be
assumsd
to
be
equal
because
the
magnitude
of
a
WER
will
probably
depsnd
on
the
sensitivity
of
the
toxicity
test
ussd
and
on
the
percsnt
effluent
in
the
site
water
(
see
Appendix
D),
both
of
which
CM
depend
on
which
WER
is
to
be
determined.
In
some
cases,
it
is
expected
that
a
larger
WER
CM
be
applied
to
the
CCC
than
to
the
CMC,
and
so
it
would
be
environmentally
conservative
to
apply
ancWERs
to
CCCs.
In
such
cases
it
is
possible
to
determine
a
cmcWER
and
apply
it
to
both
the
CMC
and
the
CCC
in
order
to
derive
a
site­
specific
CXC,
a
site­
specifik
CCC,
and
new
permit
limits.
If
these
new
pexmit
limits
are
controlled
by
the
new
site­
specific
CCC,
a
cccWER
could
be
determined
using
a
more
sensitive
test,
possibly
raising
the
site­
specific
CCC
and
the
permit
limits
again.
A
cccWER
may,
of
course,
be
determined
whenever
desired.
Unless
the
experimental
variation
is
increased,
use
of
a
cccWER
will
usually
improve
the
accuracy
of
the
resulting
site­
specific
CCC.

In
some
cases,
a
larger
WER
cannot
bs
applied
to
the
CCC
than
to
the
CRC
and
so
it
might
not
be
environmentally
conservative
to
apply
a
cmcWER
to
a
CCC
(
see
section
A.
4
of
Method
1).

5
Some
of
the
guidance
contained
herein
specifically
applies
to
situations
in
which
the
pemit
limits
were
calculated
using
steady­
state
modeling;
in
particular,
some
samples
are
to
be
obtained
when
the
actual
stream
flow
is
close
to
the
design
flow.
If
permit
limits
were
calculated
Using
dynamic
smdeling,
the
guidance
will
have
to
be
modified,
but
it
is
unclear
at
present
what
modifications
are
most
appropriate.
For
wle,
it
might
be
useful
to
determine
whether
the
magnitude
of
the
WER
is
related
to
the
flow
of
the
upstream
water
and/
or
the
effluent.

Wo
methods
are
used
to
determine
WERs.
Method
1
will
probably
be
used
to
determine
all
c!
mcWERs
and
most
CccWERs
because
it
CM
be
applied
to
situations
that
are
in
the
vicinities
of
plumes.
Because
WERs
are
likely
to
depend
on
the
concentration
of
effluent
in
the
water
and
because
the
percent
effluent
in
a
water
swle
obtained
in
the
ismediate
vicinity
of
a
plums
is
unknown,
simulated
downstream
water
is
used
so
that
the
percent
effluent
in
the
sample
is
known.
For
exas@
le,
if
a
ssmgle
that
was
supposed
to
represent
a
cwlete­
mix
situation
was
accidently
taken
in
the
plume
upstream
of
comnplete
mix,
the
sanlple
would
probably
have
a
higher
percent
effluent
and
a
higher
WI&
R
than
a
sas@
e
taken
downstream
of
coa@
ete
mix;
use
of
the
higher
WER
to
derive
a
site­
specific
criterion
for
the
complete­
mix
situation
would
result
in
underprotection.
If
the
sample
were
accidently
taken
upstream
of
cmlete
mix
but
outside
the
plume,
overprotection
would
probably
result.

Method
1
will
probably
be
used
to
determine
all
cmcWERs
and
mo8t
cccWERs
in
flowing
fresh
waters,
such
as
rivers
and
streams.
Method
1
i8
intended
to
apply
not
only
to
ordinary
rivers
and
streams
but
also
to
streams
that
some
people
might
consider
extraordinary,
such
as
streams
whose
design
flows
are
zero
and
streams
that
mane
state
and/
or
federal
agencies
refer
to
as
l
effluent­
dependentm,
%
abitat­
creating.,
or
'
effluent­
dominated'.
Method
1
is
also
used
to
determine
cmcWERs
in
such.
large
sites
as
oceans
and
large
lakes,
rebendrb,
and
estuaries
(
see
Appendix
F).

Method
2
is
used
to
determine
WEF4s
that
apply
outside
the
area
of
plumes
in
large
bodies
of
water.
Such
=
will
be
cccWERs
and
will
be
determined
using
sas@
es
of
actual
site
water
obtained
at
various
times,
locations,
and
depths
in
order
to
identify
the
range
of
WERs
that
apply
to
the
body
of
water.
These
experimentally
detenained
WERs
are
then
used
to
decide
how
many
site­
specific
criteria
should
be
derived
for
the
body
of
water
and
what
the
criterion
(
or
criteria)
should
be.
Method
2
requires
substantially
more
resources
than
Method
1.

6
The
complexity
of
each
method
increases
when
the
number
of
metals
and/
or
the
number
of
discharges
is
two
or
more:
a.
The
simplest
situation
is
when
a
WER
is
to
be
determined
for
only
one
metal
MC)
only
one
discharge
has
permit
limits
for
that
metal.
(
This
is
the
single­
metal
single­
discharge
situation.)
b.
A
more
complex
situation
is
when
a
WER
is
to
be
determined
for
only
one
metal,
but
more
than
one
discharge
has
permit
limits
for
that
metal.
(
This
is
the
single­
metal
multiple­
discharge
situation.)
c.
An
even
mOre
complex
situation
is
when
WEFU
are
to
be
determined
for
more
than
one
metal,
but
only
one
discharge
has
permit
limits
for
any
of
the
metals.
(
This
is
the
multiple­
metal
single­
discharge
situation.)
d.
The
most
complex
situation
is
when
WERs
are
to
be
determined
for
more
than
one
metal
and
more
than
one
discharge
has
permit
limits
for
some
or
all
of
the
metals.
(
This
is
the
multiple­
metal
multiple­
discharge
situation.)
WRRs
need
to
be
determined
for
each
metal
at
each
site
because
extrapolation
of
a
WRR
from
one
metal
to
another,
one
effluent
to
another,
or
one
surface
water
to
another
is
too
uncertain.

Roth
methods
work
well
in
multiple­
metal
situations,
but
special
tests
or
additional
tests
will
be
necessary
to
show
that
the
resulting
combination
of
site­
specific
criteria
will
not
be
too
toxic.
Method
2
is
better
suited
to
multiple­
discharge
situations
than
is
Method
1.
Appendix
F
provides
additional
guidance
concerning
multiple­
metal
and
multiple­
discharge
situations,
but
it
does
not
discuss
allocation
of
waste
loads,
which
is
performed
when
a
wasteload
allocation
(
WLA)
or
a
total
maximum
daily
load
(
TMDL)
is
developed
(
U.
S.
EPA
1991a).

A
total
recoverable
WER
CM
be
determined
if
the
metal
in
both
of
the
side­
by­
side
toxicity
tests
is
analyzed
using
the
total
recoverable
measurement;
similarly,
a
dissolved
WER
can
be
determined
if
the
metal
in
both
tests
is
analyzed
using
the
dissolved
measurement.
A
total
recoverable
WER
is
used
to
calculate
a
total
recoverable
site­
specific
criterion
from
an
aquatic
life
criterion
that
is
expressed
using
the
total
recoverable
measurement,
whereas
a
dissolved
WRR
is
used
to
calculate
a
dissolved
site­
specific
criterion
from
a
criterion
that
is
expressed
in
terms
of
the
dissolved
measurement.
Figure
1
illustrates
the
relationships
between
total
recoverable
and
dissolved
criteria,
WERs,
and
the
Recalculation
Procedure.

Roth
Method
1
and
Method
2
CM
be
used
to
determine
a
total
recoverable
WER
and/
or
a
dissolved
WER.
The
only
difference
in
the
experimental
procedure
is
whether
the
WER
is
based
on
measurements
of
total
recoverable
metal
or
dissolved
metal
in
the
7
test
solutions.
Both
total
recoverable
and
dissolved
measurements
are
to
bs
performed
for
all
tests
to
help
judge
the
quality
of
the
tests,
to
provide
a
check
on­
the
analytical
chemistry,
and
to
help
understand
the
results;
performing
both
wasuremsnts
also
increases
the
alternatives
available
for
use
of
the
results.
For
vie,
a
dissolved
WER
that
is
not
useful
with
a
total
recoverable
criterion
might
be
useful
in
the
future
if
a
dissolved
criterion
becomss
available.
Also,
as
explained
in
Appendix
D,
except
for
experimsntal
variation,
use
of
a
total
recoverable
WER
with
a
total
recoverable
criterion
should
produce
the
same
total
recoverable
permit
limits
as
use
of
8
dissolved
WER
with
a
dissolved
criterion;
the
internal
consistency
of
the
approaches
and
the
data
can
be
evaluated
if
both
total
recoverable
and
dissolved
criteria
and
WERs
are
determined.
It
is
expected
that
in
many
situations
total
recoverable
WEF&
will
be
larger
and
smre
variable
than
dissolved
WERs.

Traditionally,
for
practical
reasons,
the
requirssmnts
concerning
such
aspects
as
acclimation
of
test
organisms
to
test
tsnperature
and
dilution
water
have
not
been
as
stringsnt
for
toxicity
tests
on
surface
waters
and
effluents
as
for
tests
using
laboratory
dilution
water.
Because
a
WER
is
a
ratio
calculated
from
the
results
of
side­
by­
side
tests,
it
might
seem
that
acclimation
is
not
isportant
for
a
WER
as
long
as
the
organisms
and
conditions
are
idsntical
in
the
two
tests.
Because
WERs
are
used
to
adjust
aquatic
life
criteria
that
are
derivsd
fran
results
of
laboratory
tests,
the
tests
conducted
in
laboratory
dilution
water
for
the
detexmination
of
WERs
should
be
conducted
in
the
same
way
as
the
laboratory
toxicity
tests
used
in
the
derivation
of
aquatic
life
criteria.
In
the
WER
process,
the
tests
in
laboratory
dilution
water
provide
the
vital
link
bstween
national
criteria
and
site­
specific
criteria,
and
so
it
is
isportant
to
c­
are
at
least
some
results
obtained
in
the
laboratory
dilution
water
with
results
obtained
in
at
least
one
other
laboratory.

Three
important
principles
for
staking
decisions
concerning
the
methodology
for
the
side­
by­
side
tests
are:
1.
The
tests
using
laboratory
dilution
water
should
be
conducted
so
that
the
results
would
be
acceptable
for
use
in
the
derivation
of
national
criteria.
2.
As
much
as
is
feasible,
the
tests
using
site
water
should
be
conducted
using
the
same
procedures
as
the
tests
using
the
laboratory
dilution
water.
3.
All
tests
should
follow
any
special
requirements
that
are
necessary
because
the
results
are
to
be
used
to
calculate
a
WER.
Some
such
special
requirements
are
msed
because
the
criterion
for
a
rather
complex
situation
is
being
changed
based
on
few
data,
so
more
assurance
is
required
that
the
data
are
high
quality.
The
xmst
important
special
requirement
is
that
the
concentrations
of
the
metal
are
to
be
measured
using
both
the
total
recoverable
and
dissolved
methods
in
all
toxicity
tests
used
for
the
determination
of
a
WER.
This
requirement
is
necessary
because
half
of
the
tests
conducted
for
the
determination
of
WRRs
use
a
site
water
in
which
the
concentration
of
metal
probably
is
not
xiegligible.
Recause
it
is
likely
that
the
concentration
of
metal
in
the
laboratory
dilution
water
is
negligible,
assuming
that
the
concentration
in
both
waters
is
negligible
and
basing
WERs
on
the
amount
of
metal
added
would
produce
an
unnecessarily
low
value
for
the
WER.
In
addition,
WERa
are
based
on
too
few
data
to
88surne
that
ncaainal
concentrations
are
accurate.
Nominal
concentrations
obviously
cannot
be
used
if
a
dissolved
WER
is
to
be
determined.
Measured
dissolved
concentrations
at
the
beginning
and
end
of
the
test
are
used
to
judge
the
acceptability
of
the
test,
and
it
is
certainly
reasonable
to
measure
the
total
recoverable
concentration
when
the
dissolved
concentration
is
amasured.
Further,
measuring
the
concentrations
might
lead
to
an
interpretation
of
the
results
that
allows
a
substantially
better
use
of
the
WE&
S.

The
appropriate
regulatory
authority
might
recoxmen
d
that
one
or
more
conditions
be
met
when
a
WER
is
determined
in
order
to
reduce
the
mssibilitv
of
havina
to
determine
a
new
WER
later:
1.

2.

3.
Requireahnts
that
­
me
in
the­
existing
permit
concerning
VET
testing,
Toxicity
Identification
Evaluation
(
TIE),
and/
or
Toxicity
Reduction
Rvaluation
(
TRR)
(
U.
S.
EPA
1991a).
Implewntation
of
pollution
prevention
efforts,
such
as
pretreatment,
waste
minimization,
and
source
reduction.
A
dewmstration
that
applicable
technology­
based
requirements
are
being
met.

Even
if
all
re
caamsnded
conditions
are
satisfied,
determination
of
a
WER
might
not
be
possible
if
the
effluent,
upstream
water,
and/
or
downstream
water
are
toxic
to
the
test
organisms.
In
some
such
cases,
it
might
be
possible
to
determine
a
WER,
but
desired
to
determine
a
WRR
before
remediation
and
the
toxicity
is
in
the
upstream
water,
it
might
be
possible
to
use
a
laboratory
dilution
water
or
a
water
from
a
clean
tributary
in
place
of
the
upstream
water;
if
a
substitute
water
is
used,
its
water
quality
characteristics
should
be
similar
to
those
of
the
upstream
water
(
i.
e.,
the
pH
should
be
within
0.2
pH
units
and
the
hardness,

9
alkalinity,
and
concsntrationa
of
TSS
and
TOC
should
be
within
10
8
or
5
mg/
L,
whichever
is
greater,
of
those
in
the
upstream
water).
If
the
upstream
water
is
chronically
toxic,
but
not
acutely
toxic,
it
might
be
possible
to
determine
a
QllcwER
even
if
a
CCCWER
cannot
be
determined;
a
cmcWER
might
not
be
useful,
however,
if
the
petit
limits
are
controlled
by
the
CCC;
in
such
a
case,
it
would
probably
not
be
acceptable
to
assume
that
the
cxncWER
is
an
environmentally
consenmtive
estimate
of
the
cc­.
If
the
WER
is
detenninsd
using
downstream
water
and
the
toxicity
is
due
to
the
efflusnt,
tests
at
lower
concentrations
of
the
efflusnt
might
give
an
indication
of
the
amount
of
remsdiation
needed.

Besides
requiring
that
the
WER
be
valid,
the
appropriate
regulatory
authority
might
consider
inposing
other
conditions
for
the
approval
of
a
site­
specific
criterion
based
on
the
WER:
1.
Periodic
reevaluation
of
the
WER.
a.
WERs
detenninsd
in
upstream
water
take
into
account
constituents
contributed
by
point
and
nonpoint
sources
and
natural
runoff;
thus
a
WER
should
be
reevaluated
whenever
newly
implementsd
controls
or
other
changes
substantially
affect
such
factors
as
hardness,
alkalinity,
@
I,
suspended
solids,
organic
carbon,
or
other
toxic
materials.
b.
Most
WERs
determined
using
downstream
water
are
influenced
more
by
the
effluent
than
the
upstream
water.
Downstream
WEZs
should
be
reevaluated
whenever
newly
imglewnted
control&
or
other
changes
might
substantially
iqmct
the
effluent,
i.
e.,
might
inpact
the
forms
and
concentrations
of
the
metal,
hardness,
alkalinity,
pH,
suspended
solids,
organic
car&
n,
or
other
toxic
materials.
A
special
concern
is
the
possibility
of
a
shift
fraa
discharge
of
nontoxic
metal
to
discharge
of
toxic
mstal
such
that
the
concentration
of
the
metal
does
not
increase;
analytical
chemistry
might
not
detect
the
change
but
toxicity
tests
would.
Evsn
if
no
changes
are
known
to
have
occurred,
WBRs
should
be
reevaluated
periodically.
(
The
NTR
recmmmAl
that
NPDES
permits
include
periodic
dete
nninations
of
WERs
in
the
monitoring
requiremsnts.)
With
aduance
planning,
it
should
usually
bs
possible
to
perform
such
reevaluations
under
conditions
that
are
at
least
reasonably
similar
to
those
that
control
the
permit
limits
(
e.
g.,
either
design­
flow
or
high­
flow
conditions)
because
there
should
be
a
reasonably
long
period
of
time
during
which
the
reevaluation
can
be
performed.
Periodic
determination
of
WERs
should
be
designed
to
answer
questions,
not
just
generate
data.
2.
Increased
chemical
monitoring
of
the
upstream
water,
effluent,
and/
or
downstream
water,
as
appropriate,
for
water
quality
characteristics
that
probably
affect
the
toxicity
of
the
m&
al
10
(
e.
g.,
hardness,
alkalinity,
pH,
Tot,
and
TSS)
to
determine
whether
conditions
change.
The
conditions
at
the
times
the
samples
were
obtained
should
be
kept
on
record
for
reference.
The
WER
should
be
reevaluated
whenever
hardness,
alkalinity,
pH,
Tot,
and/
or
TSS
decrease
below
the
values
that
existed
when
the
WERs
were
determined.
3.
Periodic
reevaluation
of
the
environmental
fate
of
the
metal
in
the
effluent
(
see
Appendix
A).
4.
WET
testing.
5.
Instream
bioassessments.

Decisions
concerning
the
possible
imposition
of
such
conditions
should
take
into
account:
a.
The
ratio
of
the
new
and
old
criteria.
The
greater
the
increase
in
the
criterion,
the
more
concern
there
should
be
about
(
1)
the
fate
of
any
nontoxic
metal
that
contributes
to
the
WER
and
(
2)
changes
in
water
quality
that
might
occur
within
the
site.
The
imposition
of
one
or
more
conditions
should
be
considered
if
the
WER
is
used
to
raise
the
criterion
by,
for
axample,
a
factor
of
two,
and
especially
if
it
is
raised
by
a
factor
of
five
or
more.
The
significance
of
the
magnitude
of
the
ratio
can
be
judged
by
comparison
with
the
acute­
chronic
ratio,
the
factor
of
two
that
is
the
ratio
of
the
F'AV
to
the
CMC,
and
the
range
of
sensitivities
of
species
in
the
criteria
do
cument
for
the
metal
(
see
Appendix
E).
b.
The
size
of
the
site.
c.
The
size
of
the
discharge.
d.
The
rate
of
downstream
dilution.
8.
Whether
the
CMC
or
the
CCC
controls
the
pennit
limits.
When
WERs
are
determined
using
upstream
water,
conditions
on
the
use
of
a
WER
are
more
likely
when
the
water
contains
an
effluent
that
increases
the
WER
by
adding
TOC
and/
or
TSS,
because
the
WER
will
bs
larger
and
any
decrease
in
the
discharge
of
such
Tot
and/
or
TSS
might
decrease
the
WER
and
result
in
underprotection.
A
WER
determined
using
downstream
water
is
likely
to
be
larger
and
quite
dependent
on
the
composition
of
the
effluent;
there
should
be
concern
about
whether
a
change
in
the
effluent
might
result
in
underprotection
at
some
tims
in
the
future.

In
some
situations
a
discharger
might
not
want
to
or
might
not
be
allowed
to
raise
a
criterion
as
much
as
could
be
justified
by
a
WER:
1.
The
maximum
possible
increase
is
not
needed
and
raising
the
criterion
more
than
needed
might
greatly
raise
the
cost
if
a
greater
increase
would
require
more
tests
and/
or
increase
the
conditions
imposed
on
approval
of
the
site­
specific
criterion.
2.
Such
other
constraints
as
antibacksliding
or
antidegradation
requirements
or
human
health
or
wildlife
criteria
might
limit
the
amount
of
increase
regardless
of
the
magnitude
of
the
WER.

11
3.
The
permit
limits
might
be
limited
by
an
aquatic
life
criterion
that
applies
outside
the
site.
It
is
EPA
policy
that
permit
limits
cannot
be
so
high
that
they
inadequately
protect
a
portion
of
the
sams
or
a
different
body
of
water
that
is
outside
the
site;
nothing
contained
herein
changes
this
policy
in
anyway.
If
no
increase
in
the
existing
discharge
is
allowed,
the
only
use
of
a
WgR
will
be
to
determine
whether
an
existing
discharge
needs
tobe
reduced.
Thus
a
major
use
of
W
might
be
where
technology­
based
controls
allow
concentrations
in
surface
waters
to
exceed
national,
state,
or
recalculated
aquatic
life
criteria.
In
this
case,
it
might
only
bs
necessary
to
determine
that
the
WKR
is
greater
than
a
particular
value;
it
might
not
be
necessary
to
quantify
them.
When
possible,
it
might
be
desirable
to
show
that
the
wxinnsa
WER
is
grsater
than
the
WER
that
will
be
used
in
order
to
demonstrate
that
a
margin
of
safety
exists,
but
again
it
might
not
bs
necessary
to
quantify
the
smximum
WER.

In
jurisdictions
not
subject
to
the
NTR,
WERs
should
be
used
to
derive
site­
specific
criteria,
not
just
to
calculate
permit
limits,
because
data
obtained
fran
ambient
monitoring
should
be
interpreted
by
comparison
with
ambient
criteria.
(
This
is
not
a
problem
in
jurisdictions
subject
to
the
NTR
because
the
NTP
&
fines
the
smbient
criterion
as
WER
x
the
EPA
criterion'.)
If
a
WER
is
used
to
adjust
pennit
limits
without
adjusting
the
criterion,
the
permit
limits
would
allow
the
criterion
to
be
exceeded.
Thw
the
WRR
shouldbs
ussd
to
calculate
a
site­
specific
criterion,
which
should
then
bs
used
to
calculate
pennit
limits.
In
sane
states,
site­
specific
criteria
can
only
bs
adopted
as
revised
criteria
in
a
separate,
independent
water
quality
standards
review
process.
In
other
states,
site­
specific
criteria
can
be
developed
in
conjunction
with
the
NPDES
permitting
process,
as
long
as
the
adoption
of
a
site­
specific
criterion
satisfies
the
pertinent
water
quality
standards
procedural
requirements
('
i.
e.,
a
public
notice
and
a
public
hearing).
In
either
case,
site­
specific
criteria
are
to
bs
adopted
prior
to
NPDES
permit
issuance.
Moreouer,
the
EPA
Regional
Administrator
has
authority
to
approve
or
disapprove
all
new
and
revised
site­
specific
criteria
and
to
review
NPDES
permits
to
verify
cwliance
with
the
applicable
water
quality
criteria.
I
Other
aspects
of
the
use
of
WERs
in
connection
with
permit
limits,
WLAs,
and
IMDLs
are
outside
the
scope
of
this
document.
The
Technical
Support
Documen
t
(
U.
S.
EPA
1991s)
and
Prothro
(
1993)
provide
more
information
concerning
Mlaaentation
procedures.
Nothing
contained
herein
should
be
interpreted
as
changing
the
three­
part
approach
that
EPA
uses
to
protect
aquatic
life:
(
1)
numeric
chsmical­
specific
water
quality
criteria
for
individual
pollutants,
(
2)
whole
effluent
toxicity
WET)
testing,
and
(
3)
instreaxn
bioassessmsnts.

12
Even
though
there
are
similarities
between
WET
testing
and
the
determination
of
WERs,
there
are
important
differences.
For
example,
WEFU
CM
be
used
to
derive
site­
specific
criteria
for
individual
pollutants,
but
WET
testing
cannot.
The
difference
between
WET
testing
and
the
determination
of
WERs
is
less
whsn
the
toxicity
tests
used
in
the
determination
of
the
WER
are
ones
that
are
used
in
WET
testing.
If
a
WER
is
used
to
make
a
large
change
in
a
criterion,
additional
WET
testing
and/
or
instream
bioassessmsnts
are
likely
to
be
recoxnended.

A
major
problem
with
the
determination
and
use
of
aquatic
life
criteria
for
metals
is
that
no
analytical
measurement
or
combination
of
xneasuremsnts
has
yet
been
shown
to
explain
the
toxicity
of
a
metal
to
aquatic
plants,
invertebrates,
amphibians,
and
fishes
over
the
relevant
range
of
Condition8
in
surface
waters
(
see
Appendix
0).
It
is
not
just
that
insufficient
data
exist
to
justify
a
relationship;
rather,
existing
data
possibly
contradict
some
ideas
that
could
possibly
be
very
useful
if
true.
For
mle,
the
concentration
of
free
metal
ion
could
possibly
bs
a
useful
basis
for
expressing
water
quality
criteria
for
metals
if
it
could
be
feasible
and
could
bs
used
in
a
way
that
does
not
result
in
widespread
underprotection
of
aquatic
life.
Some
available
data,
however,
might
contradict
the
idea
that
ths
toxicity
of
copper
to
aquatic
organisms
is
proportional
to
the
concentration
or
the
activity
of
the
cupric
ion.
Evaluating
the
usefulness
of
any
approach
based
on
metal
speciation
is
difficult
until
it
is
known
how
many
of
the
species
of
the
metal
are
toxic,
what
the
relative
toxicities
are,
whether
they
are
additive
(
if
more
than
one
is
toxic),
and
the
quantitative
effects
of
the
factors
that
have
major
impacts
on
the
bioavailability
and/
or
toxicity
of
the
toxic
species.
Just
as
it
is
not
easy
to
find
a
useful
quantitative
relationship
between
the
analytical
chemistry
of
metals
and
the
toxicity
of
mstals
to
aquatic
life,
it
i8
al80
not
easy
to
find
a
qualitative
relationship
that
CM
be
used
to
provide
adequate
protection
for
the
aquatic
life
in
almost
all
bodies
of
water
without
providing
as
much
overprotection
for
mans
bodies
of
water
as
results
from
use
of
the
total
recoverable
and
dissolved
measurements.

The
U.
S.
EPA
cannot
ignore
the
existence
of
pollution
problems
and
delay
setting
aquatic
life
criteria
until
all
scientific
issues
have
been
adequately
resolved.
In
light
of
uncertainty,
the
agency
needs
to
derive
criteria
that
are
environmentally
conservative
in
most
bodies
of
water.
Because
of
uncertainty
concerning
the
relationship
between
the
analytical
chemistry
and
the
toxicity
of
metals,
aquatic
life
criteria
for
metals
are
expressed
in
terms
of
analytical
measuremsnts
that
result
in
the
criteria
providing
more
protection
than
necessary
for
the
aquatic
life
in
most
bodies
of
water.
The
agency
has
provided
for
the
13
use
of
WERs
to
address
the
general
conservatism,
but
expects
that
sosm
WERs
will
bs
less
than
l.
O.
because
national,
state,
and
recalculated
criteria
are
not
necessarily
environmentally
conservative
for
all
bodies
of
water.

It
has
become
obvious,
however,
that
the
determination
and
use
of
WBRs
is
not
a
smle
solution
to
the
existing
general
conservatism.
It
is
likely
that
a
pemanent
solution
will
have
to
be'based
on
an
adsquate
quantitative
sxplanation
of
how
metals
and
a&
tic
organisms
interact.
In
the
meantime,
the
use
of­
total
recoverable
and
dissolved
measurements
to
express
criteria
and
the
use
of
site­
specific
criteria
are
intended
to
provide
adequate
protection
for
almost
all
bodies
of
water
without
axcassive
overprotection
for
too
many
bodies
of
water.
Work
needs
to
continue
on
the
penaanent
solution
and,
just
in
case,
on
inproved
alternative
approaches.

Use
of
WRRs
to
derive
site­
specific
criteria
is
intended
to
allow
a
reduction
or
elimination
of
the
gsneral
overprotection
8SSoCi8ted
with
application
of
a
national
Criterion
to
individual
bodies
of
water,
but
a
major
problem
is
that
a
WER
will
rarely
bs
constant
over
time,
location,
and
depth
in
a
body
of
water
due
to
plumsa,
mixing,
and
resuspension.
It
is
possible
that
dissolved
concentration
and
WERs
will
be
less
variable
than
total
recoverable
ones.
It
might
also
be
possible
to
reduce
the
impact
of
the
heterogeneity
if
n
are
additive
across
time,
location,
and
depth
(
see
Appendix
G).
Regardless
of
what
approaches,
tools,
hypotheses,
and
assumptions
are
utilized,
variation
will
exist
and
WERs
will
have
to
bs
used
in
a
conservative
manner.
Because
of
vwiation
bstween
bodies
of
water,
national
criteria
are
derived
to
bs
environsmntally
conservative
for
most
bodies
of
water,
whereas
the
WER
procsdure,
which
is
intended
to
reduce
the
general
conServatiSm
of
national
criteria,
has
to
be
conservative
because
of
variation
among
WKF&
within
a
body
of
'
water.

The
consenratism
introduced
by
variation
among
WEFU
is
due
not
to
the
concetpt
of
WERs,
but
to
the
way
they
are
used.
The
reason
that
national
criteria
are
conservative
in
the
first
place
is
the
uncertainty
concerning
the
linkage
of
analytical
chemistry
and
.
toxiCity;
the
toxicity
of
solutions
CM
be
measured,
but
toxicity
cannot
be
smdelled
adsquately
using
available
chemical
msasurexnsnts.
Similarly,
the
currsnt
way
that
WEE&
are
used
dspends
on
a
linkage
betwssn
analytical
chemistry
and
toxicity
because
WERs
are
used
to
derive
site­
specific
criteria
that
are
expressed
in
terms
of
chemical
measurements.

Without
changing
the
amount
or
kind
of
toxicity
testing
that
is
perfonaed
when
WERs
are
determined
using
Method
2,
a
different
way
of
using
the
WERs
could
avoid
some
of
the
problans
introduced
by
the
dependence
on
analytical
chemistry.
*
The
l
sanrple­
specific
WER
approach9
could
consist
of
sampling
a
body
of
water
at
a
number
of
locations,
determining
the
WER
for
each
sample,
and
14
measuring
the
concentration
of
the
metal
in
each
sample.
Then
for
each
individual
sample,
a
quotient
would
be
calculated
by
dividing
the
concentration
of
metal
in
the
sample
by
the
product
of
the
national
criterion
times
the
WEB
obtained
for
that
sample.
Except
for
experimental
variation,
when
the
quotient
for
a
smle
is
less
than
1,
the
concentration
of
metal
in
that
sample,
is
acceptable;
when
the
quotient
for
a
sample
is
greater
than
1,
the
concentration
of
metal
in
that
sample
is
too
high.
As
a
check,
both
the
total
recoverable
measurement
and
the
dissolved
measuremsnt
should
be
used
because
they
should
provide
the
sams
answer
if
everything
is
done
correctly
and
accurately.
This
approach
CM
also
be
used
whenever
Method
1
is
used;
although
Method
1
is
used
with
simulated
downstream
water,
the
sample­
specific
WER
approach
CM
be
used
with
either
simulated
downstream
water
,
or
actual
downstream
water.'

This
sample­
specific
WER
approach
has
several
interesting
features:
1.
It
is
not
a
different
way
of
determining
WERs;
it
is
merely
a
different
way
of
using
the
WERs
that
are
determined.
2.
Variation
among
WEF&
within
a
body
of
water
is
not
a
problesn.
3.
It
eliminates
problsms
concerning
the
unknown
relationship
between
toxicity
and
analytical
chemistry.
4.
It
works
equally
well
in
areas
that
are
in
or
near
plumes
and
in
areas
that
are
away
from
plumsa.
5.
It
works
squally
well
in
single­
discharge
and
multiple­
discharge
situations.
6.
It
automatically
accounts
for
synergism,
antagonism,
and
additivity
between
toxicant8.
This
way
of
using
WERs
is
equivalsnt
to
expressing
the
national
criterion
for
a
pollutant
in
terms
of
toxicity
tests
whose
endpoints
equal
the
CMC`
and
the
CCC;
if
the
site
water
causes
less
adverse
effect
than
is
defined
to
bs
the
sndpoint,
the
concentration
of
that
pollutant
in
the
site
water
does
not
exceed
the
national
criterion.
This
sample­
specific
WER
approach
does
not
directly
fit
into
the
current
framewo
rk
wherein
criteria
are
derived
and
then
permit
limits
are
calculated
from
the
criteria.

If
the
sample­
specific
WER
approach
were
to
produce
a
number
of
quotients
that
are
greater
than
1,
it
would
seem
that
the
concentration
of
metal
in
the
discharge(
s)
should
be
reduced
enough
that
the
guotisnt
is
not
greater
than
1.
Although
this
might
sound
straightforward,
the
discharger(
s)
would
find
that
a
substantial
reduction
in
the
discharge
of
a
metal
would
not
achieve
the
intended
result
if
the
reduction
was
due
to
removal
of
nontoxic
metal.
A
chemical
monitoring
approach
that
cannot
differentiate
between
toxic
and
nontoxic
metal
would
not
detect
that
only
nontoxic
metal
had
been
removed,
but
the
sample­
specific
WER
approach
would.

15
Pigum
1:
Four
Way8
to
Dmrivo
8
Permit
Limit
16
METHOD
1:
DETERMINING
WERs
FOR
AREAS
IN
OR
NEAR
PLUMES
Method
1
is
based
on
the
determination
of
WERs
using
simulated
downstream
water
and
so
it
can
be
used
to
determine
a
WER
that
applies
in
the
vicinity
of
a
plume.
Use
of
simulated
downstream
water
ensures
that
the
concentration
of
effluent
in
the
site
water
is
known,
which
is
important
because
the
magnitude
of
the
WER
will
often
depend
on
the
concentration
of
effluent
in
the
downstream
water.
Knowing
the
concentration
of
effluent
makes
it
possible
to
quantitatively
relate
the
WER
to
the
effluent.
Method
1
can
be
used
to
determine
either
cmcWERs
or
cccWERs
or
both
in
single­
metal,
flowing
freshwater
situations,
including
streams
whose
design
flow
is
zero
and
"
effluent­
dependent"
streams
(
see
Appendix
F).
As
is
also
explained
in
Appendix
F,
Method
1
is
used
when
cmcWERs
are
determined
for
"
large
sites",
although
Method
2
is
used
when
cccWERs
are
determined
for
"
large
sites".
In
addition,
Appendix
F
addresses
special
considerations
regarding
multiple­
metal
and/
or
multiple­
discharge
situations.

Neither
Method
1
nor
Method
2
covers
all
important
methodological
details
for
conducting
the
side­
by­
side
toxicity
tests
that
are
necessary
in
order
to
determine
a
WER.
Many
references
are
made
to
information
published
by
the
U.
S.
EPA
(
1993a,
b,
c)
concerning
toxicity
tests
on
effluents
and
surface
waters
and
by
ASTM
(
1993a,
b,
c,
d,
e,
f)
concerning
tests
in
laboratory
dilution
water.
Method
1
addresses
aspects
of
toxicity
tests
that
(
a)
need
special
attention
when
determining
WERs
and/
or
(
b)
are
usually
different
for
tests
conducted
on
effluents
and
tests
conducted
in
laboratory
dilution
water.
Appendix
H
provides
additional
information
concerning
toxicity
tests
with
saltwater
species.

A.
Experimental
Design
Because
of
the
variety
of
considerations
that
have
important
implications
for
the
determination
of
a
WEB,
decisions
concerning
experimental
design
should
be
given
careful
attention
and
need
to
answer
the
following
questions:
1.
Should
WERs
be
determined
using
upstream
water,
actual
downstream
water,
and/
or
simulated
downstream
water?
2.
Should
WERs
be
determined
when
the
stream
flow
is
equal
to,
higher
than,
and/
or
lower
than
the
design
flow?
3.
Which
toxicity
tests
should
be
used?
4.
Should
a
cmcWER
or
a
cccWER
or
both
be
determined?
5.
How
should
a
FWER
be
derived?
6.
For
metals
whose
criteria
are
hardness­
dependent,
at
what
hardness
should
WERs
be
determined?
The
answers
to
these
questions
should
be
based
on
the
reason
that
WERs
are
determined,
but
the
decisions
should
also
take
into
account
some
practical
considerations.

17
1.
Should
WERs
be
determined
using
upstream
water,
actual
downstream
water,
and/
or
simulated
downstream
water?

a.
Upstream
water
provides
the
least
complicated
way
of
determining
and
using
WERs
because
plumes,
mixing
zones,
and
effluent
variability
do
not
have
to
be
taken
into
account.
Use
of
upstream
water
provides
the
least
useful
WERs
because
it
does
not
take
into
account
the
presence
of
the
effluent,
which
is
the
source
of
the
metal.
It
is
easy
to
assume
that
upstream
water
will
give
smaller
WERs
than
downstream
water,
but
in
some
cases
downstream
water
might
give
smaller
WERS
(
see
Appendix
G).
Regardless
of
whether
upstream
water
gives
smaller
or
larger
WERs,
a
WER
should
be
determined
using
the
water
to
which
the
site­
specific
criterion
is
to
apply
(
see
Appendix
Al.

b.
Actual
downstream
water
might
seem
to
be
the
most
pertinent
water
to
use
when
WERs
are
determined,
but
whether
this
is
true
depends
on
what
use
is
to
be
made
of
the
WERs.
WERs
determined
using
actual
downstream
water
CM
be
quantitatively
interpreted
using
the
sample­
specific
WER
approach
described
at
the
end
of
the
Introduction.
If,
however,
it
is
desired
to
understand
the
quantitative
implications
of
a
WER
for
an
effluent
of
concern,
use
of
actual
downstream
water
is
problematic
because
the
concentration
of
effluent
in
the
water
CM
only
be
known
approximately.

Sampling
actual
downstream
water
in
areas
that
are
in
or
near
plumes
is
especially
difficult.
The
WER
obtained
is
likely
to
depend
on
where
the
sample
is
taken
because
the
WER
will
probably
depend
on
the
percent
effluent
in
the
sample
(
see
Appendix
D).
The
sample
could
be
taken
at
the
end
of
the
pipe,
at
the
edge
of
the
acute
mixing
zone,
at
the
edge
of
the
chronic
mixing
zone,
or
in
a
completely
mixed
situation.
If
the
sample
is
taken
at
the
edge
of
a
mixing
zone,
the
composition
of
the
sample
will
probably
differ
from
one
point
to
another
along
the
edge
of
the
mixing
zone.

If
samples
of
actual
downstream
water
are
to
be
taken
close
to
a
discharge,
the
mixing
patterns
and
plumes
should
be
well
known.
Dye
dispersion
studies
(
Kilpatrick
1992)
are
commonly
used
to
determine
isopleths
of
effluent
concentration
and
complete
mix;
dilution
models
(
U.
S.
EPA
1993d)
might
also
be
helpful
when
selecting
sampling
locations.
The
most
useful
samples
of
actual
downstream
water
are
probably
those
taken
just
downstream
of
the
point
at
which
complete
mix
occurs
or
at
the
most
distant
point
that
is
within
18
the
site
to
which
the
site­
specific
criterion
is
to
apply.
When
samples
are
collected
from
a
complete­
mix
situation,
it
might
be
appropriate
to
composite
samples
taken
over
p
cross
section.
of
the
stream.
Regardless
of
where
it
is
decided
conceptually
that
a
8ample
should
be
taken,
it
might
be
difficult
to
identify
where
the
point
exists
in
the
stream
and
how
it
changes
with
flow
and
over
time.
In
addition,
if
it
i8
not
known
exactly
what
the
sample
actually
repre8ent8,
there
is
no
way
to
know
how
reproducible
the
8­
18
i8.
These
problems
make
it
difficult
to
relate
WERs
determined
in
actual
downstream
water
to'an
effluent
of
concern
because
the
concentration
of
effluent
in
the
sample
is
not
known;
this
is
not
a
problem,
however,.
if
the
sdmple­
specific
WER
approach
is
used
to
interpret
the
results.

c.
­
ated
dwtrem
would
8eem
to
be
the
most
unnatural
of
the
three
kinds
of
water,
but
it
offer8
several
important
advantage8
because
effluent
and
upstream
water
are
mixed
at
a
known
ratio.
Thi8
i8
important
because
the
magnitude
of
the
WRR
will
often
depend
on
the
concentration
of
effluentSin
the
downstream
water.
Mixtures
can
be
prepared
to
8imulate
the
ratio
of
effluent
and
upstream
water
that
uci8ts
at
the
edge
of
the
acute
mixing
zone,
at
the
edge
of
the
chronic
mixing
zone,
at
coanplete
mix,
or
at
any
other
point
of
interest.
If
desired,
a
8ample
of
effluent
can
be
mixed
with
a
smle
on
upstream
water
in
different
ratios
to
simulate
different
point8
in
a
stream.
Also,
the
ratio.
used
can
be
one
that
rimulates
condition8
at
design
flow
or
at
any
other
flow.

The
sample­
specific
WRR
approach
can
be
u8ed
with
both
actual
and
simulated
downstream
water.
Additional
quantitative
use8
can
be
made
of
WERs
determined
using
simulated
downstream
water
because
the
percent
effluent
in
the
water
is
known,
which
allows
quantitative
extrapolations
to
the
effluent.
In
addition,
Llimulated
downstream
water
can
be
used
to
determine
the
variation
in
the
WER
that
is
due
to
variation
in
the
effluent.
0
It
also
allows
comparison
of
two
or
more
effluents
and
determination
of
the
interactions
of
two
or
mOre
effluents.
Additivity
of
WRRs
can
be
studied
using
simulated
downstream
water
(
see
Appendix
G);
8tudie8
of
toxicity
within
plumes
and
studies
of
whether
increased
flow
of
upstream
water
can
increase
toxicity
are
both
studies
of
additivity
of
WERs.
Use
of
simulated
downstream
water
also
make8
it
possible
to
conduct
controlled
studies
of
changes
in
WERs
due
to
aging
and
changes
in
pH.

19
2.
In
Method
1,
therefore,
WERs
are
detexmined
using
simulated
downstream
water
that
is
prepared
by
mixing
8anples
of
effluent
and
upstream
water
in
an
appropriate
ratio.
­
8t
inlportaIltly,
Method
1
Can
be
USad
t0
determine
a
WER
that
applies
in
the
vicinity
of
a
plum
and
can
be
quantitatively
extrapolated
to
the
effluent.

Should
m
be
determined
when
the
8trem
flow
is
equal
to,
higher
than,
and/
or
lower
than
the
design
flow?

WBRS
are
used
in
the
derivation
of
8ite­
8pecific
criteria
when
it
i8
de8iredthatpennit
limit8
be
basedon
a
crituion
that
takes
into
account
the
characteristics
of
the
watu
and/
or
the
metal
at
the
site.
In
mO8t
cases,
permit
limit8
are
calculated
using
8teady­
state
models
and
are
ba8d
on
a
design
flOWa
It
is
therefore
important
that
WERE
he
adequately
protective
under
design­
flow
condition,
which
might
be
­
acted
to
require
that
some
set8
of
8~
aph8
of
l
ffluent
and
uprrtream
water
he
obtained
when
the
actual
stream
flaw
is
clo8e
to
the
derrign
flow.
Collecting
8amgle8
when
the
stream
flow
is
clo8e
to
the
de8ign
flow
will
limit
a
WRR
determination
to
the
low­
flaw
8ea8on
(
e.
g.,
from
mid­
July
to
mid­
October
in
8ome
placea)
and
to
year8
in
which
the
flow
ia
sufficiently
low.

It
58
al80
­
rtMt,
hmever,
that
WRR8
that
are
applied
at
de8ign
flow
provide
adequate
protection
at
higher
flow8.
Generalization8
concerning
the
impact
of
higher
fhw8
on
WRR8
are
difficult
becau8e
8uch
flow8
might
(
a)
reduce
hardnesrr,
alkalinity,
and
pH,
(
b)
increase
or
decrea8e
the
concentrations
of
'
1Dc
and
TSS,
(
c)
reSU8pend
toxic
and/
Or
!
mHx%
iC
metal
from
the
sediment,
and
(
d)
wa8h
additional
pollutant8
into
the
water.
Acidic
mmwmelt,
for
example,
might
lower
the
WRR
both
by
diluting
thm
WER
Md
by
reducing
the
hardness,
alkalinity,
and
@
I;
if
substantial
labile
metal
is
pre8ent,
the
WRR
might
be
lowered
more
than
the
concentration
of
the
metal,
po88ibly
r88ulting
in
increased
toxicity
at
flows
higher
than
d88ign
flow.
Ssnples
taken
at
higher
flows
might
give
­
11~:
WgRrr
because
the
concentration
of
the
effluent
is
more
dilute;
however,
total
recoverable
WERs
might
be
larger
if
the
sample
is
taken
just
after
M
event
that
greatly
increase8
the
concentration
of
TSS
and/
or
TOC
because
thi8
might
increase
both
(
1)
the
concentration
of
nontoxic
particulate
metal
in
the
water
and
(
2)
the
capacity
of
the
water
to
sorb
and
detoxify
m&
al.

WEREI
are
not
of
concern
when
the
stream
flow
is
lower
than
the
de8ign
flow
because
these
are
acknowledged
times
of
reduced
protection.
Reduced
protection
might
not
occur,
however,
if
the
WER
is
sufficiently
high
when
the
flow
is
lower
than
design
flow.

20
3.
Which
toxicity
tests
should
be
used?

a.

b.

C.

d.

e.

f.
As
explained
in
Appendix
D,
the
magnitude
of
an
experimentally
determined
WER
is
likely
to
depend
on
the
sensitivity
of
the
toxicity
test
used.
This
relationship
between
the
magnitude
of
the
WER
and
the
sensitivity
of
the
toxicity
test
is
due
to
the
aqueous
chemistry
of
metals
and
is
not
related
to
the
teat
organisms
or
the
type
of
test.
The
available
data
indicate
that
WERs
determined
with
different
tests
do
not
differ
greatly
if
the
tests
have
about
the
sams
sensitivities,
but
the
data
also
support
the
generalization
that
less
sensitive
toxicity
tests
usually
give
smaller
WERs
than
more
sensitive
test8
(
see
Appendix
D).
when
the
CCC
is
lower
than
the
CMC,
it
is
likely
that
a
larger
WE?
I
will
result
from
tests
that
are
sansitive
at
the
CCC
than
frcm
tests
that
are
sensitive
at
the
CMC.
The
considerations
concerning
the
sensitivities
of
two
tests
should
also
apply
to
two
endpoints
for
the
same
test.
For
any
lethality
test,
use
of
the
LC25
is
likely
to
result
in
a
larger
WER
than
use
of
the
X50,
although
the
difference
might
not
be
measurable
in
most
cases
and
the
LC25
is
likely
to
be
nmre
variable
than
the
X50.
Selecting
the
percent
effect
to
bs
used
to
define
the
endpoint
might
take
into
accotmt
(
a)
whether
the
endpoint
is
above
or
below
the
CMC
and/
or
the
CCC
and
(
b)
the
data
obtained
when
tests
are
conducted.
Once
the
percent
effect
is
selected
for
a
particular
test
(
e.
g.,
a
48­
hr
LCSO
with
l­
day­
old
fathead
SliIlIlOWS~,
the
same
percent
effect
aut
be
usad
whenever
that
test
is
used
to
determine
a
WER
for
that
effluent.
Similarly,
if
two
different
tests
with
the
same
species
(
e.
g.,
a
lethality
test
and
a
sublethal
test)
have
substantially
different
sensitivities,
both
a
uncWER
and
a
CCCWER
could
be
obtained
with
the
same
species.
The
primarv
toxicity
test
used
in
the
determination
of
a
WER
should
have
an
andpoint
in
laboratory
dilution
water
that
is
close
to,
but
pot
low=
than,
the
CMC
and/
or
CCC
to
which
the
WEFI
is
to
be
applied.
Because
the
endpoint
of
the
prinmy
test
in
laboratory
dilution
water
cannot
be
lower
than
the
CMC
and/
or
CCC,
the
magnitude
of
the
WER
is
likely
to
become
closer
to
I
as
the
endpoint
of
the
primary
test
becomes
closer
to
the
CMC
and/
or
CCC
(
see
Appendix
D).
The
WER
obtained
with
the
primary
test
should
be
confirmed
with
a
#
ecow
test
that
uses
a
specie8
that
is
taxonomically
different
frost
the
species
usad
in
the
primary
test.
1)
The
endpoint
of
the
secondary
test
may
be
hiohcr
Jowey
than
the
CMC,
the
CCC,
oi
the
endpoint
of
t%
primary
test.

21
2)
Because
of
the
limited
number
of
toxicity
tests
that
have
sensitivities
near
the
CMC
or
CCC
for
a
metal,
it
seems
unreasonable
to
require
that
the
two
species
be
further
apart
taxonomically
than
being
in
different
orders.
Two
different
endpoints
with
the
same
species
aast
mot
be,
wed
as
the
primary
and
secondary
tests,
even
if
one
enmint
is
lethal
and
the
other
is
sublethal.
g.
If
more
sensitive
toxicity
tests
generally
give
larger
WERS
than
less
sensitive
tests,
the
maximum
value
of
a
WER
will
usually
be
obtained
using
a
toxicity
test
whose
endpoint
in
labor&
tory
dilution
water
equals
the
CncorcK.
If
such
a
test
is
not
used,
the
maximum
possible
WKR
probably
will
not
be
obtained.
h.
No
rationale
exists
to
support
the
idea
that
different
species
or
tests
with
the
dams
sensitivity
will
produce
different
WBR8.
Because
the
mode
of
action
might
differ
fraa
species
to
species
and/
or
from
effect
to
effect,
it
is
easy
to
speculate
that
in
mans
cases
the
magnitude
of
aWERwilldependto
soms
extentonthe
species,
life
stage,
and/
or
kind
of
test,
but
no
data
are
available
to
support
conclusions
concerning
the
existence
and/
or
magnitude
of
any
such
differences.
i.
If
the.
tests
are
otherwise
acceptable,
both
cmcWEBs
and
cccWERs
msy
be
determined
using
acute
and/
or
chronic
tests
and
using
lethal
and/
or
sublethal
endpoints.
The
wrtant
consideration
is
the
sensitivity
of
the
test,
not
the
duration,
species,
life
stage,
or
adverse
effect
used.
j.
There
is
no
reason
to
use
species
that
occur
at
the
8ite;
they
may
be
used
in
the
detexmination
of
a
WER
if
desired,
but
:
1)
It
might
be
difficult
to
determine
which
of
the
species
that
occur
at
the
site
are
sensitive
to
the
metal
and
are
adaptable
to
laboratory
conditions.
2)
Species
that
occur
at
the
site
might
be
harder
to
obtain
in
sufficiant
numbars
for
conducting
toxicity
tests
over
the
testing
period.
3)
Additional
QA
tests
will
probably
be
needed
(
see
section
C.
3.
b)
because
data
are
not
likely
to
be
available
from
other
laboratories
for
comparison
with
the
results
in
laboratory
dilution
water.
k.
Because
a
WER
is
a
ratio
of
results
obtained
with
the
sasm
test
in
two
different
dilution
waters,
toxicity
tests
that
are
used
in
WET
testing,
for
example,
may
be
used,
even
if
the
national
aquatic
life
guidelines
(
U.
S.
EPA
1985)
do
not
allow
use
of
the
test
in
the
derivation
of
an
aquatic
life
criterion.
Of
course,
a
test
whose
endpoint
in
laboratory
dilution
water
is
b+
low
the
CMC
and/
or
CCC
that
is
to
be
adjusted
cannot
be
used
as
a
primary
test.

22
1.
Because
there
is
no
rationale
that
suggest
that
it
makes
any
difference
whether
the
test
is
conducted
with
a
species
that
is
warmwater
or
coldwater,
a
fish
or
an
invertebrate,
or
resident
or
nonresident
at
the­
site,
other
than
the
fact
that
less
sensitive
tests
are
likely
to
give
smaller
WERs,
such
considerations
as
the
availability
of
test
organisms
might
be
important
in
the
selection
of
the
test.
Information
in
Appendix
I,
a
criteria
do
cment
for
the
metal
of
concern
(
see
Appendix
El,
or
any
other
pertinent
source
might
be
useful
when
selecting
primary
and
secondary
tests.
m.
A
test
in
which
the
test
organisms
are
not
fed
might
give
a
different
WER
than
a
test
in
which
the
organisms
are
fed
just
because
of
the
presence
of
the
food
(
see
Appendix
D):
This
might
depend
on
the
metal,
the
type
and
amount
of
food,
and
whether
a
total
recoverable
or
dissolved
WER
is
datemined.
Different
tests
with
similar
sensitivities
are
expected
to
give
similar
WERs,
except
for
experimental
variation.
The
purpose
of
the
secondary
test
is
to
provide
information
concerning
this
assumption
and
the
validity
of
the
WER.

4.
Should
a
ascWER
or
a
CCCWER
or
both
be
detexmined?

This
question
does
not
have
to
be
answered
if
the
criterion
for
the
site
contains
either
a
CMC
or
.
a
CCC
but
not
both.
For
axamDle,
a
body
of
water
that
is
protected
for
put­
and­
take
fishing
might
have
only
a
CMC,
yhereas
a
stream
whose
design
flow
is
zero
might
have
only
a
CCC.

When
the
criterion
contains
both
a
CMC
and
a
CCC,
the
simplistic
way
to
answer
the
question
is
to
determine
whether
the
CXC
or
the
CCC
controls
the
existing
permit
limits;
which
one
is
controlling
depends
on
(
a)
the
ratio
of
the
CMC
to
the
CCC,
(
b)
whether
the
number
of
mixing
zones
is
zero,
one,
or
two,
and
(
c)
which
steady­
state
or
dynamic
model
was
used
in
the
calculation
of
the
pennit
limits.
A
better
way
to
answer
the
question
would
bs
to
also
determine
how
much
the
controlling
value
would
have
to
be
changed
for
the
other
value
to
becoms
controlling;
this
might
indicate
that
it
would
not
be
cost­
effective.
to
derive,
for
example,
a
site­
specific
CMC
(
ssCMC)
without
also
deriving
a
site­
specific
CCC
(
ssCCC1.
There
are
also
other
possibilities:
(
1)
It
might
be
appropriate
to
use
a
phased
approach,
i.
e.,
determine
either
the
cmcwER
or
the
cccWER
and
then
decide
whether
to
determine
the
other.
(
2)
It
might
be
appropriate
and
environmentally
conservative
to
determine
a
WER
that
can
be
applied
to
both
the
CMC
and
the
CCC.
(
3)
It
is
always
allowable
to
determine
and
use
both
a
ancWER
and
a
CCCWER,
although
both
can
be
determined
only
if
toxicity
tests
with
appropriate
sensitivities
are
available.

23
Because
the
phased
approach
can
always
be
used,
it
is
only,
inportant.
to
decide
whether
to
use
a
different
approach
when
its
use
might
be
cost­
effectivea
De&
ding
whether
to
use
a
different
approach
and
selecting
which
one
to
use
is
cwlex
because
a
number
of
considerations
need
to
be
taken
into
account:
:
a.
Is
the
CMC
equal
to
or
higher
than
the
CCC?
If
the
CMC
equals
the
CCC,
two
WERs
cannot
be
determined
if
they
would
be
determined
using
the
saam
site
water,
but
two
WERs.
could
be
determined
if
the
mcWER
and
the
CCCWER
would
be
determined
using
different
site
waters,
e.
g.,
waters
that
contain
different
concentrations
of
the
effluent.
b.
If
the
CM2
is
higher
than
the
CCC,
is
there
a
toxicity
te8t
whose
endpoint
in
laboratory
dilution
water
is
between
the
CMC
and
the
CCC?
If
the
CMC
is
higher
than
the
CCC
and
there
is
a
toxicity
test
whose
endpoint
in
laboratory
dilution
water
is
between
the
CMC
and
the
CCC,
both
a
cmcWER
anda
cccWER
canbe
determined.
If
the
CE
is
higher
than
the
CCC
but
no
toxicity
test
has
M
endpoint
in
laboratory
dilution
water
between
the
C24C
and
the
CCC,
two
WERs
cannot
be
determined
if
they
would
be
detemained
using
the
sams
site
water;
two
WERs
could
be
determined
if
they
were
determined
using
different
site
waters,
e.
g.,
waters
that
contain
different
.
concentrations
of
the
effluent.
c.
Was
a
steady­
state
or
a
dynamic
model
used
in
the
calculation
of
the
pexmit
limits?
I&
is
cmlex,
but
reasonably
clear,
how
to
make
a
decision
when
a
steady­
state
model
was
used,
but
it
is
not
clear
how
a
decision
should
bemade
when
a
dynamicmodelwas
used.
d.
If
a
steady­
state
model
was
used,
were
one
or
two
design
flows
used,
i.
e.,
was
the
hydrologically
based
steady­
state
method
used
or
was
the
biologically
based
steady­
state
method
used?
When
the
hydrologically
based
msthod
is
used,
one
design
flow
is
used
for
both
the
CMC
and
the
CCC,
whereas
when
the
biologically
based
method
is
used,
there
is
a
CK
design
flow
and
a
CCC
design
flow.
1
When
WERs
are
determined
using
downstream
water,
use
of
the
biologically
based
method
will
probably
cause
the
percent
effluent
in
the
site
water
used
in
the
determination
of
the
cmcWER
to
ba
different
from
the
percent
effluent
in
the
site
water
used
in
the
determination
of
the
CCCWER;
thus
the
two
WEFU
should
ba
determined
using
two
different
site
waters.
This
does
not
impact
WEFT
detemined
using
upstream
water.

24
e.
Is
there
M
acute
mixing
zone?
Is
there
a
chronic
mixing
zone?
1.
When
WEEU
are
determined
using
upstream
water,
the
Dresence
or
absence
of
mixing
zones
has
no
impact;
the
cxncWEZ4
and
the
cccWER,
will
both
be
determined
using
site
water
that
contains
zero
percent
effluent,
i.
e.,
the
two
WERs
will
be
determined
using
the
same
site
water.
2.
Even
when
downstream
water
is
used,
whether
there
is
M
acute
mixing
zone
affects
the
point
of
application
of
the
CMC
or
SSCMC,
but
it
does
not
affect
the
determination
of
any
WER.
3.
The
axistence
of
a
chronic
mixing
zone
has
important
implications
for
the
determination
of
WERs
when
downstream
water
is
used
(
see
Appendix
A)
.
When
WERs
are
determined
using
downstream
water,
the
cmcwER
should
be
determined
using
water
at
the
edge
of
the
chronic
mixing
zone,
whereas
the
cccWER
should
be
determined
using
water
froan
a
cwlete­
mix
situation.
(
If
the
biologically
based
method
is
wed,
the
two
different
design
flows
should
also
be
taken
into
account
when
determining
the
percent
effluent
that
should
be
in
the
simulated
downstream
water.)
Thus
the
percent
effluent
in
the
site
water
used
in
the
determination
of
the
cxncWEF!
will
be
different
from
the
percent
effluent
in
the
site
water
used
in
the
determination
of
the
CCCWER;
this
is
important
because
the
magnitude
of
a
WER
will
oftan
depend
substantially
on
the
percent
effluent
in
the
water
(
see
Appendix
D).
f.
In
what
situations
would
it
be
environmentally
consemative
to
determine
one
WER
and
use
it
to
adjust
both
the
cmcWER
and
the
cccWER?
Because
(
1)
the
CMC
is
never
lower
than
the
CCC
and
(
2)
a
more
sensitive
test
will
generally
give
a
WER
closer
to
1,
it
will
be
environmentally
conservative
to
use
a
cmcWER
to
adjust
a
CCC
when
there
are
no
contradicting
considerations.
In
this
case,
a
cmcWER
CM
be
determined
and
used
to
adjust
both
the
CMC
and
the
CCC.
Because
water
quality
CM
affect
the
WER,
this
approach
is
necessarily
valid
only
if
the
cmcWER
and
the
CCCWER
are
determined
in
the
same
site
water.
Other
situations
in
which
it
would
be
environmentally
conservative
to
use
one
WER
to
adjust
both
the
CMC
and
the
CCC
are
described
below.
These
considerations
have
one
set
of
implications
when
both
the
cmcWER
and
cccWER
are
to
be
determined
using
the
same
site
water,
and
Mother
set
of
ixr@
lications
when
the
two
WERs
are
to
be
determined
using
different
site
waters,
e.
g.,
when
the
site
waters
contain
different
concentrations
of
effluent.

25
When
WERS
are
determined
using
B
water,
the
same
site
water
is
used
in
the
determination
of
both
the
ancm
and
the
CCCWER.
Whenever
the
two
WERs
are
determined
in
the
sam
site
water,
any
difference
in
the
magnitude
of
the
cmcWER
and
the
cccW'ER
will
probably
be
due
to
the
sensitivities
of
the
toxicity
tests
used.
Therefore
:
a.

b.

c.

d.

e.

f.

Q*
If
more
sensitive
toxicity
tests
generally
give
larger
WERS
than
less
sensitive
tests,
the
maximum
CccWER
(
a
cccWER
determined
with
a
terrt
whose
endpoint
equals
the
CCC).
will
usually
be
larger
than
the
maximum
cmcm
because
the
CCC
is
never
higher
than
the
CMC.
Because
the
CCC
is
nwer
higher
than
the
CMC,
the
maximum
cmcWER
will
usually
be
smaller
than
the
m
ccc#
ER
and
it
will
be
environmentally
conservative
to
use
the
cmcWER
to
adjust
the
CCC.
A
CCCWER
CM
be
detemined
8apar8tely
from
a
CmcwER
only
if
there
is
a
tOXiCity
test
with
M
endpoint
in
laboratory
dilution
water
that
i8
between
the
CMC
and
the
CCC.
If
no
such
test
exists
or
CM
be
dwised,
only
a
cmcWER
can
be
determined,
but
it
CM
be
usad
to
adjust
both
the
CE
and
the
CCC.
Unless
the
experimental
variation
is
increased,
use
of
a
CCCWER,
insteadof
a
awWRR,.
to
adjust
the
CCC
till
usually
*
rove
the
accuracy
of
the
resulting
site­
specific
CCC.
Thus
a
cccWER
may
be
determined
and
used
whenever
desired,
if
a
toxicity
t88t
has
M
endpoint
in
laboratory
dilution
water
between
the
CMC
and
the
CCC.
A
CCCWER
cannot
ba
used
to
adjust
a
CMC
if
the
cccWR3
was
detexmined
using
M
endpoint
that
was
lower
than
the
CE
in
laboratory
dilution
water
because
it
will
probably
reduce
the
level
of
protection.
Even
if
there
is
a
toxicity
test
that
has
M
endpoint
in
laboratory
dilution
water
that
is
between
the
CMC
and
the
CCC,
it
is
not
necessary
to
decide
initially
whether
to
determine
a
CSCWER
and/
Or
a
CCCWER.
When
upstream
water
is
used,
it
is
always
allowable
to
determine
a
QllcwER
and
use
it
to
derive
a
site­
specific
CMC
and
a
site­
specific
CCC
and
then
decide
whether
to
determine
a
ccc­.
If
there
is
a
toxicity
test
whose
endpoint
in
laboratory
dilution
water
is
between
the
CCC
and
the,
CMC,
and
if
this
test
is
used
as
the
secondary
test
in
the
determination
of
the
cmcWER,
this
test
will
provide
information
that
should
be
very
useful
for
deciding
whether
to
detemine
a
cccWER
in
addition
to
a
cmcWEX.
Further,
if
it
is
decided
to
determine
a
CCCW'ER,
the
sams
two
tests
used
in
the
detexmination
of
the
QllcwER
could
then
be
used
in
the
determination
of
the
cccWER,
with
a
rwersal
of
their
roles
as
primary
and
secondary
tests.
Alternatively,
a
cmcWER
and
a
CCCWER
could
be
determined
simultaneously
if
both
tests
are
conducted
on
each
sample
of
site
water.

26
When
WERs
are
determined
using
m
water,
the
magnitude
of
each
WER
will
probably
depend
on
the
concentration
of
effluent
in
the
downstream
water
used
(
see
Appendix
9)
l
The
first
important
consideration
is
whether
the
design
flow
is
greater
than
zero,
and
the
second
is
whether
there
is
a
chronic
mixing
zone.
a.
If
the
design
flow
is
zero,
cmcWERs
and/
or
cccWERs
that
are
determined
for
design­
flow
conditions
will
both
be
determined
in
100
percent
effluent.
Thus
this
case
is
similar
to
using
upstream
water
in
that
both
WERs
are
determined
in
the
same
site
water.
When
WERs
are
determined
for
high­
flow
conditions,
it
will
make
a
difference
whether
a
chronic
mixing
zone
needs
to
be
taken
into
account,
which
is
the
second
consideration.
b.
If
there
is
no
chronic
mixing
zone,
both
WERs
will
be
determined
for
the
complete­
mix
situation;
this
case
is
similar
to
using
upstream
water
in
that
both
WERs
are
detexmined
using
the
same
site
water.
If
there
is
a
chronic
mixing
zone,
cmcWERs
should
be
detemined
in
the
site
water
that
exists
at
the
edge
of
the
chronic
mixing
zone,
whereas
cccWERs
should
be
determined
for
the
cmlete­
mix
situation
(
see
Appendix
A).
Thus
the
percent
effluent
will
be
higher
in
the
site
water
wed
in
the
determination
of
the
cmcWER
than
in
the
site
water
used
in
the
determination
of
the
cccwH1.
Because
a
site
water
with
a
higher
percent
effluent
will
probably
give
a
larger
WER
than
a
site
water
with
a
lower
percent
effluent,
both
a
­
and&
CccWERcan
be
determined
even
if
there
is
no
test
whose
endpoint
in
laboratory
dilution
water
is
between
the
CMC
and
the
ccc.
There
are
opposing
considerations,
however:
1)
The
site
water
used
in
the
detemination
of
the
cmcm
will
probably
have
a
higher
percent
effluent
than
the
site
water
used
in
the
determination
of
the
cccWEFt,
which
will
tend
to
cause
the
cmcWER
to
be
larger
than
the
cccWER.
2)
If
there
is
a
toxicity
test
whose
endpoint
in
laboratory
dilution
water
is
between
the
CMC
and
the
CCC,
use
of
a
more
sensitive
test
in
the
determination
of
the
CCCWER
will
tend
to
cause
the
CCCWER
to
ba
larger
than
the
cmcWER.
One
consequence
of
these
opposing
considerations
is
that
it
is
not
known
whether
use
of
the
cmcWER
to
adjust
the
CCC
would
be
environmentally
conservative;
if
this
simplification
is
not
known
to
be
conservative,
it
should
not
be
used.
Thus
it
is
wortant
whether
there
is
a
toxicity
test
whose
endpoint
in
laboratory
dilution
water
is
between
the
CMC
and
the
CCC:
a.
If
no
toxicity
test
has
an
endpoint
in
laboratory
dilution
water
between
the
CMC
and
the
CCC,
the
two
WERs
have
to
be
determined
with
the
same
test,
in
which
case
the
ancWEF4
will
probably
be
larger
because
the
27
percent
effluent
in
the
site
water
will
be
higher.
Because
of
the
difference
in
percent
effluent
in
the
site
waters
that
should
be
wed
in
the
determinations
of
the
two
WER8,
use
of
the
QllcwER
to
adjust
the
CCC
would
not
be
enviroxnnentally
corrrrervative,
but
use
of
the
cccWER
to
adjust
the
CWC
would
be
environmentally
coa8823ntive.
Although
both
WER8
could
be
determined,
it
would
also
be
acceptable
to
determine
only
the
CCCWKR
and
u8e
it
to
adjust
both
the
CMC
and
the
CCC.
b.
If
there
is
a
toxicity
test
whose
endpoint
in
laboratory
dilUtiOn
water
is
between
the
CMC
and
the
CCC,
the
two
WERs
could
be
detenained
using
different
toxicity
tests.
An
environwntally
conservative
alternative
to
deterxnining
two
WERs
would
be
to
determine
a
hybrid
WKR
by
using
(
1)
a
toxicity
test
whose
endpoint
is
above
the
Ct4C
(
i.
e.,
a
toxicity
test
that
ir
appropriate
for
the
determination
of
a
CmcwER)
and
(
2)
8ite
water
for
the
complete­
mix
situation
(
i.
e.,
8ite
water
appropriate
for
the
determination
of
cccwm
.
Ituouldbe
eavir
onmentally
conservative
to
use
thir
hybrid
WER
to
adjust
the
CbC
and
it
would
be
awiroaswntally
consenmtive
to
use
this
hybrid
WER
to
adjut
the
CCC.
Although
both
WERs
could
be
determined,
it
would
also
be
acceptable
to
determine
only
the
­
rid
WE!
R
and
use
it
to
adjust
both
the
CMC
andtheCCC­
(
This
hybrid
WER
described
here
in
paragraph
b
is
the
same
as
the
CCCWER
described
in
paragraph
a
above
in
which
no
toocicity
test
had
M
endpoint
in
laboratory
dilution
water
between
the
CMC
andtheCCC.
1
5.
EowrhouldaPWERbe
derived?

Because
of
­
imental
variation
and
variation
in
the
ccSpo8ition
of
8urface
waters
and
effluent8,
a
single
deterskination
of
a
WER
does
not
provide
sufficient
information
to
justify
adjustment
of
a
criterion.
After
a
8ufficient
nulaber
of
WER8
have
been
determined
in
M
acceptable
manner,
a
Final
Water­
Effect
Ratio
(
FWER)
ial
derived
from
the
WERs,
and
the
FWER
is
then
used
to
calculate
the
8ite­
specific
criterion.
If
both
a
site­
specific
CMC
and
a
8ite­
8pecific
CCC
are
to
be
derived,
both
a
cmcFWER
and
a
cccFWER
have
to
be
derived,
unless
M
enviroxnnentally
conservative
estimate
is
used
in
place
of
the
ancFWER
and/
or
the
ccc­.

When
a
WER
is
determined
using
upstream
water,
the
two
major
sources
of
variation
in
the
WER
are
(
a)
variability
in
the
quality
of
the
upstream
water,
much
of
which
might
be
related
to
season
and/
or
flow,
and
(
b)
experimental
20
variation.
When
a
WE3
is
determined
in
downstream
water,
the
four
major
sources
of
variation
are
(
a)
variability
in
the
quality
of
the
upstream
water,
much
of
which
might
be
related
to
season
and/
Or
flow,
(
b)
experimental
variation,
(
c)
variability
in
the
colm(
position
of
the
effluent,
and
(
d)
variability
in
the
percept
effluent
in
the
downstream
water.
Variability
and
the
possibility
of
mistakes
and
rare
events
make
it
necessary
to
try
to
comprotnise
betwaen
(
1)
providing
a
high
probability
of
adequate
protection
ahd
(
2)
placing
too
much
reliance
0"
the
smallest
experimentally
determined
WER,
which
might
reflect
experimental
variation,
a
mistake,
or
a
rare
went
rather
than
a
meaningful
difference
in
the
WER.

Various
ways
CM
be
employed
to
address
variability:
a.
Replication
CM
be
used
to
reduce
the
is@
act
of
soma
sources
of
variation
and
to
verify
the
importance
of
others.
b.
Because
variability
in
the
composition
of
the
effluent
might
contribute
substantially
to
the
variability
of
the
WER,
it
might
be
desirable
to
obtain
and
store
two
or
more
samples
of
the
effluent
at
slightly
different
times,
with
the
selection
of
the
saxrgding
times
depending
on
such
characteristics
of
the
discharge
as
the
average
retention
time,
in
case
M
unusual
WEF4
is
obtained
with
the
first
sample
used.
c.
Because
of
the
possibility
of
mistakes
and
rare
wants,
samples
of
effluent
and
upstream
water
should.
be
large
enough
that
portions
can
be
stored
for
later
testing
or
analyses
if
M
unusual
WER
is
obtained.
d.
It
might
be
possible
to
reduce
the
iqmct
of
the
variability
in
the
percent
effluent
in
the
downstream
water
by
establishing
a
relationship
between
the
WER
and
the
percent
effluent.
Confounding
of
the
sources
CM
be
a
problem
when
more
than
one
source
contributes
substantial
variability.

When
pennit
limits
are
calculated
using
a
steady­
state
model,
the
limits
are
based
on
a
design
flow,
e.
g.,
the
7410.
It
is
usually
assuwd
that
a
concentration
of
metal
in
M
effluent
that
does
not
cause
unacceptable
effects.
at
the
design
flow
will
not
cause
unacceptable
effects
at
higher
flows
because
the
smtal
is
diluted
By
the
increased
flow
of
the
upstream
water.
Dacreased
protection
might
occur,
however,
if
M
increase
in
flow
increases
toxicity
more
than
it
dilutes
the
concentration
of
metal.
When
permit
limits
are
based
on
a
national
criterion,
it
is
often
assumed
that
the
criterion
is
sufficiantly
conservative
that
M
increase
in
toxicity
till
not
ba
great
enough
to
oven&
elm
the
combination
of
dilution
and
the
assumed
conservatism,
wen
though
it
is
likely
that
the
national
criterion
is
not
overprotective
of
all
bodies
29
of
water.
When
WEIR?
are
used
to
reduce
the
assumed
conservatism,
there
is
more
concern
about
the
possibility
of
increased
toxicity
at
flows
higher
than
the
design
flow
and
it
is
ixqmrtant
to
(
1)
determine
SOWB
WERs
thal:
correspond
to
higher
flows
or
(
2)
provide
some
coMenmtism.
If
the
concentration
of
effluent
in
the
downstream
water
decreases
as
flow
increases,
wets
determined
at
higher
flows
are
likely
to
be
smaller
than
WERs
deters&
n
ed
at
design
flow
but
the
concentration
of
nrstal
will
also
be
lower.
If
the
concentration
of
TSS
increases
at
high
flcmm,
however,
both
the
WER
and
the
concentration
of
as&
al
might
increase.
If
they
are
determine&
in
an
appropriate
manner,
WERs
detexmined
at
flows
higher
than
the
design
flow
CM
be
used
in
two
ways:
a.
As
eavi
ronmntally
conservative
e8timates
of
wER8
determined
at
design
flow.
b.
m
asaeas
whether
WERs
determined
at
design
flow
will
provide
adequate
protection
at
higher
flows.

In
order
to
appropriately
take
into
account
seaso~
l
and
flow
effects
and
their
interactions,
both
ways
of
using
high­
flow
WBRs
require
that
the
downstream
water
used
in
the
detgmaination
of
theWKRbe
similar
to
that
which
actually
exist8
during
the
time
of
concern.
In
addition,
high­
flow
WERs
can
ba
usad
in
the
second
way
only
if
the
composition
of
the
downstream
watar
is
known.
To
satisfy
the
requiresmnts
that
(
a)
the
downstream
water
used
in
the
determination
of
a
WgR
be
similar
to
the
actual
water
and
(
b)
the
cmposition
of
the
downstream
water
be
known,
it
is
necessary
to
obtain
samples
of
effluent
and
upstream
water
at
the
time
of
concern
and
to
prepare
a
simulated
downstream
water
by
mixing
the
swles
at
the
ratio
of
the
flows
of
the
effluent
and
the
upstream
water
that
existed
when
the
s­
lea
were
obtained.

For
the
first
way
of
using
high­
flow
WERs,
they
are
used
directly
a8
eavi
ronmentally
consenmtive
estimates
of
the
design­
flow
WER.
For
the
second
way
of
using
high­
flow
WER8,
each
ir
u8ed
to
calculate
the
highest
concentration
of
amtal
that
could
be
in
the
effluent
without
causing
the
concentration
of
metal
in
the
downstream
water
to
exceed
the
site­
specific
criterion
that
would
be
derived
for
that
water
using
the
mcperimentally
determined
WER.
This
highest
concentration
of
m&&
l
in
the
8ffluMt
UEME)
CM
be
calculated
as:

where:
ccc
=
the
national,
state,
or
recalculated
CCC
(
or
C24C)
that
is
to
be
adjusted.

30
aFLOW
P
the
flow
of
the
effluent
that
was
the
basis
of
the
preparation
of
the
simulated
downstream
water.
This
should
be
the
flow
of
the
effluent
that
existgd
when
the
samples
were
taken.
uFLOW
=
the
flow
of
the
upstream
water
that
was
the
basis
of
the
preparation
of
the
simulated
downstream
water.
This
should
be
the
flow
of
the
upstream
water
that
existed
when
the
samples
were
taken.
uCONC
=
the
concentration
of
metal
in
the
sample
of
upstream
water
used
in
the
preparation
of
the
simulated
doivnstream
water.
In
order
to
calculate
a
HCME
from
M
experimentally
determined
WER,
the
only
information
needed
besides
the
flows
of
the
effluent
and
the
upstream
water
is
the
concentration
of
metal
in
the
upstream
water,
which
should
be
measured
anyway
in
conjunction
with
the
determination
of
the
WER.

When
a
steady­
state
model
is
used
to
derive
permit
limits,
the
limits
on
the
effluent
apply
at
all
flows;
thus,
each
HCME
CM
be
used
to
calculate
the
highest
WER
(
hwER)
that
could
be
used
to
derive
a
site­
specific
criterion
for
the
downstream
water
at
design
flow
so
that
there
would
be
adequate
protection
at
the
flow
for
which
the
HCME
was
determined.
The
hWER
is
calculated
as:

The
suffix
'
dfm
indicates
that
the
values
used
for
these
quantities
in
the
calculation
of
the
hwER
are
those
that
exist
at
design­
flow
conditions.
The
additional
datum
needed
in
order
to
calculate
the
hWER
is
the
concentration
of
metal
in
upstream
water
at
design­
flow
conditions;
if
this
is
assumed
to
be
zero,
the
hwER
will
be
environmentally
conservative.
If
a
WER
is
determined
when
uFLOW
equals
the
design
flow,
hWE3
=
WER.

The
two
ways
of
using
WERs
determined
at
flows
higher
than
design
flow
CM
be
illustrated
using
the
following
examples.
These
exsmplea
were
formulated
using
the
concept
of
additivity
of
WERs
(
see
Appendix
G).
A
WEB
determined
in
downstream
water
consists
of
two
components,
one
due
to
the
effluent
(
the
eWER)
and
one
due
to
the
upstream
water
(
the
uWER)..
If
the
eWER
and
uWER
are
strictly
additive,
when
WERs
are
determined
at
various
upstream
flows,
the
downstream
WERs
can
be
calculated
from
the
composition
of
the
downstream
water
(
the
%
effluent
and
the
%
upstream
water)
and
the
two
WERs
(
the
eWER
and
the
uWEX)
using
the
equation:

31
wgp
I
1%
effhent)
(
eWgR1
l
(%
umtrum
utter)
(
um8)

100
In
the
emsplea
below,
it
is
assumed
that:
a.
A
site­
specific
CCC
is
being
derived.
;.
2
M&
zl,;
cc
is
2
ug/
L.

d:
The
eWZR
and
uk
are
COllStMt
and
strictly
additive.
8:
The
flaw
of
the
effluent
(
eFLCW)
is
always
10
cfs.
f.
The
design
flaw
of
the
upstream
titer
(
uFT,,
Wdf)
is
40
cfr.
Therefore:

ar
((
2
w/
L)
mm
(
10
Cf8
+
wuml
­
[(
ucmc)
(
uaq
10
w/
L
.

B
m
(
ffam
(
10
cm
+
(
ucmnma
(
40
Cf8)
(
2
w/
L)
(
10
Cf8
+
40
cf8)
l
In
the
first
­
18,
the
USER
is
assumed
to
be
5
and
SO
the
upstream
site­
specific
CCC
(
wsCCC)
­
(
CCC)
(
uWW)
=
(
2
ug/
L)(
S)
­
10
UQ/
L.
whichamaM
uCONC
is
assumed
to
be
0.4
ug/
L,

water
is
that
the
assimilative
capacity
of
the
upstream
9.6
ug/
L.

40
20.0
80.0
12.000
118.4
12.00
63
13.7
06.3
9.795
140.5
14.21
90
10.0
90.0
8.500
166.4
16.80
190
5.0
95.0
6.750
262.4
26.40
490
2.0
98.0
5.700
550.4
55.20
990
1.0
99.0
5.350
1030.4
103.20
1990
0.5
99.5
5.175
1990.4
199.20
.
At
wte
Mxx
hWER
As
the
flaw
of
the
upstream
water
increases,
the
WER
decreases
to
a
limiting
value
equal
to
uWER.
Because
the
assimilative
capacity
is
greater
than
zero,
the
HCMEs
&
d
hWERs
increase
due
to
the
increased
dilution
of
the
effluent.
The
ixkrease
in
hWER
at
higher
flows
will
not
allw
any
use
of
the
assimilative
capacity
of
the
upstream
water
because
the
allowed
concentration
of
metal
in
the
effluent
is
controlled
by
the
lwest
hWER,
which
is
the
design­
flaw
hwER
in
this
­
le.
Any
WER
determined
at
a
higher
flaw
CM
be
used
as
M
MVirOIImMt&
lly
conservative
estimate
of
the
design­
flaw
WER,
and
the
hWERs
show
that
the
WER
of
12
provides
adequate
protection
at
all
flows.
Whti
uFLCW
equals
the
design
flow
of
40
cfs,
WER
=
hwER.

32
In
the
second
example,
uWER
is
assumed
to
be
1,
which
means
that
ussCCC
=
2
ug/
L.
uCONC
is
assumed
to
be
2
ug/
L,
so
that
uCONC
=
ussCCC.
The
assimilative
capacity
of
the
upstream
water
is
0
ug/
L.
.
At
Cmlete
Mix
hWER
Eizxz­
zrEff.
m
LEE
*­,

10
40
20.0
80.0
8.800
80.00
8.800
10
63
13.7
86.3
6.343
80.00
8.800
ix
190
90
10.0
90.0
95.0
4.900
2.950
80.00
8.800
10
490
2'*:
98.0
1.780
80.00
8.800
10
990
1:
o
99.0'
1.390
80.00
8.800
10
1990
0.5
99.5
1.195
80.00
8.800
All
the
WERs
in
this
example
are
lower
than
the
c­
arable
WERs
in
the
first
example
because
the
uWER
dropped
from
5
to
1;
the
limiting
value
of
the
F
at
very
high
flow
is
1.
Also,
the
HCMEs
and
hWERs
are
independent
of
flow
because
the
increased
dilution
does
not
allow
any
more
metal
to
be
discharged
when
uCONC
0:
ussCCC,
i.
e.,
when
the
assimilative
capacity
is
zero.
As
in
the
first
example,
any
WER
determined
at
a
flow
higher
than
design
flow
CM
be
used
as
M
environmentally
conservative
estimate
of
the
design­
flow
WER
and
the
hWERs
show
that
the
WER
of
8.8
determined
at
design
flow
will
provide
adequate
protection
at
all
flows
for
which
infonaation
is
available.
When
uFLOW
equals
the
design
flow
of
40
cfs,
WER
=
hWER.

In
the
third
example,
uWER
is
assumed
to
be
2,
which
Mets
that
ussccc
=
4
ug/
L.
uCONC
is
assumed
to
be
1
ug/
L;
thus
the
assimilative
capacity
of
the
upstream
water
is
3
ug/
L.

10
i8
10
10
10
10
Ez
40
E!
190
490
990
1990
ete
Bbx
4Eff.
k5
20.0
80.0
9.600
92.0
9.60
13.7
86.3.
7.206
98.9
10.29
10.0
90.0
5.800
107.0
11.10
5.0
95.0
3.900
137.0
14.10
2.0
98.0
2.760
227.0
23.10
1.0
99.0
2.380
377.0
38.10
0.5
99.5
2.190
677.0
68.10
hWER
All
the
WERs
in
this
axaz@
e
are
intermediate
between
the
comparable
WE?
U
in
the
first
two
axamples
because
the
uWER
is
now
2,
which
is
between
1
and
5;
the
limiting
value
of
the
WEZ
at
very
high
flow
is
2.
As
in
the
other
examples,
any
WER
determined
at
a
flow
higher
than
design
flow
can
be
used
as
M
environmentally
conservative
estimate
of
the
33
design­
flow
WER
and
the
hWERs
show
that
the
WER
of
9.6
determined
at
design
flow
will
provide
adequate
protection
at
all
flows
for
which
infomation
is
available.
when
UFIxks
equals
the
design
flow
of
40
cfs,
WER
=
hwm.

If
this
third
eXaUQle
i8
88mmed
to
be
subject
to
acidic
snoumalt
in
the
spring
so
that
the
eWER
and
uWER
are
less­
than­
additive
and
result
in
a
WER
of
4.8
(
rather
than
5.8)
at
a
UpIxks
of
90
cfs,
the
lthird
HCME
would
be
87
ug/
L,
and
the
third
hWBR
would
be
9.1.
This
hWER
is
lower
than
the
design­
flaw
WKR
of
9.6,
SO
the
site­
specific
criterion
would
have
to
be
derived
using
the
WER
of
9.1,
rather
than
the
design­
flaw
WER
of
9.6,
in
order
to
provide
the
intended
lwel
of
protection.
If
the
eWER
and
uWER'were
lesa­
than­
additive
only
to
the
extent
that
the
third
WER
was
5.3,
the
third
HCME
would
be
97
ug/
L
and
the
third
hWER
would
be
10.1.
In
this
case,
dilution
by
the
increased
flaw
wuld
xmre
than
compensate
for
the
WEE&
being
less­
than­
additive,
so
that
the
design­
flaw
WER
of
9.6
would
provide
adeguate
protection
at
a
uF%
W?
of
90
cfs.
Auxiliary
information
might
indicate
whether
M
unusual
WER
is
real
or
is
M
accident;
for
­
la,
if
the
hardness,
alkalinity,
and
pH
of
8noWs@
lt
are
all
law,
this
information
would
support
a
low
WER.

If
the
eWER
and
USER
were
nrore­
than­
additive
80
that
the
third
WER
was
10,
this
WER
would
not
be
M
environmentally
conservative
estimate
of
the
design­
flaw
WER.
If
a
WE2
determined
at
a
higher
flow
is
to
be
used
as
M
estimate
of
the
design­
flaw
WER
and
there
is
reason
to
beliwe
that
the
eWER
and
the
uWER
might
be
more­
than­
additive,
a
test
for
additivity
CM
be
performed
(
see
Appendix
G).

Calculating
HCMBs
and
hWERs
is
st'raightforward
if
the
WEFU
are
based
on
the
total
recoverable
msasurement.
If
they
are
based
on
the
dissolved
measurement,
it
is
necessary
to
take
into
account
the
percent
of
the
total
recoverable
metal
in
the
effluent
that
becomes
dissolved
in
the
dwnstream
water.

To
ensure
adequate
protection,
a
group
of
WERs
should
include
one
or
smre
WERs
corresponding
to
flows
near
the
design
flow,
as
well
as
one
or
more
WERs
corresponding
to
higher
flows.
a.
Calculation
of
hWERs
from
WERs
determined
at
various
flows
and
seasons
identifies
the
highest
WER
that
CM
be
used
in
the
derivation
of
a
site­
specific
criterion
and
still
provide
adequate
protection
at
all
flows
for
which
WERs
are
available.
Use
of
hWERs
eliminates
the
need
to
assums
that
WERs
determined
at
design
flow
will
provide
adequate
protection
at
higher
flows.
Because
hWERs
are
calculated
to
apply
at
design
flow,
they
34
apply
to
the
flow
on
which
the
permit
limits
are
based.
The
lwest
of
the
hWERs
ensures
adequate
protection
at
all
flows,
if
hWERs
are
available
for
a
sufficient
range
of
flows,
seasons,
and
other
conditions.
b.
Unless
ad&
tivity
is
assumed,
a
WER
cannot
ba
extrapolated
from
one
flw
to
another
and
therefore
it
is
not
possible
to
predict
a
design­
flow
WER
from
a
WER
determined
at
other
conditions.
The
largest
WER
is
likely
to
occur
at
design
flow
because,
of
the
flows
during
which
protection
is
to
be
provided,
the
design
flaw
is
the
flow
at
which
the
highest
concentration
of
effluent
will
probably
occur
in
the
dwnstream
water.
This
largest
WER
has
to
be
e%
perintentally
determined;
it
cannot
be
predicted.

The
­
18s
also
illustrate
that
if
the
concentration
of
mstal
in
the
upstream
water
is
below
the
site­
specific
criterion
for
that
water,
in
the
limit
of
infinite
dilution
of
the
effluent
with
upstream
water,
there
will
be
adequate
protection.
The
concern,
therefore,
is
for
intermediate
levels
of
dilution.
Even
if
the
assimilative
capacity
is
zero,
as
in
the
second
wle,
there
is
more
concern
at
the
lwer
or
intemsdiate
flws,
when
the
effluent
load
is
still
a
major
portion
of
the
total
load,
than
at
higher
flws
when
the
effluent
load
is
a
minor
contribution.

To
ensure
adeguate
protection
over
a
range
of
flows,
two
types
of
WESs
need
to
be
determined:
Tvps
1
Wars
are
detemined
by
obtaining
saqles
of
effluent
and
upstream
water
when
the
downstream
flow
is
between
one
and
two
tines
higher
than
what
it
would
be
under
design­
flow
conditions.
Tvpc
2
WERs
are
determined
by
obtaining
smles
of
effluent
and
upstream
water
when
the
downstream
flaw
is
between
two
andtentimes
higher
than
what
it
would
be
under
design­
flow
conditions.
The
only
di'fference
between
the
two
types
of
samples
is.
the
downstream
flaw
at
the
time
the
s­
lea
are
taken.
For
both
types
of
WERs,
the
samples
should
be
mixed
at
the
ratio
of
the
flows
that
existed
when
the
samples
were
taken
so
that
seasonal
and
flaw­
related
changes
in
the
water
quality
characteristics
of
the
upstream
water
are
properly
related
to
the
flow
at
which
they
occurred.
The
ratio
at
which
the
sas~
les
are
mixed
does
not
have
to
be
the
exact
ratio
that
existed
when
the
samp&
es
were
taken,
but
the
ratio
has
to
be
known,
which
is
why
simulated
downstream
water
is
used.
For
each
Type
1
WE&
I
and
each
Type
2
WER
that
is
detexmined,
a
hWER
is
calculated.

35
Ideally,
sufficiant
numbers
of
both
types
of
WERB
would
be
available
and
each
WER
would
be
8ufficiently
precise
and
accurate
and
the
Type
1
WERs
would
be
irufficiently
similar
that
the
PWER
could
be
the
geometric
mean
of
the
Zsrpe
1
WERS,
unless
.
tha
FWER
had
to
be
lowered
buause
of
one
or
more
hWER8.
If
M
adeguate
.
amber
of
one
or
both
types
of
WER8
is
not
available,
M
envir
onmentally
coxmervative
WER
or
hWER
8hould
be
used
as
the
FWER.

Three
'
typo
1
and/
or
m
2
WARS,
which
were
determined
urring
acceptable
procedure8
and
for
which
there
were
at
lea8tthreeweek8
between
My
two
mmpling
events,
rrutbe
available
in
order
for
a
FWR
to
be
duived.
If
three
or
mOre
are
available,
the
FWER
should
be
derived
from
the
WE~
U
Md
hWER8
using
the
loue8t
numbered
option
whose
requirements
are
ratisfied:
1.
If
there
are
two
or
more
Type
1
WER8:
a.
If
at
least
nineteen
percent
of
all
of
the
WERS
are
m
2
WBR8,
tbederivationof
theFWERdepends
on
the
propertie
of
the
T!
ypa
1
WER8:
1)
If
the
range
of
the
TYpa
1
WERa
i8
not
greater
than
a
factor
of
5
Md/
or
the
range
of
the
ratio8
of
the
Typa
1
WER
to
the
concentration
of
metal
in
the
simulated
downstream
watu
ir
not
greater
than
a
factor
of
5,
the
FWER
ir
the
1­
r
of
(
a)
the
adjustad
geometric
maan
(
8ee
Figure
2)
of
all
of
the
Type
1
m
and
(
b)
the
louert
hWER.
2)
If
the
range
of
the
m
1
W3R8
is
greater
than
a
factor
of
5
and
the
range
of
the
ratios
of
the
w
1
WER
to
the
concentration
of
metal
in
the
simulated
dcwmtream
watu
is
greater
than
a
factor
of
5,
the
FWER
is
the
lowest
of
(
a)
the
lowest
w
1
WER,
(
b)
the
louemt
hWER,
Mb
(
c)
the
geaoetric
maan
of
all
the
Tvpa
1
and
lLpc
2
WE&
B,
unless
an
analyris
of
the
joint
probabilitiu
of
the
ouurren
Cm
of
WEIZII
Md
metal
concentration8
indicate8
that
a
highu
WER
would
still
provide
the
level
of
protection
intended
by
the
crituion.
(
EPA
intMd8
to
provide
guidance
coacuning
8uch
M
analysis.)
b.
If
leslr
than
niaeteen
parcent
of
all
of
the
WERs
are
Type
2
WERs,
the
FWBR
is
the
lower
of
(
1)
the
lowut
Type
1
WER
and
(
2)
the
lowert
hWER.
2.
If
there
is
one
Type
1
WER,
the
FWER
is
the
lowest
of
(
a)
the
Type
1
WER,
(
b)
the
lowe8t
hWER,
and
(
c)
the
geometric
maan
of
all
of
the
Type
1
and
Type
2
WERs.
3.
If
there
are
no
Type
1
WEM,
the
FWER
is
the
lower
of
(
a)
the
lowest
3!
ype
2
WER
Md
(
b)
the
lowest
hWER.
If
fewer
than
three
WER8
are
available
and
a
rite­
8pecific
criterion
is
to
be
derived
using
a
WER
or
a
FWER,
the
WER
or
FWER
has
to
be
assumed
to
be
1.
Examples
of
deriving
FWER8
using
these
option8
are
presented
in
Figure
3.

36
The
options
are
designed
to
ensure
that:
a.
The
options
apply
equally
well
to
ordinary
flowing
waters
and
to
streams
whose
design
flow
is
zero.
b.
The
requirements
for
deriving
the
FWER
as
something
other
thari
the
lowest
WEX
are
not
too
stringent.
c.
The
probability
is
high
that
the
criterion
will
be
adequately
protective
at
all
flows,
regardless
of
the
aumunt
of
data
that
are
available.
d.
The
generation
of
both
types
of
WERs
is
encouraged
because
environmental
conservatism
is
built
in
if
both
types
of
WERs
are
not
available
in
acceptable
numbers.
e.
The
amount
of
conservatism
decreases
as
the
quality
and
quantity
of
the
available
data
increase.
The
requirement
that
three
WERs
be
available
is
based
on
a
judgment
that
fewer
WERs
will
not
provide
sufficient
information.
The
requirement
that
at
least
nineteen
percent
of
all
of
the
available
WERs
be
Type
2
WERS
is
based
on
a
judgment
concerning
what
constitutes
an
adequate
mix
of
the
two
types
of
WERS:
when
there
are
five
or
more
WERs,
at
least
one­
fifth
should
be
Type
2
WERS.

Because
each
of
these
options
for
deriving
a
FWER
is
expected
to
provide
adequate
protection,
anyone
who
desires
to
detemine
a
FWER
can
generate
three
or
more
appropriate
WERs
and
u8e
the
option
that
corresponds
to
the
WERs
that
are
available.
The
options
that
utilize
the
least
useful
WEF&
are
expected
to
provide
adequate
protection
because
of
the
way
the
FWER
is
derived
from
the
WERS.
It
is
intended
that,
on
the
average,
Option
la
will
result
in
the
highest
FW&
R,
and
so
it
is
ret
oamended
that
data
generation
should
be
designed
to
satisfy
the
requirements
bf
this
option
if
possible.
For
example,
if
two
Type
1
WERs
have
been
determined,
determining
a
third
Qpe
1
WER
will
require
use
of
Option
lb,
whereas
determining
a
Type
2
WER
will
require
use
of
Option
la.

Calculation
of
the
FWEF4
as
M
adjusted
geometric
mean
raises
three
i88ues:
a.
The
level
of
protection
would
be
greater
if
the
lowest
WER,
rather
than
M
adjusted
xne~,
were
used
as
the
FWEFI.
Although
true,
the
intended
level
of
protection
is
provided
by
the
national
aquatic
life
criterion
derived
according
to
the
national
guidelines;
when
sufficient
data
are
available
and
it
is
clear
how
the
data
should
be
used,
there
is
no
reason
to
add
a
substantial
margin
of
safety
and
thereby
change
the
intended
level
of
protection.
Use
of
M
adjusted
geometric
mean
is
acceptable
if
sufficient
data
are
available
concerning
the
WER
to
demonstrate
that
the
adjusted
geometric
mean
will
provide
the
intended
level
of
protection.
Use
of
the
lowest
of
three
or
more
WERs
would
be
justified,
if,
for
example,
the
criterion
had
37
been
lowered
to
protect
a
comnercially
important
species
and
a
WER
determined
with
that
species
was
1­
r
than
WEF&
determined
with
other
species.
b.
The
level
of
protection
would
be
greater
if
the
adjustment
wa8
to
a
probability
of
0.95
rather
than
to
a
probability
of
0.70.
As
above,
the
intended
level
of
protection
is
provided
by
the
national
aquatic
life
criterion
derived
according
to
the
national
guidelines.
There
is
no
need
to
substantially
increase
the
level
of
protection
when
qite­
8pecific
criteria
are
derived.
c.
It
would
be
easier
to
use
the
more
cmmnon
arithmetic
wan,
especially
becawe
the
geowtric
xae~
usually
does
not
provide
much
xmrc
protection
than
the
arithmetic
mean.
Although
true,
ume
of
the
geanetric
mean
rather
than
the
arithmetic
WM
i8
justified
on
the
basis
of
statistics
and
mathemtics;
use
of
the
gntric
mean
is
also
consistent
with
the
intended
level
of
protection.
U8e
of
the
arithmetic
WM
is
amropriatewhen
thevaluaa
can
range
franmimm
infinity
to
plu8
infinity.
The
geaaetric
xne~
(
GM)
i8
equivalent
to
using
the
arithmetic
WM
of
the
logarithm8
of
the
value8.
WDFU
cannot
be
negative,
but
the
logarithm
of
WEEUS
CM.
The
di8tribution
of
the
logarithms
of
WERa
i8
therefore
more
likely
to
be
normally
distributed
than
is
the
distribution
of
the
WER8.
Thus,
it
is
better
to
use
the
GM
of
WERS.
In
addition,
when
dealing
with
quotienta,
use
of
the
GM
reduce8
a
rguwnts
about
the
correct
way
to
do
sane
calculations
because
the
saw
answer
is
obtained
in
dif
farantwayr.
For
axanple,
if
WERl
­
Wl)/(
Dl)
and
WER2
=
(
N2)/(
02),
then
the
GM
of
WERl
and
WER2
gives
the
saw
value
a8
[(
GM
of
Nl
and
N2)/(
GM
of
01
and
DZ)]
and
also
equals
the
square
root
of
([(
Nl)
W2)
l/
t(
Dl)
(
D2)
l).

Anytime
the
FWER
is
derived
as
the
lowest
of
a
series
of
experimentally
determined
#
VERa
and/
or
hWER8,
the
magnitude
of
the
FWER
will
depend
at
least
in
part
on
experimental
variation.
There
are
at
least
three
ways
that
the
influence
of
experimental
variation
on
the
FWER
CM
be
reduced:
a.
A
WER
determined
with
a
primary
test
CM
be
replicated
and
the
geometric
mean
of
the
replicates
wed
as
the
value
of
the
WER
for
that
detemination.
Then
the
FWER
would
be
the
lowest
of
a
number
of
geometric
means
rather
than
the
lowest
of
a
number
of
individual
WERS.
To
be
true
replicates,
the
replicate
determinations
of
a
WER
should
not
be
based
on
the
saqm
test
in
laboratory
dilution
water,
the
same
8­
1~
of
site
water,
or
the
same
sample
of
effluent.
b.
If,
for
example,
Option
3
is
to
be
used
with
three
Type
2
WERs
and
the
endpoints
of
both
the
primary
and
38
secondary
tests
in
laboratory
dilution
water
are
above
the
CMC
and/
or
CCC
to
which
the
WER
is
to
apply,
WERS
can
be
determined
with
both
the
primary
and
secondary
tests
for
each
of
the
three
sampling
times.
For
each
sapling
tune,
the
geometric
mean
of
the
WER
obtained
with
the
primary
test
and
the
WER
obtained
with
the
secondary
test
could
be
calculated;
then
the
lowest
of
these
three
geometric
means
could
be
used
as
the
FWER.
The
three
WERs
cannot
consist
of
some
WERs
determined
with
one
of
the
tests
and
some
WEFU
determined
with
the
other
test;
similarly
the
three
b
cannot
consist
of
a
combination
of
individual
WERs
obtained
with
the
prixnaxy
and/
or
secondary
tests
and
geometric
means
of
results
of
primary
and
Secondary
tests.
c.
AS
mentioned
above,
because
the
variability
of
the
effluent
might
contribute
substantially
to
the
variability
of
the
WERs,
it
might
be
desirable
to
obtain
and
store
more
than
one
sample
of
the
effluent
when
a
WER
is
to
be
determined
in
case
M
unusual
WER
is
obtained'with
the
first
sample
used.
Examples
of
the
first
and
second
ways
of
reducing
the
impact
of
experimsntal
variation
are
presented
in
Figure
4.
The
availability
of
these
alternatives
does
not
mean
that
they
are
necessarily
cost­
effective.

6.
For
metals
whose
criteria
are
hardness­
dependent,
at
what
hardness
should
WERs
be
determined?

The
issue
of
hardness
bears
on
such
topics
as
acclimation
of
test
organisms
to
the
site
water,
adjustment
of
the
hardness
of
the
site
water,
and
how
M
experimsntally
determined
WER
should
be
used.
If
all
WERs
were
determined
at
design­
flow
conditions,
it
might
seem
that
all
WERs
should
be
determined
at
the
design­
flow
hardness.
Sow
permit
limits,
however,
are
not
based
on
the
hardness
that
is
most
likely
to
occur
at
design
flow;
in
addition,
conducting
all
tests
at
design­
flow
conditions
provides
no
infoxmation
concerning
whether
adequate
protection
will
be
provided
at
other
flows.
Thus,
unless
the
hardnesses
of
the
upstream
water
and
the
effluent
are
similar
and
do
not
vary
with
flow,
the
hardness
of
the
site
water
will
not.
be
the
same
for
all
WER
determinations.

Because
the
toxicity
tests
should
be
begun
within
36
hours
after
the
samples
of
effluent
and
upstream
water
are
collected,
there
is
little
time
to
acclimate
organisms
to
a
sample­
specific
hardness.
One
alternative
would
be
to
acclimate
the
organisms
to
a
preselected
hardness
and
then
adjust
the
hardness
of
the
site
water,
but
adjusting
the
hardness
of
the
site
water
might
have
various
effects
on
the
toxicity
of
the
metal
due
to
competitive
binding
and
ionic
impacts
on
the
test
organisms
and
on
the
speciation
39
of
the
metal;
lowering
hardness
without
also
diluting
the
WER
is
especially
problematic.
The
least
objectionable
approach
is
to
acclimate
the
organisms
to
a
laboratory
dilution
water
with
a
hardness
in
the
range
of
50
to
150
mg/
L
and
thezi
u8e
this
water
as
the
laboratory
dilution
water
when
the
WER
is
determined.
In
this
way,
the
test
organisms
will
be
acclimated
to
the.
laboratory
dilution
water'as
specified
by
ASTM
(
1993a,
b,
c,
d,
e).

Te8t
organisms
may
be
aCCliWted
to.
the
site
water
for
a
short
time
as
long
as
this
does
not
cause
the
tests
to
bsgin
more
than
36
hour8
after
the
sauples.
were
collected.
Regardless
of
what
acclimation
procedure
is
used,
the
organisms
used
for
the
toxicity
test
conducted
using
site
water
are
unlikely
to
be
acclimated
as
well
as
would
be
desirable.
This
is
a
general
problem
with
toxicity
tests
conducted
in
site
water
(
U.
S.
EPA
1993a,
b,
c;
ASTM
1993f),
and
its
impact
on
the
result8
of
tests
is
unknown.

For
the
practical
reasons
given
above,
M
experimentally
detezmined
WER
will
u8wlly
be
a
ratio
of
endpoints
determined
at
two
differsnt
hardnesses
and
will
thus
include
contributions
from
a
variety
of
differences
between
the
two
waters,
including
hardness.
The
disadvantagds
of
differing
hardnesses
are
that
(
a)
the
test
organisms
probably
will
not
be
adeguately
acclimated
to
site
water
and
(
b)
additional
calculations
will
be
needed
to
account
for
the
differing
hardnesses;
the
advantages
are
that
it
allows
the
generation
of
data
concerning
the
adequacy
of
protection
at
various
flows
of
ugstrsam
water
and
it
provides
a
way
of
overcoming
two
problems
with
the
hardness
sguation8:
(
1)
it
is
not
known
hckJ
applicable
they
are
to
hardnesses
outside
the
range
of
25
to
400
mg/
L
and
(
2)
it
is
not
known
how
applicable
they
are
to
unu8ual
combinations
of
hardness,
alkalinity,
and
pH
or
to
unusual
ratios
of
calcium
and
magnesium.

The
additional
calculations
that
ure
necessazy'to
account
for
the
differing
hardnesses
will
also
overcome
the
shortcomings
of
the
hardness
equations.
The
purpose
of
determining
a
Ww
is
to
determine
how
much
metal
can
bein
a
site
water
without
lowering
the
intsnded
level
of
protection.
Each
mtperimentally
determined
WER
is
inherently
refersncsd
to
the
hardness
of
the
laboratory
dilution
water
that
was
used
in
the
determination
of
the
WER,
but
the
hardness
eguation
CM
be
u8ed
to
calculate
adjusted
WERs
that
are
referenced
to
other
hardnesses
for
the
laboratory
dilution
water.
When
used
to
adjust
WERs,
a
hardness
eguation
for
a
CMC
or
CCC
CM
bs
used
to
reference
a
WER
to
any
hardness
for
a
laboratory
dilution
water,
whether
it
is
inside
or
outside
the
range
of
25
to
400
mg/
L,
because
any
inappropriateness
in
the
equation
40
will
be
automatically
compensated
for
when
the
adjusted
WER
is
used
in
the
derivation
of
a
FWER
and
permit
limits.

For
example,
the
hardness
equation
for
the
freshwater
CMC
for
copper
gives
CMCs
of
9.2,
18,
and
34
ug/
L
at
hardnesses
of
SO,
100,
and
201)
mg/
L,,
respectively.
If
acute
toxicity
tests
with
wiodw
geti­
gave
an
EC50
of
18
ug/
L
using
a
laboratory
dilution
watsr
with
a
hardness
of
100
!
ag/
L
and
M
EC50
of
532.2
;~/
k;
c
site
water,
the
resulting
WER
would
be
29.57.
assumed
that,
within
experimental
variation,
EC508
of
9.2
and
34
ug/
L
and
WERs
of
57.85
and
15.65
would
have
been
obtained
if
laboratory
dilution
waters
with
bardnesses
of
50
and
200
mg/
L,
reSpeCtiVely,
had
been
used,
because
the
EC50
of
532.2
ug/
L
obtained
in
the
site
water
does
not
depend
on
what
water
is
used
for
the
laboratory
dilution
water.
The
WERs
of
57.85
and
15.65
CM
be
considered
to
be
adjusted
WERs
that
were
extrapolated
from
the
experimentally
determined
WER
using
the
hardness
equation
for
the
copper
CMC.
If
used
correctly,
the
experimentally
determined
WER
and
all
of
the
adjusted
WERs
will
result
in
the
same
permit
limits
because
they
are
internally
consistent
and
are
all
based
on
the
EC50
of
532.2
ug/
L
that
was
obtained
in
site
water.

A
hardness
eguation
for
copper
CM
be
used
to
adjust
the
WER
if
the
hardness
of
the
laboratory
dilution
water
used
in
the
determination
of
the
WER
is
in
the
range
of
25
to
400
mg/
L
(
preferably
in
the
range
of
about
40
to
250
mg/
L
because
most
of
the
data
used
to
derive
the
eguation
are
in
this
range).
However,
the
hardness
equation
CM
be
used
to
adjust
WERs
to
hardnesses
outside
the
range
of
25
to
400
mg/
L
because
the
basis
of
the
adjusted
WBR
doe8
not
change
the
fact
that
the
EC50
obtained
in
site
water
was
532.2
ug/
L.
If
the
hardness
of
the
site
water
was
16
mg/
L,
the
hardness
eguation
would
predict
an
EC50
of
3.153
.
.
.
1
Similarly,
if
the
hardness
of
the
site
water
had
been
447
mg/
L,
the
hardness
equation
would
predict
an
EC50
of
72.66
ug/
L,
with
a
corresponding
adjusted
WER
of
7.325.
If
the
hardness
of
447
mg/
L
were
due
to
M
effluent
that
contained
Cal&
m
chloride
and
the
alkalinity
and
pH
of
the
site
water
were
what
would
usually
occur
at
a
hardness
of
50
mg/
L
rather
than
400
mg/
L,
any
inappropriateness
in
the
calculated
EC50
of
72.66
ug/
L
will
be
coanpensated
for
in
the
adjusted
WER
of
7.325,
because
the
adjusted
WER
is
based
on
the
EC50
of
532.2
ug/
L
that
was
obtained
using
the
site
water.

41
In
the
above
examples
it
was.
assuqed
that
at
a
hardness
of
100
mg/
L
the
EC50
for
c.
setxcw
tgualled
the
CMC,
which
is
a
very
reasonable
simplifying
assumption.
If,
hmever,
the
WER
had
been
detexmined
with
the
nmre.
resistant
aw
and
EC508
of
50
ug/
L
and
750
ug/
L
had
been
obtained
using
a
laboratory
dilution
water
and
a
site
water,
respectively,
the
CMC
given
by
the
hardness
equation
could
not
bs
used
a8
the
predicted
ECSO.
A
new
squation
would
have
to
be
derived
by
changing
the
intercept
80
that
the
nsw
equation
gives
an
EC50
of
50
ug/
L
at
a
hardness
of
100
mg/
L;
this
new
equation
could
then
ba
umd
to
calculate
adjusted
ECSOs,
which
could
then
be
uued
to
calculate
corresponding
adjusted
WERs:

Hudneso
Cl&
L)

f:

100
200
447
EC50
JJasuu
­

26.022
8.894
28.82
84.33
50.000+
15.00.
96.073
7.01
204.970
3.66
The
values
marked
with
M
asterisk
are
the
assumed
ucpu%
mentally
determined
values;
the
others
were
calculated
frun
these
values.
At
each
hardness
the
product
of
the
EC50
time8
the
WER
equals
750
ug/
L
because
all
of
the
WKRs
are
based
on
the
sanm
EC50
obtained
using
site
water.
Thus
use
of
the
WER
allows
application
of
the
hardness
equation
for
a
metal
to
conditions
to
which
it
othemise
might
not
be
applicable.

and
the
flows
of
the
effluent
and
upstream
water
were
9
and
73
cfs,
respctively,
when
the
samples
were
collected,
the
HCME
calculated
from
the
WER
of
15.00
would
be:

flgqPr
(
17.73
U&
L)
(
15)(
9
l
73
cm
­
(
1
w/
t)
(
73
Cf@)
9
cfs
­
2415
L&
L
because
the
CMC
is
17.73
ug/
L
at
a
hardness
of
100
mg/~.
(
The
value
of
17.73
ug/
L
is
used
for
the
CMC
instead
of
18
ug/
L
to
reduce
roundoff
error
in
this
example.)
If
the
hardness
of
the
site
water
was
actually
447
ug/
L,
the
HCKE
could
also
be
calculated
using
the
WEFI
of
3.66
and
the
CK
of
72.66
ug/
L
that
would
be
obtained
froan
the
CMC
hardness
equation:

42
AYQUE=
(
72.66
up/
L)
(
3.66)
(
9
l
73
MS)
­
(
1
UP/
L)
(
73
cfs)
9
cfs
­
241s
ug/
L
.

Either
WER
can
be
used
in
the
calculation
of
the
HCME
as
long
as
the
Q&
and
the
WER
correspond
to
the
same
hardness
and
therefore
to
each
other,
because:

(
17.73
t&
L)
(
15)
­
(
72.66
f&
L)
(
3.66)
.

Although
the
HCME
will
be
correct
as
long
as
the
hardness,
C&
K;.
and
WER
correspond
to
each
other,
the
WER
used
in
the
derivation
of
the
FWER
mast
be
the
one
that
is
calculated
using
a
hardness
equation
to
be
conrpatible
with
the
hardness
of
the
site
water.
If
the
hardness
of
the
site
water
was
447
ug/
L,
the
WER
used
in
the
derivation
of
the
FWER
has
to
be
3.66;
therefore,
the
simplest
approach
is
to
calculate
the
HCME
using
the
WER
of
3.66
and
the
corresponding
CMC
of
72.66
ug/
L,
because
these
correspond
to
the
hardness
of
447
ug/
L,
which
is
the
hardness
of
the
site
water.

In
contrast,
the
hWER
should
be
calculated
using
the
CMC
that
corresponds
to
the
design
hardness.
If
the
design
hardness
is
50
mg/
L,
the
corresponding
CMC
is
9.2
ug/
L.
If
the
design
flows
of
the
effluent
and
the
upstream
water
are
9
and
20
cfs,
respectively,
and
the
concentration
of
metal
in
upstream
water
at
design
conditions
is
1
ug/
L,
the
hWER
obtained
from
the
WER
determined
using
the
site
water
with
a
hardness
of
447
mg/
L
would
be:

hm
*
(
2414
w/
L)
(
9
cm
+
cl
MS/
L)
(
20
Cf8)
*
*
1
l
%
.
(
9.2
l&
t)
(
9
Cf8
+
20
Cf8)

None
of
these
calculations
provides
a
way
of
extrapolating
a
WER
from
one
site­
water
hardness
to
another.
The
only
extrapolations
that
are
possible
are
from
one
hardness
of
laboratory
dilution
water
to
another;
the
adjusted
WERs
are
based
on
predicted
toxicity
in
laboratory
dilution
water,
but
they
are
all
based
on
measured
toxicity
in
site
water.
If
a
WER
is
to
apply
to
the
design
flow
and
the
design
hardness,
one
or
more
toxicity
tests
have
to
be
conducted
using
samples
of
effluent
and
upstream
water
obtained
under
design­
flow
conditions
and
mixed
at
the
design­
flow
ratio
to
produce
the
design
hardness.
A
WER
that
is
specifically
appropriate
to
design
conditions
cannot
be
based
on
predicted
toxicity
in
site
water;
it
has
to
be
based
on
measured
toxicity
in
site
water
that
corresponds
to
design­
flow
conditions.
The
situation
is
more
cosnplicated
if
the
design
hardness
is
not
the
hardness
that
is
most
likely
to
occur
when
effluent
and
upstream
water
are
mixed
at
the
ratio
of
the
design
flows.

43
B.
Background
Information
and
Initial
Decisions
1.

2.

3.

4.

5.

6.

7.

8.
Infoxmation
should
be
obtained
concerning
the
effluent
and
the
operating
and
discharge
schedules
of
the
discharger.

The
spatial
extent
of
the
site
to
which
the
WER
and
the
rite­
specific
criterion
are
intended
to
apply
should
be
defined
(
see
Appendix
A).
Information
concerning
tributaries,
the
plume,
and
the
point
of
complete
mix
8botid
be
obtained.
Dilution
models
(
U.
S.
EPA
19936)
and
dye
dispersion
studies
(
Kilpatrick
1992)
might
provide
information
that
is
useful
for
defining
sites
for
ancwBB8.

If
the
Recalculation
Procedure
(
see
Appendix
B)
is
to
IX
used,
it
should
be
perfoxmed.

Pertinent
information.
conccrning
the
calculation
of
the
permit
limits
should
bs
obtained:
a.
What
are
the
design
flows,
i.
e.,
the
flow
of
the
upstream
water
(
e.
g.,
7910)
and
the
flow
of
the
;
szzt
that
are,
used
in
the
calculation
of
the.
psmit
(
The
deslm
flows
for
the
CMC
and
CCC
rmght
be
the
same
or
diffcrsnt.)
b.
Is
there
a
CMC
(
acute)
mixing
zone
and/
or
a
CCC
(
chronic)
mixing
cone?
c.
What
are
the
dilution(
s)
at
the
edge(
s)
of
the
mixing
zone(
s)?
d.
If
the
criterion
is
hardness­
dependent,
what
is
the
hardness
on
which
the
permit
limits
are
based?
Is
this
a
hardness
that
is
likely
to
occur
under
design­
flow
conditioas?

It
should
be
decidsd
whether
to
determine
a
amHER
and/
or
a
CCCWER.

The
water
quality
criteria
docuwn
t
(
see
Appendix
E)
that
88­
88
as
the
basis
of
the
aquatic
life
criterion
should
be
read
to
identify
any
chemical
or
toxicological
properties
of
the
mstal
that
are
relevant.

If
the
WEB
is
being
determined
by
or
for
a
discharger,
it
will
probably'bs
desirable
to
decide
what
is
the
smallest
WBR
that
is
desired
by
the
discharger
(
e.
g.,
the
smallest
Ww
that
wuld
not
require
a
reduction
in
the
amount
of
metal
discharged).
This
@
smallest
desired
WER'
might
be
useful
when
deciding
whether
to
detezmine
a
WER.
If
a
WER
is
determined,
this
gsmallest
desired
WEE!'
might
be
useful
when
selecting
the
range
of
concentrations
to
be
tested
in
the
site
water.

Information
should
be
read
concerning
health
and
safety
considerations
regarding
collection
and
handling
of
44
effluent
and
surface
water
samples
and
conducting
toxicity
tests
(
U.
S.
EPA
1993a;
ASTM
1993a).
Information
should
also
be
read
concerning
safety
and
handling
of
the
metallic
salt
that
will
be
used
in
the
preparation
of
the
stock
solution.

9.
The
proposed
work
should
be
.
discussed
with
the
appropriate
regulatory
authority
(
and
possibly
the
Water
Management
Division
of
the
EPA
Regional
Office)
before
deciding
how
to
proceed
with
the
development
of
a
detailed
wrkplan.

10.
Plans
should
be
made
to
perform
one
or
more
rangefinding
tests
in
both
laboratory
dilution
water
and
site
water
(
see
section
G.
7).

C.
Selecting
Primary
and
Secondary
Tests
1.
For
each
WER
(
cmcWER
and/
Or
cccWER1
to
be
determined,
the
primary
and
secondary
tests
should
be
selected
using
the
rationale
presented
in
section
A.
3,
the
information
in
~
ppsndix
I,
the
information
in
the
criteria
document
for
the
metal
(
see
Appsndix
E),
and
any
other
pertinent
information
that
is
available.
When
a
specific
test
species
is
not
specified,
also
select
the
species.
Because
at
least
three
WERs
wst
be
determined
with
the
primary
test,
but
only
one
mt
be
determined
with
the
secondary
test,
selection
of
the
tests
might
be
influenced
by
the
availability
of
the
species
(
and
the
life
stage
in
so1~
e
cases)
during
the
planned
testing
period.
a.
The
description
of
a
@
test'
specifies
not
only
the
test
species
and
the
duration
of
the
test
but
also
the
life
stage
of
the
species
and
the
adverse
effect
on
which
the
results
are
to
be
based,
all
of
which
can
have
a
major
impact
on
the
sensitivity
of
the
test.
b.
The
endpoint
(
e.
g.,
LCSO,
ECSO,
IC50)
of
the
priznaxy
test
in
laboratory
dilution
water
should
be
as
close
as
possible,
but
it
mst
mot
be
below,
the
CMC
and/
or
CCC
to
which
the
WER
is
to
bs
applied,
because
for
any
two
tests,
the
test
that
has
the
lower
endpoint
is
likely
to
give
the
higher
WER
(
see
Appendix
D).
NOTE:
If
both
the
Recalculation
Procedure
and
a
WER
are
to
be
used
in
the
derivation
of
the
site­
specific
criterion,
the
Recalculation
Procedure
mt
be
completed
first
because
the
recalculated
CRC
and/
or
CCC
mast
be
used
in
the
selection
of
the
primary
and
secondary
tests.
c.
The
endpoint
(
e.
g.,
LCSO,
ECSO,
ICSO)
of
the
secondary
test
in
laboratory
dilution
water
should
be
as
close
as
possible,
but
may
bs
above
or
below,
the
CRC
and/
or
CCC
to
which
the
WER
is
to
be
applied.

45
2.
1)
Because
few
toxicity
tests
have
endpoints
close
to
the
CMC
and
CCC
and
because
the
major
use
of
the
secondary
test
is
confirmation~(
see
section
1.7.
b),
the
sndpoint
of
the
secondary
test
may
be
below
the
CMCorCCC.
If
the
endpoint
of
the
secondary
test
in
h
¶
bOratOIy
dilUtiOn3vStSr
i8
SbOVS
the
CMC
and/
or
CCC,
it
might
be
possible
to
use
the
results
to
reduce
the
*
act
of
experimsntal
variation
(
see
Figure
4).
If
the
endpoint
of
the
primary
test
in
laboratory
dilution
water
is
above
the
CMC
and
the
endpoint
of
the
secondary
test
is
bstwwn
the
CMC
and
CCC,
it
should
be
po88ible
to
determine
both
a
CCCWER
and
a
CllmER
using
the
I­
two
tests.
2)
It
is
often
desirable
to
conduct
the
secondary
test
when
the
first
primary
test
i8
conducted
in
case
the
results
are
surprising;
conducting
both
tests
the
first
tiw
also
makes
it
possible
to
interchange
the
primary
and
secondary
tests,
if
desired,
without
increasing
the
number
of
tests
that
need
to
bs
conducted.
(
If
results
of
one
or
more
rangefinding
tests
are
not
available,
it
might
be
desirable
to
wait
and
conduct
the
8eCoIXhWy
test
when
more
information
is
available
concerning
the
laboratory
dilution
water
and
the
site
water.)

The
primary
and
secondary
tests
mast
be
conducted
with
species
in
different
tsxonoxnic
m;
at
least
one
species
mast
be
an
animal
and,
when
feasible,
one
species
should
be
a
vertebrate
and
the
other
should
bs
M
invertebrate.
A
plant
cannot
be
used
if
nutrients
and/
or
chelators
need
to
bs
added
to
either
or
both
dilution
waters
in
order
to
determine
the
WER.
It
is
desirable
to
use
a
test
and
species
for
which
the
rate
of
success
is
known
to
bs
high
and
for
which
the
test
organisms
are
readily
available.
(
If
the
WER
is
to
bs
used
with
a
recalculatsd
CMC
and/
or
CCC,
the
species
used
in
the
primary
and
secondary
tests
do
I&
have
to
bs
on
the
list
of
species
that
are
used
to
obtain
the
recalculated
CMC
and/
Or
CCC.)

3.
There
are
advantages
to
using
tests
suggested
in
Appendix
I
or
other
tests
of
comparable
sensitivity
for
which
data
are
available
from
one
or
more
other
laboratories.
a.
A
good
indication
of
the
sensitivity
of
the
test
is
available.
This
helps
ensure
that
the
endpoint
in
laboratory
dilution
water
is
close
to
the
CMC
and/
or
CCC
and
aids
in
the
selection
of
concentrations
of
the
mstal
to
be
ussd
in
the
rangefinding
and/
or
definitive
toxicity
tests
in
laboratory
dilution
water.
Tests
with
other
species
such
as
specie8
that
occur
at
the
site
may
be
used,
but
it
is
sometimes
more
difficult
to
obtain,
hold,
and
test
such
species.

46
b.
When
a
WEB
is
determined
and
used,
the
results
of
the
tests
in
laboratory
dilution
water
provide
the
connection
between
the
data
used
in
the
derivation
of
the
national
criterion
and
the
data
obtained
in.
site
water,
i.
e.,
the
results
in
laboratory
dilution
water
are
a
vital
link
in
the
deivation
and
use
of
a
WEP.
It
is,
therefore,
important
to
be
able
to
judge
the
quality
of
the
results
in
laboratory
dilution
water.
Comparison
of
results
with
data
froan
other
laboratories
evaluates
all
aspects
of
the
test
methodology
simultaneously,
but
for
the
determination
of
WERs,
the
most
important
aspect
is
the
quality
of
the
laboratory
dilution
water
because
the
dilution
water
is
the
most
important
difference
between
the
two
side­
by­
side
tests
from
which
the
WER
is
calculated.
Thus,
two
tests
aut
be
conducted
for
which
data
are
available
on
the
metal
of
concern
in
a
laboratory
dilution
water
ftom
at
least
one
other
laboratory.
If
both
the
primary
and
secondary
tests
are
ones
for
which
acceptable
data
are
available
from
at
least
one
other
laboratory,
these
are
the
only
two
tests
that
have
to
be
conducted.
If,
however,
the
primary
and/
or
secondary
tests
are
ones
for
which
no
results
are
already
available
for
the
metal
of
concern
from
another
laboratory,
the
first
or
second
time
a
WER
is
determined
at
least
two
additional
tests
mast
be
conducted
in
the
laboratory
dilution
water
in
addition
to
the
tests
that
are
conducted
for
the
determination
of
WERs
(
see
sections
F.
5
and
1.5).
1)
For
the
determination
of
a
WER,
data
are
not
required
for
a
reference
toxicant
with
either
the
primary
test
or
the
secondary
test
because
the
above
requirement
provides
similar
data
for
the
metal
for
which
the
WER
is
actually
being
determined.
2)
See
Section
I.
5
concerning
interpretation
of
the
results
of
these
tests
before
additional
tests
are
conducted.

D.
Acquiring
and
Acclimating
Test
Organisms
1.
The
test
organisms
should
be
obtained,
cultured,
held,
acclimated,
fed,
and
handled
as
recoamsn
ded
by
the
U.
S.
EPA
(
1993a,
b,
c)
and/
or
by
ASTM
(
1993a,
b,
c,
d,
e).
All
test
organisms
mast
bs
accsptably
acclimatsd
to
a
laboratory
dilution
water
that
satisfies
the
requirements
given
in
sections
F.
3
and
F­
4;
M
appropriate
number
of
the
organisms
may
be
randomly
or
impartially
removed
from
the
laboratory
dilution
water
and
placed
in
the
site
water
when
it
becomes
available
in
order
to
acclimate
the
organisms
to
the
site
water
for
a
while
just
before
the
tests
are
begun.

47
2.
The
organisms
used
in
a
pair
of
side­
by­
side
tests
wst
be_
draw
from
the
sams
population
and
tested
under
identical
conditions.

E.
Collecting
and
Handling
UpStreul,
Water
and
Effluent
1.

2.
Upstream
water
will
usually
be
mixed
with
effluent
to
prepare
simulated
downstream
water.
Upstream
water
may
also
be
used
as
a
site
water
if
a
F
is
to
be
determined
using
upstream
water
in
addition
to
or
instead
of
detenaining
a
WER
using
dowstream
water.
The
samples
of
upstream'watu
mat
bs
representative;
they
amt
sot
be
unduly
affected
by
recent
runoff
events
(
or
othsr
erosion
or
resuspension
events)
that
cause
higher
levels
of
TSS
than
would
normally
bs
present,
unless
there
is
particular
concern
about
such
conditions.

The
sample
of
effluent
used
in
the
determination
of
a
WEP
mst
be
represmtative;
it
mst
be
collected
during
a
period
when
the
discharger
is
operating
normally.
Selection
of
the
date
and
tims
of
sampling
of
the
effluent
should
take
into
account
the
discharge
pattun
of
the
discharger.
It
might
be
appropriate
to
collect
effluent
samples
during
the
middle
of
the
week
to
allow
for
reestablishment
of
steady­
state
conditions
after
shutdowns
for
weekends
and
holidays;
alternatively,
if
end­
of­
the­
week
slug
discharges
are
routine,
they
should
probably
be
evaluated.
As
msationsd
above,
because
the
variability
of
the
effluent
might
contribute
substantially
to
the
variability
of
the
WEF&,
it
might
be
desikable
to
obtain
and
store
more
than
one
saxrple
of
the
effluent
when
WEFU
are
to
bs
determined
in
case
M
unusual
WER
is
obtained
withths
firstsanple
used.

3.
When
samples
of
rite
water
and
effluent
are
Collected
for
the
determination
of
the
WERs
with
the
primary,
test,
there
rdttg
&
between
o?
e
saspling,
evsnt
It
1s
desirable
to
obtain
samples
zn
at
least
two
diiferent
seasons
and/
or
during
times
of
probable
differences
in
the
characteristics
of
the
site
water
and/
or
effluent.

4.
Samples
of
upstream
water
and
effluent
mast
be
collected,
transported,
handlsd,
and
stored
as
ret
onmsnded
by
the
U.
S.
EPA
(
1993a1.
For
example,
samples
of
effluent
should
usually
bs
composites,
but
grab
saaples
are
acceptable
if
the
residence
time
of
the
effluent
is
sufficiently
long.
A
sufficient
volums
should
bs
obtainsd
so
that
soms
CM
be
storsd
for
additional
testing
or
analyses
if
M
unusual
WER
is
obtained.
Samples
mast
be
stored
at
0
to
4OC
in
the
dark
with
no
air
space
in
the
sample
container.

48
F.
Laboratory
Dilution
Water
5.

6.

7.

8.
At
the
time
of
collection,
the
flow
of
both
the
upstream
water
and
the
effluent
must
be
either
measured
or
estimated
by
means
of
correlation
with
a
nearby
U.
S.
G.
S.
gauge,
the
pH
of
both
upstream
water
and
effluent
must
be
measured,
and
samples
of
both
upstream
water
and
effluent
should
be
filtered
for
measurement
of
dissolved
metals.
Hardness,
TSS,
7K,
and
totai
recoverable
and
dissolved
metal
must
be
measured
in
both
the
effluent
and
the
upstream
water.
Any
other
water
quality
characteristics,
such
as
total
dissolved
solids
('
IDS)
and
conductivity,
that
are
monitored
monthly
or
more
often
by
the
permittee
and
reported
in
the
Discharge
Monitoring
Report
mast
also
be
measured.
These
and
the
other
measurements
provide
information
concerning
the
representativeness
of
the
samples
and
the
variability
of
the
upstream
water
and
effluent.

'
Chain
of
custodym
procedures
(
U.
S.
EPA
1991b)
should
be
used
for
all
samples
of
site
water
and
effluent,
especially
if
the
data
might
be
involved
in
a
legal
proceeding.

Tests
mast
be
begun
within
36
hours
after
the
collection
of
the
samples
of
the
effluent
and/
or
the
site
water,
except
that
tests
may
be
begun
more
than
36
hours
after
the
collection
of
the
samples
if
it
would
require
an
inordinate
amount
of
resources
to
transport
the
samples
to
the
laboratory
and
begin
the
tests
within
36
hours.

If
acute
and/
or
chronic
tests
are
to
be
conducted
with
daphnids
and
if
the
sample
of
the
site
water
contains
predators,
the
site
water
wt
be
filtered
through
a
37­
p
sieve
or
screen
to
remove
predators.

1.
The
laboratory
dilution
water
mast
satisfy
the
requirements
given
by
U.
S.
EPA
(
1993a,
b,
c)
or
ASI'M
(
1993a,
b,
c,
d,
e).
The
laboratory
dilution
water
mast
be
a
ground
water,
surface
water,
reconstituted
water,
diluted
mineral
water,
or
dechlorinated
tap
water
that
has
been
demonstrated
to
be
acceptable
to
aquatic
organisms.
If
a
surface
water
is
used
for
acute
or
chronic
tests
with
daphnids
and
if
predators
are
observed
in
the
sample
of
the
water,
it
mast
be
filtered
through
a
37­
p
sieve
or
screen
to
remove
the
predators.
Water
prepared
by
such
treatments
as
deionization
and
reverse
osmosis
must
not
be
used
as
the
laboratory
dilution
water
unless
salts,
mineral
water,
hypersaline
brine,
or
sea
salts
are
added
as
ret
ommended
by
U.
S.
EPA
(
1993a)
or
ASTM
(
1993a).

49
2.

3.

4.

5.
The
concentrations
of
both
Tot
and
TSS
mast
be
less
than
5
xng/
L.

The
hardness
of
the
laboratory
dilution
water
should
be
bat­
M
50
and
15Omg/
L
andautbebetween
40
and
220
mg/
L.
If
the
criterion
for
the
metal
is
hardness­
dependent,
the
hardness
of
the
laboratory
dilution
water
mst
mot
be
above
the
hardness
of
the
site
water,
unless
the
hardness
of
the
site
water
is
below
50
xng/
L.

The
alkalinity
and
pH
of
the
laboratory
dilution
water
mat
bs
appropriate
for
its
hardness;
values
for
alkalinity
and
pH
that
are
appropriate
for
soam
hardnesses
are
given
by
U.
S.
EPA
(
1993a)
and
ASTM
(
1993a);
other
corresponding
values
should
be
determined
by
interpolation.
Alkalinity
should
be
adjusted
using
sodium
bicarboMte,
and
pH
should
be
adjusted
using
aeration,
sodi\
lm
&&
oxide,
and/
or
sulfuric
acid.

It
would
seem
reasonable
that,
before
any
samples
of
site
water
or
effhmnt
are
collected,
the
toxicity
tests
that
are
to
bs
conducted
in
the
laboratory
dilution
water
for
coaparison
with
results
of
the
sass
tests
froan
other
laboratories
(
see
sections
C.
3.
h
and
1.5)
should
be
conducted.
These
should
be
performed
at
the
hardness,
alkalinity,
and
pH
specified
in
sections
F.
3
and
F.
4.

G.
Conducting
Tests
1.

2.

3.

4.
There
mst
be
no
differences
bstween
the
side­
by­
side
tests
other
than
the
composition
of
the
dilution
water,
the
concentrations
of
xnetal
tested,
and
possibly
the
water
in
which
the
test
organisuns
are
acclimated
just
prior
to
the
beginning
of
the
tests.

More
than
one
test
using
site
water
may
be
conducted
side­
by­
side
with
a
test
using
laboratory
dilution
water;
the
one
test
in
laboratory
dilution
water
will
be
used
in
the
calculation
of
several
WERs,
which
msans
that
it
is
very
inportant
that
that
one
test
be
acceptable.

Facilities
for
conducting
toxicity
tests
should
be
set
up
and
test
chambers
should
be
selected
and
cleaned
as
recoamnended
by
the
U.
S.
EPA
(
1993a,
b,
c)
and/
or
ASTM
(
1993a,
b,
c,
d,
el.

A
stock
solution
should
be
prepared
using
M
inorganic
salt
that
is
highly
soluble
in
water.
a.
The
salt
does
not
have
to
be
one
that
was
used
in
tests
that
were
used
in
the
derivation
of
the
national
criterion.
Nitrate
salts
are
generally
acceptable;

50
5.
chloride
and
sulfate
salts
of
many
metals
are
also
acceptable
(
see
Appendix
J).
It
is
usually
desirable
to
avoid
use
of
a
hygroscopic
salt.
The
salt
used
should
Illeef
A.
C.
S.
specifications
for
reagent­
grade,
if
such
specifications
are
available;
use
of
a
better
grade
is
usually
not
worth
the
extra
cost.
No
salt
should
be
used
until
information
concerning
safety
and
handling
has
been
read.
b.
The
stock
solution
may
be
acidified
(
using
metal­
free
nitric
acid)
only
as
necessary
to
get
the
metal
into
solution.
c.
The
same
stock
solution
must
be
used
to
add
metal
to
all
tests
conducted
at
one
time.

For
tests
suggested
in
Appendix
I,
the
appendix
presents
the
recoxmen
ded
duration
and
whether
the
static
or
renewal
technique
should
be
used;
additional
information
is
available
in
the
references
cited
in
the
appendix.
Regardless
of
whether
or
not
or
how
often
test
solutions
are
renewed
when
these
tests
are
conducted
for
other
purposes,
the
following
guidance
applies
to
all
tests
that
are
conducted
for
the
determination
of
WERs:
a.
The
renewal
technique
mast
be
used
for
tests
that
last
longer
than
48
hr.
b.
If
the
concentration
of
dissolved
metal
decreases
by
more
than
50
b.
in
48
hours
in
static
or
renewal
tests,
the
test
solutions
mat
be
renewed
every
24
hours.
Similarly,
if
the
concentration
of
dissolved
oxygen
becomes
too
low,
the
test
solutions
mast
be
renewed
every
24
hours.
If
one
test
in
a
pair
of
tests
is
a
renewal
test,
both
tests
mast
be
renewal
tests.
c.
When
test
solutions
are
to
be
renewed,
the
new
test
solutions
mat
be
prepared
from
the
original
unspiked
effluent
and
water
samples
that
have
been
stored
at
0
to
4OC
in
the
dark
with
no
air
space
in
the
sample
container.
d.
The
static
technique
may
be
used
for
tests
that
do
not
last
longer
than
48
hours
unless
the
above
specifications
require
use
of
the
renewal
technique.
If
a
test
is
used
that
is
not
suggested
in
Appendix
I,
the
duration
and
technique
ret
onmended
for
a
comparable
test
should
be
used.

6.
Recomnendations
concerning
temperature,
loading,
feeding,
dissolved
oxygen,
aeration,
disturbance,
and
controls
given
by
the
U.
S.
EPA
(
1993a,
b,
c)
and/
or
ASTM
(
1993a,
b,
c,
d,
e)
mast
be
followed.
The
procedures
that
are
used
mast
be
used
in
both
of
the
side­
by­
side
tests.

7.
To
aid
in
the
selection
of
the
concentrations
of
metals
that
should
be
used
in
the
test
solutions
in
site
water,
a
static
rangefinding
test
should
be
conducted
for
8
to
96
51
hours,
using
a
dilution
factor
of
10
(
or
0.1)
or
3.2
(
or
0.32)
increasing
from
about
a
factor
of
10
below
the
value
of
the
endpoint
given
in
the
criteria
document
for
the
metal
or
in
Appendix
I
of
this
document
for
tests
with
newly
hatched
fathead
minnows.
If
the
test
is
not
in
the
criteria
do
cument
and
no
other
data
are
available,
a
mean
acute
value
or
other
data
for
a
taxonomically
similar
species
should
be
used
as
the
predicted
value.
This
rangefinding
test
will
provide
information
concerning
the
concentrations
that
should
be
used
to
bracket
the
endpoint
in
the
definitive
test
and
will
provide
information
cacerni.
ng
whether
the
control
survival
will
be
acceptable.
If
dissolved
metal
is
measured
in
one
or
more
treatumnts
at
the
beginning
and
end
of
the
rangefinding
test,
these
data
will
indicate
whether
the
concentration
should
be
expected
to
decrease
by
more
than
50
8
during
the
definitive
tecrt.
The
rangefinding
test
may
be
conducted
in
either
of
two
ways:
a.

b.
It
mny
be
conducted
using
the
samples
of
effluent
and
site
water
that
will
be
used
in
the
definitive
test.
IIA
this
case,
the
duration
of
the
rangefinding
test
8hould
be
as
long
as
possible
within
the
limitation
that
the
definitive
test
mat
begin
within
36
hours
after
the
sar@
es
of
effluent
and/
or
site
water
were
collected,
except
as
per
section
E.
7.
It
may
be
conducted
using
one
set
of
samples
of
effluent
and
upstream
water
with
the
definitive
tests
being
conducted
using
sasqles
obtained
­
at
a
later
date.
In
thi8
case
the
rangefinding
test
might
give
better
results
because
it
can
last
longer,
but
there
is
the
possibility
that
the
quality
of
the
effluent
and/
or
site
water
might
change.
Chemic&
l
analyses
for
hardness
and
pH
might
indicate
whether
any
major
changes
occurred
from
one
sas&
e
to
the
next.
Rangefinding
tests
are
especially
desirable
before
the
first
set
of
toxicity
tests.
It
might
be
desirable
to
conduct
rangefinding
tests
before
each
individual
determination
of
a
WER
to
obtain
additional
information
concerning
the
effluent,
dilution
water,
organisms,
etc.,
before
each
set
of
8ide­
by­
side
tests
are
begun.

0.
Several
considerations
are
important
in
the
selection
of
the
dilution
factor
for
definitive
tests.
Use
of
concentrations
that
are
close
together
will
reduce
the
uncertainty
in
the
WER
but
will
require
more
concentrations
to
cover
a
range
within
which
the
endpoints
might
occur.
Because
of
the
resources
necessary
to
determine
a
WER,
it
is
important
that
endpoints
in
both
dilution
waters
be
obtained
whenever
a
set
of
side­
by­
side
t88tS
are
conducted.
Because
static
and
renewal
tests
can
be
used
to
determine
WERs,
it
is
relatively
easy
to
use
nmre
treatments
than
would
be
used
in
flow­
through
tests.

52
The
dilution
factor
for
total
recoverable
metal
mat
be
between
0.65
and
0.99,
and
the
recomended
factor
is
0.7.
Although
factors
between
0.75
and
0.99
may
be
used,
their
use
will
prob&
ly
not
be
cost­
effective.
Because
there
is
likely
to
be
inore
uncertainty
in
the
predicted
value
of
the
endpoint
in
site
water,
6'
or
7
concentrations
are
recomnended
in
the
laboratory.
dilution
water,
and
6
or
9
in
the
simulated
downstream
water,
at
a
dilution
factor
of
0.7.
It
might
be
desirable
to
use
even
more
treatments
in
the
first
of
the
WEB
determinations,
because
the
design
of
subsequent
tests
can
be
based
on
the
results
of
the
first
tests
if
the
site
water,
laboratory
dilution
water,
and
test
organisms
do
not
change
too
much.
The
cost
of
adding
treatnmnts
can
be
minimized
if
the
concentration
of
metal
is
measured
only
in
sanrples
from
treatments
that
will
be
used
in
the
calculation
of
the
endpoint.

9.
Each
test
mast
contain
a
dilution­
water
control.
The
number
of
test
organisms
intended
to
be
exposed
to
8ach
treatment,
including
the
controls,
maat
bc
at
least
20.
It
is
desirable
that
the
organisms
be
distributed
between
two
or
xmre
test
chambers
per
treatmnt.
If
test
organisms
are
not
randomly
assigned
to
the
test
chambers,
they
must
be
assigned
impartially
(
U.
S.
EPA
1993a;
ASTM
1993a)
between
all
test
ChaarbcrS
for
a
pair
Of
side­
by­
side
tests.
For
ocarqple,
it
is
not
acceptable
to
assign
20
organisms
to
one
treatment,
and
then
assign
20
organisms
to
another
treatmnt,
etc.
Similarly,
it
is
not
acceptable
to
assign
all
the
organisms
to
the
test
using
one
of
the
dilution
waters
and
then
assign
organisms
to
the
test
using
the
other
dilution
water.
The
test
chasibers
should
be
assigned
to
location
in
a
totally
random
arrangesmnt
or
in
a
randomized
block
design.

10.
For
the
test
using
site
water,
one
of
the
following
procedures
should
be
used
to
prepare
the
test
solutions
for
the
test
chambers
and
the
l
chemistry
controlam
(
se8
section
H.
l):
a.
Thoroughly
mix
the
sample
of
the
effluent
and
place
the
same
known
voluxm
of
the
effluent
in
each
test
chamber;
add
the
necessary
mount
of
metal,
which
will
be
different
for
each
treatment;
mix
thoroughly;
let
stand
for
2
to
4
hours;
add
the
necessary
amount
of
upstream
water
to
each
test
chanber;
mix
thoroughly;
let
stand
for
1
to
3
hours.
b.
Add
the
necessary
amount
of
metal
to
a
large
sample
of
the
effluent
and
also
maintain
an
unspiked
sample
of
the
effluent;
perform
serial
dilution
using
a
graduated
cylinder
and
the
well­
mixed
spiked
and.
unspiked
sacqples
of
the
effluent;
let
stand
for
2
to
4
hours:
add
the
necessary
amount
of
upstream
water
to
each
test
chamber;
mix
thoroughly;
let
stand
for
1
to
3
hours.

53
c.
Prepare
a
large
volume
of
simulated
downstream
water
by
mixing
effluent
and
upstream
wbter
in
the
desired
m
ratio;
place
the
same
known
volume
of
the
simulated
downstream
water
in
each
test
chamber;
,
add
the
.
necessary
amount
of
metal,
which
will
be
different
for
each
treatment;
mix
thoroughly
and
let
stand
for
1
to
3
hours.
4.
Prepare
a
large
volume
of
simulated
downstream
water
by
mixing
effluent
and
upstream
water
in
the
desired
ratio;
divide
it
into
two
portions;
prepare
a
large
v&
mm
of
the
highest
test
conctitration
of
metal
using
one
portion
of
the
simulated
downstream
water;
perform
serial
dilution
using
a
graduated
cylinder
and
the
well­
mixed
spiked
and
wspiked
samples
of
the
simulated
downstream
water:
let
stand
for
1
to
3
hours.
procedures
.
a0
and
.
bm
allow
the
metal
to
equilibrate
set
with
the
effluent
before
the
solution
is
diluted
witbupstreamwater.

11.
For
the
test
using
the
laboratory
dilution
water,
either
of
the
following
procedures
may
be
used
to
prepare
the
test
solutions
for
the
test
chambers
and
the
'
chemistry
controls.
(
see
section
H.
1):
a.
Place
the
same
known
volume
of
the
laboratory
dilution
water
in
each
test
chamber;
add
the
necessary
amunt
of
metal,
which
will
be
different
for
each
treatment;
mix
thoroughly;
let
stand
for
1
to
3
hours.
b.
Prepare
a
large
volume
of
the
highest
test
concentration
in
the
laboratory
dilution
water;
perfom
serial
dilution
using
a
graduated
cylinder
and
the
well­
mixed
spiked
and
unspiked
samples
of
the
laboratory
dilution
water;
let
stand
for
1
to
3
hours.

12.
The
test
organisms,
which
have
been
acclimated
as
per
section
D.
l,
maat
be
added
to
the
test
chambers
for
the
site­
by­
side
tests
at
the
same
time.
The
time
at
which
the
test
organisms
are
placed
in
the
test
chambers
is
defined
as
the
beginning
of
the
tests,
which
maat
be
within
36
hours
of
the
collection
of
the
samples,
except
as
per
section
E.
7.

13.
Observe
the
test
organisms
and
record
the
effects
and
­
tans
as
specified
by
the
U.
S.
EPA
(
1993a,
b,
cI
and/
or
ASTM
(
1993a,
b,
c,
d,
e).
Especially
note
whether
the
effects,
mtcms,
and
time
course
of
toxicity
are
the
sams
in
the
side­
by­
side
tests.

14.
Whenever
solutions
are
renewed,
sufficient
solution
should
be
prepared
to
allow
for
chemical
analyses.

54
H.
Chemical
and
Other
Measurements
1.

2.

3.

4.
To
reduce
the
possibility
of
contamination
of
test
solutions
before
or
during
tests,
thexmmeters
and.
probes
for
measuring
pH
and
dissolved
oxygen
must
xmt
be
placed
in
test
chambers
that
will
provide
data
concerning
effects
on
test
organisms
or
data
concerning
the
concentration
of
the
metal.
Thus
measurements
of
pH,
dissolved
oxygen,
and
temperature
before
or
during
a
test
aast
be
performed
either
on
l
chexnistry
controls.
that
contain
test
organisms
and
are
fed
the
same
as
the
other
test
chambers
or
on
aliquots
that
are
removed
from
the
test
chambers.
The
other
measurements
may
be
performed
on
the
actual
test
solutions
at
the
beginning
and/
or
end
of
the
test
or
the
renewal.

Hardness
(
in
fresh
water)
or
salinity
(
in
salt
water),
pH,
alkalinity,
TSS,
and
T0C
mat
be
measured
on
the
upstream
water,
the
effluent,
the
simulated
and/
or
actual
downstream
water,
and
the
laboratory
dilution
water.
Measurement
of
conductivity
and/
or
total
dissolved
solids
(
TDS)
is
recofmen
dad
in
fresh
water.

Dissolved
oxygen,
pH,
and
tqerature
aut
be
measured
during
the
test
at
the
times
specified
by
the
U.
S.
EPA
(
1993a,
b,
c)
and/
or
ASTM
(
1993a,
b,
c,
d,
e).
The
measurements
must
be
performed
on
the
same
schedule
for
both
of
the
side­
by­
side
tests.
Measurements
mat
be
perfomed
on
both
the
chemistry
controls
and
actual
test
solutions
at
the
end
of
the
test.

Roth
total
recoverable
and
dissolved
metal
mast
be
measured
in
the
upstream
water,
the
effluent,
and
appropriate
test
solutions
for
each
of
the
tests.
a.
The
analytical
measurements
should
be
sufficiently
sensitive
and
precise
that
variability
in
analyses
will
not
greatly
increase
the
variability
of
the
WERs.
If
the
detection
limit
of
the
analytical
method
that
will
be
used
to
detexmine
the
metal
is
greater
than
one­
tenth
of
the
CCC
oz
CRC
that
is
to
be
adjusted,
the
analytical
method
should
probably
be
improved
or
replaced
(
see
Appendix
C).
If
additional
sensitivity
is
needed,
it
is
.
often
useful
to
separate
the
metal
from
the
matrix
because
this
will
simultaneously
concentrate
the
metal
and
remove
interferences.
Replicate
analyses
should
be
performed
if
necessary
to
reduce
the
impact
of
analytical
variability.
1)
EPA
methods
(
U.
S..
EPA
1983b,
1991c)
should
usually
be
used
for
both
total
recoverable
and
dissolved
measurements,
but
in
some
cases
alternate
methods
might
have
to
be
used
in
order
to
achieve
the
necessary
sensitivity.
Approval
for
use
of
55
b.

C.

d.
alternate
methods
is
to
be
requested
from
the
appropriate
regulatory
authority.

it
is
often
p88ibh
to
Store
samples
and
then
analyze
only
those
that
are
needed
to
calculate
the
results
of
the
toxicity
tests.
For
dichotomous
data
(
e.
g.,
either­
or
data;
data
concerning
survival),
the
metal
in
the
following
rrut
he
measured:
1)
all
concentrations
in
which
some,
but
not
all,
of
the
test
organisms
were
adversely
affected.
2)
the
highest
concentration
that
did
not
adversely
affect
say
test
organisms.
3)
the
lowest
concentration
that
adversely
affected
all
of
the
te8t
organisms.
4)
the
controls.
For
data
that
are
not
dichotomous
(
i.
e.,
for
count
and
continuous
data),
the
metal
in
the
controls
and
in
the
treatments
that
define
the
concentration­
effect
curve
mat
be
measured;
measurement
of
the
concentrations
of
metals
in
other
treatments
is
desirable.
In
each
treatment
in
which
the
concentration
of
metal
is
to
he
measured,
&
th
w
total
recovq,
g&
le
Qog
:
1)

2)
Samples
mat
be
taken
for
measurement
of
total
r8COVerable
metal
once
for
a
static
test,
and
once
for
each
renewal
for
renewal
tests;
in
renewal
tests,
the
samples
are
to
be
taken
after
the
organisms
have
bees
transferred
to
the
new
test
solutions.
When
total
recoverable
metal
is
measured
in
a
test
chamber,
the
whole
solution
in
the
chamber
mast
be
mixed
before
the
sample
is
taken
for
analysis;
the
solution
in
the
test
chamber
mat
mot
be
acidified
before
the
sample
is
taken.
The
sample
mat
be
acidified
after
it
is
placed
in
the
8­
1~
container.
Dissolved
metal
rout
he
measured
at
the
beginning
and
end
of
each
static
test;
in
a
renewal
test,
the
dissolved
metal
mat
he
measured
at
the
beginning
of
the
test
and
just
before
the
solution
is
renewed
the
first
time.
when
dissolved
metal
is
measured
in
a
test
chamber,
the
whole
solution
in
the
test
chamber
mat
be
mixed
before
a
sufficient
axmunt
is
removed
for
filtration;
the
solution
in
the
test
chamber
mat
sot
be
acidified
before
the
sample
is
taken.
The
sample
mast
be
filtered
within
one
hour
after
it
is
taken,
and
the
filtrate
mut
be
acidified
after
filtration.

56
5.
Replicates,
matrix
spikes,
and
other.
QA/
QC
checks
mat
be
performed
as
required
by
the
U.
S.
EPA
(
1983a,
1991c).

I.
Calculating
and
Interpreting
the
Results
1.
To
prevent
roundoff
error
in
subsequent
calculations,
at
least
four
significant
digits
must
be
retained
in
all
endpoints,
WRRs,
and
FWERS.
This
requirement
is
not
based
on
mathematics
or
statistics
and
does
not
reflect
the
precision
of
the
value;
its
purpose
is
to
minimize
concern
about
the
effects
of
rounding
off
on
a
site­
specific
criterion.
All
of
these
numbers
are
intermediate
values
in
the
calculation
of
permit
limits
and
should
not
be
rounded
off
as
if
they
were
values
of
ultimate
concern.

2.
Evaluate
the
acceptability
of
each
toxicity
test
individually.
a.
If
the
procedures
used
deviated
from
those
specified
above,
particularly
in
terms
of
acclimation,
randomization,
temperature
control,
measurement
of
metal,
and/
or
disease
or
disease­
treatment,
the
test
should
be
rejected;
if
deviations
were
numerous
and/
or
substantial,
the
test
mat
be
rejected.
b.
Most
tests
are
unacce#
table
if
more
than
10
percent
of
the
organisms
in
the
controls
were
adversely
affected,
but
the
limit
is
higher
for
some
tests;
for
the
tests
reconmended
in
Appendix
I,
the
references
given
should
be
consulted.
c.
If
M
LC50
or
EC50
is
to
be
calculated:
1)
The
percent
of
the
organisms
that
were
adversely
affected
must
have
been
less
than
50
percent,
and
should
have
been
less
than
37
percent,
in
at
least
one
treatment
other
than
the
control.
21
In
laboratory
dilution
water
the
percent
of
the
organisms
that
were
adversely
affected
mast
have
been
greater
than
50
percent,
and
should
have
bees
greater
than
63
percent,
in
at
least
one
treatment.
In
site
water
the
percent
of
the
organisms
that
were
adversely
affected
should
have
been
greater
than
63
percent
in
at
least
one
treatment.
(
The
LC50
or
EC50
may
be
a
l
greater
than'
or
'
less
than'
value
in
site
water,
but
not
in
laboratory
dilution
water.)
3)
If
there
was
M
inversion
in
the
data
(
i.
e.,
if
a
lower
concentration
killed
or
affected
a
greater
percentage
of
the
organisms
than
a
higher
concentration),
it
mast
not
have
involved
more
than
two
concentrations
that
killed
or
affected
between
20
and
80
percent
of
the
test
organisms.
If
M
endpoint
other
than
M
LC5O
or
EC50
is
used
or
if
Abbott's
formula
is
used,
the
above
requirements
will
have
to
be
modified
accordingly.

57
3.
d.
Determine
whether
there
was
anything
unusual
about
the
test
results
that
would
stake
them
questionable.
e.
If
solutions
were
not
renewed
every
24
hours,
the
concentration
of
dissolved
metal
mast
rrat
have
.
decreased
by
smre
than
50
percent
from
the
beginning
to
the
end
of
a
static
test
or
from
the
beginning
to
the
end
of
a
renewal
in
a
renewal
test
in
test
concentrations
that
were
used
in
the
calculation
of
the
re8ults
of
the
test.

Determine
whether
the
effects,
mtoms,
and
time
course
of
toxicity
­
8
the
same
in
the
side­
by­
side
tests
in
the
site
water
and
the
laboratory
dilution
water.
For
­
18,
did
nmrtality
occur
in
one
acute
test,
but
Wilization
in
the
other?
Did
most
deaths
occur
before
24
hour8
in
one
test,
but
after
24
hours
in
the
other?
In
8ublethal
test8,
was
the
most
sensitive
effect
the
same
in
both
tests?
If
the
effects,
8ysQtmU,
and/
or
time
course
of
too&
city
were
different,
it
might
indicate
that
the
test
is
questionable
or
that
additivity,
synergism,
or
aatagonha
occurred
in
site
water.
Such
information
might
be
particularly
u8eful
when
comparing
tests
that
produced
ummually
low
or
high
WERs
with
tests
that
produced
moderate
WERs.

4.
Calculate
the
rerults
of
each
test:
a.
If
the
data
for
the
most
sensitive
effect
are
dichotwurr,
the
endpoint
mmt
be
calculated
as
a
LC50,
X50,
X25,
EC25,
etc.,
u8ing
method8
described
by
the
U.
S.
EPA
(
1993a)
or
ASTM
(
1993a).
If
two
or
mOrc
treatments
affected
betwemn
0
and
100
percent
in
both
tests
in
a
side­
by­
side
pair,
probit
analysis
mast
be
u8ed
to
calculate
results
of
both
tests,
unless
the
probit
model
is
rejected
by
the
goodness
of
fit
test
in
one
or
both
of
the
acute
tests.
If
probit
analysis
cannot
be
used,
either
because
fewer
than
two
percentage8
are
between
0
and
100
percent
or
because
the
model
doe8
not
fit
the
data,
cooqputational
interpolation
mat
be
used
(
see
Figure
5);
graphical
interpolation
mm+
mot
be
used.
1)
The
8au1e
endpoint
(
LC50,
EC25,
etc.)
and
the
ssnm
coaqmtational
method
mat
be
used
for
both
tests
wedin
the
calculation
of
a
WER.
2)
The
selection
of
the
percentage
used
to
define
the
endpoint
might
be
influenced
by
the
percent
effect
that
occurred
in
the
tests
and
the
correspondence
with
the
CCC
and/
Or
CK.
3)
If
no
treatment
killed
or
affected
more
than
50
percent
of
the
test
organisms
and
the
test
was
otherwise
acceptable,
the
IL50
or
EC50
should
be
reported
to
be
greater
than
the
highest
test
concentration.

50
4)
If
no
treatment
other
than
the
control
killed
or
affected
lesa
than
50
percent
of
the
test
organisms
and
the
test
was
otherwise
acceptable,
the
LC50
or
EC50
should
be
reported
to
be
less
than
the
lowest
test
concentration.
b.
If
the
data
for
the
most
sensitive
effect
are
not
dichotormus,
the
endpoint
auf
be
calculated
using
a
regression­
type
method
(
Hoekstra
and
Van
Ewijk
1993;
Stephan
and
Rogers
19851,
such
as
linear
interpolation
(
U.
S.
EPA
1993b,
c)
or
a
nonlinear
regression
method
(
Barnthouse
et
al.
1987;
Suter
et
al.
1987;
Bruce
and
Versteeg
1992).
The
selection
of
the
percentage
used
to
define
the
endpoint
might
be
influenced
by
the
percent
effect
that
occurred
in
the
tests
and
the
correspondence
with
the
CCC
and/
Or
CMC.
The
endpoints
in
the
side­
by­
side
tests
mast
be
based
on
the
am
amount
of
the
saw
adverse
effect
so
that
the
WER
is
a
ratio
of
identical
endpoints.
The
8aXthe
ccqmtational
method
mast
be
used
for
both
tests
used
in
the
calculation
of
the
WER.
c.
Both
total
recoverable
and
dissolved
results
should
be
calculated
for
each
test.
d.
Results
should
be
based
on
the
time­
weighted
average
measured
metal
concentrations
(
see
Figure
6).

5.
The
acceptability
of
the
laboratory
dilution
water
mast
be
evaluated
by
c­
ring
results
obtained
with
two
sensitive
tests
using
the
laboratory
dilution
water
with
results
that
were
obtained
using
a
comparable
laboratory
dilution
water
is
one
or
more
other
laboratories
(
see
sections
C.
3.
b
and
F.
5).
a.
If,
after
taking
into
account
any
known
effect
of
hardness
on
toxicity,
the
new
values
for
the
endpoints
of
m
of
the
tests
are
(
1)
more
than
a
factor
of
1.5
higher
than
the
respective
Mets
of
the
values
from
the
other
laboratories
or
(
2)
more
than
a
factor
of
1.5
lower
than
the
respective
xne~
s
of
values
froan
the
other
laboratories
or
(
3)
lower
than
the
respective
lowest
values
available
from
other
laboratories
or
(
4)
higher
than
the
respective
highest
values
available
from
other
laboratories,
the
new
and
old
data
mast
bc
carefully
evaluated
to
determine
whether
the
laboratory
dilution
water
used
in
the
WER
detexmination
was
acceptable.
For
exaqle,
there
might
have
been
M
error
in
the
chemical
measurements,
which
might
INM
that
the
results
of
all
tests
performed
in
the
WER
determination
need
to
be
adjusted
and
that
the
WER
would
not
change.
It
is
also
possible
that
the
metal
is
more
or
less
toxic
in
the
laboratory
dilution
water
used
in
the
WER
determination.
Further,
if
the
new
data
were
based
on
measured
concentrations
but
the
old
data
were
based
on
nominal
concentrations,
the
new
data
59
should
probably
be
considered
to
be
better
than
the
old.
Evaluation
of
results
of
any
other
toxicity
tests
on
the
same
or
a
different
metal
using
the
same
laboratory,
dilution
water
might
be
useful.
b.
If,
after
taking
into
account
any
known
effect
of
hardness
on
toxicity,
the;
new
values
for
the
endpoints
of
the
two
tests
are
not
either
p
or
m
m
in
comq?
arison
than
data
frm
other
laboratories
(
as
per
section
a
above)
and
if
both
of
the
new
values
are
within
a
factor
of
2
of
the
respective
means
of
the
previously
available
values
or
are
within
the
ranges
of
the
values,
the
laboratory
dilution
water
used
in
the
WER
determination
iS
acceptable.
c.
A
contiol
chart
approach
may
be
used
if
sufficient
data
are
available.
d.
If
the
cosgmrisons
do
not
indicate
that
the
laboratory
dilution
water,
test
method,
etc.,
are
acceptable,
the
tests
probably
rhould
be
considered
unacceptable,
unles8.
other
toxicity
data
are
available
to
indicate
that
they
are
acceptable.
Csison
of
re8ults
of
tests
between
laboratories
provides
a
check
on
all
aspects
of
the
test
procedure;
the
a@
msis
here
i8
on
the
quality
of
the
laboratory
dilution
water
because
all
other
aspects
of
the
side­
by­
side
tests
onwhichtheWERis
basedmutbethc
same,
except
pos8ibly
for
the
concentrations
of
metal
used
and
the
acclimation
ju8t
prior
to
the
beginning
of
the
tests.

6.
If
all
tlhe
necersary
tests
and
the
laboratory
dilution
water
are
acceptable,
a
WER
mast
be
calculated
by
dividing
the
endpoint
obtained
using
site
water
by
the
endpoint
obtained
u8ing
laboratory
dilution
water.
a.
If
both
a
pri.
UUy
t88t
and
a
SeCOn&
Iy
test
were
conducted
uriag
both
waters,
WERs
amt
be
calculated
for
both
t88t8.
b.
Both
total
recoverable
and
dissolved
WERs
mat
be
calculated.
c.
If
the
detection
limit
of
the
analytical
method
used
to
measure
the
smtal
is
above
the
endpoint
in
laboratory
dilution
water,
the
detection
limit
mast
be
used
as
the
endpoint,
which
will
result
in
a
lower
WER
than
would
be
obtained
if
the
actual
concentration
had
been
Bmasurad.
If
the
detection
limit
of
the
analytical
method
used
is
above
the
endpoint
in
site
water,
a
WER
cannot
be
determined.

7.
Inve8tigation
of
the
WER.
a.
The
results
of
the
chemical
measuresmnts
of
hardness,
alkalinity,
pH,
TSS,
Tot,
total
recoverable
metal,
dissolved
metal,
etc.,
on
the
effluent
and
the
upstream
water
shouldbe
examined
and
c­
red
with
previously
available
values
for
the
effluent
and
upstream
water,

60
respectively,
to
determine
whether
the
samples
were
representative
and
to
get
some
indication
of
the
variability
in
the
composition,
especially
as
it
might
affect
the
toxicity
of
the
metal
and
the
WER,
and
to
see
if
the
WER
correlates
with
one
or
more
of
the
measurements.
b.
The
WERs
obtained
with
the
primary
and
secondary
tests
should
be
compared
to
determine
whether
the
WER
obtained
with
the
secondary
test
confirmed
the
WER
obtained
with
the
primary
test.
Equally
sensitive
tests
are
expected
to
give
WERs
that
are
similar
(
e.
g.,
within
a
factor
of
31,
whereas
a
test
that
is
less
sensitive
will
probably
give
a
smaller
WER
than
a
more
sensitive
test
(
see
Appendix
D).
Thus
a
WER
obtained
with
a
primary
test
is
considered
confirmed
if
either
or
both
of
the
following
are
true:
1)
the
WERs
obtained
with
the
primary
and
secondary
tests
are
within
a
factor
of
3.
2)
the
test,
regardless
of
whether
it
is
the
primary
or
secondary
test,
that
gives
a
higher
endpoint
in
the
laboratory
dilution
water
also
gives
the
larger
WRR.
If
the
WER
obtained
with
the
secondary
test
does
not
confirm
the
WER
obtained
with
the
primary
test,
the
results
should
be
investigated.
In
addition,
WERs
probably
should
be
determined
using
both
tests
the
next
time
samples
are
obtained
and
it
would
be
desirable
to
determine
a
WER
using
a
third
test.
It
is
also
important
to
evaluate
what
the
results
imply
about
the
protectiveness
of
any
proposed
site­
specific
criterion.
C.
If
the
WER
is
larger
than
5,
it
should
be
investigated.
11
If
the
endpoint
obtained
using
the
laboratory
dilution
water
was
lower
than
previously
reported
lowest
value
or
was
more
than
a
factor
of
two
lower
than
an
existing
Species
Mean
Acute
Value
in
a
criteria
document,
additional
tests
in
the
laboratory
dilution
water
are
probably
desirable.
2)
If
a
total
recoverable
WER
was
larger
than
5
but
the
dissolved
WER
was
not,
is
the
metal
one
whose
WER
is
likely
to
be
affected
by
TSS
and/
or
TQC
and
was
the
concentration
of
TSS
and/
or
TOC
high?
Was
there
a
substantial
difference
between
the
total
recoverable
and
dissolved
concentrations
of
the
metal
in
the
downstream
water?
3)
If
both
the
total
recoverable
and
dissolved
WERs
were
larger
than
5,
is
it
likely
that
there
is
nontoxic
dissolved
metal
in
the
downstream
water?
d.
The
adverse
effects
and
the
time­
course
of
effects
in
the
side­
by­
side
tests
should
be
compared.
If
they
are
different,
it
might
indicate
that
the
site­
water
test
is
questionable
or
that
additivity,
synergism,
or
antagonism
occurred
in
the
site
water.
This
might
be
especially
important
if
the
WER
obtained
with
the
61
secondary
test
did
not
confirm
the
WER
obtained
with
the
primary
test
or
if
the
WER
was
very
large
or
small.

J.
Reporting
the
Results
0.
If
at
least
one
WER
determined
with
the
primary
test
was
confirmed
by
a
WER
that
was
simultaneously
determined
with
the
secondary
test,
the
cmcFWER
and/
or
the
CCCFWER
should
be
derived
as
described
in
section
A.
5.

9.
All
data
generated
during
the
determination
of
the
WER
should
be
examined
to
see
if
there
are
any
implications
for
the
national
or
site­
specific
aquatic
life
criterion.
a.
If
there
are
data
for
a
species
for
which
data
were
not
previously
available
or
unusual
data
for
a
species
for
which
data
were
available,
need
to
be
revised.
the
national
criterion
might
b.
If
the
primary
test
gives
an
LC50
or
EC50
in
laboratory
dilution
water
that
is
the
same
as
the
national
CMC,
the
resulting
site&
specific
CRC
should
be
similar
to
the
LC50
that
was
obtained
with
the
primary
test
using
downstream
water.
Such
relationships
might
serve
as
a
check
on
the
applicability
of
the
use
of
WERs.
c.
If
data
indicate
that
the
site­
specific
criterion
would
not
adequately
protect
a
critical
species,
the
site­
specific
criterion
probably
should
be
lowered.

A
report
of
the
experimental
determination
of
a
WER
to
the
appropriate
regulatory
authority
mu8t
include
the
following:
1.

2.

3.

4.

5.

6.

7.
Name(
s)
of
the
investigator(
s),
name
and
location
of
the
laboratory,
and
dates
of
initiation
and
termination
of
the
tests.
A
description
of
the
laboratory
dilution
water,
including
source,
preparation,
and
any
demonstrations
that
an
aquatic
species
can
survive,
grow,
and
reproduce
in
it.
The
name,
location,
and
description
of
the
discharger,
a
description
of
the
effluent,
and
the
design
flows
of
the
effluent
and
the
upstream
water.
A
description
of
each
sampling
station,
date,
and
time,
with
an
explanation
of
why
they
were
selected,
and
the
flows
of
the
upstream
water
and
the
effluent
at
the
time
the
samples
were
collected.
The
procedures
used
to
obtain,
transport,
and
store
the
samples
of
the
upstream
water
and
the
effluent.
Any
pretreatment,
such
as
filtration,
of
the
effluent,
site
water,
and/
or
laboratory
dilution
water.
Results
of
all
chemical
and
physical
measurements
on
upstream
water,
effluent,
actual
and/
or
simulated
downstream
water,
and
laboratory
dilution
water,
including
hardness
(
or
salinity),
alkalinity,
pH,
and
concentrations
of
total
recoverable
metal,
dissolved
metal,
TSS,
and
TOC.

62
8.

9.

10.

11.

12.

13.

14.

15.
Description
of
the
experimental
design,
test
chambers,
depth
and
volume
of
solution
in
the
chambers,
loading
and
lighting,
and
numbers
of
organisms
and
chambers
per
treatment.
Source
and
grade
of
the
metallic
salt,
and
how
the
stock
solution
was
prepared,
including
any
acids
or
bases
used.
Source
of
the
test
organisms,
scientific
name
and
how
verified,
age,
life
stage,
means
and
ranges
of
weights
and/
or
lengths,
observed
diseases,
treatments,
holding
and
acclimation
procedures,
and
food.
The
average
and
range
of
the
temperature,
pH,
hardness
(
or
salinity),
and
the
concentration
of
dissolved
oxygen
(
as
8
saturation
and
as
mg/
L)
during
acclimation,
and
the
method
used
to
measure
them.
The
following
must
be
presented
for
each
toxicity
test:
a.
The
average
and
range
of
the
measured
concentrations
of
dissolved
owgen,
as
%
saturation
and
as
mg/
L.
b.
The
average
and
range
of
the
test
temperature
and
the
method
used
to
measure
it.
c.
The
schedule
for
taking
samples
of
test
solutions
and
the
methods
used
to
obtain,
prepare,
and
store
them.
d.
A
surgery
table
of
the
total
recoverable
and
dissolved
concentrations
of
the
metal
in
each
treatment,
including
all
controls,
in
which
they
were
measured.
e.
A
summary
table
of
the
values
of
the
toxicological
variable(
s)
for
each
treatment,
including
all
controls,
in
sufficient
detail
to
allow
an
independent
statistical
analysis
of
the
data.
f.
The
endpoint
and
the
method
used
to
calculate
it.
g.
Comparisons
with
other
data
obtained
by
conducting
the
same
test
on
the
same
metal
using
laboratory
dilution
water
in
the
same
and
different
laboratories;
such
data
may
be
from
a
criteria
document
or
from
another
source.
h.
Anything
unusual
about
the
test,
any
deviations
from
the
procedures
described
above,
and
any
other
relevant
information.
All
differences,
other
than
the
dilution
water
and
the
concentrations
of
metal
in
the
test
solutions,
between
the
side­
by­
side
tests
using
laboratory
dilution
water
and
site
water.
Comparison
of
results
obtained
with
the
primary
and
secondary
tests.
The
WER
and
an
explanation
of
its
calculation.

A
report
of
the
derivation
of
a
FWER
mU8t
include
the
following:
1.
A
report
of
the
determination
of
each
WER
that
was
determined
for
the
derivation
of
the
FWER;
all
WERs
determined
with
secondary
tests
Prrmst
be
reported
along
with
all
WERs
that
were
determined
with
the
primary
test.

63
2.
The
design
flow
of
the
upstream
water
and
the
effluent
and
the
hardness
used
in
the
derivation
of
the
permit
limits
if
the
criterion
for
the
metal
is
hardness­
dependent.
3.
A
sussnary
table
must
be
presented
that
contains
the
following
for
each
WER
that
was
derived:
a.
the
value
of
the
WER
and
the
two
endpoints
from
which
it
was
calculated.
b.
the
hWER
calculated
from
the
WER.
c.
the
test
and
species
that
was
used.
d.
the
date
the
samples
of
effluent
and
site
water
were
collected.
e.
the
flows
of
the
effluent
and
upstream
water
when
the
samples
were
taken.
f.
the
following
information
concerning
the
laboratory
dilution
water,
effluent,
upstream
water,
and
actual
and/
or
simulated
downstream
water:
hardness
(
salinity),
alkalinity,
pH,
and
concentrations
of
total
recoverable
metal,
dissolved
metal,
TSS,
and
TQC.
4.
A
detailed
explanation
of
how
the
F'WEF!
was
derived
from
the
WERS
that
are
in
the
suna~
ry
table.

64
METHOD
2:
DETERMINING
cccWERS
FOR
AREAS
AWAY
FROM
PLUMES
Method
2
might
be
viewed
as
a
simple
process
wherein
samples
of
site
water
are
obtained
from
locations
within
a
large
body
of
fresh
or
salt
water
(
e.
g.,
an
ocean
or
a
large
lake,
reservoir,
or
estuary),
a
WER
is
determined
for
each
sample,
and
the
FWER
is
calculated
as
the
geometric
mean
of
some
or
all
of
the
WERs.
In
reality,
Method
2
is
not
likely
to
produce
useful
results
unless
substantial
resources
are
devoted
to
planning
and
conducting
the
study.
Most
sites
to
which
Method
2
is
applied
will
have
long
retention
times,
complex
mixing
patterns,
and
a
number
of
dischargers.
Because
metals
are
persistent,
the
long
retention
times
mean
that
the
sites
are
likely
to
be
defined
to
cover
rather
large
areas;
thus
such
sites
will
herein
be
referred
to
generically
as
"
large
sites".
Despite
the
differences
between
them,
all
large
sites
require
similar
special
considerations
regarding
the
determination
of
WERs.
Because
Method
2
is
based
on
samples
of
actual
surface
water
(
rather
than
simulated
surface
water),
no
sample
should
be
taken
in
the
vicinity
of
a
plume
and
the
method
should
be
used
to
determine
cccWERs,
not
cmcWERs.
If
WERs
are
to
be
determined
for
more
than
one
metal,
Appendix
F
should
be
read.

Method
2
uses
many
of
the
same
methodologies
as
Method
1,
such
as
those
for
toxicity
tests
and
chemical
analyses.
Because
the
sampling
plan
is
crucial
to
Method
2
and
the
plan
has
to
be
based
on
site­
specific
considerations,
this
description
of
Method
2
will
be
more
qualitative
than
the
description
of
Method
1.

Method
2
is
based
on
use
of
actual
surface
water
samples,
but
use
of
simulated
surface
water
might
provide
information
that
is
useful
for
some
purposes:
1.
It
might
be
desirable
to
compare
the
WERs
for
two
discharges
that
contain
the
same
metal.
This
might
be
accomplished
by
selecting
an
appropriate
dilution
water
and
preparing
two
simulated
surface
waters,
one
that
contains
a
known
concentration
of
one
effluent
and
one
that
contains
a
known
concentration
of
the
other
effluent.
The
relative
magnitude
of
the
two
WERs
is
likely
to
be
more
useful
than
the
absolute
values
of
the
WERs
themselves.
2.
It
might
be
desirable
to
determine
whether
the
eWER
for
a
particular
effluent
is
additive
with
the
WER
of
the
site
water
(
see
Appendix
G).
This
can
be
studied
by
determining
WERs
for
several
different
known
concentrations
of
the
effluent
in
site
water.
3.
An
event
such
as
a
rain
might
affect
the
WER
because
of
a
change
in
the
water
quality,
but
it
might
also
reduce
the
WER
just
by
dilution
of
refractory
metal
or
TSS.
A
proportional
decrease
in
the
WER
and
in
the
concentration
of
the
metal
(
such
as
by
dilution
of
refractory
metal)
will
not
result
in
underprotection;
if,
however,
dilution
decreases
the
WER
65
proportionally
more
than
it
decreases
the
concentration
of
metal
in
the
downstream
water,
underprotection
is
likely
to
occur.
This
is
essentially
a
determination
of
whether
the
WER
is
additive
when
the
effluent
is
diluted
with
rain
water
(
see
Appendix
G).
4.
An
event
that
increases
TSS
might
increase
the
total
recoverable
concentration
of
the
metal
and
the
total
recoverable
WER
without
having
much
effect
on
either
the
dissolved
concentration
or
the
dissolved
WER.
In
all
four
cases,
the
use
of
simulated
surface
water
is
useful
because
it
allows
for
the
determination
of
WERs
using
known
concentrations
of
effluent.

An
important
step
in
the
determination
of
any
WER
is
to
define
the
area
to
be
included
in
the
site.
The
major
principle
that
should
be
applied
when
defining
the
area
is
the
same
for
all
sites:
The
site
should
be
neither
too
small
nor
too
large.
If
the
area
selected
is
too
small,
permit
limits
might
be
unnecessarily
controlled
by
a
criterion
for
an
area
outside
the
site,
whereas
too
large
an
area
might
unnecessarily
incorporate
spatial
complexities
that
are
not
relevant
to
the
discharge(
s)
of
concern
and
thereby
unnecessarily
increase
the
cost
of
determining
the
WER.
Applying
this
principle
is
likely
to
be
more
difficult
for
large
sites
than
for
flowing­
water
sites.

Because
WERs
for
large
sites
will
usually
be
determined
using
actual,
rather
than
simulated,
surface
water,
there
are
five
major
considerations
regarding
experimental
design
and
data
analysis:

1.
Total
recoverable
WERs
at
large
sites
might
vary
so
much
across
time,
location,
and
depth
that
they
are
not
very
useful.
An
assumption
should
be
developed
that
an
appropriately
defined
WER
will
be
much
more
similar
across
time,
location,
and
depth
within
the
site
than
will
a
total
recoverable
WER.
If
such
an
assumption
cannot
be
used,
it
is
likely
that
either
the
FWER
will
have
to
be
set
equal
to
the
lowest
WER
and
be
overprotective
for
most
of
the
site
or
separate
site­
specific
criteria
will
have
to
be
derived
for
two
or
more
sites.
a.
One
assumption
that
is
likely
to
be
worth
testing
is
that
the
dissolved
WER
varies
much
less
across
time,
location,
and
depth
within
a
site
than
the
total
recoverable
WER.
If
the
assumption
proves
valid,
a
dissolved
WER
can
be
applied
to
a
dissolved
national
water
quality
criterion
to
derive
a
dissolved
site­
specific
water
quality
criterion
that
will
apply
to
the
whole
site.
b.
A
second
assumption
that
might
be
worth
testing
is
that
the
WER
correlates
with
a
water
quality
characteristic
such
as
TSS
or
TOC
across
time,
location,
and
depth.
c.
Another
assumption
that
might
be
worth
testing
is
that
the
dissolved
and/
or
total
recoverable
WER
is
mostly
due
to
66
2.
nontoxic
metal
rather
than
to
a
water
quality
characteristic.
that
reduces
toxicity.
If
this
is
true
and
if
there
is
variability
in
the
WER,
the
WER
will
correlate
with
the
concentration
of
metal
in
the
site
water.
.
This
is
similar
to
the
first
assumption,
but
this
one
can
allow
use
of
both
total
recoverable
and
dissolved
WRRs,
whereas
the
first
one
only
allows
use
of
a
dissolved
WER.
If
WERs
are
too
variable
to
be
useful
and
no
way
can
be
found
to
deal
with
the
variability,
additional
sampling
will
probably
be
required
in
order
to
develop
a
WER
and/
or
a
site­
specific
water
quality
criterion
that
is
either
(
a)
spatially
and/
or
temporally
dependent
or
(
b)
constant
and
environmentally
conservative
for
nearly
all
conditions.

An
experimental
design
should
be
developed
that
tests
whether
the
assumption
is
of
practical
value
across
the
range
of
conditions
that
occur
at
different
times,
locations,
and
depths
within
the
site.
Each
design
has
to
be
formulated
individually
to
fit
the
specific
site.
The
design
should
try
to
take
into
account
the
times,
locations,
and
depths
at
which
the
extremes
of
the
physical,
chemical,
and
biological
conditions
occur
within
the
site,
which
will
require
detailed
information
concerning
the
site.
In
addition,
the
experimental
design
should
balance
available
resources
with
the
need
for
adequate
sampling.
a.
Selection
of
the
number
and
timing
of
sampling
events
should
take
into
account
seasonal,
weekly,
and
daily
considerations.
Intensive
sampling
should
occur
during
the
two
most
extreme
seasons,
with
confirmatory
sampling
during
the
other
two
seasons.
Selection
of
the
day
and
time
of
sample
collection
should
take
into
account
the
discharge
schedules
of
the
major
industrial
and/
or
municipal
discharges.
For
example,
it
might
be
appropriate
to
collect
samples
during
the
middle
of
the
week
to
allow
for
reestablishment
of
steady­
state
conditions
after
shutdowns
for
weekends
and
holidays;
alternatively,
end­
of­
the­
week
slug
discharges
are
routine
in
some
situations.
In
coastal
sites,
the
tidal
cycle
might
be
important
if
facilities
discharge,
for
example,
over
a
four­
hour
period
beginning
at
slack
high
tide.
Because
the
highest
concentration
of
effluent
in
the
surface
water
probably
occurs
at
ebb
tide,
determination
of
WERs
using
site
water
samples
obtained
at
this
time
might
result
in
inappropriately
large
WERs
that
would
result.
in
underprotection
at
other
times;
samples
with
unusually
large
WERs
might
be
especially
useful
for
testing
assumptions.
The
importance
of
each
consideration
should
be
determined
on
a
case­
by­
case
basis.
b.
Selection
of
the
number
and
locations
of
stations
to
be
sampled
within
a
sampling
event
should
consider
the
site
as
a
whole
and
take
into
account
sources
of
water
and
discharges,
mixing
patterns,
and
currents
(
and
tides
in
coastal
areas).
If
the
site
has
been
adequately
67
characterized,
an
acceptable
design
can
probably
be
developed
using
existing
information
concerning
(
1)
sources
of
the
metal
and
other
pollutants
and
(
2)
the
spatial
and
temporal
distribution
of
concentrations
of
the
met&
l
and
water
quality
factors
that
might
affect
the
toxicity
of
the
metal.
Samples
should
not
be
taken
within
or
near
mixing
zones
or
plumes
of
dischargers;
dilution
models
(
U.
S.
EPA
1993)
and
dye
dispersion
studies
(
Kilpatrick
1992)
can
indicate
areas
that
should
definitely
be
avoided.
mRs,
current
charts,
hydrodynamic
models,
and
water
quality
models
used
to
allocate
waste
loads
and
derive
permit
limits
are
likely
to
be
helpful
when
determining
when
and
where
to
obtain
site­
water
samples.
Available
information
might
provide
an
indication
of
the
acceptability
of
site
water
for
testing
selected
species.
The
larger
and
more
complex
the
site,
the
greater
the
number
of
sampling
locations
that
will
be
needed.
c.
In
addition
to
determining
the
horizontal
location
of
each
sampling
station,
the
vertical
location
(
i.
e.,
depth)
of
the
sampling
point
needs
to
be
selected.
Known
mixing
regimes,
the
presence
of
vertical
stratification
of
TSS
and/
or
salinity,
concentration
of
metal,
effluent
plumes,
tolerance
of
test
species,
and
the
need
to
obtain
samples
of
site
water
that
span
the
range
of
site
conditions
should
be
considered
when
selecting
the
depth
at
which
the
sample
is
to
be
taken.
Some
decisions
concerning
depth
cannot
be
made
until
information
is
obtained
at
the
time
of
sampling;
for
example,
a
conductivity
meter,
salinometer,
or
transmissometer
might
be
useful
for
determining
where
and
at
what
depth
to
collect
samples.
Turbidity
might
correlate
with
TSS
and
both
might
relate
to
the
toxicity
of
the
metal
in
site
water;
salinity
can
indicate
whether
the
test
organisms
and
the
site
water
are
compatible.
Because
each
site
is
unique,
specific
guidance
cannot
be
given
here
concerning
either
the
selection
of
the
appropriate
number
and
locations
of
sampling
stations
within
a
site
or
the
frequency
of
sampling.
All
available
information
concerning
the
site
should
be
utilized
to
ensure
that
the
times,
locations,
and
depths
of
samples
span
the
range
of
water
quality
characteristics
that
might
affect
the
toxicity
of
the
metal:
a.
High
and
low
concentrations
of
TSS.
b.
High
and
low
concentrations
of
effluents.
c.
Seasonal
effects.
d.
The
range
of
tidal
conditions
in
saltwater
situations.
The
sampling
plan
should
provide
the
data
needed
to
allow
an
evaluation
of
the
usefulness
of
the
assumption(
s)
that
the
qerimental
design
is
intended
to
test.
Statisticians
should
play
a
key
role
in
experimental
design
and
data
analysis,
but
professional
judgment
that
takes
into
account
pertinent
biological,
chemical,
and
toxicological
considerations
is
at
least
as
important
as
rigorous
statistical
analysis
when
68
interpreting
the
data
and
determining
the
degree
to
which
the
data
correspond
to
the
assumption(
s).

3.
The
details
of
each
sampling
design
should
be
formulated
with
the
aid
of
people
who
understand
the
site
and
people
who
have
a
working
knowledge
of
WERs.
Because
of
the
complexity
of
designing
a
WER
study
for
large
sites,
the
design
team
should
utilize
the
combined
expertise
and
experience
of
individuals
from
the
appropriate
EPA
Region,
states,
municipalities,
dischargers,
environmental
groups,
and
others
who
can
constructively
contribute
to
the
design
of­
the
study.
Building
a
team
of
cooperating
aquatic
toxicologists,
aquatic
chemists,
limnologists,
oceanographers,
water
quality
modelers,
statisticians,
individuals
from
other
key
disciplines,
as
well
as
regulators
and
those
regulated,
who
have
knowledge
of
the
site
and
the
site­
specific
procedures,
is
central
to
success
of
the
derivation
of
a
WER
for
a
large
site.
Rather
than
submitting
the
workplan
to
the
appropriate
regulatory
authority
(
and
possibly
the
Water
Management
Division
of
the
EPA
Regional
Office)
for
comment
at
the
end,
they
should
be
members
of
the
team
from
the
beginning.

4.
Data
from
one
sampling
event
should
always
be
analyzed
prior
to
the
next
sampling
event
with
the
goal
of
improving
the
sampling
design
as
the
study
progresses.
For
example,
if
the
toxicity
of
the
metal
in
surface
water
samples
is
related
to
the
concentration
of
TSS,
a
water
quality
characteristic
such
as
turbidity
might
be
measured
at
the
time
of
collection
of
water
samples
and
used
in
the
selection
of
the
concentrations
to
be
used
in
the
WER
toxicity
tests
in
site
water.
At
a
minimum,
the
team
that
interprets
the
results
of
one
sampling
event
and
plans
the
next
should
include
an
aquatic
toxicologist,
a
metals
chemist,
a
statistician,
and
a
modeler
or
other
user
of
the
data.

5.
The
final
interpretation
of
the
data
and
the
derivation
of
the
FWER(
s)
should
be
performed
by
a
team.
Sufficient
data
are
likely
to
be
available
to
allow
a
quantitative
estimate
of
experimental
variation,
differences
between
species,
and
seasonal
differences.
It
will
be
necessary
to
decide
whether
one
site­
specific
criterion
can
be
applied
to
the
whole
area
or
whether
separate
site­
specific
criteria
need
to
be
derived
for
two
or
more
sites.
The
interpretation
of
the
data
might
produce
two
or
more
alternatives
that
the
appropriate
regulatory
authority
could
subject
to
a
cost­
benefit
analysis.

Other
aspects
of
the
determination
of
a
WER
for
a
large
site
are
likely
to
be
the
same
as
described
for
Method
1.
For
example:
a.
WERs
should
be
determined
using
two
or
more
sensitive
species;
the
suggestions
given
in
Appendix
I
should
be
considered
when
selecting
the
tests
and
species
to
be
used.

69
b.
Chemical
analyses
of
site
water,
laboratory
dilution
water,
and
test
solutions
should
follow
the
requirements
for
the
specific
test
used
and
those
given
in
this
document.
c.
If
tests
in
many
surface
water
samples
are
compared
to.
one
test
in
a
laboratory
dilution
water,
it
is
very
important
that
that
one
test
be
acceptable.
Use
of
(
1)
rangefinding
tests,
(
2)
additional
treatments
beyond
the
standard
five
concentrations
plus
controls,
and
(
3)
dilutions
that
are
functions
of
the
known
concentration­
effect
relationships
obtained
with
the
toxicity
test
and
metal
of
concern
will
help
ensure
that
the
desired
endpoints
and
WEF&
can
be
calculated.
d.
Measurements
of
the.
concentrations
of
both
total
recoverable
and
dissolved
metal
should
be
targeted
to
the
test
concentrations
whose
data
will
be
used
in
the
calculation
of
the
endpoints.
e.
Samples
of
site
water
and/
or
effluent
should
be
collected,
handled,
and
transported
so
that
the
tests
can
begin
as
soon
as
is
feasible.
f.
If
the
large
site
is
a
saltwater
site,
the
considerations
presented
in
Appendix
H
ought
to
be
given
attention.

70
Figure
2:
Calculating
an
Mjuotd
Geometric
Mean
Where
n
=
the
number
of
experimentally
determined
WERs
in
a
set,
the
'
adjusted
geometric
mean.
of
the
set
is
calculated
as
follows:

a.
Take
the
logarithm
of
each
of
the
WERs.
The
logarithms
can
be
to
any
base,
but
natural
logarithms
(
base
e)
are
preferred
for
reporting
purposes.
b.
Calculate
x'
=
the
arithmetic
mean
of
the
logarithms.
c.
Calculate
8
=
the
sample
standard
deviation
of
the
logarithms:

d.
Calculate
SE
=
the
standard
error
of
the
arithmetic
mean:
SE
­
s/
fi
.
e.
Calculate
A
I??­
(&,,
I
(
SE),
where
f+,,
is
the
value
of
Student's
t
statistic
for
a
one­
sided
probability
of
0.70
with
n
­
1
degrees
of
freedom.
The
values
of
to.,
for
some
common
degrees
of
freedom
(
df)
are:

df
to.
1
0.727
0.617
3
0.584
4
0.569
5
0.559
6
0.553
7
0.549
8
0.546
9
0.543
10
0.542
11
0.540
12
0.539
The
values
of
to,,
for
more
degrees
of
freedom
are
available,
for
example,
on
page
T­
5
of
Natrella
(
1966).
f.
Take
the
antilogarithm
of
A.

This
adjustment
of
the
geometric
mean
accounts
for
the
fact
that
the
means
of
fifty
percent
of
the
sets
of
WERs
are
expected
to
be
higher
than
the
actual
mean;
using
the
one­
sided
value
of
t
for
0.70
reduces
the
percentage
to
thirty.

71
Figtar
3:
An
­
10
Dwivation
of
a
FWER
tis
example
assumes.
that
cccWEF&
were
determined
monthly.
using
simulated
downstream
water
that
was
prepared
by
mixing
upstream
water
with
effluent
at
the
ratio
that
existed
when
the
samples
were
obtained.
Also,
the
flow
of
the
effluent
is
always
10
cfs,
and
the
design
flow
of
the
upstream
water
is
40
cfs.
(
Therefore,
the
downstream
flow
at
design­
flow
conditions
is
50
cfs.)
The
concentration
of
metal
in
upstream
water
at
design
flow
is
0.4
ug/
L,
and
the
CCC
is
2
ug/
L.
Each
FWER
is
derived
from
the
WRS
and
hWERs
that
are
available
through
that
month.

Month
March
April
­
Y
June
July
Aug.
Sept.
Oct.
Nov.
Dec.
Jan.
Feb.
10
10
10
ii
10
10
10
10
10
10
10
850
289
300
430
120
05
t:
150
110
180
244
0.8
;­;:
826.4
82.80
0.6
341.5
34.31
0.6
5:
aE
341.6
34.32
0.6
5.7=
475.8
47.74
0.4
7.0=
177.2
17.88
0.4
10.5'
196.1
19.77
0.4
12.0'
118.4
12.00
0.4
11.0'
119.2
12.08
0.4
7.5=
234.0
23.56
0.4
c
z­;
c
79.6
8.12
0.6
6:
lc
251.4
25.30
0.6
295.2
29.68
hWEX
Neither
Type
1
nor
Type
2;
the
downstream
flow
(
i.
e.,
of
the
eFLOW
and
the
UFLOW)
is
>
500
cfs.
the
sum
The
total
number
of
available
Type
1
and
Type
2
WERs
is
less
than
3.
A
Type
2
WER;
the
downstream
flow
is
between
100
and
500
cfs.
No
Type
1
WER
is
available;
the
FWER
is
the
lower
of
the
lowest
Type
2
WER
and
the
lowest
hWER.
A
Type
1
WER;
the
downstream
flow
is
between
50
and
iO0
cfs.
One
Type
1
WEFI
is
available;
the
FWER
is
the
geometric
mean
of
all
Type
1
and
Type
2
WERs.
Two
or
more
Type
1
WERs
are
available
and
the
range
is
less
than
a
factor
of
5;
the
FWER
is
the
adjusted
geometric
mean
(
see
Figure
2)
of
the
Type
1
WERs,
because
all
the
hWERs
are
higher.
TWo
or
more
Type
1
WERs
are
available
and
the
range
is
not
greater
than
a
factor
of
5;
the
FWEFI
is
the
lowest
hWER
because
the
lowest
hWER
is
lower
than
the
adjusted
geometric
mean
of
the
Type
1
WEF&.
l.
ob
l.
ob
l..
Ob
5.7d
5.7d
6.80'
10.69"
10.880
10.880
8.12h
8.12"
8.12'

72
Figure
4:
Roduciag
the
Impact
of
Bacpertital
Variation
When
the
FWER
is
the
lowest
of,
for
example,
three
WERs,
the
impact
of
experimental
variation
can
be
reduced
by
conducting
additional
primary
tests.
Xf
the
endpoint
of
the
secondary
test
is
above
the
CMC
or
CCC
to
which
the
FWER
is
to
be
applied,
the
additional
tests
can
also
be
conducted
with
the
secondary
test.

Month
April
WY
June
Lowest
Month
April
WY
June
Lowest
Case
1
Case
1
(
Primary
Test)

4.801
2.552
9.164
2.552
(
Primary
Test)

4.801
2.552
9.164
Case
3
(
Primary
(
Second.
Geo.
Test)
Test)
Mean
4.801
3.163
3.897
2.552
5.039
3.586
9.164
7.110
8.072
3.586
Case
2
(
Primary
Geometric
Test1
Mean
3.565
4.137
4.190
3.270
6.736
7.857
3.270
Case
4
(
P;
riyry
(
Second.
Geo.
t)
Test)
Mean
4.801
3.163
3.897
2.552
2.944
2.741
9.164
7.110
8.072
2.741
uses
the
individual
WERs
obtained
with
the
primary
test
for
the
three
months,
and
the*
PWER
is
the
lowest
of
the
three
WERS.
In
Case
2,
duplicate
primary
tests
were
conducted
in
each
month,
so
that
a
geometric
mean
could
be
calculated
for
each
month;
the
PWER
is
the
lowest
of
the
three
geometric
means.

In
Cases
3
and
4,
both
a
primary
test
and
a
secondary
test
were
conducted
each
month
and
the
endpoints
for
both
tests
in
laboratory
dilution
water
are
above
the
CMC
or
CCC
to
which
the
PWER
is
to
be
applied.
In
both
of
these
cases,
therefore,
the
FWER
is
the
lowest
of
the
three
geometric
means.

The
availability
of
these
alternatives
does
not
mean
that
they
are
necessarily
cost­
effective.

73
Fiw
5:
Calculating
an
I&
SO
(
or
BCSO)
by
Interpolation
When
fewer
than
two
treatments
kill
some
but
not
all
of
the
exposed
test
organisms,
a
statistically
sound
estimate
of
an
~~
50
cannot
be
calculated.
Some
programs
and
methods
produce
LCSOs
when
there
are
fewer
than
two
'
partial
kills',
but
such
results
are
obtained
using
interpolation,
not
statistics.
If
(
a)
a
test
is
otherwise
acceptable,
(
b)
a
sufficient
number
of
organisms
are
exposed
to
each
treatment,
and
(
c)
the
concentrations
are
sufficiently
close
together,
a
test
with
zero
or
one
partial
kill
can
provide
all
the
information
that
is
needed
concerning
the
LCSO.
in
LCSO
calculated
by
interpolation
should
probably
be
called
an
'
approximate
LCSO'
to
acknowledge
the
lack
of
a
statistical
basis
for
its
calculation,
but
this
does
not
imply
that
such
an
LCSO
provides
no
useful
toxicological
information.
If
desired,
the
binomial
test
can
be
used
to
calculate
a
statistically
sound
probability
that
the
true
LCSO
lies
between
two
tested
concentrations
(
Stephan
1977).

Although
more
complex
interpolation
methods
can
be
used,
they
will
not
produce
a
more
useful
LCSO
than
the
method
described
here.
Inversions
in
the
data
between
two
test
concentrations
should
be
removed
by
pooling
the
mortality
data
for
those
two
concentrations
and
calculating
a
percent
mortality
that
is
then
assigned
to
both
concentrations.
Logarithms
to
a
base
other
than
10
can
be
used
if
desired.
If
Pl
and
P2
are
the
percentages
of
the
test
organisms
that
died
when
exposed
to
concentrations
Cl
and
C2,
respectively,
and
if
Cl
<
c2,
Pl
<
P2,
0
s
Pl
5
50,
and
50
5
P2
s
100,
then:

c
=
Log
Cl
+
Puog
c2
­
Log
Cl)

Lc50
­
lo=

If
Pl
=
0
and
P2
=
100,
tcso­
m
If
Pl
=
P2
If
Pl
=
50,
Lcso=
m
=
50,
LCSO
=
Cl.
If
P2
=
50,
LCSO
=
c2.
If
Cl
=
4
mg/
L,
C2
=
7
rag/
L,
Pl
=
15
%,
and
P2
=
100
%,
then
LCSO
=
5.036565
mg/
L.

Besides
the
mathematical
requirements
given
above,
the
following
toxicological
recommendations
are
given
in
sections
G.
8
and
1.2:
a.
0.65
<
Cl/
C2
<
0.99.
b.
0
5
Pl
<
37.
c.
63
<
P2
S
100.

74
tiguro
6:
Calculating
a
Time­
Weighted
Average
If
a
sampling
plan
(
e.
g.,
for
measuring
metal
in
a
treatment
in
a
toxicity
test)
is
designed
so
that
a
series
of
values
are
obtained
over
time
in
such
a
way
that
each
value
contains
the
same
amount
of
information
(
i.
e.,
represents
the
same
amount
of
time),
then
the
most
meaningful
average
is
the
arithmetic
average.
In
most
cases,
however,
when
a
series
of
values
is
obtained
over
time,
some
values
contain
more
information
than
others;
in
these
cases
the
most
meaningful
average
is
a
time­
weighted
average
(
TWA).
If
each
value
contains
the
same
amount
of
information,
the
arithmetic
average
will
equal
the
TWA.

A
TWA
is
obtained
by
multiplying
each
value
by
a
weight
and
then
dividing
the
sum
of
the
products
by
the
sum
of
the
weights.
The
simplest
approach
is
to
let
each
weight
be
the
duration
of
time
that
the
sample
represents.
Except
for
the
first
and
last
samples,
the
period
of
time
represented
by
a
sample
starts
halfway
to
the
previous
sample
and
ends
halfway
to
the
next
sample.
The
period
of
time
represented
by
the
first
sample
starts
at
the
beginning
of
the
test,
and
the
period
of
time
represented
by
the
last
sample
ends
at
the
end
of
the
test.
Thus
for
a
96­
hr
toxicity
test,
the
sum
of
the
weights
will
be
96
hr.

The
following
are
hypothetical
examples
of
grab
samples
taken
from
96­
hr
flow­
through
tests
for
two
comon
sampling
regimes:

Sampling
Cont.
Weight
Product
Time­
weighted
average
time
(
hr)
(
mu/
L)
(
hr)
(
hr)
(
mu/
L)
(
ma/
L)

0
12
48
576
96
14
48
96
i%

2:
8
12
6
24
1::
48
7
24
168
72
9
24
216
96
8
u
96
1248/
96
=
13.00
720/
96
=
7.500
When
all
the
weights
are
the
same,
the
arithmetic
average
equals
the
TWA.
Similarly,
if
only
one
sample
is
taken,
both
the
arithmetic
average
and
the
TWA
equal
the
value
of
that
sample.

The
rules
are
more
complex
for
composite
samples
and
for
samples
from
renewal
tests.
In
all
cases,
however,
the
sampling
plan
can
be
designed
so
that
the
TWA
equals
the
arithmetic
average.

75
REFERENCES
ASTM.
1993a.
Guide
for
Conducting
Acute
Toxicity
Tests
with
Fishes,
Macroinvertebrates,
and
Amphibians.
Standard
E729.
American
Society
for
Testing
and
Materials,
Philadelphia,
PA.

ASTM.
1993b.
Guide
for
Conducting
Static
Acute
Toxicity
Tests
Starting
with
Embryos
of
Four
Species
of
Saltwater
Bivalve
Molluscs.
Standard
E724.
American
Society
for
Testing
and
Materials,
Philadelphia,
PA.

ASTM.
1993c.
Guide
for
Conducting
Renewal
Life­
Cycle
Toxicity
Tests
with
Daphnia
magna.
Standard
E1193.
American
Society
for
Testing
and
Materials,
Philadelphia,
PA.

ASTM.
1993d.
Guide
for
Conducting
Early
Life­
Stage
Toxicity
Tests
with
Fishes.
Standard
E1241.
American
Society
for
Testing
and
Materials,
Philadelphia,
PA.

ASTM.
1993e.
Guide
for
Conducting
Three­
Brood,
Renewal
Toxicity
Tests
with
Ceriodaphnia
dubia.
Standard
E1295.
American
Society
for
Testing
and
Materials,
Philadelphia,
PA.

ASTM.
1993f.
Guide
for
Conducting
Acute
Toxicity
Tests
on
Aqueous
Effluents
with
Fishes,
Macroinvertebrates,
and
Amphibians.
Standard
E1192.
American
Society
for
Testing
and
Materials,
Philadelphia,
PA.

Barnthouse,
L.
W.,
G.
W.
Suter,
A.
E.
Rosen,
and
J.
J.
Beauchamp.
1987.
Estimating
Responses
of
Fish
Populations
to
Toxic
Contaminants.
Environ.
Toxicol.
Chem.
6:
811­
824.

Bruce,
R.
D.,
and
D.
J.
Versteeg.
1992.
A
Statistical
Procedure
for
Modeling
Continuous
Toxicity
Data.
Environ.
Toxicol.
Chem.
11:
1485­
1494.

Hoekstra,
J.
A.,
and
P.
H.
Van
Ewijk.
1993.
Alternatives
for
the
No­
Observed­
Effect
Level.
Environ.
Toxicol.
Chem.
12:
187­
194.

Kilpatrick,
F.
A.
1992.
Simulation
of
Soluble
Waste
Transport
and
Buildup
in
Surface
Waters
Using
Tracers.
Open­
File
Report
92­
457.
U.
S.
Geological
Survey,
Books
and
Open­
File
Reports,
Box
25425,
Federal
Center,
Denver,
CO
80225.

Natrella,
M.
G.
1966.
Experimental
Statistics.
National
Bureau
of
Standards
Handbook
91.
(
Issued
August
1,
1963;
reprinted
October
1966
with
corrections).
U.
S.
Government
Printing
Office,
Washington,
DC.

76
Prothro,
M.
G.
1993.
Memorandum
titled
"
Office
of
Water
Policy
and
Technical
Guidance
on
Interpretation
and
Implementation
of
Aquatic
Life
Metals
Criteria'.
October
1.

Stephan,
C.
E.
1977.
Methods
for
Calculating
an
LC50.
In:
Aquatic
Toxicology
and
Hazard
Evaluation.
(
F.
L.
Mayer
and
J.
L.
Hamelink,
eds.)
ASTM
STP
634.
American
Society
for
Testing
and
Materials,
Philadelphia,
PA.
pp.
65­
84.

Stephan,
C.
E.,
and
J.
W.
Rogers.
1985.
Advantages
of
Using
Regression
Analysis
to
Calculate
Results
of
Chronic
Toxicity
Tests.
In:
Aquatic
Toxicology
and
Hazard
Assessment:
Eighth
Symposium.
(
R.
C.
Bahner
and
D.
J.
Hansen,
eds.)
ASTM
STP
891.
American
Society
for
Testing
and
Materials,
Philadelphia,
PA.
pp.
328­
338.

Suter,
G.
W.,
A.
E.
Rosen,
E.
Linder,
and
D.
F.
Parkhurst.
1987.
Endpoints
for
Responses
of
Fish
to
Chronic
Toxic
Exposures.
Environ.
Toxicol.
Chem.
6:
793­
809.

U.
S.
EPA.
1983a.
Water
Quality
Standards
Handbook.
Office
of
Water
Regulations
and
Standards,
Washington,
DC.

U.
S.
EPA.
1983b.
Methods
for
Chemical
Analysis
of
Water
and
Wastes.
EPA­
600/
4­
79­
020.
National
Technical
Information
Service,
Springfield,
VA.

U.
S.
EPA.
1984.
Guidelines
for
Deriving
Numerical
Aquatic
Site­
Specific
Water
Quality
Criteria
by
Modifying
National
Criteria.
EPA­
600/
3­
84­
099
or
PB85­
121101.
National
Technical
Information
Service,
Springfield,
VA.

U.
S.
EPA.
1985.
Guidelines
for
Deriving
Numerical
National
Water
Quality
Criteria
for
the
Protection
of
Aquatic
Organisms
and
Their
Uses.
PB85­
227049.
National
Technical
Information
Service,
Springfield,
VA.

U.
S.
EPA.
1991a.
Technical
Support
Document
for
Water
Quality­
based
Toxics
Control.
EPA/
505/
2­
90­
001
or
PB91­
127415.
National
Technical
Information
Service,
Springfield,
VA.

U.
S.
EPA.
1991b.
Manual
for
the
Evaluation
of
Laboratories
Performing
Aquatic
Toxicity
Tests.
EPA/
600/
4­
90/
031.
National
Technical
Information
Service,
Springfield,
VA.

U.
S.
EPA.
1991c.
Methods
for
the
Determination
of
Metals
in
Environmental
Samples.
EPA­
600/
4­
91­
010.
National
Technical
Information
Service,
Springfield,
VA.

77
U.
S.
EPA.
1992.
Interim
Guidance
on
Interpretation
and
Implementation
of
Aquatic
Life
Criteria
for
Metals.
Office
of
Science
and
Technology,
Health
and
Ecological
Criteria
Division,
Washington,
DC.

Uf;
S.
EPA.
1993a.
Methods
for
Measuring
the
Acute
Toxicity
of
Effluents
and
Receiving
Waters
to
Freshwater
and
Marine
Organisms.
Fourth
Edition.
EPA/
600/
4­
90/
027F.
National
Technical
Information
Service,
Springfield,
VA.

U.
S.
EPA.
1993b.
Short­
term
Methods
for
Estimating
the
Chronic
Toxicity
of
Effluents
and
Receiving
Waters
to
Freshwater
Organisms.
Third
Edition.
EPA/
600/
4­
91/
002.
National
Technical
Information
Service,
Springfield,
VA.

U.
S.
EPA.
1993c.
Short­
Term
Methods
for
Estimating
the
Chronic
Toxicity
of
Effluents
and
Receiving
Waters
to
Marine
and
Estuarine
Organisms.
Second
Edition.
EPA/
600/
4­
91/
003.
National
Technical
Information
Service,
Springfield,
VA.

U.
S.
EPA.
19938.
Dilution
Models
for
Effluent
Discharges.
Second
Edition.
EPA/
600/
R­
93/
139.
National
Technical
Information
Service,
Springfield,
VA.

78
mix
A:
comparison
of
WERs
Datormined
Using
Upstram
and
Downstream
Wator
The
'
Interim
Guidance'
concerning
metals
(
U.
S.
EPA
1992)
made
a
fundamental
change
in
the
way
WERs
should
be
experimentally
determined
because
it
changed
the
source
of
the
site
water.
The
earlier
guidance
(
U.
S.
EPA
1983,1984)
required
that
upstream
water
be
used
as
the
site
water,
whereas
the
newer
guidance
(
U.
S.
EPA
1992)
recommended
that
downstream
water
be
used
as
the
site
water.
The
change
in
the
source
of
the
site
water
was
merely
an
acknowledgement
that
the
WER
that
applies
at
a
location
in
a
body
of
water
should,
when
possible,
be
determined
using
the
water
that
occurs
at
that
location.

Because
the
change
in
the
source
of
the
dilution
water
was
expected
to
result
in
an
increase
in
the
magnitude
of
many
WERs,
interest
in
and
concern
about
the
determination
and
use
of
WERs
increased.
When
upstream
water
was
the
required
site
water,
it
was
expected
that
WERs
would
generally
be
low
and
that
the
determination
and
use
of
WERs
could
be
fairly
simple.
After
downstream
water
became
the
recommended
site
water,
the
determination
and
use
of
WERs
was
examined
much
more
closely.
It
was
then
realized
that
the
determination
and
use
of
upstream
WERs
was
more
complex
than
originally
thought.
It
was
also
realized
that
the
use
of
downstream
water
greatly
increased
the
complexity
and
was
likely
to
increase
both
the
magnitude
and
the
variability
of
many
WERs.
Concern
about
the
fate
of
discharged
metal
also
increased
because
use
of
downstream
water
might
allow
the
discharge
of
large
amounts
of
metal
that
has
reduced
or
no
toxicity
at
the
end
of
the
pipe.
The
probable
increases
in
the
complexity,
magnitude,
and
variability
of
WERs
and
the
increased
concern
about
fate,
increased
the
importance
of
understanding
the
relevant
issues
as
they
apply
to
WERs
determined
using
both
upstream
water
and
downstream
water.

A.
Characteristics
of
the
Site
Water
The
idealized
concept
of
an
upstream
water
is
a
pristine
water
that
is
relatively
unaffected
by
people.
In
the
real
world,
however,
many
upstream
waters
contain
naturally
occurring
ligands,
one
or
more
effluents,
and
materials
from
nonpoint
sources;
all
of
these
might
impact
a
WER.
If
the
upstream
water
receives
an
effluent
containing
Tot
and/
or
TSS
that
contributes
to
the
WER,
the
WER
will
probably
change
whenever
the
quality
or
quantity
of
the
TOC
and/
or
TSS
changes.
In
such
a
case,
the
determination
and
use
of
the
WER
in
upstream
water
will
have
some
of
the
increased
complexity
associated
with
use
of
downstream
water
and
some
of
the
concerns
associated
with
multiple­
discharge
situations
(
see
Appendix
F)
l
The
amount
of
complexity
will
depend
greatly
on
the
79
mxnber
and
type
of
upstream
point
and
nonpoint
sources,
the
frequency
and
magnitude
of
fluctuations,
and
whether
the
WER
is
being
determined
above
or
below
the
point
of
complete
mix
of
the
upstream
sources.

Downstream
water
is
a
mixture
of
effluent
and
upstream
water,
each
of
which
can
contribute
to
the
WEFT,
and
so
there
are
two
cosaponents
to
a
WER
determined
in
downstream
water:
the
effluent
corqmnent
and
the
upstream
component.
The
existence
of
1.

2.

3.

4.

5.

6.
these
two
cmponents
has
the
following
implications:
WERS
determined
using
downstream
water
are
likely
to
be
larger
and
more
variable
than
WEFIs
determined
using
upstream
water.
The
effluent
component
should
be
applied
only
where
the
effluent
occurs,
which
has
implications
concerning
implementation.
The
magnitude
of
the
effluent
caflponent
of
a
WER
will
depend
on
the
concentration
of
effluent
in
the
downstream
water.
(
A
consequence
of
this
is
that
the
effluent
component
will
be
zero
where
the
concentration
of
effluent
is
zero,
which
is
the
point
of
item
2
above.)
The
magnitude
of
the
effluent
component
of
a
WER
is
likely
to
vary
as
the
composition
of
the
effluent
varies.
Compared
to
upstream
water,
many
effluents
contain
higher
concentrations
of
a
wider
variety
of
substances
that
can
impact
the
toxicity
of
metals
in
a
wider
variety
of
ways,
and
so
the
effluent
component
of
a
WER
can
be
due
to
a
variety
of
chemical
effects
in
addition
to
such
factors
as
hardness,
alkalinity,
pH,
and
humic
acid.
Because
the
effluent
component
might
be
due,
in
whole
or
in
part,
to
the
discharge
of
refractory
metal
(
see
Appendix
D)
i
the
WER
cannot
be
thought
of
simply
as
being
caused
by
the
effect
of
water
quality
on
the
toxicity
of
the
metal.
Dealing
with
downstream
WERs
is
so
much
simpler
if
the
effluent
WER
(
eWER1
and
the
upstream
WER
(
WER)
are
additive
that
it
is
desirable
to
understand
the
concept
of
additivity
of
WERS,
its
experimental
determination,
and
its
use
(
see
Appendix
G).

B.
The
Implications
of
Mixing
Zones.

When
WERs
are
determined
using
upstream
water,
the
presence
or
absence
of
mixing
zones
has
no
impact;
the
cmcWER
and
the
cccWER
will
both
be
determined
using
site
water
that
contains
zero
percent
of
the
effluent
of
concern,
i.
e.,
the
two
WERs
will
be
determined
using
the
same
site
water.

When
WERs
are
determined
using
downstream
water,
the
magnitude
of
each
WER
will
probably
depend
on
the
concentration
of
effluent
in
the
downstream
water
used
(
see
Appendix
D).
The
concentration
of
effluent
in
the
site
water
will
depend
on
80
where
the
sample
is
taken,
which
will
not
be
the
same
for
the
cmcWER
and
the
cccWJ3R
if
there
are
mixing
zone(
s).
Most,
if
not
all,
discharges
have
a
chronic
(
CCC)
mixing
zone;
many,
but
not
all,
also
have
an
acute
(
CMC)
mixing
zone.
The
CMC
applies
at
all
points
except
those
inside
a
CMC
mixing
zone;
thus
if
there
is
no
CMC
mixing
zone,
the
CMC
applies
at
the
end
of
the
pipe.
The
CCC
applies
at
all
points
outside
the
CCC
mixing
zone.
It
is
generally
assumed
that
if
permit
limits
are
based
on
a
point
in
a
stream
at
which
both
the
CMC
and
the
CCC
apply,
the
CCC
will
control
the
permit
limits,
although
the
CMC
might
control
if
different
averaging
periods
are
appropriately
taken
into
account.
For
this
discussion,
it
will
be
assumed
that
the
same
design
flow
(
e.
g.,
7910)
is
used
for
both
the
CMC
and
the
CCC.

If
the
cmcwER
is
to
be
appropriate
for
use
inside
the
chronic
mixing
zone,
but
the
CCCWER
is
to
be
appropriate
for
use
outside
the
chronic
mixing
zone,
the
concentration
of
effluent
that
is
appropriate
for
use
in
the
determination
of
the
two
WERs
will
not
be
the
same.
Thus
even
if
the
same
toxicity
test
is
used
in
the
determination
of
the
uncWER
and
the
CCCWER,
the
two
WERs
will
probably
be
different
because
the
concentration
of
effluent
will
be
different
in
the
two
site
waters
in
which
the
WERs
are
determined.

If
the
CMC
is
only
of
concern
within
the
CCC
mixing
zone,
the
highest
relevant
concentration
of
metal
will
occur
at
the
edge
of
the
CMC
mixing
zone
if
there
is
a
CMC
mixing
zone;
the
highest
concentration
will
occur
at
the
end
of
the
pipe
if
there
is
no
CMC
mixing
zone.
In
contrast,
within
the
CCC
mixing
zone,
the
lowest
cmcWER
will
probably
occur
at
the
outer
edge
of
the
CCC
mixing
zone.
Thus
the
greatest
level
of
protection
would
be
provided
if
the
cmcWER
is
determined
using
water
at
the
outer
edge
of
the
CCC
mixing
zone,
and
then
the
calculated
site­
specific
CMC
is
applied
at
the
edge
of
the
CMC
mixing
zone
or
at
the
end
of
the
pipe,
depending
on
whether
there
is
an
acute
mixing
zone.
The
cmcWER
is
likely
to
be
lowest
at
the
outer
edge
of
the
CCC
mixing
zone
because
of
dilution
of
the
effluent,
but
this
dilution
will
also
dilute
the
metal.
If
the
cmcWER
is
determined
at
the
outer
edge
of
the
CCC
mixing
zone
but
the
resulting
site­
specific
CMC
is
applied
at
the
end
of
the
pipe
or
at
the
edge
of
the
CMC
mixing
zone,
dilution
is
allowed
to
reduce
the
WER
but
it
is
not
allowed
to
reduce
the
concentration
of
the
metal.
This
approach
is
environmentally
conservative,
but
it
is
probably
necessary
given
current
implementation
procedures.
(
The
situation
might
be
more
complicated
if
the
WER
is
higher
than
the
eWER
or
if
the
two
WERs
are
less­
than­
additive.)

A
comparable
situation
applies
to
the
CCC.
Outside
the
CCC
mixing
zone,
the
CMC
and
the
CCC
both
apply,
but
it
is
assumed
that
the
CMC
can
be
ignored
because
the
CCC
will
be
more
81
restrictive.
The
CCCWER
should
probably
be
determined
for
the
complete­
mix
situation,
but
the
site­
specific
CCC
will
have
to
be
met
at
the
edge
of
the
CCC
mixing
zone.
Thus
dilution
of
the
WER
from
the
edge
of
the
CCC
mixing
zone
to
the
point
of
complete
mix
is
taken
into
account,
but
dilution
of
the
metal
is
not.

If
there
is
neither
an
acute
nor
a
chronic
mixing
zone,
both
the
CMC
and
the
CCC
apply
at
the
end
of
the
pipe,
but
the
CCC
should
still
be
determined
for
the
complete­
mix
situation.

C.
Definition
of
site.

In
the
general
context
of
site­
specific
criteria,
a
.
site.
may
be
a
state,
region,
watershed,
waterbody,
segment
of
a
waterbody,
category
of
water
(
e.
g.,
ephemeral
streams),
etc.,
but
the
site­
specific
criterion
is
to
be
derived
to
provide
adequate
protection
for
the
entire
site,
however
the
site
is
defined.
Thus,
when
a
site­
specific
criterion
is
derived
using
the
Recalculation
Procedure,
all
species
that
'
occur
at
the
site.
need
to
be
taken
into
account
when
deciding
what
species,
if
any,
are
to
be
deleted
from
the
dataset.
Similarly,
when
a
site­
specific
criterion
is
derived
using
a
WER,
the
WER
is
to
be
adequately
protective
of
the
entire
site.
If,
for
example,
a
site­
specific
criterion
is
being
derived
for
an
estuary,
WERs
could
be
determined
using
samples
of
the
surface
water
obtained
from
various
sampling
stations,
which,
to
avoid
confusion,
should
not
be
called
'
sites'.
If
all
the
WERs
were
sufficiently
similar,
one
site­
specific
criterion
could
be
derived
to
apply
to
the
whole
estuary.
If
the
WERs
were
sufficiently
different,
either
the
lowest
WER
could
be
used
to
derive
a
site­
specific
criterion
for
the
whole
estuary,
or
the
data
might
indicate
that
the
estuary
should
be
divided
into
two
or
more
sites,
each
with
its
own
criterion.

The
major
principle
that
should
be
applied
when
defining
the
area
to
be
included
in
the
site
is
very
simplistic:
The
site
should
be
neither
too
small
nor
too
large.
1.
Small
sites
are
probably
appropriate
for
cmcWERs,
but
usually
are
not
appropriate
for
CCCWERS
because
metals
are
persistent,
although
some
oxidation
states
are
not
persistent
and
some
metals
are
not
persistent
in
the
water
column.
For
cccWERs,
the
smaller
the
defined
site,
the
more
likely
it
is
that
the
permit
limits
will
be
controlled
by
a
criterion
for
an
area
that
is
outside
the
site,
but
which
could
have
been
included
in
the
site
without
substantially
changing
the
WER
or
increasing
the
cost
of
determining
the
WER.
2.
Too
large
an
area
might
unnecessarily
increase
the
cost
of
determining
the
WER.
As
the
size
of
the
site
increases,

82
the
spatial
and
temporal
variability
is
likely
to
increase,
which
will
probably
increase
the
number
of
water
samples
in
which
WERs
will
need
to
be
determined
before
a
site­
specific
criterion
can
be
derived.
3.
Events
that
import
or
resuspend
TSS
and/
or
TW
are
likely
to
increase
the
total
recoverable
concentration
of
the
metal
and
the
total
recoverable
WER
while
having
a
much
smaller
effect
on
the
dissolved
concentration
and
the
dissolved
WER.
Where
the
concentration
of
dissolved
metal
is
substantially
more
constant
than
the
concentration
of
total
recoverable
metal,
the
site
can
probably
be
much
larger
for
a
dissolved
criterion
than
for
a
total
recoverable
criterion.
If
one
criterion
is
not
feasible
for
the
whole
area,
it
might
be
possible
to
divide
it
into
two
or
more
sites
with
separate
total
recoverable
or
dissolved
criteria
or
to
make
the
criterion
dependent
on
a
water
quality
characteristic
such
as
TSS
or
salinity.
4.
Unless
the
site
ends
where
one
body
of
water
meets
another,
at
the
outer
edge
of
the
site
there
will
usually
be
an
instantaneous
decrease
in
the
allowed
concentration
of
the
metal
in
the
water
column
due
to
the
change
from
one
criterion
to
another,
but
there
will
not
be
an
instantaneous
decrease
in
the
actual
concentration
of
metal
in
the
water
column.
The
site
has
to
be
large
enough
to
include
the
transition
zone
in
which
the
actual
concentration
decreases
so
that
the
criterion
outside
the
site
is
not
exceeded.
It
is,
of
course,
possible
in
some
situations
that
relevant
distant
conditions
(
e.
g.,
a
lower
downstream
pH)
will
necessitate
a
low
criterion
that
will
control
the
permit
limits
such
that
it
is
pointless
to
determine
a
WER.

When
a
WER
is
determined
in
upstream
water,
it
is
generally
assumed
that
a
downstream
effluent
will
not
decrease
the
WER.
It
is
therefore
assumed
that
the
site
can
usually
cover
a
rather
large
geographic
area.

When
a
site­
specific
criterion
is
derived
based
on
WERs
determined
using
downstream
water,
the
site
should
not
be
defined
in
the
same
way
that
it
would
be
defined
if
the
WER
were
determined
using
upstream
water.
The
eWER
should
be
allowed
to
affect
the
site­
specific
criterion
wherever
the
effluent
occurs,
but
it
should
not
be
allowed
to
affect
the
criterion
in
places
where
the
effluent
does
not
occur.
In
addition,
insofar
as
the
magnitude
of
the
effluent
component
at
a
point
in
the
site
depends
on
the
concentration
of
effluent,
the
magnitude
of
the
WER
at
a
particular
point
will
depend
on
the
concentration
of
effluent
at
that
point.
To
the
extent
that
the
eWER
and
the
WER
are
additive,
the
WER
and
the
concentration
of
metal
in
the
plume
will
decrease
proportionally
(
see
Appendix
G).

83
D.
The
variability
in
the
experimental
determination
of
a
WER.
When
WERs
are
determined
using
downstream
water,
the
following
considerations
should
be
taken
into
account
when
the
site
is
defined:
1.
If
a
site­
specific
criterion
is
derived
using
a
WER
that
applies
to
the
complete­
mix
situation,
the
upstream
edge
of
the
site
to
which
this
criterion
applies
should
be
the
point
at
which
complete
mix
actually
occurs.
If
the
site
to
which
the
complete­
mix
WEZ
is
applied
starts
at
the
end
of
the
pipe
and
extends
all
the
way
across
the
stream,
there
will
be
an
area
beside
the
plume
that
will
not
be
adequately
protected
by
the
site­
specific
criterion.
2.
Upstream
of
the
point
of
complete
mix,
it
will
usually
be
protective
to
apply
a
site­
specific
criterion
that
was
derived
using
a
WER
that
was
determined
using
upstream
water.
3.
The
plums
might
bs
an
area
in
which
the
concentration
of
metal
could
exceed
a
site­
specific
criterion
without
causing
toxicity
because
of
simultaneous
dilution
of
the
metal
and
the
eWER.
The
fact
that
the
plume
is
much
larger
than
the
mixing
zone
might
not
be
important
if
there
is
no
toxicity
within
the
plume.
As
long
as
the
concentration
of
metal
in
100
%
effluent
does
not
exceed
that
allowed
by
the
additive
portion
of
the
eWER,
from
a
toxicologic&
l
standpoint
neither
the
size
nor
the
definition
of
the
plume
needs
to
be
of
concern
because
the
metal
will
not
cause
toxicity
within
the
plume.
If
there
is
no
toxicity
within
the
plume,
the
area
in
the
plums
might
be
like
a
traditional
mixing
zone
in
that
the
concentration
of
metal
exceeds
the
site­
specific
criterion,
but
it
would
be
different
from
a
traditional
mixing
zone
in
that
the
level
of
protection
is
not
reduced.

Special
considerations
are
likely
to
be
necessary
in
order
to
take
into
account
the
ewER
when
defining
a
site
related
to
multiple
discharges
(
see
Appendix
F).

When
a
WER
is
determined
using
upstream
water,
the
two
major
sources
of
variation
in
the
WEB
are
(
a)
variability
in
the
quality
of
the
site
water,
which
might
be
related
to
season
and/
or
flow,
and
(
b)
experimental
variation.
Ordinary
day­
to­
day
variation
will
account
for
some
of
the
variability,
but
seasonal
variation
is
likely
to
be
more
important.

As
explained
in
Appendix
D,
variability
in
the
concentration
of
nontoxic
dissolved
metal
will
contribute
to
the
variability
of
both
total
recoverable
WERs
and
dissolved
WERs;
variability
in
the
concentration
of
nontoxic
particulate
metal
will
contribute
to
the
variability
in
a
total
recoverable
WER,
but
not
to
the
variability
in
a
dissolved
WER.
Thus,
dissolved
84
WERs
are
expected
to
be
less
variable
than
total
recoverable
WERs,
especially
where
events
conunonly
increase
TSS
and/
or
Tot.
In
some
cases,
therefore,
appropriate
use
of
analytical
chemistry
can
greatly
increase
the
usefulness
of
the
experimental
determination
of
WERs.
The
concerns
regarding
variability
are
increased
if
an
upstream
effluent
contributes
to
the
WER.

When
a
WER
is
determined
in
downstream
water,
the
four
major
sources
of
variability
in
the
WER
are
(
a)
variability
in
the
quality
of
the
upstream
water,
which
might
be
related
to
season
and/
or
flow,
(
b)
experimental
variation,
(
c)
variability
in
the
composition
of
the
effluent,
and
(
d)
variability
in
the
ratio
of
the
flows
of
the
upstream
water
and
the
effluent.
The
considerations
regarding
the
first
two
are
the
same
as
for
WERs
determined
using
upstream
water;
because
of
the
additional
sources
of
variability,
WERs
determined
using
downstream
water
are
likely
to
be
more
variable
than
WEXs
determined
using
upstream
water.

It
would
be
desirable
if
a
sufficient
number
of
WEZRs
could
be
determined
to
define
the
variable
factors
in
the
effluent
and
in
the
upstream
water
that
contribute
to
the
variability
in
WERs
that
are
determined
using
downstream
water.
Not
only
is
this
likely
to
be
very
difficult
in
most
cases,
but
it
is
also
possible
that
the
WER
will
be
dependent
on
interactions
between
constituents
of
the
effluent
and
the
upstream
water,
i.
e.,
the
eWER
and
uWER
might
be
additive,
more­
than­
additive,
or
less­
than­
additive
(
see
Appendix
G).
When
interaction
occurs,
in
order
to
completely
understand
the
variability
of
WERs
determined
using
downstream
water,
sufficient
tests
would
have
to
be
conducted
to
determine
the
means
and
variances
of:
the
effluent
component
of
the
WER.
t:
the
upstream
component
of
the
WER.
C.
any
interaction
between
the
two
components.
An
interaction
might
occur,
for
example,
if
the
toxicity
of
a
metal
is
affected
by
pH,
and
the
pH
and/
or
the
buffering
capacity
of
the
effluent
and/
or
the
upstream
water
vary
considerably.

An
increase
in
the
variability
of
WERs
decreases
the
usefulness
of
any
one
WER.
Compensation
for
this
decrease
in
usefulness
can
be
attempted
by
determining
WERs
at
more
times;
although
this
will
provide
more
data,
it
will
not
necessarily
provide
a
proportionate
increase
in
understanding.
Rather
than
determining
WERs
at
more
times,
a
better
use
of
resources
might
be
to
obtain
more
information
concerning
a
smaller
number
of
specially
selected
occasions.

It
is
likely
that
some
cases
will
be
so
complex
that
achieving
even
a
reasonable
understanding
will
require
unreasonable
resources.
In
contrast,
some
WERs
determined
using
the
85
methods
presented
herein
might
be
relatively
easy
to
understand
if
appropriate
chemical
measurements
are
performed
when
WERs
are
determined.
1.
If
the
variation
of
the
total
recoverable
WER
is
substantially
greater
than
the
variation
of
the
comparable
dissolved
WER,
there
is
probably
a
variable
and
substantial
concentration
of
particulate
nontoxic
metal.
It
might
be
advantageous
to
use
a
dissolved
WER
just
because
it
will
have
less
variability
than
a
total
recoverable
WER.
2.
If
the
total
recoverable
and/
or
dissolved
WER
correlates
with
the
total
recoverable
and/
or
dissolved
concentration
of
metal
in
the
site
water,
it
is
likely
that
a
substantial
percentage
of
the
metal
is
nontoxic.
In
this
case
the
WER
will
probably
also
depend
on
the
concentration
of
effluent
in
the
site
water
and
on
the
concentration
of
metal
in
the
effluent.
These
approaches
are
more
likely
to
be
useful
when
WERs
are
determined
using
downstream
water,
rather
than
upstream
water,
unless
both
the
magnitude
of
the
WER
and
the
concentration
of
the
metal
in
the
upstream
water
are
elevated
by
an
upstream
effluent
and/
or
events
that
increase
TSS
and/
or
Tot.

Both
of
these
approaches
can
be
applied
to
WERs
that
are
determined
using
actual
downstream
water,
but
the
second
can
probably
provide
much
better
information
if
it
is
used
with
WERs
determined
using
simulated
downstream
water
that
is
prepared
by
mixing
a
sample
of
the
effluent
with
a
sample
of
the
upstream
water.
In
this
way
the
composition
and
characteristics
of
both
the
effluent
and
the
upstream
water
can
be
determined,
and
the
exact
ratio
in
the
downstream
water
isknown.

Use
of
simulated
downstream
water
is
also
a
way
to
study
the
relation
between
the
WER
and
the
ratio
of
effluent
to
upstream
water
at
one
point
in
time,
which
is
the
most
direct
way
to
test
for
additivity
of
the
ewER
and
the
uWER
(
see
Appendix
G).
This
can
be
viewed
as
a
test
of
the
assumption
that
WERs
determined
using
downstream
water
will
decrease
as
the
concentration
of
effluent
decreases.
If
this
assumption
is
true,
as
the
flow
increases,
the
concentration
of
effluent
in
the
downstream
water
will
decrease
and
the
WER
will
decrease.
Obtaining
such
information
at
one
point
in
tims
is
useful,
but
confirmation
at
one
or
more
other
times
would
be
much
more
useful
.

E.
The
fate
of
metal
that
has
reduced
or
no
toxicity.

Metal
that
has
reduced
or
no
toxicity
at
the
end
of
the
pipe
might
be
more
toxic
at
some
time
in
the
future.
For
example,
metal
that
is
in
the
water
column
and
is
not
toxic
now
might
become
more
toxic
in
the
water
column
later
or
might
move
into
86
the
sediment
and
become
toxic.
If
a
WER
allows
a
surface
water
to
contain
as
much
toxic
metal
as
is
acceptable,
the
WER
would
not
be
adequately
protective
if
metal
that
was
nontoxic
when
the
WER
was
determined
became
toxic
in
the
water.
column,
unless
a
compensating
change
occurred.
Studies
of
the
fate
of
metals
need
to
address
not
only
the
changes
that
take
place,
but
also
the
rates
of
the
changes.

Concern
about
the
fate
of
discharged
metal
justifiably
raises
concern
about
the
possibility
that
metals
might
contaminate
sediments.
The'possibility
of
contamination
of
sediment
by
toxic
and/
or
nontoxic
metal
in
the
water
column
was
one
of
the
concerns
that
led
to
the
establishment
of
EPA's
sediment
quality
criteria
program,
which
is
developing
guidelines
and
criteria
to
protect
sediment.
A
separate
program
was
necessary
because
ambient
water
quality
criteria
are
not
designed
to
protect
sediment.
Insofar
as
technology­
based
controls
and
water
quality
criteria
reduce
the
discharge
of
metals,
they
tend
to
reduce
the
possibility
of
contamination
of
sediment.
Conversely,
insofar
as
WERs
allow
an
increase
in
the
discharge
of
metals,
they
tend
to
increase
the
possibility
of
contamination
of
sediment.

When
WERs
are
determined
in
upstream
water,
the
concern
about
the
fate
of
metal
with
reduced
or
no
toxicity
is
usually
small
because
the
WERs
are
usually
small.
In
addition,
the
factors
that
result
in
upstream
WERs
being
greater
than
1.0
usually
are
(
a)
natural
organic
materials
such
as
humic
acids
and
(
b)
water
quality
characteristics
such
as
hardness,
alkalinity,
and
pH.
It
is
easy
to
assume
that
natural
organic
materials
will
not
degrade
rapidly,
and
it
is
easy
to
monitor
changes
in
hardness,
alkalinity,
and
pH.
Thus
there
is
usually
little
concern
about
the
fate
of
the
metal
when
WERs
are
determined
in
upstream
water,
especially
if
the
WER
is
small.
If
the
WER
is
large
and
possibly
due
at
least
in
part
to
an
upstream
effluent,
there
is
more
concern
about
the
fate
of
metal
that
has
reduced
or
no
toxicity.

When
WERs
are
determined
in
downstream
water,
effluents
are
allowed
to
contain
virtually
unlimited
amounts
of
nontoxic
particulate
metal
and
nontoxic
dissolved
metal.
It
would
seem
prudent
to
obtain
some
data
concerning
whether
the
nontoxic
metal
might
become
toxic
at
some
time
in
the
future
whenever
(
1)
the
concentration
of
nontoxic
metal
is
large,
(
2)
the
concentration
of
dissolved
metal
is
below
the
dissolved
national
criterion
but
the
concentration
of
total
recoverable
metal
is
substantially
above
the
total
recoverable.
national
criterion,
or
(
3)
the
site­
specific
criterion
is
substantially
above
the
national
criterion.
It
would
seem
appropriate
to:
a.
Generate
some
data
concerning
whether
'
fate'
(
i.
e.,
environmental
processes)
will
cause
any
of
the
nontoxic
metal
to
become
toxic
due
to
oxidation
of
organic
matter,

87
oxidation
of
sulfides,
etc.
For
example,
a
WER
could
be
determined
using
a
sample
of
actual
or
simulated
downstream
water,
the
sample
aerated
for
a
period
of
time
(
e.
g.,
two
weeks),
the
pH
adjusted
if
necessary,
and
another
TIER
determined.
If
aeration
reduced
the
WER,
shorter
and
longer
periods
of
aeration
could
be
used
to
study
the
rate
of
change.
b.
Determine
the
effect
of
a
change
in
water
quality
characteristics
on
the
WER;
for
example,
determine
the
effect
of
lowering
the
pH
on
the
WER
if
influent
lowers
the
pH
of
the
downstream
water
within
the
area
to
which
the
site­
specific
criterion
is
to
apply.
c.
Determine
a
WEX
in
actual
downstream
water
to
demonstrate
whether
downstream
conditions
change
sufficiently
(
possibly
due
to
degradation
of
organic
matter,
multiple
dischargers,
etc.)
to
lower
the
WER
more
than
the
concentration
of
the
metal
is
lowered.
If
environmental
processes
cause
nontoxic
metal
to
become
toxic,
it
is
important
to
determine
whether
the
time
scale
involves
days,
weeks,
or
years.

When
WERs
are
determined
using
downstream
water,
the
site
water
contains
effluent
and
the
WER
will
take
into
account
not
only
the
constituents
of
the
upstream
water,
but
also
the
toxic
and
nontoxic
metal
and
other
constituents
of
the
effluent
as
they
exist
after
mixing
with
upstream
water.
The
determination
of
the
WER
automatically
takes
into
account
any
additivity,
synergism,
or
antagonism
between
the
metal
and
components
of
the
effluent
and/
or
the
upstream
water.
The
effect
of
Calcium,
magnesium,
and
various
heavy
metals
on
competitive
binding
by
such
organic
materials
as
humic
acid
is
also
taken
into
account.
Therefore,
a
site­
specific
criterion
derived
using
a
WER
is
likely
to
be
more
appropriate
for
a
site
than
a
national,
state,
or
recalculated
criterion
not
only
because
it
takes
into
account
the
water
quality
characteristics
of
the
site
water
but
also
because
it
takes
into
account
other
constituents
in
the
effluent
and
upstream
water.

Determination
of
WERs
using
downstream
water
causes
a
general
increase
in
the
complexity,
magnitude,
and
variability
of
WEI&,
and
an
increase
in
concern
about
the
fate
of
metal
that
has
reduced
or
no
toxicity
at
the
end
of
the
pipe.
In
addition,
there
are
some
other
drawbacks
with
the
use
of
downstream
water
in
the
determination
of
a
WER:
1.
It
might
serve
as
a
disincentive
for
some
dischargers
to
remove
any
more
organic
carbon
and/
or
particulate
matter
than
required,
although
WEF&
for
some
metals
will
not
be
related
to
the
concentration
of
Tot
or
TSS.

88
2.
If
conditions
change,
a
WER
might
decrease
in
the
future.
This
is
not
a
problem
if
the
decrease
is
due
to
a
reduction
in
nontoxic
metal,
but
it
might
be
a
problem
if
the
decrease
is
due
to
a
decrease
in
TOC
or
TSS
or
an
increase
in
competitive
binding.
3.
If
a
WBR
is
determined
when
the
effluent
contains
refractory
metal
but
a
change
in
operations
results
in
the
discharge
of
toxic
metal
in
place
of
refractory
metal,
the
site­
specific
criterion
and
the
permit
limits
will
not
provide
adequate
protection.
In
most
cases
chemical
monitoring
probably
will
not
detect
such
a
change,
but
toxicological
monitoring
probably
will.

Use
of
WERs
that
are
determined
using
downstream
water
rather
than
upstream
water
increases:
1.
The
importance
of
understanding
the
various
issues
involved
in
the
determination
and
use
of
WERs.
2.
The
importance
of
obtaining
data
that
will
provide
understanding
rather
than
obtaining
data
that
will
result
in
the
highest
or
lowest
WER.
3.
The
appropriateness
of
site­
specific
criteria.
4.
The
resources
needed
to
determine
a
WBR.
5.
The
resources
needed
to
use
a
WER.
6.
The
resources
needed
to
monitor
the
acceptability
of
the
downstream
water.
A
WER
determined
using
upstream
water
will
usually
be
smaller,
less
variable,
and
simpler
to
implement
than
a
WER
determined
using
downstream
water.
Although
in
some
situations
a
downstream
WBR
might
be
smaller
than
an
upstream
WER,
the
important
consideration
is
that
a
WER
should
be
determined
using
the
water
to
which
it
is
to
apply.

Reference%

U.
S.
EPA.
1983.
Water
Quality
Standards
Handbook.
Office
of
Water
Regulations
and
Standards,
Washington,
DC.

U.
S.
EPA.
1984.
Guidelines
for
Deriving
Numerical
Aquatic
Site­
Specific
Water
Quality
Criteria
by
Modifying
National
Criteria.
EPA­
600/
3­
84­
099
or
PB85­
121101.
National
Technical
Information
Service,
Springfield,
VA.

U.
S.
EPA.
1992.
Interim
Guidance
on
Interpretation
and
Implementation
of
Aquatic
Life
Criteria
for
Metals.
Office
of
Science
and
Technology,
Health
and
Ecological
Criteria
Division,
Washington,
DC.

89
Appmdix
B:
Tha
Roorlenrlrtion
Procodurm
NOTE:
The
National
Toxics
Rule
(
NTR)
does
not
allow
use
of
the
Recalculation
Procedure
in
the
derivation
of
a
site­
specific
criterion.
Thus
nothing
in
this
appendix
applies
to
jurisdictions
that
are
subject
to
the
NTR.

The
Recalculation
Procedure
is
intended
to
cause
a
site­
specific
criterion
to
appropriately
differ
from
a
national
aquatic
life
criterion
if
justified
by
demonstrated
pertinent
toxicological
differences
between
the
aquatic
species
that
occur
at
the
site
and
those
that
were
used
in
the
derivation
of
the
national
criterion.
There
are
at
least
three
reasons
why
such
differences
might
exist
between
the
two
sets
of
species.
First,
the
national
dataset
contains
aquatic
species
that
are
sensitive
to
many
pollutants,
but
these
and
comparably
sensitive
species
might
not
occur
at
the
site.
Second,
a
species
that
is
critical
at
the
site
might
be
sensitive
to
the
pollutant
and
require
a
lower
criterion.
(
A
critical
species
is
a
species
that
is
comnercially
or
recreationally
important
at
the
site,
a
species
that
exists
at
the
site
and
is
listed
as
threatened
or
endangered
under
section
4
of
the
Endangered
Species
Act,
or
a
species
for
which
there
is
evidence
that
the
loss
of
the
species
from
the
site
is
likely
to
cause
an
unacceptable
impact
on
a
commercially
or
recreationally
important
species,
a
threatened
or
endangered
species,
the
abundances
of
a
variety
of
other
species,
or
the
structure
or
function
of
the
coxmuunity.)
Third,
the
species
that
occur
at
the
site
might
represent
a
narrower
mix
of
species
than
those
in
the
national
dataset
due
to
a
limited
range
of
natural
environmental
conditions.
The
procedure
presented
here
is
structured
so
that
corrections
and
additions
can
be
made
to
the
national
dataset
without
the
deletion
process
being
used
to
take
into
account
taxa
that
do
and
do
not
occur
at
the
site;
in
effect,
this
procedure
makes
it
possible
to
update
the
national
aquatic
life
criterion.

The
phrase
'
occur
at
the
site.
includes
the
species,
genera,
families,
orders,
classes,
and
phyla
that:
are
usually
present
at
the
site.
k:
are
present
at
the
site
only
seasonally
due
to
migration.
C.
are
present
intermittently
because
they
periodically
return
to
or
extend
their
ranges
into
the
site.
d.
were
present
at'the
site
in
the
past,
are
not
currently
present
at
the
site
due
to
degraded
conditions,
and
are
expected
to
return
to
the
site
when
conditions
improve.
e.
are
present
in
nearby
bodies
of
water,
are
not
currently
present
at
the
site
due
to
degraded
conditions,
and
are
expected
to
be
present
at
the
site
when
conditions
improve.
The
taxa
that
'
occur
at
the
site"
cannot
be
determined
merely
by
sampling
downstream
and/
or
upstream
of
the
site
at
one
point
in
time.
'
Occur
at
the
site'
does
not
include
taxa
that
were
once
90
present
at
the
site
but
cannot
exist
at
the
site
now
due
to
permanent
physical
alteration
of
the
habitat
at
the
site
resulting
from
dams,
etc.

The
definition
of
the
.
site'
can
be
extremely
important
when
using
the
Recalculation
Procedure.
For
example,
the
number
of
taxa
that
occur
at
the
site
will
generally
decrease
as
the
size
of
the
site
decreases.
Also,
if
the
site
is
defined
to
be
very
small,
the
permit
limit
might
be
controlled
by
a
criterion
that
applies
outside
(
e.
g.,
downstream
of)
the
site.

Note:
If
the
variety
of
aquatic
invertebrates,
amphibians,
and
fishes
is
so
limited
that
species
in
fewer
than
eicrht
ies
occur
at
the
site,
the
general
Recalculation
Procedure
is
not
applicable
and
the
following
special
version
of
the
Recalculation
Procedure
must
be
used:
1.
Data
muat
be
available
for
at
least
one
species
in
each
of
the
families
that
occur
at
the
site.
2.
The
lowest
Species
Mean
Acute
Value
that
is
available
for
a
species
that
occurs
at
the
site
must
be
used
as
the
FAV.
3.
The
site­
specific
CMC
and
CCC
must
be
calculated
as
described
below
in
part
2
of
step
E,
which
is
titled
.
Determination
of
the
CMC
and/
or
CCC'.

The
concept
of
the
Recalculation
Procedure
is
to
create
a
dataset
that
is
appropriate
for
deriving
a
site­
specific
criterion
by
modifying
the
national
dataset
in
some
or
all
of
three
ways:
a.
Correction
of
data
that
are
in
the
national
dataset.
b.
Addition
of
data
to
the
national
dataset.
c.
Deletion
of
data
that
are
in
the
national
dataset.
All
corrections
and
additions
that
have
been
approved
by
U.
S.
EPA
are
required,
whereas
use
of
the
deletion
process
is
optional.
The
Recalculation
Procedure
is
more
likely
to
result
in
lowering
a
criterion
if
the
net
result
of
addition
and
deletion
is
to
decrease
the
number
of
genera
in
the
dataset,
whereas
the
procedure
is
more
likely
to
result
in
raising
a
criterion
if
the
net
result
of
addition
and
deletion
is
to
increase
the
number
of
genera
in
the
dataset.

The
Recalculation
Procedure
consists
of
the
following
steps:
A.
Corrections
are
made
in
the
national
dataset.
B.
Additions
are
made
to
the
national
dataset.
C;
The
deletion
process
may
be
applied
if
desired.
D.
If
the
new
dataset
does
not
satisfy
the
applicable
Minimum
Data
Requirements
(
MDRs),
additional
pertinent
data
mumt
be
generated;
if
the
new
data
are
approved
by
the
U.
S.
EPA,
the
Recalculation
Procedure
mwt
be
started
again
at
step
B
with
the
addition
of
the
new
data.
E.
The
new
CMC
or
CCC
or
both
are
determined.
F.
A
report
is
written.
Each
step
is
discussed
in
more
detail
below.

91
A.
Corrections
1.
Only
corrections
approved
by
the
U.
S.
EPA
may
be
made.
2.
The
concept
of
.
correction.
includes
removal
of
data
that
should
not
have
been
in
the
national
dataset
in
the
first
place.
The
concept
of
.
correction"
does
not
include
removal
of
a
datum
from
the
national
dataset
just
because
the
quality
of
the
datum
is
claimed
to
be
suspect.
If
additional
data
are
available
for
the
same
species,
the
U.
S.
EPA
will
decide
which
data
should
be
used,
based
on
the
available
guidance
(
U.
S.
EPA
1985);
also,
data
based
on
measured
concentrations
are
usually
preferable
to
those
based
on
nominal
concentrations.
3.
Two
kinds
of
corrections
are
possible:
a.
The
first
includes
those
corrections
that
are
known
to
and
have
been
approved
by
the
U.
S.
EPA;
a
list
of
these
will
be
available
from
the
U.
S.
EPA.
b.
The
second
includes
those
corrections
that
are
submitted
to
the
U.
S.
EPA
for
approval.
If
approved,
these
will
be
added
to
EPA's
list
of
approved
corrections.
4.
Selective
corrections
are
not
allowed.
All
corrections
on
EPA's
newest
list
mrmt
be
made.

B.
Additioag
1.
2.

3.

P
Only
additions
approved
by
the
U.
S.
EPA
may
be
made.
Wo
kinds
of
additions
are
possible:
a.
The
first
includes
those
additions
that
are
known
to
and
have
been
approved
by
the
U.
S.
EPA;
a
list
of
these
will
be
available
from
the
U.
S.
EPA.
b.
The
second
includes
those
additions
that
are
submitted
to
the
U.
S.
EPA
for
approval.
If
approved,
these
will
be
added
to
EPA's
list
of
approved
additions.
Selective
additions
are
not
allowed.
All
additions
on
EPA's
newest
list
mast
be
made.

.
&.
me
Deletion
Process
The
basic
principles
are:
1.
Additions
and
corrections
muat
be
made
as
per
steps
A
and
B
above,
before
the
deletion
process
is
performed.
2.
Selective
deletions
are
not
allowed.
If
any
species
is
to
be
deleted,
the
deletion
process
described
below
mamt
be
applied
to
all
species
in
the
national
dataset,
after
any
necessary
corrections
and
additions
have
been
made
to
the
national
dataset.
The
deletion
process
specifies
which
species
muat
be
deleted
and
which
species
aunt
not
be
deleted.
Use
of
the
deletion
process
is
optional,
but
no
deletions
are
optional
when
the
deletion
process
is
used.
3.
Comprehensive
information
must
be
available
concerning
what
species
occur
at
the
site;
a
species
cannot
be
deleted
based
92
on
incomplete
information
concerning
the
species
that
do
and
do
not
satisfy
the
definition
of
'
occur
at
the
site'.
4.
Data
might
have
to
be
generated
before
the
deletion
process
is
begun:
a.
Acceptable
pertinent
toxicological
data
mwt
be
available
for
at
least
one
species
in
each
class
of
aquatic
plants,
invertebrates,
amphibians,
and
fish
that
contains
a
species
that
is
a
critical
species
at
the
site.
b.
For
each
aquatic
plant,
invertebrate,
amphibian,
and
fish
species
that
occurs
at
the
site
and
is
listed
as
threatened
or
endangered
under
section
4
of
the
Endangered
Species
Act,
data
munt
be
available
or
be
generated
for
an
acceptable
Surrogate
Species.
Data
for
each
surrogate
species
must
be
used
as
if
they
are
data
for
species
that
occur
at
the
site.
If
additional
data
are
generated
using
acceptable
procedures
(
U.
S.
EPA
1985)
and
they
are
approved
by
the
U.
S.
EPA,
the
Recalculation
Procedure
must
be
started
again
at
step
B
with
the
addition
of
the
new
data.
5.
Data
might
have
to
be
generated
gfter
the
deletion
process
is
completed.
Even
if
one
or
more
species
are
deleted,
there
still
are
MDRs
(
see
step
D
below)
that
muet
be
satisfied.
If
the
data
remaining
after
deletion
do
not
satisfy
the
applicable
MDRs,
additional
toxicity
tests
must
be
conducted
using
acceptable
procedures
(
U.
S.
EPA
1985)
so
that
all
MDRs
are
satisfied.
If
the
new
data
are
approved
by
the
U.
S.
EPA,
the
Recalculation
Procedure
mumt
be
started
again
at
step
B
with
the
addition
of
new
data.
6.
Chronic
tests
do
not
have
to
be
conducted
because
the
national
Final
Acute­
Chronic
Ratio
(
FACR)
may
be
used
in
the
derivation
of
the
site­
specific
Final
Chronic
Value
(
FCV).
If
acute­
chronic
ratios
(
ACRs)
are
available
or
are
generated
so
that
the
chronic
MDRs
are
satisfied
using
only
species
that
occur
at
the
site,
a
site­
specific
FACR
may
be
derived
and
used
in
place
of
the
national
FACR.
Because
a
FACR
was
not
used
in
the
derivation
of
the
freshwater
CCC
for
cadmium,
this
CCC
can
only
be
modified
the
same
way
as
a
FAV;
what
is
acceptable
will
depend
on
which
species
are
deleted.

If
any
species
are
to
be
deleted,
the
following
deletion
process
at
be
applied:
a.
Obtain
a
copy
of
the
national
dataset,
i.
e.,
tables
1,
2,
and
3
in
the
national
criteria
document
(
see
Appendix
E).
b.
Make
corrections
in
and/
or
additions
to
the
national
dataset
as
described
in
steps
A
and
B
above.
c.
Group
all
the
species
in
the
dataset
taxonomically
by
phylum,
class,
order,
family,
genus,
and
species.
d.
Circle
each
species
that
satisfies
the
definition
of
.
occur
at
the
site'
as
presented
on
the
first
page
of
this
appendix,
and
including
any
data
for
species
that
are
surrogates
of
threatened
or
endangered
species
that
occur
at
the
site.

93
e.
Use
the
following
step­
wise
process
to
determine
which
of
the
uncircled
species
mumt
be
deleted
and
which
muat
not
be
deleted:

1.
Does
the
genus
occur
at
the
site?
If
'
No',
gotostep2.
If
.
Yes.,
are
there
one
or
more
species
in
the
genus
that
occur
at
the
site
but
are
not
in
the
dataset?
If"
No',
go
to
step
2.
If
'
Yes',
retain
the
uncircled
species.*

2.
Does
the
family
occur
at
the
site?
If
'
No',
go
to
step
3.
If
'
Yes.,
are
there
one
or
more
genera
in
the
family
that
occur
at
the
site
but
are
not
in
the
dataset?
If
'
No',
go
to
step
3.
If
'
Yes',
retain
the
uncircled
species.*

3.
Does
the
order
occur
at
the
site?
If
'
No',
go
to
step
4.
If
'
Yes',
does
the
dataset
contain
a
circled
species
that
is
in
the
same
order?
If
'
No',
retain
the
uncircled
species.*
If
'
Yes',
delete
the
uncircled
species.*

4.
Does
the
class
occur
at
the
site?
If
'
No',
go
to
step
5.
If
.
Yesm,
does
the
dataset
contain
a
circled
species
that
is
in
the
same
class?
If
'
No',
retain
the
uncircled
species.*
If
'
Yes',
delete
the
uncircled
species.+

5.
Does
the
phylum
occur
at
the
site?
If
'
No',
delete
the
uncircled
species.*
If
WYesm,
does
the
dataset
contain
a
circled
species
that
is
in
the
same
phylum?
If
.
No',
retain
the
uncircled
species.+
If
.
Yes.,
delete
the
uncircled
species.+

l
=
Continue
the
deletion
process
by
starting
at
step
1
for
another
uncircled
species
unless
all
uncircled
species
in
the
dataset
have
been
considered.

The
species
that
are
circled
and
those
that
are
retained
constitute
the
site­
specific
dataset.
(
An
example
of
the
deletion
process
is
given
in
Figure
Bl.)

This
deletion
process
is
designed
to
ensure
that:
a.
Each
species
that
occurs
both
in
the
national
dataset
and
at
the
site
also
occurs
in
the
site­
specific
dataset.

94
b.
Each
species
that
occurs
at
the
site
but
does
not
occur
in
the
national
dataset
is
represented
in
the
site­
specific
dataset
by
u
species
in
the
national
dataset
that
are
in
the
same
genus.
C.
Each
genus
that
occurs
at
the
site
but
does
not
occur
in
the
national
dataset
is
represented
in
the
site­
specific
dataset
by
all
genera
in
the
national
dataset
that
are
in
the
same
fay.
d.
Each
order,
class,
and
phylum
that
occurs
both
in
the
national
dataset
and
at
the
site
is
represented
in
the
site­
specific
dataset
by
the
one
or
more
species
in
the
national
dataset
that
are
most
closely
related
to
a
species
that
occurs
at
the
site.

D.
Checkina
the
Minimum
Data
Reuuirements
The
initial
MDRs
for
the
Recalculation
Procedure
are
the
same
as
those
for
the
derivation
of
a
national
criterion.
If
a
specific
requirement
cannot
be
satisfied
after
deletion
because
that
kind
of
species
does
not
occur
at
the
site,
a
taxonomically
similar
species
muat
be
substituted
in
order
to
meet
the
eight
MDRs:

If
no
species
of
the
kind
required
occurs
at
the
site,
but
a
species
in
the
same
order
does,
the
MDR
can
only
be
satisfied
by
data
for
a
species
that
occurs
at
the
site
and
is
in
that
order;
if
no
species
in
the
order
occurs
at
the
site,
but
a
species
in
the
class
does,
the
MDR
can
only
be
satisfied
by
data
for
a
species
that
occurs
at
the
site
and
is
in
that
class.
If
no
species
in
the
same
class
occurs
at
the
site,
but
a
species
in
the
phylum
does,
the
MDR
can
only
be
satisfied
by
data
for
a
species
that
occurs
at
the
site
and
is
in
that
phylum.
If
no
species
in
the
same
phylum
occurs
at
the
site,
any
species
that
occurs
at
the
site
and
is
not
used
to
satisfy
a
different
MDR
can
be
used
to
satisfy
the
MDR.
If
additional
data
are
generated
using
acceptable
procedures
(
U.
S.
EPA
1985)
and
they
are
approved
by
the
U.
S.
EPA,
Recalculation
Procedure
must
be
started
again
at
step
the
addition
of
the
new
data.

If
fewer
than
eight
families
of
aquatic
invertebrates,
amphibians,
and
fishes
occur
at
the
site,
a
Species
Mean
Value
murt
be
available
for
at
least
one
species
in
each
the
B
with
Acute
of
the
families
and
the
special
version
of
the
Recalculation
Procedure
described
on
the
second
page
of
this
appendix
must
be
used.

E.
Determinina
the
CMC
and/
or
CCC
1.
Determining
the
FAV:
a.
If
the
eight
family
MDRs
are
satisfied,
the
site­
specific
FAV
muat
be
calculated
from
Genus
Mean
Acute
Values
using
95
the
procedure
described
in
the
national
aquatic
life
guidelines
(
U.
S.
EPA
1985).
b.
If
fewer
than
eight
families
of
aquatic
invertebrates,
amphibians,
and
fishes
occur
at
the
site,
the
lowest
Species
Mean
Acute
Value
that
is
available
for
a
species
that
occurs
at
the
site
mumt
be
used
as
the
FAV,
as
per
the
special
version
of
the
Recalculation
Procedure
described
on
the
second
page
of
this
appendix.
2.
The
site­
specific
CMC
nut
be
calculated
by
dividing
the
site­
specific
FAV
by
2.
The
site­
specific
FCV
rout
be
calculated
by
dividing
the
site­
specific
FAV
by
the
national
FACR
(
or
by
a
site­
specific
FACR
if
one
is
derived).
(
Because
a
FACR
was
not
used
to
derive
the
national
CCC
for
cadmium
in
fresh
water,
the
site­
specific
CCC
equals
the
site­
specific
FCV.)
3.
The
calculated
FAV,
CMC,
and/
or
CCC
mrmt
be
lowered,
if
necessary,
to
(
1)
protect
an
aquatic
plant,
invertebrate,
amphibian,
or
fish
species
that
is
a
critical
species
at
the
site,
and
(
2)
ensure
that
the
criterion
is
not
likely
to
jeopardize
the
continued
existence
of
any
endangered
or
threatened
species
listed
under
section
4
of
the
Endangered
Species
Act
or
result
in
the
destruction
or
adverse
modification
of
such
species'
critical
habitat.

.
.
Fe
Wrltlna
the
ReDort
The
report
of
the
results
of
use
of
the
Recalculation
Procedure
mm+
include:
1.
A
list
of
all
species
of
aquatic
invertebrates,
amphibians,
and
fishes
that
are
known
to
'
occur
at
the
site',
along
with
the
source
of
the
information.
2.
A
list
of
all
aquatic
plant,
invertebrate,
amphibian,
and
fish
species
that
are
critical
species
at
the
site,
including
all
species
that
occur
at
the
site
and
are
listed
as
threatened
or
endangered
under
section
4
of
the
Endangered
Species
Act.
3.
A
site­
specific
version
of
Table
1
from
a
criteria
document
produced
by
the
U.
S.
EPA
after
1984.
4.
A
site­
specific
version
of
Table
3
from
a
criteria
document
produced
by
the
U.
S.
EPA
after
1984.
5.
A
list
of
all
species
that
were
deleted.
6.
me
new
calculated
FAV,
CMC,,
and/
or
CCC.
7.
The
lowered
FAV,
CMC,
and/
or
CCC,
if
one
or
more
were
lowered
to
protect
a
specific
species.

U.
S.
EPA.
1985.
Guidelines
for
Deriving
Numerical
National
Water
Quality
Criteria
for
the
Protection
of
Aquatic
Organisms
and
Their
Uses.
PB85­
227049.
National
Technical
Information
Service,
Springfield,
VA.

96
Figure
81:
An
Xxapqple
of
thm
Dolotion
Procwm
IWing
Three
Phyla
SPECIES
THAT
ABE
IN
THE
THREE
phvlum
Class
Order
Annelida
Hirudin.
Bhynchob.
Bryozoa
(
No
species
in
this
Chordata
Osteich.
Cyprinif.
Chordata
Osteich.
Cyprinif.
Chordata
Osteich.
Cyprinif.
Chordata
Osteich.
Cyprinif.
Chordata
Osteich.
Salmonif.
Chordata
Osteich.
Percifor.
Chordata
Osteich.
Percifor.
Chordata
Amphibia
Caudata
PHYLAANDOCCURATTHE
SITE
Familv
SDecies
Glossiph.
Gloss+.
complanata
phylum
occur
at
the
site.)
Cyprinid.
Carassius
auratus
Cyprinid.
Notropis
anogenus
Cyprinid.
Phoxinus
eos
Catostom.
Carpiodes
carpio
Osmerida.
Osmerus
mordax
Centrarc.
Lepomis
cyanellus
Centrarc.
Legomis
humilis
Ambystom.
Ambystoma
gracile
SPECIES
THAT
ABE
IN
THE
THREE
phvlum
Class
Order
PHYLA
AND
IN
THE
NATIONAL
DATASET
Familv
Annelida
Oligoch.
HaplOtaX.
Bryozoa
Phylact.
­­­
Chordata
Cephala.
Petromyz.
Chordata
Osteich.
Cyprinif.
Chordata
Osteich.
Cyprinif.
Chordata
Osteich.
Cyprinif.
Chordata
Osteich.
Cyprinif.
Chordata
Osteich.
Cyprinif.
Chordata
Osteich.
Cyprinif.
Chordata
Osteich.
Cyprinif.
Chordata
Osteich.
Salmonif.
Chordata
Osteich.
Percifor.
Chordata
Osteich.
Percifor.
Chordata
Osteich.
Percifor.
Chordata
Amphibia
Anura
Tubifici.
LoDhODod.
Petromyz.
Cyprinid.
Cyprinid.
Cyprinid.
Cyprinid.
Cyprinid.
Cyprinid.
Catostom.
Salmonid.
Centrarc.
Centrarc.
Percidae
Pigidae
mecies
Coda
Tubifextubifex
P
Loghopod.
carteri
D
Petromyzon
marinus
D
Carassius
auratus
S
Notropis
hudsonius
G
Notropis
stramineus
G
Phoxinus
eos
S
Phoxinus
oreas
D
Tinca
tinca
D
Ictiobus
bubalus
F
Oncorhynchus
mykiss
0
Lepomis
cyanellus
S
Legomis
macrochirus
G
Perca
flavescens
D
Xenopus
laevis
C
Explanations
of
Codes:
s=
retained
because
this
Species
occurs
at
the
site.
G
=
retained
because
there
is
a
species
in
this
Genus
that
occurs
at
the
site
but
not
in
the
national
dataset.
F
=
retained
because
there
is
a
genus
in
this
Family
that
occurs
at
the
site
but
not
in
the
national
dataset.
o=
retained
because
this
Order
occurs
at
the
site
and
is
not
represented
by
a
lower
taxon.
c=
retained
because
this
Class
occurs
at
the
site
and
is
not
represented
by
a
lower
taxon.
P
=
retained
because
this
Phylum
occurs
at
the
site
and
is
not
represented
by
a
lower
taxon.
D
=
deleted
because
this
species
does
not
satisfy
any
of
the
requirements
for
retaining
species.

97
AppMdix
C:
alaidaaco
ConcorPing
the
Us.
of
'
Clam
Toohniquo8=
Ipd
QA/
QC
when
Meamuing
Tram
Dfotalm
Note:
This
version
of
this
appendix
contains
more
information
than
the
version
that
was
Appendix
B
of
Prothro
(
1993).

Recent
information
(
Shiller
and
Boyle
1987;
Windom
et
al.
1991)
has
raised
questions
concerning
the
quality
of
reported
concentrations
of
trace
metals
in
both
fresh
and
salt
(
estuarine
and
marine)
surface
waters.
A
lack
of
awareness
of
true
ambient
concentrations
of
metals
in
fresh
and
salt
surface
waters
can
be
both
a
cause
and
a
result
of
the
problem.
The
ranges
of
dissolved
metals
that
are
typical
in
surface
waters
of
the
United
States
away
from
the
immediate
influence
of
discharges
(
Bruland
1983;
Shiller
and
Boyle
1985,1987;
Trefry
et
al.
1986;
Windom
et
al.
1991)
are:

Metal
Salt
water
Fresh
water
(
uu/
L)
(
w/
L)

Cadmium
0.01
to
0.2
0.002
to
0.08
Copper
0.1
to
3.
0.4
to
4.
Lead
0.01
to
1.
0.01
to
0.19
Nickel
0.3
to
5.
1.
to
2.
Silver
0.005
to
0.2
­­­­­­­­­­­­­
Zinc
0.1
to
15.
0.03
to
5.

The
U.
S.
EPA
(
1983,1991)
has
published
analytical
methods
for
monitoring
metals
in
waters
and
wastewaters,
but
these
methods
are
inadequate
for
determination
of
ambient
concentrations
of
some
metals
in
some
surface
waters.
Accurate
and
precise
measurement
of
these
low
concentrations
requires
appropriate
attention
to
seven
areas:
1.
Use
of
.
clean
techniques'
during
collecting,
handling,
storing,
preparing,
and
analyzing
samples
to
avoid
contamination.
2.
Use
of
analytical
methods
that
have
sufficiently
low
detection
limits.
3.
Avoidance
of
interference
in
the
quantification
(
instrumental
analysis)
step.
4.
Use
of
blanks
to
assess
contamination.
5.
Use
of
matrix
spikes
(
sample
spikes)
and
certified
reference
materials
(
CRMs)
to
assess
interference
and
contamination.
6.
Use
of
replicates
to
assess
precision.
7.
Use
of
certified
standards.
In
a
strict
sense,
the
term
#
clean
techniques'
refers
to
techniques
that
reduce
contamination
and
enable
the
accurate
and
precise
measurement
of
trace
metals
in
fresh
and
salt
surface
waters.
In
a
broader
sense,
the
term
also
refers
to
related
issues
concerning
detection
limits,
quality
control,
and
quality
98
assurance.
Documenting
data
quality
demonstrates
the
amount
of
confidence
that
can
be
placed
in
the
data,
whereas
increasing
the
sensitivity
of
methods
reduces
the
problem
of
deciding
how
to
interpret
results
that
are
reported
to
be
below
detection.
limits.

This
aDDendiX
is
written
for
those
analvtical
laboratories
that
.
want
aldance
concernina
wavs
to
lower
detection
limits.
increase
accu
acv.
a
d/
o
.
c
e
s
Drecision.
The
ways
to
achieve
these
goal:
are
ti
in&:
isz
:
hz
sensitivity
of
the
analytical
methods,
decrease
contamination,
and
decrease
interference.
Ideally,
validation
of
a
procedure
for
measuring
concentrations
of
metals
in
surface
water
requires
demonstration
that
agreement
can
be
obtained
using
completely
different
procedures
beginning
with
the
sampling
step
and
continuing
through
the
quantification
step
(
Bruland
et
al.
1979),
but
few
laboratories
have
the
resources
to
compare
two
different
procedures.
Laboratories
can,
however,
(
a)
use
techniques
that
others
have
found
useful
for
improving
detection
limits,
accuracy,
and
precision,
and
(
b)
document
data
quality
through
use
of
blanks,
spikes,
CRMs,
replicates,
and
standards.
.
.
NJoth=
contained
or
not
contained
in
this
aDDendiX
adds
to
o
subtracts
from
anv
reaulatorv
reouirement
set
forth
in
other
EPA
docume
ts
conce
'
a
analyses
of
metals.
A
WER
can
be
acceptably
determ?
ned
withE?
the
use
of
clean
techniques
as
long
as
the
detection
limits,
accuracy,
and
precision
are
acceptable.
No
QA/
QC
requirements
beyond
those
that
apply
to
measuring
metals
in
effluents
are
necessary
for
the
determination
of
WEI&.
The
word
'
must'
is
not
used
in
this
appendix.
Some
items,
however,
are
considered
so
important
by
analytical
chemists
who
have
worked
to
increase
accuracy
and
precision
and
lower
detection
limits
in
trace­
metal
analysis
that
'
mhouldn
is
in
bold
print
to
draw
attention
to
the
item.
Most
such
items
are
emphasized
because
they
have
been
found
to
have
received
inadequate
attention
in
some
laboratories
performing
trace­
metal
analyses.

In
general,
in
order
to
achieve
accurate
and
precise
measurement
of
a
particular
concentration,
both
the
detection
limit
and
the
blanks
should
be
less
than
one­
tenth
of
that
concentration.
Therefore,
the
term
.
metal­
free'
can
be
interpreted
to
mean
that
the
total
amount
of
contamination
that
occurs
during
sample
collection
and
processing
(
e.
g.,
from
gloves,
sample
containers,
labware,
sampling
apparatus,
cleaning
solutions,
air,
reagents,
etc.)
is
sufficiently
low
that
blanks
are
less
than
one­
tenth
of
the
lowest
concentration
that
needs
to
be
measured.

Atmospheric
particulates
can
be
a
major
source
of
contamination
(
Moody
1982;
Adeloju
and
Bond
1985).
The
term
'
class­
100"
refers
to
a
specification
concerning
the
amount
of
particulates
in
air
(
Moody
1982)
;
although
the
specification
says
nothing
about
the
composition
of
the
particulates,
generic
control
of
particulates
can
greatly
reduce
trace­
metal
blanks.
Except
during
collection
99
of
samples,
initial
cleaning
of
equipment,
and
handling
of
samples
containing
high
concentrations
of
metals,
all
handling
of
samples,
sample
containers,
labware,
and
sampling
apparatus
should
be
performed
in
a
class­
100
bench,
room,
or
glove
box.

Neither
the
'
ultraclean
techniques'
that
might
be
necessary
when
trace
analyses
of
mercury
are
performed
nor
safety
in
analytical
laboratories
is
addressed
herein.
Other
documents
should
be
consulted
if
one
or
both
of
these
topics
are
of
concern.

v*
l
.
.
*

Measurement
of
trace
metals
in
surface
waters
should
take
into
account
the
potential
for
contamination
during
each
step
in
the
process.
Regardless
of
the
specific
procedures
used
for
collection,
handling,
storage,
preparation
(
digestion,
filtration,
and/
or
extraction),
and
quantification
(
instrumental
analysis),
the
general
principles
of
contamination
control
should
be
a.

b.

C.

d.

e.

f.
applied.
Some
specific
reconnnendations
are:
Powder­
free
(
non­
talc,
class­
1001
latex,
polyethylene,
or
polyvinyl
chloride
(
PVC,
vinyl)
gloves
should
be
worn
during
all
steps
from
sample
collection
to
analysis.
(
Talc
seems
to
be
a
particular
problem
with
zinc;
gloves
made
with
talc
cannot
be
decontaminated
sufficiently.)
Gloves
should
only
contact
surfaces
that
are
metal­
free;
gloves
should
be
changed
if
even
suspected
of
contamination.
The
acid
used
to
acidify
samples
for
preservation
and
digestion
and
to
acidify
water
for
final
cleaning
of
labware,
sampling
apparatus,
and
sample
containers
should
be
metal­
free.
The
quality
of
the
acid
used
should
be
better
than
reagent­
grade.
Each
lot
of
acid
should
be
analyzed
for
the
metal(
s)
of
interest
before
use.
The
water
used
to
prepare
acidic
cleaning
solutions
and
to
rinse
la&
are,
sample
containers,
and
sampling
apparatus
may
be
prepared
by
distillation,
deionization,
or
reverse
osmosis,
and
&
aould
be
demonstrated
to
be
metal­
free.
The
work
area,
including
bench
tops
and
hoods,
should
be
cleaned
(
e.
g.,
washed
and
wiped
dry
with
lint­
free,
class­
100
wipes)
frequently
to
remove
contamination.
All
handling
of
samples
in
the
laboratory,
including
filtering
and
analysis,
should
be
performed
in
a
class­
100
clean
bench
or
a
glove
box
fed
by
particle­
free
air
or
nitrogen;
ideally
the
clean
bench
or
glove
box
should
be
located
within
a
class­
100
clean
room.
Labware,
reagents,
sampling
apparatus,
and
sample
containers
should
never
be
left
open
to
the
atmosphere;
they
should
be
stored
in
a
class­
100
bench,
covered
with
plastic
wrap,
stored
in
a
plastic
box,
or
turned
upside
down
on
a
clean
surface.
Minimizing
the
time
between
cleaning
and
using
will
help
minimize
contamination.

100
g.
Separate
sets
of
sample
containers,
labware,
and
sampling
apparatus
should
be
dedicated
for
different
kinds
of
samples,
.
surface
water
samples
effluent
samples,
etc.
h.
Gog&
oid
contamination
of
clean
rooms,
samples
that
contain
very
high
concentrations
of
metals
and
do
not
require
use
of
@
clean
techniques'
should
not
be
brought
into
clean
rooms.
i.
Acid­
cleaned
plastic,
such
as
high­
density
polyethylene
(
HDPE)
,
low­
density
polyethylene
(
LDPE),
or
a
fluoroplastic,
should
be
the
only
material
that
ever
contacts
a
sample,
except
possibly
during
digestion
for
the
total
recoverable
measurement.
1.
Total
recoverable
samples
can
be
digested
in
some
plastic
containers.
2.
HDPE
and
LDPE
might
not
be
acceptable
for
mercury.
3.
Even
if
acidified,
samples
and
standards
containing
silver
should
be
in
amber
containers.
j.
All
labware,
sample
containers,
and
sampling
apparatus
8hould
be
acid­
cleaned
before
use
or
reuse.
1.
Sample
containers,
sampling
apparatus,
tubing,
membrane
filters,
filter
assemblies,
and
other
labware
rhould
be
soaked
in
acid
until
metal­
free.
The
amount
of
cleaning
necessary
might
depend
on
the
amount
of
contamination
and
the
length
of
time
the
item
will
be
in
contact
with
samples.
For
example,
if
an
acidified
sample
will
be
stored
in
a
sample
container
for
three
weeks,
ideally
the
container
should
have
been
soaked
in
an
acidified
metal­
free
solution
for
at
least
three
weeks.
2.
It
might
be
desirable
to
perform
initial
cleaning,
for
which
reagent­
grade
acid
may
be
used,
before
the
items
are
taken
into
a
clean
room.
For
most
metals,
items
should
be
either
(
a)
soaked
in
10
percent
concentrated
nitric
acid
at
50
°
C
for
at
least
one
hour,
or
(
b)
soaked
in
50
percent
concentrated
nitric
acid
at
room
temperature
for
at
least
two
days;
for
arsenic
and
mercury,
soaking
for
up
to
two
weeks
at
50
°
C
in
10
percent
concentrated
nitric
acid
might
be
required.
For
plastics
that
might
be
damaged
by
strong
nitric
acid,
such
as
polycarbonate
and
possibly
HDPE
and
LDPE,
soaking
in
10
percent
concentrated
hydrochloric
acid,
either
in
place
of
or
before
soaking
in
a
nitric
acid
solution,
might
be
desirable.
3.
Chromic
acid
should
not
be
used
to
clean
items
that
will
be
used
in
analysis
of
metals.
4.
Final
soaking
and
cleaning
of
sample
containers,
labware,
and
sampling
apparatus
should
be
performed
in
a
class­
100
clean
room
using
metal­
free
acid
and
water.
The
solution
in
an
acid
bath
8hould
be
analyzed
periodically
to
demonstrate
that
it
is
metal­
free.
k.
Labware,
sampling
apparatus,
and
sample
containers
should
be
stored
appropriately
after
cleaning:
1.
After
the
labware
and
sampling'apparatus
are
cleaned,
they
may
be
stored
in
a
clean
room
in
a
weak
acid
bath
prepared
using
metal­
free
acid
and
water.
Before
use,
the
items
101
should
be
rinsed
at
least
three
times
with
metal­
free
water.
After
the
final
rinse,
the
items
should
be
moved
ixmnediately,
with
the
open
end
pointed
down,
to
a
class­
100
clean
bench.
Items
may
be
dried
on
a
class­
100
clean
bench;
items
nhould
not
be
dried
in
an
oven
or
with
laboratory
towels.
The
sampling
apparatus
should
be
assembled
in
a
class­
100
clean
room
or
bench
and
double­
bagged
in
metal­
free
polyethylene
zip­
type
bags
for
transport
to
the
field;
new
bags
are
usually
metal­
free.
2.
After
sample
containers
are
cleaned,
they
should
be
filled
with
metal­
free
water
that
has
been
acidified
to
a
pH
of
2
with
metal­
free
nitric
acid
(
about
0.5
mL
per
liter)
for
storage
until
use.
1.
Labware,
sampling
apparatus,
and
sample
containers
Should
be
rinsed
and
not
rinsed
with
sample
as
necessary
to
prevent
high
and
low
bias
of
analytical
results
because
acid­
cleaned
plastic
will
sorb
some
metals
from
unacidified
solutions.
1.
Because
samples
for
the
dissolved
measurement
are
not
acidified
until
after
filtration,
all
sampling
apparatus,
sample
containers,
labware,
filter
holders,
membrane
filters,
etc.,
that
contact
the
sample
before
or
during
filtration
should
be
rinsed
with
a
portion
of
the
solution
and
then
that
portion
discarded.
2.
For
the
total
recoverable
measurement,
labware,
etc.,
that
contact
the
sample
onlv
before
it
is
acidified
should
be
rinsed
with
sample,
whereas
items
that
contact
the
sample
after
it
is
acidified
should
not
be
rinsed.
For
example,
the
sampling
apparatus
should
be
rinsed
because
the
sample
will
not
be
acidified
until
it
is
in
a
sample
container,
but
the
sample
container
should
not
be
rinsed
if
the
sample
will
be
acidified
in
the
sample
container.
3.
If
the
total
recoverable
and
dissolved
measurements
are
to
be
performed
on
the
same
sample
(
rather
than
on
two
samples
obtained
at
the
same
time
and
place),
all
the
apparatus
and
labware,
including
the
sample
container,
should
be
rinsed
before
the
sample
is
placed
in
the
sample
container;
then
an
unacidified
aliguot
should
be
removed
for
the
total
recoverable
measurement
(
and
acidified,
digested,
etc.)
and
an
unacidified
aliguot
should
be
removed
for
the
dissolved
measurement
(
and
filtered,
acidified,
etc.)
(
If
a
container
is
rinsed
and
filled
with
sample
and
an
unacidified
aliguot
is
removed
for
the
dissolved
measurement
and
then
the
solution
in
the
container
is
acidified
before
removal
of
an
aliguot
for
the
total
recoverable
measurement,
the
resulting
measured
total
recoverable
concentration
might
be
biased
high
because
the
acidification
might
desorb
metal
that
had
been
sorbed
onto
the
walls
of
the
sample
container;
the
amount
of
bias
will
depend
on
the
relative
volumes
jnmlved
and
on
the
amount
of
sorption
and
desorption.)
m.
Field
samples
Should
be
collected
in
a
manner
that
eliminates
the
potential
for
contamination
from
sampling
platforms,

102
n.

0.

P*

Q*

r.

S.

t.
probes,
etc.
Exhaust
from
boats
and
the
direction
of
wind
and
water
currents
should
be
taken
into
account.
The
people
who
collect
the
samples
&
ould
be
specifically
trained
on
how
to
collect
field
samples.
After
collection,
all
handling
of
samples
in
the
field
that
will
expose
the
sample
to
air
ahould
be
performed
in
a
portable
class­
100
clean
bench
or
glove
box.
Samples
should
be
acidified
(
after
filtration
if
dissolved
metal
is
to
be
measured)
to
a
pH
of
less
than
2,
except
that
the
pH
should
be
less
than
1
for
mercury.
Acidification
should
be
done
in
a
clean
room
or
bench,
and
so
it
might
be
desirable
to
wait
and
acidify
samples
in
a
laboratory
rather
than
in
the
field.
If
samples
are
acidified
in
the
field,
metal­
free
acid
can
be
transported
in
plastic
bottles
and
poured
into
a
plastic
container
from
which
acid
can
be
removed
and
added
to
samples
using
plastic
pipettes.
Alternatively,
plastic
automatic
dispensers
can
be
used.
Such
things
as
probes
and
thermometers
l
hould
xaot
be
put
in
samples
that
are
to
be
analyzed
for
metals.
In
particular,
pH
electrodes
and
mercury­
in­
glass
thermometers
should
not
be'
used
if
mercury
is
to
be
measured.
If
pH
is
measured,
it
should
be
done
on
a
separate
aliguot.
Sample
handling
should
be
minimized.
For
example,
instead
of
pouring
a
sample
into
a
graduated
cylinder
to
measure
the
volume,
the
sample
can
be
weighed
after
being
poured
into
a
tared
container,
which
is
less
likely
to
be
subject
to
error
than
weighing
the
container
from
which
the
sample
is
poured.
(
For
saltwater
samples,
the
salinity
or
density
should
be
taken
into
account
if
weight
is
converted
to
volume.)
Each
reagent
used
mhould
be
verified
to
be
metal­
free.
If
metal­
free
reagents
are
not
commercially
available,
removal
of
metals
will
probably
be
necessary.
For
the
total
recoverable
measurement,
samples
should
be
digested
in
a
class­
100
bench,
not
in
a
metallic
hood.
If
feasible,
digestion
should
be
done
in
the
sample
container
by
acidification
and
heating.
The
longer
the
time
between
collection
and
analysis
of
samples,
the
greater
the
chance
of
contamination,
loss,
etc.
Samples
should
be
stored
in
the
dark,
preferably
between
0
and
4OC
with
no
air
space
in
the
sample
container.

Achievina
low
detection
limits
a.
Extraction
of
the
metal
from
the
sample
can
be
extremely
useful
if
it
simultaneously
concentrates
the
metal
and
eliminates
potential
matrix
interferences.
For
example,
ammonium
I­
pyrrolidinedithiocarbamate
and/
or
diethylammonium
diethyldithiocarbamate
can
extract
cadmium,
copper,
lead,
nickel,
and
zinc
(
Bruland
et
al.
1979;
Nriagu
et
al.
1993).
b.
The
detection
limit
should
be
less
than
ten
percent
of
the
lowest
concentration
that
is
to
be
measured.

103
a.
Potential
interferences
should
be
assessed
for
the
specific
instrumental
analysis
technique
used
and
for
each
metal
to
be
measured.
b.
If
direct
analysis
is
used,
the
salt
present
in
high­
salinity
saltwater
samples
is
likely
to
cause
interference
in
most
instrumental
techniques.
c.
As
stated
above,
extraction
of
the
metal
from
the
sample
is
particularly
useful
because
it
simultaneously
concentrates
the
metal
and
eliminates
potential
matrix
interferences.

a.
A
laboratory
(
procedural,
method)
blank
consists
of
filling
a
sample
container
with
analyzed
metal­
free
water
and
processing
(
filtering,
acidifying,
etc.)
the
water
through
the
laboratory
procedure
in
exactly
the
same
way
as
a
sample.
A
laboratory
blank
should
be
included
in
each
set
of
ten
or
fewer
samples
to
check
for
contamination
in
the
laboratory,
and
should
contain
less
than
ten
percent
of
the
lowest
concentration
that
is
to
be
measured.
Separate
laboratory
blanks
&
ould
be
processed
for
the
total
recoverable
and
dissolved
measurements,
if
both
measurements
are
performed.
b.
A
field
(
trip)
blank
consists
of
filling
a
sample
container
with
analyzed
metal­
free
water
in
the
laboratory,
taking
the
container
to
the
site,
processing
the
water
through
tubing,
filter,
etc.,
collecting
the
water
in
a
sample
container,
and
acidifying
the
water
the
same
as
a
field
sample.
A
field
blank
mhould
be
processed
for
each
sampling
trip.
Separate
field
blanks
should
be
processed
for
the
total
recoverable
measurement
and
for
the
dissolved
measurement,
if
filtrations
are
performed
at
the
site.
Field
blanks
l
hauld
be
processed
in
the
laboratory
the
same
as
laboratory
blanks.

Assessinu
accuracy
a.
A
calibration
curve
should
be
determined
for
each
analytical
run
and
the
calibration
should
be
checked
about
every
tenth
sample.
Calibration
solutions
should
be
traceable
back
to
a
certified
standard
from
the
U.
S.
EPA
or
the
National
Institute
of
Science
and
Technology
(
NIST).
b.
A
blind
standard
or
a
blind
calibration
solution
hould
be
included
in
each
group
of
about
twenty
samples.
c.
At
least
one
of
the
following
should
be
included
in
each
group
of
about
twenty
samples:
1.
A
matrix
spike
(
spiked
sample;
the
method
of
known
additions).

104
2.
A
CPM,
if
one
is
available
in
a
matrix
that
closely
approximates
that
of
the
samples.
Values
obtained
for
the
CPM
should
be
within
the
published
values.
The
concentrations
in/
blind
standards
and
solutions,
spikes,
and
CRMs
should
not
be
more
than
5
times
the
median
concentration
expected
to
be
present
in
the
samples.

a.
A
sampling
replicate
should
be
included
with
each
set
of
samples
collected
at
each
sampling
location.
b.
If
the
volume
of
the
sample
is
large
enough,
replicate
analysis
of
at
least
one
sample
l
hould
be
performed
along
with
each
group
of
about
ten
samples.

special
considerations
concernina
the
dissolved
measurement
Whereas
total
recoverable
measurements
are
especially
subject
to
contamination
during
digestion,
dissolved
measurements
are
subject
to
both
loss
and
contamination
during
filtration.
a.
Because
acid­
cleaned
plastic
sorbs
metal
from
unacidified
solutions
and
because
samples
for
the
dissolved
measurement
are
not
acidified
before
filtration,
all
sampling
apparatus,
sample
containers,
labware,
filter
holders,
and
membrane
filters
that
contact
the
sample
before'or
during
filtration
should
be
conditioned
by
rinsing
with
a
portion
of
the
solution
and
discarding
that
portion.
b.
Filtrations
should
be
performed
using
acid­
cleaned
plastic
filter
holders
and
acid­
cleaned
membrane
filters.
Samples
should
not
be
filtered
through
glass
fiber
filters,
even
if
the
filters
have
been
cleaned
with
acid.
If
positive­
pressure
filtration
is
used,
the
air
or
gas
should
be
passed
through
a
0.2­
v
in­
line
filter;
if
vacuum
filtration
is
used,
it
should
be
performed
on
a
class­
100
bench.
C.
Plastic
filter
holders
should
be
rinsed
and/
or
dipped
between
filtrations,
but
they
do
not
have
to
be
soaked
between
filtrations
if
all
the
samples
contain
about
the
same
concentrations
of
metal.
It
is
best
to
filter
samples
from
low
to
high
concentrations.
A
membrane
filter
&
ould
not
be
used
for
more
than
one
filtration.
After
each
filtration,
the
membrane
filter
should
be
removed
and
discarded,
and
the
filter
holder
should
be
either
rinsed
with
metal­
free
water
or
dilute
acid
and
dipped
in
a
metal­
free
acid
bath
or
rinsed
at
least
twice
with
metal­
free
dilute
acid;
finally,
the
filter
holder
ohould
be
rinsed
at
least
twice
with
metal­
free
water.
d.
For
each
sample
to
be
filtered,
the
filter
holder
and
membrane
filter
should
be
conditioned
with
the
sample,
i.
e.,
an
initial
portion
of
the
sample
ohould
be
filtered
and
discarded.

105
The
accuracy
and
precision
of
the
dissolved
measurement
should
be
assessed
periodically.
A
large
volume
of
a
buffered
solution
(
such
as
aerated
0.05
N
sodium
bicarbonate
for
analyses
in
fresh
water
and
a
combination
of
sodium
bicarbonate
and
sodium
chloride
for
analyses
in
salt
water)
should
be
spiked
so
that
the
concentration
of
the
metal
of
interest
is
in
the
range
of
the
low
concentrations
that
are
to
be
measured.
Sufficient
samples
should
be
taken
alternately
for
(
a)
acidification
in
the
same
way
as
after
filtration
in
the
dissolved
method
and
(
b)
filtration
and
acidification
using
the
procedures
specified
in
the
dissolved
method
until
ten
samples
have
been
processed
in
each
Way.
The
concentration
of
metal
in
each
of
the
twenty
samples
should
then
be
determined
using
the
same
analytical
procedure.
The
means
of
the
two
groups
of
ten
measurements
should
be
within
10
percent,
and
the
coefficient
of
variation
for
each
group
of
ten
should
be
less
than
20
percent.
Any
values
deleted
as
outliers
should
be
acknowledged.

To
indicate
the
quality
of
the
data,
reports
of
results
of
measurements
of
the
concentrations
of
metals
should
include
a
description
of
the
blanks,
.
spikes,
C??
Ms,
replicates,
and
standards
that
were
run,
the
number
run,
and
the
results
obtained.
All
values
deleted
as
outliers
&
aould
be
acknowledged.

The
items
presented
above
are
some
of
the
important
aspects
of
'
clean
techniques';
some
aspects
of
quality
assurance
and
quality
control
are
also
presented.
This
is
not
a
definitive
treatment
of
these
topics;
additional
information
that
might
be
useful
is
available
in
such
publications
as
Patterson
and
Settle
(
1976),
Zief
and
Mitchell
(
1976),
Bruland
et
al.
(
1979),
Moody
and
Beary
(
1982),
Moody
(
1982),
Bruland
(
1983),
Adeloju
and
Bond
(
1985),
Berman
and
Yeats
(
1985),
Byrd
and
Andreae
(
1986),
Taylor
(
1987),
Sakamoto­
Arnold
(
1987),
Tramontano
et
al.
(
1987),
Puls
and
Barcelona
(
1989),
Windom
et
al.
(
1991),
U.
S.
EPA
(
1992),
Horowitz
et
al.
(
1992),
and
Nriagu
et
al.
(
1993).

106
References
Adeloju,
S.
B.,
and
A.
M.
Bond.
1985.
Influence
of
Laboratory
Environment
on
the
Precision
and
Accuracy
of
Trace
Element
Analysis.
Anal.
Chem.
57:
1728­
1733.

Berman,
S.
S.,
and
P.
A.
Yeats.
1985.
Sampling
of
Seawater
for
Trace
Metals.
CRC
Reviews
in
Analytical
Chemistry
16:
1­
14.

Bruland,
K.
W.,
R.
P.
Franks,
G.
A.
Knauer,
and
J.
H.
Martin.
1979.
Sampling
and
Analytical
Methods
for
the
Determination
of
Copper,
Cadmium,
Zinc,
and
Nickel
at
the
Nanogram
per
Liter
Level
in
Sea
Water.
Anal.
Chim.
Acta
105:
233­
245.

Bruland,
K.
W.
1983.
Trace
Elements
in
Sea­
water.
In:
Chemical
Oceanography,
Vol.
8.
(
J.
P.
Riley
and
R.
Chester,
eds.)
Academic
Press,
.
New
York,
NY.
pp.
157­
220.

Byrd,
J.
T.,
and
M.
O.
Andreae.
1986.
Dissolved
and
Particulate
Tin
in
North
Atlantic
Seawater.
Marine
Chem.
19:
193­
200.

Horowitz,
A.
J.,
K.
A.
Elrick,
and
M.
R.
Colberg.
1992.
The
Effect
of
Membrane
Filtration
Artifacts
on
Dissolved
Trace
Element
Concentrations.
Water
Res.
26:
753­
763.

Moody,
J.
R.
1982.
NBS
Clean
Laboratories
for
Trace
Element
Analysis.
Anal.
Chem.
54:
1358A­
1376A.

Moody,
J.
R.,
and
E.
S.
Beary.
1982.
Purified
Reagents
for
Trace
Metal
Analysis.
Talanta
29:
1003­
1010.

Nriagu,
J.
O.,
G.
Lawson,
H.
K.
T.
Wong,
and
J.
M.
Azcue.
1993.
A
Protocol
for
Minimizing
Contamination
in
the
Analysis
of
Trace
Metals
in
Great
Lakes
Waters.
J.
Great
Lakes
Res.
19:
175­
182.

Patterson,
C.
C.,
and
D.
M.
Settle.
1976.
The
Reduction
in
Orders
of
Magnitude
Errors
in
Lead
Analysis
of
Biological
Materials
and
Natural
Waters
by
Evaluating
and
Controlling
the
Extent
and
Sources
of
Industrial
Lead
Contamination
Introduced
during
Sample
Collection
and
Processing.
In:
Accuracy
in
Trace
Analysis:
Sampling,
Sample
Handling,
Analysis.
(
P.
D.
LaFleur,
ed.)
National
Bureau
of
Standards
Spec.
Publ.
422,
U.
S.
Government
Printing
Office,
Washington,
DC.

Prothro,
M.
G.
1993.
Memorandum
titled
'
Office
of
Water
Policy
and
Technical
Guidance
on
Interpretation
and
Implementation
of
Aquatic
Life
Metals
Criteria'.
October
1.

Puls,
R.
W.,
and
M.
J.
Barcelona.
1989.
Ground
Water
Sampling
for
Metals
Analyses.
EPA/
540/
4­
89/
001.
National
Technical
Information
Service,
Springfield,
VA.

107
Sakamoto­
Arnold,
C.
M.,
A.
K.
Hanson,
Jr.,
D.
L.
Huizenga,
and
D.
R.
Kester.
1987.
Spatial
and
Temporal
Variability
of
Cadmium
in
Gulf
Stream
Warm­
core
Rings
and
Associated
Waters.
J.
Mar.
Res.
45:
201­
230.

Shiller,
A.
M.,
and
E.
Boyle.
1985.
Dissolved
Zinc
in
Rivers.
Nature
317:
49­
52.

Shiller,
A.
M.,
and
E.
A.
Boyle.
1987.
Variability
of
Dissolved
Trace
Metals
in
the
Mississippi
River.
Geochim.
Cosmochim.
Acta
51:
3273­
3277.

Taylor,
J.
K.
1987.
Quality
Assurance
of
Chemical
Measurements.
Lewis
Publishers,
Chelsea,
MI.

Tranmntano,
J.
M.,
J.
R.
Scudlark,
and
T.
M.
Church.
1987.
A
Method
for
the
Collection,
Handling,
in
Precipitation.
and
Analysis
of
Trace
Metals
Environ.
Sci.
Technol.
21:
749­
753.

Trefzy,
J.
H.,
T.
A.
Nelsen,
R.
P.
Trocine,
S.
Metz.,
and
T.
W.
Vetter.
1986.
Delta
System.
Trace
Metal
Fluxes
through
the
Mississippi
River
Rapp.
P.­
v.
Reun.
Cons.
int.
Explor.
Mer.
186:
277­
288.

U.
S.
EPA.
1983.
Methods
for
Chemical
Analysis
of
Water
and
Wastes.
EPA­
600/
4­
79­
020.
National
Technical
Information
Service,
Springfield,
VA.
Sections
4.1.1,
4.1.3,
and
4.1.4
U.
S.
EPA.
1991.
Methods
for
the
Determination
of
Metals
in
Environmental
Samples.
EPA­
600/
4­
91­
010.
National
Technical
Information
Service,
Springfield,
VA.

U.
S.
EPA.
1992.
Evaluation
of
Trace­
Metal
Levels
in
Ambient
Waters
and
Tributaries
to
New
York/
New
Jersey
Harbor
for
Waste
Load
Allocation.
Prepared
by
Battelle
Ocean
Sciences
under
Contract
No.
68­
C8­
0105.

Windom,
H.
L.,
J.
T.
Byrd,
R.
G.
Smith,
and
F.
Huan.
1991.
Inadequacy
of
NASQAN
Data
for
Assessing
Metals
Trends
in
the
Nation's
Rivers.
Environ.
Sci.
Technol.
25:
1137­
1142.
(
Also
see
the
c­
t
and
response:
Environ.
Sci.
Technol.
25:
1940­
1941.)

Fief,
M.,
and
J.
W.
Mitchell.
1976.
Contamination
Control
in
Trace
Element
Analysis.
Chemical
Analysis
Series,
Vol.
47.
Wiley,
New
York,
NY.

108
Appendix
D:
Relationships
between
WXRm
and
the
Chamimtry
and
Toxicology
of
M&
ala
The
aquatic
toxicology
of
metals
is
complex
in
part
because
the
chemistry
of
metals
in
water
is
complex.
Metals
usually
exist
in
surface
water
in
various
combinations
of
particulate
and
dissolved
forms,
some
of
which
are
toxic
and
some
of
which
are
nontoxic.
In
addition,
all
toxic
forms
of
a
metal
are
not
necessarily
equally
toxic,
and
various
water
quality
characteristics
can
affect
the
relative
concentrations
and/
or
toxicities
of
some
of
the
forms.

The
toxicity
of
a
metal
has
sometimes
been
reported
to
be
proportional
to
the
concentration
or
activity
of
a
specific
species
of
the
metal.
For
example,
Allen
and
Hansen
(
1993)
surmnarized
reports
by
several
investigators
that
the
toxicity
of
copper
is
related
to
the
free
cupric
ion,
but
other
data
do
not
support
a
correlation
(
Erickson
1993a).
For
example,
Borgmann
(
1983),
Chapman
and
McCrady
(
19771,
and
French,
and
Hunt
(
1986)
found
that
toxicity
expressed
on
the
basis
of
cupric
ion
activity
varied
greatly
with
pH,
and
Cowan
et
al.
(
1986)
concluded
that
at
least
one
of
the
copper
hydroxide
species
is
toxic.
Further,
chloride
and
sulfate
salts
of
calcium,
magnesium,
potassium,
and
sodium
affect
the
toxicity
of
the
cupric
ion
(
Nelson
et
al.
1986).
Similarly
for
aluminum,
Wilkinson
et
al.
(
1993)
concluded
that
'
mortality
was
best
predicted
not
by
the
free
Al'*
activity
but
rather
as
a
function
of
the
sum
I:([
A13']
+
[
AlF'*])"
and
that
.
no
longer
can
the
reduction
of
Al
toxicity
in
the
presence
of
organic
acids
be
interpreted
simply
as
a
consequence
of
the
decrease
in
the
free
Al'*
concentration'.

Until
a
model
has
been
demonstrated
to
explain
the
quantitative
relationship
between
chemical
and
toxicological
measurements,
aquatic
life
criteria
should
be
established
in
an
environmentally
conservative
manner
with
provision
for
site­
specific
adjustment.
Criteria
should
be
expressed
in
terms
of
feasible
analytical
measurements
that
provide
the
necessary
conservatism
without
substantially
increasing
the
cost
of
implementation
and
site­
specific
adjustment.
Thus
current
aquatic
life
criteria
for
metals
are
expressed
in
terms
of
the
total
recoverable
measurement
and/
or
the
dissolved
measurement,
rather
than
a
measurement
that
would
be
more
difficult
to
perform
and
would
still
require
empirical
adjustment.
The
WER
is
operationally
defined
in
terms
of
chemical
and
toxicological
measurements
to
allow
site­
specific
adjustments
that
account
for
differences
between
the
toxicity
of
a
metal
in
laboratory
dilution
water
and
in
site
water.

109
Fo­
of
Me­
u
Even
if
the
relationship
of
toxicity
to
the
forms
of
metals
is
not
understood
well
enough
to
allow
setting
site­
specific
water
quality
criteria
without
using
empirical
adjustments,
appropriate
use
and
interpretation
of
WERs
requires
an
understanding
of
how
changes
in
the
relative
concentrations
of
different
forms
of
a
metal
might
affect
toxicity.
Because
WERs
are
defined
on
the
basis
of
relationships
between
measurements
of
toxicity
and
measurements
of
total
recoverable
and/
or
dissolved
metal,
the
toxicologically
relevant
distinction
is
between
the
forms
of
the
metal
that
are
toxic
and
nontoxic
whereas
the
chemically
relevant
distinction
is
between
the
forms
that
are
dissolved
and
particulate.
'
Dissolved
metal'
is
defined
here
as
'
metal
that
passes
through
either
a
0.45­
p
or
a
0.40­
p
membrane
filter'
and
l
particulate
metalm
is
defined
as
.
total
recoverable
metal
minus
dissolved
metal'.
Metal
that
is
in
or
on
particles
that
pass
through
the
filter
is
operationally
defined
as
'
dissolved'.

In
addition,
some
species
of
metal
can
be
converted
from
one
form
to
another.
Some
conversions
are
the
result
of
reeguilibration
in
response
to
changes
in
water
quality
characteristics
whereas
others
are
due
to
such
fate
processes
as
oxidation
of
sulfides
and/
or
organic
matter.
Reequilibration
usually
occurs
faster
than
fate
processes
and
probably
results
in
any
rapid
changes
that
are
due
to
effluent
mixing
with
receiving
water
or
changes
in
pH
at
a
gill
surface.
To
account
for
rapid
changes
due
to
reeguilibration,
the
terms
'
labile'
and
'
refractory'
will
be
used
herein
to
denote
metal
species
that
do
and
do
not
readily
convert
to
other
species
when
in
a
noneguilibrium
condition,
with
'
readily'
referring
to
substantial
progression
toward
equilibrium
in
less
than
about
an
hour.
Although
the
toxicity
and
lability
of
a
form
of
a
metal
are
not
merely
yes/
no
properties,
but
rather
involve
gradations,
a
simple
classification
scheme
such
as
this
should
be
sufficient
to
establish
the
principles
regarding
how
WEF&
are
related
to
various
operationally
defined
forms
of
metal
and
how
this
affects
the
determination
and
use
of
WERs.

Figure
Dl
presents
the
classification
scheme
that
results
from
distinguishing
forms
of
metal
based
on
analytical
methodology,
toxicity
tests,
and
lability,
as
described
above.
Metal
that
is
not
measured
by
the
total
recoverable
measurement
is
assumed
to
be
sufficiently
nontoxic
and
refractory
that
it
will
not
be
further
considered
here.
Allowance
is
made
for
toxicity
due
to
particulate
metal
because
some
data
indicate
that
particulate
metal
might
contribute
to
toxicity
and
bioaccumulation,
although
other
data
imply
that
little
or
no
toxicity
can
be
ascribed
to
particulate
metal
(
Erickson
1993b).
Even
if
the
toxicity
of
particulate
metal
is
not
negligible
in
a
particular
situation,
a
dissolved
criterion
will
not
be
underprotective
if
the
dissolved
criterion
was
derived
using
a
dissolved
WER
(
see
below)
or
if
there
are
sufficient
compensating
factors.

110
­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­
Figure
01:
A
Schamo
for
ClauifyiPg
Formm
of
Metal
in
Water
Total
recoverable
metal
Dissolved
Nontoxic
Labile
Refractory
Toxic
Labile
Particulate
Nontoxic
Labile
Refractory
Toxic
Labile
Metal
not
measured
by
the
total
recoverable
measurement
­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­

Not
only
can
some
changes
in
water
quality
characteristics
shift
the
relative
concentrations
of
toxic
and
nontoxic
labile
species
of
a
metal,
some
changes
in
water
quality
can
also
increase
or
decrease
the
toxicities
of
'
the
toxic
species
of
a
metal
and/
or
the
sensitivities
of
aquatic
organisms.
Such
changes
might
be
caused
by
(
a)
a
change
in
ionic
strength
that
affects
the
activity
of
toxic
species
of
the
metal
in
water,
(
b)
a
physiological
Rffect
whereby
an
ion
affects
the
permeability
of
a
membrane
and
thereby
alters
both
uptake
and
apparent
toxicity,
and
(
c)
toxicological
additivity,
synergism,
or
antagonism
due
to
effects
within
the
organism.

Another.
possible
complication
is
that
a
form
of
metal
that
is
toxic
to
one
aquatic
organism
might
not
be
toxic
to
another.
Although
such
differences
between
organisms
have
not
been
demonstrated,
the
possibility
cannot
be
ruled
out.

The
Imoortance
of
Lailitv
The
only
common
metal
measurement
that
can
be
validly
extrapolated
from
the
effluent
and
the
upstream
water
to
the
downstream
water
merely
by
taking
dilution
into
account
is
the
total
recoverable
measurement.
A
major
reason
this
measurement
is
so
useful
is
because
it
is
the
only
measurement
that
obeys
the
law
of
mass
balance
(
i.
e.,
it
is
the
only
measurement
that
is
conservative).
Other
metal
measurements
usually
do
not
obey
the
law
of
mass
balance
because
they
measure
some,
but
not
all,
of
the
labile
species
of
metals.
A
measurement
of
refractory
metal
111
would
be
conservative
in
terms
of
changes
in
water
quality
characteristics,
but
not
necessarily
in
regards
to
fate
processes;
such
a
measurement
has
not
been
developed,
however.

Permit
limits
apply
to
effluents,
whereas
water
quality
criteria
apply
to
surface
waters.
If
permit.'
limits
and
water
quality
criteria
are
both
expressed
in
terms
of
total
recoverable
metal,
extrapolations
from
effluent
to
surface
water
only
need
to
take
dilution
into
account
and
can
be
performed
as
mass
balance
calculations.
If
either
permit
limits
or
water
quality
criteria
or
both
are
expressed
in
terms
of
any
other
metal
measurement,
lability
needs
to
be
taken
into
account,
even
if
both
are
expressed
in
terms
of
the
same
measurement.

Extrapolations
concerning
labile
species
of
metals
from
effluent
to
surface
water
depend
to
a
large
extent
on
the
differences
between
the
water
quality
characteristics
of
the
effluent
and
those
of
the
surface
water.
Although
equilibrium
models
of
the
speciation
of
metals
can
provide
insight,
the
interactions
are
too
complex
to
be
able
to
make
useful
nonempirical
extrapolations
from
a
wide
variety
of
effluents
to
a
wide
variety
of
surface
waters
of
either
(
a)
the
speciation
of
the
metal
or
(
b)
a
metal
measurement
other
than
total
recoverable.

Rnpirical
extrapolations
can
be
performed
fairly
easily
and
the
most
cormon
case
will
probably
occur
when
petit
limits
are
based
on
the
total
recoverable
measurement
but
water
quality
criteria
are
based
on
the
dissolved
measurement.
The
empirical
extrapolation
is
intended
to
answer
the
question
What
percent
of
the
total
recoverable
metal
in
the
effluent
becomes
dissolved
in
the
downstream
water?'
This
question
can
be
answered
by:
a.
Collecting
samples
of
effluent
and
upstream
water.
b.
Measuring
total
recoverable
metal
and
dissolved
metal
in
both
samples.
c.
Combining
aliquots
of
the
two
samples
in
the
ratio
of
the
flows
when
the
sanrples
were
obtained
and
mixing
for
an
appropriate
period
of
time
under
appropriate
conditions.
d.
Measuring
total
recoverable
metal
and
dissolved
metal
in
the
mixture.
An
example
is
presented
in
Figure
D2.
This
percentage
cannot
be
extrapolated
from
one
metal
to
another
or
from
one
effluent
to
another.
The
data
needed
to
calculate
the
percentage
will
be
obtained
each
time
a
WER
is
determined
using
simulated
downstream
water
if
both
dissolved
and
total
recoverable
metal
are
measured
in
the
effluent,
upstream
water,
and
simulated
downstream
water.

The
interpretation
of
the
percentage
is
not
necessarily
as
straightforward
as
might
be
assumed.
For
example,
some
of
the
metal
that
is
dissolved
in
the
upstream
water
might
sorb
onto
particulate
matter
in
the
effluent,
which
can
be
viewed
as
a
detoxification
of
the
upstream
water
by
the
effluent.
Regardless
of
the
interpretation,
the
described
procedure
provides
a
simple
112
way
of
relating
the
total
recoverable
concentration
in
the
effluent
to
the
concentration
of
concern
in
the
downstream
water.
Because
this
empirical
extrapolation
can
be
used
with
any
analytical
measurement
that
is
chosen
as
the
basis
for
expression
of
aquatic
life
criteria,
use
of
the
total
recoverable
measurement
to
express
permit
limits
on
effluents
does
not
place
any
restrictions
on
which
analytical
measurement
can
be
used
to
express
criteria.
Further,
even
if
both
criteria
and
permit
limits
are
expressed
in
terms
of
a
measurement
such
as
dissolved
metal,
an
empirical
extrapolation
would
still
be
necessary
because
dissolved
metal
is
not
likely
to
be
conservative
from
effluent
to
downstream
water.

Merits
of
Total
Recoverable
and
Dissolved
WERs
and
Criteria
A
WER
is
operationally
defined
as
the
value
of
an
endpoint
obtained
with
a
toxicity
test
using
site
water
divided
by
the
value
of
the
same
endpoint
obtained
with
the
same
toxicity
test
using
a
laboratory
dilution
water.
Therefore,
just
as
aquatic
life
criteria
can
be
expressed
in
terms
of
either
the
total
recoverable
measurement
or
the
dissolved
measurement,
so
can
WERs.
A
pair
of
side­
by­
side
toxicity
tests
can
produce
both
a
total
recoverable
WER
and
a
dissolved
WER
if
the
metal
in
the
test
solutions
in
both
of
the
tests
is
measured
using
both
methods.
A
total
recoverable
WER
is
obtained
by
dividing
endpoints
that
were
calculated
on
the
basis
of
total
recoverable
metal,
whereas
a
dissolved
WER
is
obtained
by
dividing
endpoints
that
were
calculated
on
the
basis
of
dissolved
metal.
Because
of
the
way
they
are
determined,
a
total
recoverable
WER
is
used
to
calculate
a
total
recoverable
site­
specific
criterion
from
a
national,
state,
or
recalculated
aquatic
life
criterion
that
is
expressed
using
the
total
recoverable
measurement,
whereas
a
dissolved
WER
is
used
to
calculate
a
dissolved
site­
specific
criterion
from
a
national,
state,
or
recalculated
criterion
that
is
expressed
in
terms
of
the
dissolved
measurement.

In
terms
of
the
classification
scheme
given
in
Figure
Dl,
the
basic
relationship
between
a
total
recoverable
national
water
quality
criterion
and
a
total
recoverable
WER
is:

l
A
total
recoverable
criterion
treats
allthe
toxic
and
nontoxic
metal
in
the
site
water
as
if
its
average
toxicity
were
the
same
as
the
average
toxicity
of
all
the
toxic
and
nontoxic
metal
in
the
toxicity
tests
in
laboratory
dilution
water
on
which
the
criterion
is
based.

l
A
total
recoverable
WER
is
a
measurement
of
the
actual
ratio
of
the
ave=
e
toxicities
of
the
total
recoverable
metal
and
replaces
the
assumption
that
the
ratio
is
1.

113
Similarly,
the
basic
relationship
between
a
dissolved
national
.
crlterloa
and
a
dissolved
m
is:

l
A
dissolved
criterion
treats
all
the
toxic
and
nontoxic
dissolved
metal
in
the
site
water
as
if
its
average
toxicity
were
the
same
as
the
average
toxicity
of
all
the
toxic
and
nontoxic
dissolved
metal
in
the
toxicity
tests
in
laboratory
dilution
water
on
which
the
criterion
is
based.

l
A
dissolved
m
is
a
measurement
of
the
actual
ratio
of
the
average
toxicities
of
the
dissolved
metal
and
replaces
the
assumption
that
the
ratio
is
1.
In
both
cases,
use
of
a
criterion
without
a
WER
involves
measurement
of
toxicity
in
laboratory
dilution
water
but
only
prediction
of
toxicity
in
site
water,
whereas
use
of
a
criterion
with
a
wER
involves
measurement
of
toxicity
in
both
laboratory
dilution
water
and
site
water.

When
WERs
are
used
to
derive
site­
specific
criteria,
the
total
recoverable
and
dissolved
approaches
are
inherently
consistent.
They
are
consistent
because
the
toxic
effects
caused
by
the
metal
in
the
toxicity
tests
do
not
depend
on
what
chemical
measurements
are
performed;
the
same
number
of
organisms
are
killed
in
the
acute
lethality
tests
regardless
of
what,
if
any,
measurements
of
the
concentration
of
the
metal
are
made.
The
only
difference
is
the
chemical
measurement
to
which
the
toxicity
is
referenced.
Dissolved
WERs
can
be
derived
from
the
same
pairs
of
toxicity
tests
from
which
total
recoverable
WERs
are
derived,
if
the
metal
in
the
tests
is
measured
using
both
the
total
recoverable
and
dissolved
measurements.
Both
approaches
start
at
the
same
place
(
i.
e.,
the
amount
of
toxicity
observed
in
laboratory
dilution
water)
and
end
at
the
same
place
(
i.
e.,
the
amount
of
toxicity
observed
in
site
water).
The
combination
of
a
total
recoverable
criterion
and
WER
accomplish
the
same
thing
as
the
combination
of
a
dissolved
criterion
and
WER.
By
extension,
whenever
a
criterion
and
a
WER
based
on
the
same
measurement
of
the
metal
are
used
together,
they
will
end
up
at
the
same
place.
Because
use
of
a
total
recoverable
criterion
with
a
total
recoverable
m
ends
up
at
exactly
the
same
place
as
use
of
a
dissolved
criterion
with
a
dissolved
m,
whenever
one
WER
is
determined,
both
should
be
determined
to
allow
(
a)
a.
check
on
the
analytical
chemistry,
(
b)
use
of
the
inherent
internal
consistency
to
check
that
the
data
are
used
correctly,
and
(
c)
the
option
of
using
either
approach
in
the
derivation
of
permit
limits.
.
An
examination
of
how
the
two
approaches
(
the
total
recoverable
approach
and
the
dissolved
approach)
address
the
four
relevant
forms
of
metal
(
toxic
and
nontoxic
particulate
metal
and
toxic
and
nontoxic
dissolved
metal)
in
laboratory
dilution
water
and
in
site
water
further
explains
why
the
two
approaches
are
inherently
consistent.
Here,
only
the
way
in
which
the
two
approaches
address
each
of
the
four
forms
of
metal
in
site
water
will
be
considered:

114
a.
Toxic
dissolved
metal:
This
form
contributes
to
the
toxicity
of
the
site
water
and
is
measured
by
both
chemical
measurements.
If
this
is
the
only
form
of
metal
present,
the
two
WERs
will
be
the
same.
b.
Nontoxic
dissolved
metal:
This
form
does
not
contribute
to
the
toxicity
of
the
site
water,
but
it
is
measured
by
both
chemical
measurements.
If
this
is
the
only
form
of
metal
present,
the
two
WERs
will
be
the
same.
(
Nontoxic
dissolved
metal
can
be
the
only
form
present,
however,
only
if
all
of
the
nontoxic
dissolved
metal
present
is
refractory.
If
any
labile
nontoxic
dissolved
metal
is
present,
equilibrium
will
require
that
some
toxic
dissolved
metal
also
be
present.)
c.
Toxic
particulate
metal:
This
form
contributes
to
the
toxicological
measurement
in
both
approaches;
it
is
measured
by
the
total
recoverable
measurement,
but
not
by
the
dissolved
measurement.
Even
though
it
is
not
measured
by
the
dissolved
measurement,
its
presence
is
accounted
for
in
the
dissolved
approach
because
it
increases
the
toxicity
of
the
site
water
and
thereby
decreases
the
dissolved
WER.
It
is
accounted
for
because
it
makes
the
dissolved
metal
appear
to
be
more
toxic
than
it
is.
Most
toxic
particulate
metal
is
probably
not
toxic
when
it
is
particulate;
it
becomes
toxic
when
it
is
dissolved
at
the
gill
surface
or
in
the
digestive
system;
in
the
surface
water,
however,
it
is
measured
as
particulate
metal.
d.
Nontoxic
particulate
metal:
This
form
does
not
contribute
to
the
toxicity
of
the
site
water;
it
is
measured
by
the
total
recoverable
measurement,
but
not
by
the
dissolved
measurement.
Because
it
is
measured
by
the
total
recoverable
measurement,
but
not
by
the
dissolved
measurement,
it
causes
the
total
recoverable
WER
to
be
higher
than
the
dissolved
WER.
In
addition
to
dealing
with
the
four
forms
of
metal
similarly,
the
WERs
used
in
the
two
approaches
comparably
take
synergism,
antagonism,
and
additivity
into
account.
Synergism
and
additivity
in
the
site
water
increase
its
toxicity
and
therefore
decrease
the
WER;
in
contrast,
antagonism
in
the
site
water
decreases
toxicity
and
increases
the
WER.

Each
of
the
four
forms
of
metal
is
appropriately
taken
into
account
because
use
of
the
WERs
makes
the
two
approaches
internally
consistent.
In
addition,
although
experimental
variation
will
cause
the
measured
WERs
to
deviate
from
the
actual
WERs,
the
measured
WERs
will
be
internally
consistent
with
the
data
from
which
they
were
generated.
If
the
percent
dissolved
is
the
same
at
the
test
endpoint
in
the
two
waters,
the
two
WERs
will
be
the
same.
If
the
percent
of
the
total
recoverable
metal
that
is
dissolved
in
laboratory
dilution
water
is
less
than
100
percent,
changing
from
the
total
recoverable
measurement
to
the
dissolved
measurement
will
lower
the
criterion
but
it
will
115
comparably
lower
the
denominator
in
the
WER,
thus
increasing
the
WER.
If
the
percent
of
the
total
recoverable
metal
that
is
dissolved
in
the
site
water
is
less
than
100
percent,
changing
from
the
total
recoverable
measurement
to
the
dissolved
.
measurement
will
lower
the
concentration
in
the
site
water
that
is
to
be
compared
with
the
criterion,
but
it
also
lowers
the
numerator
in
the
WER,
thus
lowering
the
WER.
Thus
when
WEI&
are
used
to
adjust
criteria,
the
total
recoverable
approach
and
the
dissolved
approach
result
in
the
same
interpretations
of
concentrations
in
the
site
water
(
see
Figure
D3)
and
in
the
same
maximum
acceptable
concentrations
in
effluents
(
see
Figure
D4).

Thus,
if
WEFU
are
based
on
toxicity
tests
whose
endpoints
equal
the
CMC
or
CCC
and
if
both
approaches
are
used
correctly,
the
two
measurements
will
produce
the
same
results
because
each
WER
is
based
on
measurements
on
the
site
water
and
then
the
WER
is
used
to
calculate
the
site­
specific
criterion
that
applies
to
the
site
water
when
the
same
chemical
measurement
is
used
to
express
the
site­
specific
criterion.
The
eguivalency
of
the
two
approaches
applies
if
they
are
based
on
the
same
sample
of
site
water.
When
they
are
applied
to
multiple
samples,
the
approaches
can
differ
depending
on
how
the
results
from
replicate
samples
are
used:
a.
If
an
appropriate
averaging
process
is
used,
the
two
will
be
equivalent.
b.
If
the
lowest
value
is
used,
the
two
approaches
will
probably
be
equivalent
only
if
the
lowest
dissolved
WER
and
the
lowest
total
recoverable
WER
were
obtained
using
the
same
sample
of
site
water.

There
are
several
advantages
to
using
a
dissolved
criterion
even
when
a
dissolved
WER
is
not
used.
In
some
situations
use
of
a
dissolved
criterion
to
interpret
results
of
measurements
of
the
concentration
of
dissolved
metal
in
site
water
might
demonstrate
that
there
is
no
need
to
determine
either
a
total
recoverable
WER
or
a
dissolved
WER.
This
would
occur
when
so
much
of
the
total
recoverable
metal
was
nontoxic
particulate
metal
that
even
though
the
total
recovdrable
criterion
was
exceeded,
the
corresponding
dissolved
criterion
was
not
exceeded.
The
particulate
metal
might
come
from
an
effluent,
a
resuspension.
event,
or
runoff
that
washed
particulates
into
the
body
of
water.
In
such
a
situation
the
total
recoverable
WER
would
also
show
that
the
site­
specific
criterion
was
not
exceeded,
but
there
would
be
no
need
to
determine
a
WER
if
the
criterion
were
expressed
on
the
basis
of
the
dissolved
measurement.
If
the
variation
over
time
in
the
concentration
of
particulate
metal
is
much
greater
than
the
variation
in
the
concentration
of
dissolved
metal,
both
the
total
recoverable
concentration
and
the
total
recoverable
WER
are
likely
to
vary
so
much
over
time
that
a
dissolved
criterion
would
be
much
more
useful
than
a
total
recoverable
criterion.

116
Use
of
a
dissolved
criterion
without
a
dissolved
WER
has
three
disadvantages,
however:
1.
Nontoxic
dissolved
metal
in
the
site
water
is
treated
as
if
it
is
toxic.
2.
Any
toxicity
due
to
particulate
metal
in
the
site
water
is
ignored.
3.
Synergism,
antagonism,
and
additivity
in
the
site
water
are
not
taken
into
account.
Use
of
a
dissolved
criterion
with
a
dissolved
WER
overcomes
all
three
problems.
For
example,
if
(
a)
the
total
recoverable
concentration
greatly
exceeds
the
total
recoverable
criterion,
(
b)
the
dissolved
concentration
is
below
the
dissolved
criterion,
and
(
c)
there
is
concern
about
the
possibility
of
toxicity
of
particulate
metal,
the
determination
of
a
dissolved
WER
would
demonstrate
whether
toxicity
due
to
particulate
metal
is
measurable.

Similarly,
use
of
a
total
recoverable
criterion
without
a
total
recoverable
WER
has
three
comparable
disadvantages:
1.
Nontoxic
dissolved
metal
in
site
water
is
treated
as
if
it
is
toxic.
2.
Nontoxic
particulate
metal
in
site
water
is
treated
as
if
it
is
toxic.
3.
Synergism,
antagonism,
and
additivity
in
site
water
are
not
taken
into
account.
Use
of
a
total
recoverable
criterion
with
a
total
recoverable
WER
overcomes
all
three
problems.
For
example,
determination
of
a
total
recoverable
WER
would
prevent
nontoxic
particulate
metal
(
as
well
as
nontoxic
dissolved
metal)
in
the
site
water
from
being
treated
as
if
it
is
toxic.

pelationshim
between
WERs
and
the
Forms
of
Metal%

Probably
the
best
way
to
understand
what
WERs
can
and
cannot
do
is
to
understand
the
relationships
between
WERs
and
the
forms
of
metals.
A
WER
is
calculated
by
dividing
the
concentration
of
a
metal
that
corresponds
to
a
toxicity
endpoint
in
a
site
water
by
the
concentration
of
the
same
metal
that
corresponds
to
the
same
toxicity
endpoint
in
a
laboratory
dilution
water.
using
the
classification
scheme
given
in
Figure
Dl:
Therefore,

mR=
R,,
+
N,
+
T,
+
dV'
+
AT,
Rt
+
NL
+
Tt
+
AN,
+
AT,

The
subscripts
IS.
and
.
t'
denote
site
water
and
laboratory
dilution
water,
respectively,
and:

R
=
the
concentration
of
Befractory
metal
in
a
water.
definition,
(
By
all
refractory
metal
is
nontoxic
metal.)

117
N
=
the
concentration
of
Nontoxic
labile
metal
in
a
water.

T
=
the
concentration
of
xoxic
labile
metal
in
a
water.

AN
=
the
concentration
of
metal
added
during
a
WEX
determination
that
is
Nontoxic
labile
metal'after
it
is
added.

AT
=
the
concentration
of
metal
added
during
a
WER
determination
that
is
sxic
labile
metal
after
it
is
added.

For
a
total
recoverable
WER;
each
of
these
five
concentrations
includes
both
particulate
and
dissolved
metal,
if
both
are
present;
for
a
dissolved
WER
only
dissolved
metal
is
included.

Because
the
two
side­
by­
side
tests
use
the
same
endpoint
and
are
conducted
under
identical
conditions
with
comparable
test
organisms,
T,+
AT,
­
Tt
+
AT,
when
the
toxic
species
of
the
metal
are
equally
toxic
in
the
two
waters.
If
a
difference
in
water
quality
causes
one
or
more
of
the
toxic
species
of
the
metal
to
be
mOre
toxic
in
one
water
than
the
other,
or
causes
a
shift
in
the
ratios
of
various
toxic
species,
we
can
define
II=
Tn
+
AT,
T&+
AT,
'

Thus
H
is
a
multiplier
that
accounts
for
a
proportional
increase
or
decrease
in
the
toxicity
of
the
toxic
forms
in
site
water
as
compared
to
their
toxicities
in
laboratory
dilution
water.
Therefore,
the­
general
WE3
equation
is:

WERt
Rs+
N,+
aN,+
H(
TL+~
TL)
R,+
N,+
aN,+
(
T,+
AT,)
l
Several
things
are
obvious
from
this
equation:
1.
A
WER
should
not
be
thought
of
as
a
simple
ratio
such
as
H.
H
is
the
ratio
of
the
toxicities
of
the
toxic
species
of
the
metal,
whereas
the
WER
is
the
ratio
of
the
sum
of
the
toxic
and
the
nontoxic
species
of
the
metal.
Only
under
a
very
specific
set
of
conditions
will
wg~
­
H.
If
these
conditions
are
satisfied
and
if,
in
addition,
H=
1,
then
wg~
­
I.
Although
it
might
seem
that
all
of
these
conditions
will
rarely
be
satisfied,
it
is
not
all
that
rare
to
find
that
an
experimentally
determined
WER
is
close
to
1.
2.
When
the
concentration
of
metal
in
laboratory
dilution
water
is
negligible,
RL
­
NL
=
Tt
­
0
and
wm­
RH
+
NB
+
dV"
+
HAT,)
A&
+
AT,
.

118
Even
though
laboratory
dilution
water
is
low
in
Tot
and
TSS,
when
metals
are
added
to
laboratory
dilution
water
in
toxicity
tests,
ions
such
as
hydroxide,
carbonate1
and
chloride
react
with
some
metals
to
form
some
particulate
species
and.
some
dissolved
species,
both
of
which
might
be
toxic
or
nontoxic.
The
metal
species
that
are
nontoxic
contribute
to
A&,
whereas
those
that
are
toxic
contribute
to
AT,.
Hydroxide,
carbonate,
chloride,
Tot,
and
TSS
can
increase
AN'.
Anything
that
causes
AN',
to
differ
from
AN'
will
cause
the
WER
to
differ
from
1.
3.
Refractory
metal
and
nontoxic
labile
metal
in
the
site
water
above
that
in
the
laboratory
dilution
water
will
increase
the
WER.
Therefore,
if
the
WER
is
determined
in
downstream
water,
rather
than
in
upstream
water,
the
WER
will
be
increased
by
refractory
metal
and
nontoxic
labile
metal
in
the
effluent.
Thus
there
are
three
major
reasons
why
WERs
might
be
larger
or
smaller
than
1:
a.
The
toxic
species
of
the
metal
might
be
more
toxic
in
one
water
than
in
the
other,
i.
e.,
H
l
1.
b.
AN
might
be
higher
in
one
water
than
in
the
other.
C.
R
and/
or
N
might
be
higher
in
one
water
than
in
the
other.

The
last
reason
might
have
great
practical
importance
in
some
situations.
When
a
WER
is
determined
in
downstream
water,
if
most
of
the
metal
in
the
effluent
is
nontoxic,
the
WER
and
the
endpoint
in
site
water
will
correlate
with
the
concentration
of
metal
in
the
site
water.
In
addition,
they
will
depend
on
the
concentration
of
metal
in
the
effluent
and
the
concentration
of
effluent
in
the
site
water.
This
correlation
will
be
best
for
refractory
metal
because
its
toxicity
cannot
be
affected
by
water
quality
characteristics;
even
if
the
effluent
and
upstream
water
are
quite
different
so
that
the
water
quality
characteristics
of
the
site
water
depend
on
the
percent
effluent,
the
toxicity
of
the
refractory
metal
will
remain
constant
at
zero
and
the
portion
of
the
WER
that
is
due
to
refractory
metal
will
be
additive.

The
Denendence
of
WERs
on
the
Sensitivitv
of
Toxicity
Tests
It
would
be
desirable
if
the
magnitude
of,
the
WER
for
a
site
water
were
independent
of
the
toxicity
test
used
in
the
determination
of
the
WER,
so
that
any
convenient
toxicity
test
could
be
used.
It
can
be
seen
from
the
general
WER
equation
that
the
WER
will
be
independent
of
the
toxicity
test
only
if:

which
would
require
that
R,­
N,
­
A&
­
RL
­
NL
­
tiL
­
0.
(
It
would
be
easy
to
assume
that
Tt
­
0,
but
it
can
be
misleading
in
some
situations
to
make
more
simplifications
than
are
necessary.)

119
This
is
the
simplistic
concept
of
a
WER
that
would
be
advantageous
if
it
were
true,
but
which
is
not
likely
to
be
true
very
often.
Any
situation
in
which
one
or
more
of
the
terms
is
greater
than
zero
can
cause
the
WER
to
depend
on
the
sensitivity
of
the
toxicity
test,
although
the
difference
in
the
WERs
might
be
small.

Two
situations
that
might
be
common
can
illustrate
how
the
WER
can
depend
on
the
sensitivity
of
the
toxicity
test.
For
these
illustrations,.
there
is
no
advantage
to
assuming
that
H=
1,
so
a
will
be
retained
for
generality.
1.

2.
The
simplest
situation
is
when
R,>
0,
i.
e.,
when
a
substantial
concentration
of
refractory
metal
occurs
in
the
site
water.
If,
for
simplification,
it
is
assumed
that
N,­
aN,=
RL=
NL­
tiL­
Or
then:

Nmi­
R,
+
H(
T,
+
AT,)
I
R6
(
T,
+
AT,)
(
TL
+
AT,)
+
H.

The
quantity
T,+
AT,
obviously
changes
as
the
sensitivity
of
the
toxicity
test
changes.
When
R,­
0,
then
MER­
H
and
the
WER
is
independent
of
the
sensitivity
of
the
toxicity
test.
when
R#>
0,
then
the
WER
will
decrease
as
the
sensitivity
of
the
test
decreases
because
TL
+
AT,
will
increase.

More
complicated
situations
occur
when
(
N,
+
a~,)
>
0.
If,
for
simplification,
it
is
assumed
that
R6
­
RL
­
N,
­
AN,
­
0,
then:

=
­
(
N6
+
d7,)
+
H(
T,
+
A?`,)
­
(
4
+
fl#)
(
T
+
AT
+
H.
L
L)
(
T,
+
AT,)

a.
If
(
N,
+
AN,)
>
0
because
the
site
water
contains
a
substantial
concentration
of
a
complexing
agent
that
has
an
affinity
for
the
metal
and
if
corqplexation
converts
toxic
metal
into
nontoxic
metal,
the
complexation
reaction
will
control
the
toxicity
of
the
solution
(
Allen
1993).
A
complexation
cume
can
be
graphed
in
several
ways,
but
the
S­
shaped
curve
presented
in
Figure
DS
is
most
convenient
here.
The
vertical
axis
is
'%
uncomplexed',
which
is
assumed
to
correlate
with
'%
toxic'.
then
the
'%
nontoxic'.
The
ratio
of
toxic
metal
is:

%
nontoxic,
%
ccnl@
ad
%
toxic
%
Llncoilplm!
Bd
For
the
complexed
nontoxic
metal:
The
'%
complexed'
is
nontoxic
metal
to
­
v.

vt
concentration
ofnontoxdcmetal
concentration
of
toxicmetal
l
120
In
the
site
water,
the
concentration
of
complexed
nontoxic
metal
is
(
N,
+
ti6)
and
the
concentration
of
toxic
metal
is
(
T,+
AT,),
so
that:.

(
Iy,
+
u#)
(
N4
+
a#)
v6
­
(
T6
+
AT,)
­
H(
T,
+
AT,)
l
and
wBRN
V&(
TL
+
AT,)
+
H(=,
+
AT,)

(
TL
+
AT,)
­
vfi+
H=
H(
V,+
1)
.

If
the
WRR
is
determined
using
a
sensitive
toxicity
test
so
that
the
%
uncomplexed
(
i.
e.,
the
%
toxic)
is
10
%,
then
v,­
(
90
%)/(
I0
%)
­
9,
whereas
if
a
less
sensitive
test
is
used
so
that
the
%
uncomplexed
is
50
%,
then
v,­
(
50
%)/(
50
%)
­
1.
Therefore,
if
a
portion
of
the
WER
is
due
to
a
complexing
agent
in
the
site
water,
the
magnitude
of
the
WER
can
decrease
as
the
sensitivity
of
the
toxicity
test
decreases
because
the
%
uncomplexed
will
decrease.
In
these
situations,
the
largest
WER
will
be
obtained
with
the
most
sensitive
toxicity
test;
progressively
smaller
WERs
will
be
obtained
with
less
sensitive
toxicity
tests.
The
magnitude
of
a
WER
will
depend
not
only
on
the
sensitivity
of
the
toxicity
test
but
also
on
the
concentration
of
the
complexing
agent
and
on
its
binding
constant
(
complexation
constant,
stability
constant).
In
addition,
the
binding
constants
of
most
complexing
agents
depend
on
pH.

If
the
laboratory
dilution
water
contains
a
low
concentration
of
a
complexing
agent,

VL
­
ru,
+
flL
TL
+
AT,

and
mR­
v&
T,
+
AT,)
+
H(
T,
+
A?',)
=
v,$
+
H
I
H(
v,
+
1)
.

VL(
TL
+
AT,)
+
tTL
+
AT,)
v,
+
1
v,
+
1
The
binding
constant
of
the
complexing
agent
in
the
laboratory
dilution
water
is
probably
different
from
that
of
the
complexing
agent
in
the
site
water.
Although
changing
from
a
more
sensitive
test
to
a
less
sensitive
test
will
decrease
both
V,
and
v,,
the
amount
of
effect
is
not
likely
to
be
proportional.

If
the
change
from
a
more
sensitive
test
to
a
less
sensitive
test
were
to
decrease
v,
proportionately
more
than
v,,
the
change
could
result
in
a
larger
WER,
rather
121
than
a
smaller
WER,
as
resulted
in
the
case
above
when
it
was
assumed
that
the
laboratory
dilution
water
did
not
contain
any
complexing
agent.
This
is
probably
most
likely
to
occur
if
H?
1
and
if
v#
<
V,,
which
would
mean
that
lygR<
l.
Although
this
is
likely
to
be
a
rare
situation,
it
does
demonstrate
again
the'importance
of
determining
WERS
using
toxicity
tests
that
have
endpoints
in
laboratory
dilution
water
that
are
close
to
the
CMC
or
CCC
to
which
the
WER
is
to
be
applied.

b.
If
(
N,+
AN,)
>
0
because
the
site
water
contains
a
substantial
concentration
of
an
ion
that
will
precipitate
the
metal
of
concern
and
if
precipitation
converts
toxic
metal
into
nontoxic
metal,
the
precipitation
reaction
will
control
the
toxicity
of
the
solution.
The
'
precipitation
curve.
given
in
Figure
D6
is
analogous
to
the
'
coxnplexation
curve.
given
in
Figure
D5;
in
the
precipitation
curve,
the
vertical
axis
is
'%
dissolved.,
which
is
assumed
to
correlate
with
'
8
toxic'.
If
the
endpoint
for
a
toxicity
test
is
below
the
solubility
limit
of
the
precipitate,
(
N,+
AN,)
­
Or
whereas
if
the
endpoint
for
a
toxicity
test
is
above
the
solubility
limit,
(
N,
+
ti#)
>
0.
If
WERs
are
determined
with
a
series
of
toxicity
tests
that
have
increasing
endpoints
that
are
above
the
solubility
limit,
the
WER
will
reach
a
maximum
value
and
then
decrease.
The
magnitude
of
the
WER
will
depend
not
only
on
the
sensitivity
of
the
toxicity
test
but
also
on
the
concentration
of
the
precipitating
agent,
the
solubility
limit,
and
the
solubility
of
the
precipitate.

Thus,
depending
on
the
composition
of
the
site
water,
a
WER
obtained
with
an
insensitive
test
might
be
larger,
smaller,
or
similar
to
a
WER
obtained
with
a
sensitive
test.
Because
of
the
range
of
possibilities
that
exist,
the
best
toxicity
test
to
use
in
the
experimental
determination
of
a
WEX
is
one
whose
endpoint
in
laboratory
dilution
water
is
close
to
the
CMC
or
CCC
that
is
to
be
adjusted.
This
is
the
rationale
that
was
used
in
the
selection
of
the
toxicity
tests
that
are
suggested
in
Appendix
I.

The
available
data
indicate
that
a
less
sensitive
toxicity
test
usually
gives
a
smaller
WER
than
a
more
sensitive
test
(
Hansen
1993a).
Thus,
use
of
toxicity
tests
whose
endpoints
are
higher
than
the
CMC
or
CCC
probably
will
not
result
in
underprotection;
in
contrast,
use
of
tests
whose
endpoints
are
substantially
below
the
CMC
or
CCC
might
result
in
underprotection.

The
factors
that
cause
~~
and
(
N,
+
AN,)
to
be
greater
than
zero
are
all
external
to
the
test
organisms;
they
are
chemical
effects
that
affect
the
metal
in
the
water.
The
magnitude
of
the
WER
is
therefore
expected
to
depend
on
the
toxicity
test
used
only
in
regard
to
the
sensitivity
of
the
test.
If
the
endpoints
for
two
122
different
tests
occur
at
the
same
concentration
of
the
metal,
the
magnitude
of
the
WERs
obtained
with
the
two
tests­
should
be
the
same;
they
should
not
depend
on
(
a)
the
duration
of
the
test,
(
b)
whether
the
endpoint
is
based
on
a
lethal
or
sublethal
effect,
or
(
c)
whether
the
species
is
a
vertebrate
or
an
invertebrate.

Another
interesting
consequence
of
the
chemistry
of
complexation
is
that
the
%
uncosnplexed
will
increase
if
the
solution
is
diluted
(
Allen
and
Hansen
1993).
The
concentration
of
total
metal
will
decrease
with
dilution
but
the
%
uncomplexed
will
increase.
The
increase
will
not
offset
the
decrease
and
so
the
concentration
of
uncomplexed
metal
will
decrease.
Thus
the
portion
of
a
WER
that
is
due
to
complexation
will
not
be
strictly
additive
(
see
Appendix
G),
but
the
amount
of
nonadditivity
might
be
difficult
to
detect
in
toxicity
studies
of
additivity.
A
similar
effect
of
dilution
will
occur
for
precipitation.

The
illustrations
presented
above
were
simplified
to
make
it
easier
to
understand
the
kinds
of
effects
that
can
occur.
The
illustrations
are
qualitatively
valid
and
demonstrate
the
direction
of
the
effects,
but
real­
world
situations
will
probably
be
so
much
mOre
complicated
that
the
various
effects
cannot
be
dealt
with
separately.

.
Pther
ProDertles
of
WQ&

1.
Because
of
the
variety
of
factors
that
can
affect
WERs,
no
rationale
exists
at
present
for
extrapolating
WERs
from
one
metal
to
another,
from
one
effluent
to
another,
or
from
one
surface
water
to
another.
Thus
WERs
should
be
individually
determined
for
each
metal
at
each
site.

2.
The
most
important
infonaation
that
the
determination
of
a
WER
provides
is
whether
simulated
and/
or
actual
downstream
water
adversely
affects
test
organisms
that
are
sensitive
to
the
metal.
A
WER
cannot
indicate
how
much
metal
needs
to
be
removed
from
or
how
much
metal
can
be
added
to
an
effluent.
a.
If
the
site
water
already
contains
sufficient
metal
that
it
is
toxic
to
the
test
organisms,
a
WER
cannot
be
determined
with
a
sensitive
test
and
so
an
insensitive
test
will
have
to
be
used.
Even
if
a
WER
could
be
determined
with
a
sensitive
test,
the
WER
cannot
indicate
how
much
metal
has
to
be
removed.
For
example,
if
a
WER
indicated
that
there
was
20
percent
too
much
metal
in
an
effluent,
a
30
percent
reduction
by
the
discharger
would
not
reduce
toxicity
if
only
nontoxic
metal
was
removed.
Thenext
WER
determination
would
show
that
the
effluent
still
contained
too
much
metal.
Removing
metal
is
useful
only
if
the
metal
removed
is
toxic
metal.
Reducing
the
total
recoverable
concentration
does
not
necessarily
reduce
toxicity.

123
3.
b.
If
the
simulated
or
actual
downstream
water
is
not
toxic,
a
WER
can
be
determined
and
used
to
calculate
how
much
additional
metal
the
effluent
could
contain
and
still
be
acceptable.
Because
an
unlimited
amount
of
refracmry
metal
can
be
added
to
the
effluent
without
affecting
the
organisms,
what
the
WER
actually
determines
is
how
much
additional
toxic
metal
can.
be
added
to
the
effluent.

The
effluent
component
of
nearly
all
WEFU
is
likely
to
be
due
mostly
to
either
(
a)
a
reduction
in
toxicity
of
the
metal
by
TSS
or
'
pot,
or
(
b)
the
presence
of
refractory
metal.
For
both
of
these,
if
the
percentage
of
effluent
in
the
downstream
water
decreases,
the
magnitude
of
the
WER
will
usually
decrease.
If
the
water
quality
characteristics
of
the
effluent
and
the
upstream
water
are
quite
different,
it
is
possible
that
the
interaction
will
not
be
additive;
this
can
affect
the
portion
of
the
WER
that
is
due
to
reduced
toxicity
caused
by
sorption
and/
or
binding,
but
it
cannot
affect
the
portion
of
the
WER
that
is
due
to
refractory
metal.

4.
Test
organisms
are
fed
during
some
toxicity
tests,
but
not
during
others;
it
is
not
clear
whether
a
WER
determined
in
a
fed
test
will
differ
from
a
WER
determined
in
an
unfed
test.
Whether
there
is
a
difference
is
likely
to
depend
on
the
metal,
the
type
and
amount
of
food,
and
whether
a
total
recoverable
or
dissolved
WER
is
determined.
This
can
be
evaluated
by
determining
two
WERs
using
a
test
in
which
the
organisms
usually
are
not
fed
­
one
WER
with
no
food
added
to
the
tests
and
one
with
food
added
to
the
tests.
Any
effect
of
food
is
probably
due
to
an
increase
in
TOC
and/
or
TSS.
If
food
increases
the
concentration
of
nontoxic
metal
in
both
the
laboratory
dilution
water
and
the
site
water,
the
food
will
probably
decrease
the
WER.
Because
complexes
of
metals
are
usually
soluble,
cqlexation
is
likely
to
lower
both
total
recoverable
and
dissolved
WEF&;
sorption
to
solids
will
probably
reduce
only
total
recoverable
WERs.
The
food
might
also
affect
the
acute­
chronic
ratio.
Any
feeding
during
a
test
should
be
limited
to
the
minimum
necessary.

The
acceptable
WERs
found
by
Brungs
et
al.
(
1992)
were
total
recoverable
WERs
that
were
determined
in
relatively
clean
fresh
water.
These
WERs
ranged
from
about
1
to
15
for
both
copper
and
cadmium,
whereas
they
ranged
from
about
0.7
to
3
for
zinc.
The
few
WEF&
that
were
available
for
chromium,
lead,
and
nickel
ranged
from
about
1
to
6.
Both
the
total
recoverable
and
dissolved
WERs
for
copper
in
New
York
harbor
range
from
about
0.4
to
4
with
most
of
the
WERs
being
between
1
and
2
(
Hansen
1993b).

124
Figure
D2:
An
&
amp10
of
the
Empirical
S%
trapolation
Proces8
Assume
the
following
hypothetical
effluent
and
upstream
water:

Effluent:
T,:
100
ug/
L
D6:
10
ug/
L
U*:
24
cfs
Upstream
water:
T,:
40
ug/
L
Do:
38
ug/
L
Pa:
48
cfs
Downstream
water:
Tn
:
60
ug/
L
4.8:
36
ug/
L
QD:
72
cfs
where:

T
=
concentration
of
D
=
concentration
of
u
=
flow.
(
10
8
dissolved)

(
95
%
dissolved)

(
60
%
dissolved)

total
recoverable
metal.
dissolved
metal.

The
subscripts
E,
U,
and
D
signify
effluent,
upstream
water,
and
downstream
water,
respectively.

By
conservation
of
flow:
&­
Q6+
PO.

By
conservation
of
total
recoverable
metal:
TIpo­
Tp,+
Trpo.

If
P
=
the
percent
of
the
total
recoverable
metal
in
the
effluent.
that
becomes
dissolved
in
the
downstream
water,

pI
loo(~
D­
DI&)

W6
'

For
the
data
given
above,
the
percent
of
the
total
recoverable
metal
in
the
effluent
that
becomes
dissolved
in
the
downstream
water
is:

p
I
100
[
(
36
ug/
t)
(
72
CfB)
­
(
38
W/
t)
(
48
Cfd
1
.
32
Q
,
(
100
u&
L)
(
24
Cf8)

which
is
greater
than
the
10
%
dissolved
in
the
effluent
and
less
than
the
60
%
dissolved
in
the
downstream
water.

125
Figure
03:
The
Intormal
Coa8imtuxcy
of
the
Two
Approachem
The
internal
consistency
of
the
total
recoverable
and
dissolved
approaches
can
be
illustrated
by
considering
the
use
of
WERS
to
interpret
the
total
recoverable
and.
dissolved
concentrations
of
a
metal
in
a
site
water.
For
this
hypothetical
example,
it
will
be
assumed
that
the
national
CCCs
for
the
metal
are:
200
ug/
L
as
total
recoverable
metal.
160
ug/
L
as
dissolved
metal.
It
will
'
also
be
assumed
that
the
concentrations
of
the
metal
in
the
site
water
are:
300
ug/
L
as
total
recoverable
metal.
120
ug/
L
as
dissolved
metal.
The
total
recoverable
concentration
in
the
site
water
exceeds
the
national
CCC,
but
the
dissolved
concentration
does
not.

The
following
results
might
be
obtained
if
WERs
are
determined:

atorv
.
.
Dilution
Water
Total
recoverable
LCSO
=
400
ug/
L.
%
of
the
total
recoverable
metal
that
is
dissolved
=
80.
(
This
is
based
on
the
ratio
of
the
national
CCCs,
which
were
determined
in
laboratory
dilution
water.)
Dissolved
LCSO
=
320
ug/
L.

Total
recoverable
LCSO
=
620
ug/
L.
%
of
the
total
recoverable
metal
that
is
dissolved
=
40.
(
This
is
based
on
the
data
given
above
for
site
water).
Dissolved
LCSO
=
248
ug/
L.

Total
recoverable
WEB
=
(
620
ug/
L)/(
400
ug/
L)
=
1.55
Dissolved
WER
=
(
248
ug/
L)/(
320
ug/
L)
=
0.775
Tmalrocowrabl~
WgR
nissolvedclrgR
91.55.
0.775
lab
water
%
dissolvd
*
80
N
2
site
hater
%
dfssolwd
40
Total
recoverable
ssCCC
=
(
200
ug/
L)(
l.
Sf)
=
310
ug/
L.
Dissolved
ssCCC
=
(
160
ug/
L)(
O.
775)
=
124
ug/
L.

Both
concentrations
in
site
water
are
below
the
respective
sscccs.

126
In
contrast,
the
following
results
might
have
been
obtained
when
the
WERs
were
determined:

In
Laboratorv
Dilution
Wate
Total
recoverable
LCSO
=
r400
ug/
L.
%
of
the
total
recoverable
metal
that
is
dissolved
=
80.
Dissolved
LCSO
=
320
ug/
L.

In
Site
ate
TotalWre&
erable
LC50
=
580
ug/
L.
%
of
the
total
recoverable
metal
that
is
dissolved
=
40.
Dissolved
LC50
=
232
ug/
L.

S
Total
recoverable
WER
=
(
580
ug/
L)/(
QoO
ug/
L)
=
Dissolved
WER
=
(
232
ug/
L)/(
320
ug/
L)
=
0.725
Checkina
the
Calculations
mta1
reBcoverable
WER
1.45
t­
9
lab
water
0
dissolwd
Dissolved
WHZ
0.725
site
water
0
ciissolv8d
1.45
980.2
40
Site­
snecific
CCCs
(
ssCCCs)

Total
recoverable
ssCCC
=
(
200
ug/
L)
(
1.45)
=
290
ug/
L.
Dissolved
ssCCC
=
(
160
ug/
L)
(
0.725)
=
116
ug/
L.

In
this
case,
both
respective
ssCCCs.
concentrations
in
site
water
are
above
the
In
each
case,
both
approaches
resulted
in
the
same
concerning
whether
the
concentration
in
site
water
site­
specific
criterion.
conclusion
exceeds
the
The
two
key
assumptions
are:
1.
The
ratio
of
total
recoverable
metal
to
laboratory
dilution
water
when
the
WERs
.
the
ratio
of
the
national
CCCs.
2.
The
ratio
of
total
recoverable
metal
to
site
water
when
the
WERs
are
determined
the
concentrations
reported
in
the
site
Differences
in
the
ratios
that
are
outside
experimental
variation
will
cause
problems
dissolved
metal
in
are
determined
equals
dissolved
metal
in
equals
the
ratio
of
water.
the
range
of
for
the
derivation
of
site­
specific
criteria
and,
therefore,
with
the
internal
consistency
of
the
two
approaches.

127
Figure
01:
The
AOplicrtion
of
tha
SW0
Approrchm
Hypothetical
upstream
water
and
effluent
will
be
used
to
demonstrate
the
equivalence
of
the
total
recoverable
and
dissolved
approaches.
The
upstream
water
and
the
effluent
will
be
assumed
to
have
specific
properties
in
order
to
allow
calculation
of
the
properties
of
the
downstream
water,
which
will
be
assumed
to
be
a
1:
l
mixture
of
the
upstream
water
and
effluent.
It
will
also
be
assumed
that
the
ratios
of
the
forms
of
the
metal
in
the
upstream
water
and
in
the
effluent
do
not
change
when
the
total
recoverable
concentration
changes.

YEfFEergiF
=
3
cfs)
Refractory
par&
ulate:
400
ug/
L
200
ug/
L
Toxic
dissolved:
200
ug/
L
(
50
%
dissolved)

Bffluea
(
Flow
=
3
cfs)
Total
recoverable:
440
ug/
L
Refractory
particulate:
396
ug/
L
Labile
nontoxic
particulate:
44
ug/
L
Toxic
dissolved:
0
ug/
L
(
0
%
dissolved)
(
The
labile
nontoxic
particulate,
which
is
10
%
of
the
total
recoverable
in
the
effluent,
becomes
toxic
dissolved
in
the
downstream
water.)

(
Flow
=
6
cfs)
Downstream
wateq
Total
recoverable:
420
ug/
L
Refractory
particulate:
298
ug/
L
Toxic
dissolved:
122
ug/
L
(
29
%
dissolved)

The
values
for
the
downstream
water
are
calculated
from
the
values
for
the
upstream
water
and
the
effluent:
Total
recoverable:
[
3(
400)
+
3(
440)]/
6
=
420
ug/
L
Dissolved:
[
3(
200)
+
3(
44+
0)]/
6
=
122
ug/
L
Refractory
particulate:
[
3(
200)
+
3(
396))/
6
=
298
ug/
L
Assumed
National
CCC
(
CCC)
Total
recoverable
tn300
ug/
L
Dissolved
=
240
ug/
L
128
Uostream
site­
snecific
CCC
(
ussCCC)

Assume:
Dissolved
CCCWER
=
1.2
Dissolved
ussCCC
=
(
1.2)(
240
ug/
L)
=
288
ug/
L
By
calculation:
TR
ussCCC
=
(
288
ug/
L)/(
O.
S)
=
576
ug/
L
Total
recoverable
CCCWER
=
(
576
ug/
L)/(
300
ug/
L)
=
1.92
30:
ccc
ECCWER
ussccc
co
.
Total
recoverable:
ug/
L
1.92
576
ug/
L
Dissolved:
400n:
g,
L
240
ug/
L
1.2
288
ug/
L
200
ug/
L
%
dissolved
80
%
­­­­
50
%
50
%
Neither
concentration
exceeds
its
respective
ussCCC.

lmalreowrablemR
91.92.
lab
water
0
dissolwd
Dissolved
HER
Sit8
diseolwd
N
1.2
water
%
80
50
N1
l
6
Downstream
site­
soecific
CCC
(
dssCCC)

Assume:
Dissolved
cccWER
=
1.8
Dissolved
dssCCC
=
(
1.8)
(
240
ug/
L)
=
432
ug/
L
By
calculation:
TR
dssCCC
=
((
432
ug/
L­
I(
200
ug/
L)/
2l)/
O.
ll+((
400
ug/
L)/
2)
=
3520
ug/
L
This
calculation
determines
the
amount
of
dissolved
metal
contributed
by
the
effluent,
accounts
for
the
fact
that
ten
percent
of
the
total
recoverable
metal
in
the
effluent
becomes
dissolved,
and
adds
the
total
recoverable
metal
contributed
by
the
upstream
flow.
Total
recoverable
cccWER
=
(
3520
ug/
L)/(
300
ug/
L)
=
11.73
ccc
Total
recoverable:
30:
ug/
L
cccl+
mR
dssCCC
co
c.
11.73
Dissolved:
3520
ug/
L
420nug,
L
240
ug/
L
1.80
432
ug/
L
%
dissolved
122
ug/
L
80
%
­­­­
12.27
%
29
%
Neither
concentration
exceeds
its
respective
dssCCC.

Total
recowrable
HER
80
Dissolved
MER
N
11.73
N
lab
water
0
disoolwd
1.80
site
­
­
water
0
dissolwd
12.27
­
6.52
Calculatina
the
Maximum
Accentable
Concentration
in
the
Effluent
Because
neither
the
total
recoverable
concentration
nor
the
dissolved
concentration
in
the
downstream
water
exceeds
its
respective
site­
specific
CCC,
the
concentration
of
metal
in
the
effluent
could
be
increased.
Under
the
assumption
that
the
ratios
of
the
two
forms
of
the
metal
in
the
effluent
do
not
change
when
the
total
recoverable
concentration
changes,
the
maximum
acceptable
concentration
of
total
recoverable
metal
in
the
effluent
can
be
calculated
as
follows:

129
Starting
with
the
total
recoverable
dssCCC
of
3520
ug/
L
(
6
cfs)
(
3520
w/
L)
­
(
3
cfs)(
400
up/
=)
3
cfs
­
6640
ug/
L
Starting
with
the
dissolved
dssCCC
of
432
ug/
L
(
6
cfs)
(
432
u&
L)
­
(
3
cf8)
(
400
ug/=)
(
o­
5)
N
6640
ug,~
(
3
cfs)
(
0.10)

Total
recoverable:

(
3
cfs)
(
6640
u#/
L)
+
(
3
cfs)
(
400
ug/
='
6
cfs
­
3520
u&
L.

Dissolved:

(
3
cf8)
(
6640
ug/
L)
(
0.10)
+
(
3
Cf8)
(
400
Ug/
L)(
0*
50)
6
cfs
­
432
ug/
L
.

The
value
of
0.10
is
used
because
this
is
the
percent
of
the
total
recoverable
metal
in
the
effluent
that
becomes
dissolved
in
the
downstream
water.

The
values
of
3520
ug/
L
and
432
ug/
L
equal
the
downstream
site­
specific
CCCs
derived
above.

Another
Wav
to
Calculate
the
Maximum
Accentable
Concentration
The
maximum
acceptable
concentration
of
total
recoverable
metal
in
the
effluent
can
also
be
calculated
from
the
dissolved
dssCCC
of
432
ug/
L
using
a
partition
coefficient
to
convert
from
the
dissolved
dssCCC
of
432
ug/
L
to
the
total
recoverable
dssCCC
of
3520
ug/
L:

16
cfi]
[
4;
21g;
L
­
(
3
cfs)
(
400
I&
L)]
.
3
cfs
­
6640
tag/
L.

Note
that
the
value
used
for
the
partition
coefficient
in
this
calculation
is
0.1227
(
the
one
that
applies
to
the
downstream
water
when
the
total
recoverable
concentration
of
metal
in
the
effluent
is
6640
ug/
L),
not
0.29
(
the
one
that
applies
when
the
concentration
of
metal
in
the
effluent
is
only
420
ug/
L).
The
three
ways
of
calculating
the
maximum
acceptable
concentration
give
the
same
result
if
each
is
used
correctly.

130
Figure
D5:
A
Qenoralimd
Coqplexation
Cufpe
The
curve
is
for
a
constant
concentration
of
the
complexing
ligand
and
an
increasing
concentration
of
the
metal.

100
h
W
ti
.
.
.

.
.
.

.

.

.

.

.

.

.

.

.
.
.
.
I
I
I
LOG
OF
CONCENTRATION
OF
METAL
131
?
iguro
D6t
A
Gumralir~
Precipitation
CURO
The
cume
is
for
a
constant
concentration
of
the
precipitating
ligand
and
an
increasing
concentration
of
the
metal.

loo­*
l
l
.

l
0
I
I
I
I
LOG
OF
CONCENTRATION
OF
METAL
132
Allen,
H.
E.
1993.
Importance
of
Metal
Speciation
to
Toxicity.
Proceedings
of
the
Water
Environment
Federation
Workshop
on
Aguatic
Life
Criteria
for
Metals.
Anaheim,
CA.
pp.
55­
62.

Allen,
H.
E.,
and
D.
J.
Hansen.
1993.
The
Importance
of
Trace
Metal
Speciation
to
Water
Quality
Criteria.
Paper
presented
at
Society
for
Environmental
Toxicology
and
Chemistry.
Houston,
TX.
November
15.

Borgmanxi,
U.
1983.
Metal
Speciation
and
Toxicity
of
Free
Metal
Ions
to
Aquatic
Biota.
IN:
Aquatic
Toxicology.
(
J.
O.
Nriagu,
ed.)
Wiley,
New
York,
NY.

Brungs,
W.
A.,
T.
S.
Holderman,
and
M.
T.
Southerland.
1992.
Synopsis
of
Water­
Effect
Ratios
for
Heavy
Metals
as
Derived
for
Site­
Specific
Water
Quality
Criteria.
U.
S.
EPA
Contract
68­
CO­
0070.

Chapman,
G.
A.,
and
J.
K.
McCrady.
1977.
Copper
Toxicity:
A
Question
of
Form.
In:
Recent
Advances
in
Fish
Toxicology
Tubb,
ed.)
EPA­
600/
3­
77­
085
or
PB­
273
500.
National
T&~~
c~
l
.
Information
Service,
Springfield,
VA.
pp.
132­
151.

Erickson,
R.
1993a.
Memorandum
to
C.
Stephan.
July
14.

Erickson,
R.
1993b.
Memorandum
to
C.
Stephan.
November
12.
.
French,
P.,
and
D.
T.
E.
Hunt.
1986.
The
Effects
of
Inorganic
Complexing
upon
the
Toxicity
of
Copper
to
Aquatic
Organisms
(
Principally
Fish).
IN:
Trace
Metal
Speciation
and
Toxicity
to
Aguatic
Organisms
­
A
Review.
(
D.
T.
E.
Hunt,
ea.
1
Report
TR
247.
Water
Research
Centre,
United
Kingdom.

Hansen,
D.
J.
1993a.
Memorandum
to
C.
E.
Stephan.
April
29.

Hansen,
D.
J.
1993b.
Memorandum
to
C.
E.
Stephan.
October
6.

Nelson,
H.,
D.
Benoit,
R.
Erickson,
V.
Mattson,
and
J.
Lindberg.
1986.
The
Effects
of
Variable
Hardness,
pH,
Alkalinity,
Suspended
Clay,
and
Humics
on
the
Chemical
Speciation
and
Aquatic
Toxicity
of
Copper.
PB86­
171444.
National
Technical
Information
Service,
Springfield,
VA.

Wilkinson,
K.
J.,
P.
M.
Bertsch,
C.
H.
Jagoe,
and
P.
G.
C.
Campbell.
1993.
Surface
Complexation
of
Aluminum
on
Isolated
Fish
Gill
Cells.
Environ.
Sci.
Technol.
27:
1132­
1138.

133
Appmdix
St
U.
S.
SPA
Aquatic
Lifo
Criteria
Dccunntm
for
Hotrlm
Aluminum
Antimony
Arsenic
Beryllium
Cadmium
chromium
Copper
Lead
Mercury
Nickel
Selenium
Silver
Thallium
Zinc
EPA
440/
S­
86­
008
EPA
440/
S­
80­
020
EPA
440/
S­
84­
033
EPA
440/
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80­
024
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440/
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84­
032
EPA
440/
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029
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440/
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84­
031
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440/
S­
84­
027
EPA
440/
S­
84­
026
EPA
440/
S­
86­
004
EPA
440/
S­
87­
006
EPA
440/
S­
80­
071
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440/
S­
80­
074
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440/
S­
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003
PB88­
245998
PB81­
117319
PB85­
227445
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117350
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227031
PB85­
227478
PB85­
227023
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227437
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NTIS)
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Road
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TEL:
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134
­
ix
F:
Coruidorrtionm
Concoming
Multiple­
Metal,
Ibultiplo­
Dischugo,
axad
Special
Flowing­
Water
SitUatiOM
Multinle­
Metal
Situation9
Both
Method
1
and
Method
2
work
well
in
multiple­
metal
situations,
although
the
amount
of
testing
required
increases
as
the
number
of
metals
increases.
The
major
problem
is
the
same
for
both
methods:
even
when
addition
of
two
or
more
metals
individually
is
acceptable,
simultaneous
addition
of
the
two
or
more
metals,
each
at
its
respective
maximum
acceptable
concentration,
might
be
unacceptable
for
at
least
two
reasons:
1.
Additivity
or
synergism
might
occur
between
metals.
2.
More
than
one
of
the
metals
might
be
detoxified
by
the
same
complexing
agent
in
the
site
water.
When
WERs
are
determined
individually,
each
metal
can
utilize
all
of
the
complexing
capacity;
when
the
metals
are
added
together,
however,
they
cannot
simultaneously
utilize
all
of
the
complexing
capacity.
Thus
a
discharger
might
feel
that
it
is
cost­
effective
to
try
to
justify
the
lowest
site­
specific
criterion
that
is
acceptable
to
the
discharger
rather
than
trying
to
justify
the
highest
site­
specific
criterion
that
the
appropriate
regulatory
authority
might
approve.

There
are
two
options
for
dealing
with
the
possibility
of
additivity
and
synergism
between
metals:
a.

b.
WERs
could
be
developed
using
a
mixture
of
the
metals
but
it
might
be
necessary
to
use
several
primary
toxicity
tests
depending
on
the
specific
metals
that
are
of
interest.
Also,
it
might
not
be
clear
what
ratio
of
the
metals
should
be
used
in
the
mixture.
If
a
WER
is
determined
for
each
metal
individually,
one
or
more
additional
toxicity
tests
must
be
conducted
at
the
end
to
show
that
the
combination
of
all
metals
at
their
proposed
new
site­
specific
criteria
is
acceptable.
Acceptability
must
be
demonstrated.
with
each
toxicity
test
that
was
used
as
a
primary
toxicity
test
in
the
determination
of
the
WEXs
for
the
individual
metals.
Thus
if
a
different
primary
test
was
used
for
each
metal,
the
number
of
acceptability
tests
needed
would
equal
the
number
of
metals.
It
is
possible
that
a
toxicity
test
used
as
the
primary
test
for
one
metal
might
be
more
sensitive
than
the
'
CMC
(
or
CCC)
for
another
metal
and
thus
might
not
be
usable
in
the
combination
test
unless
antagonism
occurs.
When
a
primary
test
cannot
be
used,
an
acceptable
alternative
test
must
be
used.
The
second
option
is
nreferred
because
it
is
more
definitive;
it
provides
data
for
each
metal
individually
and
for
the
mixture.
The
first
option
leaves
the
possibility
that
one
of
the
metals
is
antagonistic
towards
another­
so
that
the
toxicity
of
the
mixture
would
increase
if
the
metal
causing
the
antagonism
were
not
present.

135
Because
the
National
Toxics
Rule
WTR)
incorporated
WERs
into
the
aquatic
life
criteria
for
some
metals,
it
might
be
envisioned
that
more
than
one
criterion
could
apply
to
a
metal
at
a
site
if
different
investigators
obtained
different
WERs
for
the
same
metal
at
the
site.
.
.
.
.
.
;
EI1
mzqdictrons
sublect
to
the
NTR.
as
well,
l
.
ens,
EPA
intends
that
Dere
s­
d
be
.
no
­
re
one
w&
szaon
for
a
Dolluta
t
at
a
Dolnt
in
a
bide
Thus
whenever
a
site­
specificncriterion
is
to
be
of
water.
derived
using
a
WER
at
a
site
at
which
more
than
one
discharger
has
permit
limits
for
the
same
metal,
it
is
important
that
all
dischargers
work
together
with
the
appropriate
regulatory
authority
to
develop
a
workplan
that
is
designed
to
derive
a
site­
specific
criterion
that
adequately
protects
the
entire
site.

Method
2
is
ideally
suited
for
taking
into
account
more
than
one
discharger.

Method
1
is
straightforward
if
the
dischargers
are
sufficiently
far
downstream
of
each
other
that
the
stream
can
be
divided
into
a
separate
site
for
each
discharger.
Method
1
can
also
be
fairly
straightforward
if
the
WERs
are
additive,
but
it
will
be
complex
if
the
WERs
are
not
additive.
Deciding
whether
to
use
a
simulated
downstream
water
or
an
actual
downstream
water
can
be
difficult
in
a
flowing­
water
multiple­
discharge
situation.
Use
of
actual
downstream
water
can
be
complicated
by
the
existence
of
multiple
mixing
zones
and
plumes
and
by
the
possibility
of
varying
discharge
schedules;
these
same
problems
exist,
however,
if
effluents
from
two
or
more
discharges
are
used
to
prepare
simulated
downstream
water.
Dealing
with
a
multiple­
discharge
situation
is
much
easier
if
the
WERs
are
additive,
and
use
of
simulated
downstream
water
is
the
best
way
to
determine
whether
the
WERs
are
additive.
Taking
into
account
all
effluents
will
take
into
account
synergism,
antagonism,
and
additivity.
If
one
of
the
discharges
stops
or
is
modified
substantially,
however,
it
will
usually
be
necessary
to
determine
a
new
WER,
except
possibly
if
the
metal
being
discharged
is
refractory.
Situations
concerning
intermittent
and
batch
discharges
need
to
be
handled
on
a
case­
by­
case
basis.

Method
1
is
intended
to
apply
not
only
to
ordinary
rivers
and
streams
but
also
to
streams
that
some
people
might
consider
'
special.,
such
as
streams
whose
design
flows
are
zero
and
streams
that
some
state
and/
or
federal
agencies
might
refer
to
as
'
effluent­
dependent',
'
habitat­
creating',
'
effluent­
dominated',
etc.
(
Due
to
differences
between
agencies,
some
streams
whose
design
flows
are
zero
are
not
considered
'
effluent­
dependent',

136
etc.,
and
some
'
effluent­
dependent'
streams
have
design
flows
that
are
greater
than
zero.)
The
application
of
Method
1
to
these
kinds
of
streams
has
the
following
implications:
1.
If
the
design
flow..
is
zero,
at
least
some
WRRs
ought
ti
be
determined
in
100%
effluent.
2.
If
thunderstorms,
etc.,
occasionally
dilute
the
effluent
substantially,
at
least
one
WER
should
be
determined
in
diluted
effluent
to
assess
whether
dilution
by
rainwater
might
result
in
underprotection
by
decreasing
the
WRR
faster
than
it
decreases
the
concentration
of
the
metal.
This
might
occur,
for
example,
if
rainfall
reduces
hardness,
alkalinity,
and
pH
substantially.
This
might
not
be
a
concern
if
the
WER
demonstrates
a
substantial
margin
of
safety.
3.
If
the
site­
specific
criterion
is
substantially
higher
than
the
national
criterion,
there
should
be
increased
concern
about
the
fate
of
the
metal
that
has
reduced
or
no
toxicity.
Even
if
the
WRR
demonstrates
a
substantial
margin
of
safety
(
e.
g.,
if
the
site­
specific
criterion
is
three
times
the
national
criterion,
but
the
experimentally
determined
WE%
is
111,
it
might
be
desirable
to
study
the
fate
of
the
metal.
4.
If
the
stream
merges
with
another
body
of
water
and
a
site­
specific
criterion
is
desired
for
the
merged
waters,
another
WER
needs
to
be
determined
for
the
mixture
of
the
waters.
5.
Whether
WET
testing
is
reguired
is
not
a
WER
issue,
although
WET
testing
might
be
a
condition
for
determining
and/
or
using
aWER.
6.
A
concern
about
what
species
should
be
present
and/
or
protected
in
a
stream
is
a
beneficial­
use
issue,
not
a
WER
issue,
although
resolution
of
this
issue
might
affect
what
species
should
be
used
if
a
WER
is
determined.
(
If
the
Recalculation
Procedure
is
used,
determining
what
species
should
be
present
and/
or
protected
is
obviously
important.)
7.
Human
health
and
wildlife
criteria
and
other
issues
might
restrict
an
effluent
more
than
an
aquatic
life
criterion.
Although
there
are
no
scientific
reasons
why
'
effluent­
dependent',
etc.,
streams
and
streams
whose
design
flows
are
zero
should
be
subject
to
different
guidance
than
other
streams,
a
regulatory
decision
(
for
example,
see
40
CFR
131)
might
require
or
allow
some
or
all
such
streams
to
be
subject
to
different
guidance.
For
example,
it
might
be
decided
on
the
basis
of
a
use
attainability
analysis
that
one
or
more
constructed
streams
do
not
have
to
comply
with
usual
aquatic
life
criteria
because
it
is
decided
that
the
water
quality
in
such
streams
does
not
need
to
protect
sensitive
aquatic
species.
Such
a
decision
might
eliminate
any
further
concern
for
site­
specific
aquatic
life
criteria
and/
or
for
WET
testing
for
such
streams.
The
water
quality
might
be
unacceptable
for
other
reasons,
however.

In
addition
to
its
use
with
rivers
and
streams,
Method
1
is
also
appropriate
for
determining
cmcWERs
that
are
applicable
to
near­
field
effects
of
discharges
into
large
bodies
of
fresh
or
salt
water,
such
as
an
ocean
or
a
large
lake,
reservoir,
or
estuary:

137
a.

b.
The
near­
field
effects
of
a
pipe
that
extends
far
into
a
large
body
of
fresh
or
salt
water
that
has
a
current,
such
as
an
ocean,
can
probably
best
be
treated
the
same
as
a
single
discharge
into
a
flowing
stream.
For
example,
if
a
mixing
zone
is
defined,
the
concentration
of
effluent
at
the
edge
of
the
mixing
zone
might
be
used
to'define
how
to
prepare
a
simulated
site
water.
A
dye
dispersion
study
(
Kilpatrick
1992)
might
be
useful,
but
a
dilution
model
(
U.
S.
EPA
1993)
is
likely
to
be
a
more
cost­
effective
way
of
obtaining
information
concerning
the
amount
of
dilution
at
the
edge
of
the
mixing
zone.
The
near­
field
effects
of
a
single
discharge
that
is
near
a
shore
of
a
large
body
of
fresh
or
salt
water
can
also
probably
best
be
treated
the
same
as
a
single
discharge
into
a
flowing
stream,
especially
if
there
is
a
definite
plume
and
a
defined
mixing
zone.
The
potential
point
of
impact
of
near­
field
effects
will
often
be
an
embayment,
bayou,
or
estuary
that
is
a
nursery
for
fish
and
invertebrates
and/
or
contains
conmercially
important
shellfish
beds.
Because
of
their
importance,
these
areas
should
receive
special
consideration
in
the
determination
and
use
of
a
WER,
taking
into
account
sources
of
water
and
discharges,
mixing­
patterns,
and
currents
(
and
tides
in
coastal
areas).
The
current
and
flushing
patterns
in
estuaries
can
result
in
increased
pollutant
concentrations
in
confined
embayments
and
at
the
terminal
up­
gradient
portion
of
the
estuary
due
to
poor
tidal
flushing
and
exchange.
Dye
dispersion
studies
(
Kilpatrick
1992)
can
be
used
to
determine
the
spatial
concentration
of
the
effluent
in
the
receiving
water,
but
dilution
models
(
U.
S.
EPA
1993)
might
not
be
sufficiently
accurate
to
be
useful.
Dye
studies
of
discharges
in
near­
shore
tidal
areas
are
especially
complex.
Dye
injection
into
the
discharge
should
occur
over
at
least
one,
and
preferably
two
or
three,
complete
tidal
cycles;
subsequent
dispersion
patterns
should
be
monitored
in
the
ambient
water
on
consecutive
tidal
cycles
using
an
intensive
sampling
regime
over
time,
location,
and
depth.
Information
concerning
dispersion
and
the
comunity
at
risk
can
be
used
to
define
the
appropriate
mixing
zone(
s),
which
might
be
used
to
define
how
to
prepare
simulated
site
water.

Kilpatrick,
F.
A.
19'
92.
Simulation
of
Soluble
Waste
Transport
and
Buildup
in
Surface
Waters
Using
Tracers.
Open­
File
Report
92­
457.
U.
S.
Geological
Survey,
Books
and
Open­
File
Reports,
Box
25425,
Federal
Center,
Denver,
CO
80225.

U.
S.
EPA.
1993.
Dilution
Models
for
Effluent
Discharges.
Second
Edition.
EPA/
600/
R­
93/
139.
National
Technical
Information
Service,
Springfield,
VA.

138
­
ix
0:
Additivity
and
the
Two
Compoxmnta
of
a
WER
Determined
Using
Downstrwm
Water
The
Concent
of
Additivitv
of
WERs
In
theory,
whenever
samples
of
effluent
and
upstream
water
are
taken,
determination
of
a
WRR
in
100
%
effluent
would
quantify
the
effluent
WRR
(
eWER)
and
determination
of
a
WRR
in
100
%
upstream
water
would
quantify
the
upstream
WRR
(
uWRR);
determination
of
WRRs
in
known
mixtures
of
the
two
samples
would
demonstrate
whether
the
eWER
and
the
uWRR
are
additive.
For
example,
if
eWRR
=
40,
uWRR
=
5,
and
the
two
WERs
are
additive,
a
mixture
of
20
%
effluent
and
80
%
upstream
water
would
give
a
WER
of
12,
except
possibly
for
experimental
variation,
because:

2O(
emR)
+
8O(
umR)
t
20(
40)
+
80(
S)
t
800
+
400
100
100
100
dE!
L~
2
100
l
Strict
additivity
of
an
eWRR
and
an
uWER
will
probably
be
rare
because
one
or
both
WERs
will
probably
consist
of
a
portion
that
is
additive
and
a
portion
that
is
not.
The
portions
of
the
ewER
and
uWER
thut
are
due
to
refractory
metal
will
be
strictly
additive,
because
a
change
in
water
quality
will
not
make
the
metal
more
or
less
toxic.
In
contrast,
metal
that
is
nontoxic
because
it
is
complexed
by
a
complexing
agent
such
as
RDTA
will
not
be
strictly
additive
because
the
%
uncomplexed
will
decrease
as
the
solution
is
diluted;
the
amount
of
change
in
the
%
uncomplexed
will
usually
be
small
and
will
depend
on
the
concentration
and
the
binding
constant
of
the
complexing
agent
(
see
Appendix
D).
Whether
the
nonrefractory
portions
of
the
UWER
and
eWRR
are
additive
will
probably
also
depend
on
the
differences
between
the
water
quality
characteristics
of
the
effluent
and
the
upstream
water,
because
these
will
determine
the
water
quality
characteristics
of
the
downstream
water.
If,
for
example,
85
%
of
the
eWRR
and
30
%
of
the
uWER
are
due
to
refractory
metal,
the
WRR
obtained
in
the
mixture
of
20
%
effluent
and
80
%
upstream
water
could
range
from
8
to
12.
The
WRR
of
8
would
be
obtained
if
the
only
portions
of
the
eWER
and
uWRR
that
are
additive
are
those
due
to
refractory
metal,
because:

ZO(
O.
85)
(
eMEN
+
8OtO.
30)
(
WER)
­
20(
0.85)
(
40)
+
SO(
O.
30)
(
5)
=
8
100
100
The
WRR
could
be
as
high
as`
12
depending
on
the
percentages
of
the
other
portions
of
the
WERs
that
are
also
additive.
Even
if
the
eWER
and
uWER
are
not
strictly
additive,
the
concept
of
additivity
of
WRRs
can
be
useful
insofar
as
the
eWER
and
uWER
are
partially
additive,
i.
e.,
insofar
as
a
portion
of
at
least
one
of
the
WERs
is
additive.
In
the
example
given
above,
the
WER
determined
using
downstream
water
that
consisted
of
20
%
effluent
139
and
80
%
upstream
water
would
be
12
if
the
eWER
and
uwER
were
strictly
additive;
the
downstream
WER
would
be
less
than
12
if
the
eWER
and
uWER
were
partially
additive.

The
major
advantage
of
additivity
of
WERs
can
be
dmnstrated
using
the
effluent
and
upstream
water
that
were
used
above.
TO
sir&
ify
this
illustration,
the
acute­
chronic
ratio
will
be
assumed
to
be
large,
and
the
eWEF!
of
40
and
the
uWER
of
5
will
be
assumed
to
be
cccWERs
that
will
be
assumed
to
be
due
to
refractory
metal
and
will
therefore
be
strictly
additive.
In
addition,
the
coap?
lete­
mix
downstream
water
at
design­
flow
conditions
will
be
assumed
to
be
20
8
effluent
and
80
%
upstream
water,
so
that
the
downstream
WER
will
be
12
as
calculated
above
for
strict
additivity.

Because
the
ewER
and
the
USER
are
cccWERs
and
are
strictly
additive,
this
metal
will
cause
neither
acute
nor
chronic
toxicity
in
downstream
water
if
(
a)
the
concentration
of
metal
in
the
effluent
is
less
than
40
times
the
CCC
and
(
b)
the
concentration
of
metal
in
the
upstream
water
is
less
than
5
times
the
CCC.
As
the
effluent
is
diluted
by
mixing
with
upstream
water,
both
the
ewER
and
the
concentration
of
metal
will
be
diluted
simultaneously;
proportional
dilution
of
the
metal
and
the
eWER
will
prevent
the
metal
from
causing
acute
or
chronic
toxicity
at
any
dilution.
When
the
upstream
flow
equals
the
design
flow,
the
WER
in
the
plume
will
decrease
from
40
at
the
end
of
the
pipe
to
12
at
complete
mix
as
the
effluent
is
diluted
by
upstream
water;
because
this
WER
is
due
to
refractory
metal,
neither
fate
processes
nor
changes
in
water
quality
characteristics
will
affect
the
WER.
When
stream
flow
is
higher
or
lower
than
design
flow,
the
complete­
mix
WER
will
be
lower
or
higher,
respectively,
than
12,
but
toxicity
will
not
occur
because
the
concentration
of
metal
will
also
be
lower
or
higher.

If
the
eWER
and
the
uWER
are
strictly
additive
and
if
the
national
CCC
is
1
mg/
L,
the
following
conclusions
are
valid
when
the
concentration
of
the
metal
in
100
%
effluent
is
less
than
40
xng/
L
and
the
concentration
of
the
metal
in
100
%
upstream
water
is
less
than
5
mg/
L:
1.
This
metal
will
not
cause
acute
or
chronic
toxicity
in
the
upstream
water,
in
100
%
effluent,
in
the
plume,
or
in
downstream
water.
2.
There
is
no
need
for
an
acute
or
a
chronic
mixing
zone
where
a
lesser
degree
of
protection
is
provided.
3.
If
no
mixing
zone
exists,
there
is
no
discontinuity
at
the
edge
of
a
mixing
zone
where
the
allowed
concentration
of
metal
decreases
instantaneously.
These
results
also
apply
to
partial
additivity
as
long
as
the
concentration
of
metal
does
not
exceed
that
allowed
by
the
amount
140
of
additivity
that
exists.
It
would
be
more
difficult
to
take
into
account
the
portions
of
the
eWER
and
uWER
that
are
not
additive.

The
concept
of
additivity
becomes
unimportant
when
the
ratios,
concentrations
of
the
metals,
or
WRRs
are
very
different.
For
example,
if
eWER
=
40,
uWER
=
5,
and
they
are
additive,
a
mixture
of
1
%
effluent
and
99
%
upstream
water
would
have
a
WRR
of
5.35.
Given
the
reproducibility
of
toxicity
tests
and
WERs,
it
would
be
extremely
difficult
to
distinguish
a
WER
of
5
from
a
WER
of
5.35.
In
cases
of
extreme
dilution,
rather
than
experimentally
determining
a
WER,
it
is
probably
acceptable
to
use
the
limiting
WER
of
5
or
to
calculate
a
WRR
if
additivity
has
been
demonstrated.

Traditionally
it
has
been
believed
that
it
is
environmentally
conservative
to
use
a
WER
determined
in
upstream
water
(
i.
e.,
the
uWRR)
to
derive
a
site­
specific
criterion
that
applies
downstream
(
i.
e.,
that
applies
to
areas
that
contain
effluent).
This
belief
is
probably
based
on
the
assumption
that
a
larger
WER
would
be
obtained
in
downstream
water
that
contains
effluent,
but
the
belief
could
also
be
based
on
the
assumption
that
the
uWER
is
additive.
It
is
possible
that
in
some
cases
neither
assumption
is
true,
which
means
that
using
a
uWRR
to
derive
a
downstream
site­
specific
criterion
might
result
in
underprotection.
It
seems
likely,
however,
that
WRRs
determined
using
downstream
water
will
usually
be
at
least
as
large
as
the
uWER.

Several
kinds
of
concerns
about
the
use
of
WERs
are
actually
concerns
about
additivity:
1.
Do
WERs
need
to
be
determined
at
higher
flows
in
addition
to
being
determined
at
design
flow?
2,
Do
WERs
need
to
be
determined
when
two
bodies
of
water
mix?
3.
Do
WERs
need
to
be
determined
for
each
additional
effluent
in
a
multiple­
discharge
situation.
In
each
case,
the
best
use
of
resources
might
be
to
test
for
additivity
of
WERs.

Mixinu
Zones
In
the
example
presented
above,
there
would
be
no
need
for
a
regulatory
mixing
zone
with
a
reduced
level
of
protection
if:
1.
The
eWRR
is
always
40
and
the
concentration
of
the
metal
in
100
%
effluent
is
always
less
than
40
mg/
L.
2.
The
uWER
is
always
5
and
the
concentration
of
the
metal
in
100
%
upstream
water
is
always
less
than
5
mg/
L.
3.
The
WERs
are
strictly
additive.
If,
however,
the
concentration
exceeded
40
mg/
L
in
100
%
effluent,
but
there
is
some
assimilative
capacity
in
the
upstream
water,
a
regulatory
mixing
zone
would
be
needed
if
the
discharge
were
to
be
allowed
to
utilize
some
or
all
of
the
assimilative
141
capacity.
The
concept
of
additivity
of
WERs
can
be
used
to
calculate
the
maximum
allowed
concentration
of
the
metal
in
the
effluent
if
the
eWER
and
the
uWER
are
strictly
additive.

If
the
concentration
of
metal
in
the
upstream
water
never
exceeds
0.8
mg/
L,
the
discharger
might
want'to
determine
how
much
above
40
mg/
L
the
concentration
could
be
in
100
%
effluent.
If,
for
example,
the
downstream
water
at
the
edge
of
the
chronic
mixing
zone
under
design­
flow
conditions
consists
of
70
%
effluent
and
30
b
upstream
water,
the
WER
that
would
apply
at
the
edge
of
the
mixing
zone
would
be:

70(
e)
+
3o(
uwim)
­
70(
40)
+
30(
s)
I
2800
+
150
I
100
100
100
2g
.
5
.

Therefore,
the
maximum
concentration
allowed
at
this
point
would
be
29.5
lng/
L.
If
the
concentration
of
the
metal
in
the
upstream
water
was
0.8
m/
L,
the
maximum
concentration
allowed
in
100
%
effluent
would
be
41.8
mg/
L
because:

70(
41.8
a&
L)
l
3OfO.
8
n&
L)
I
2926
m/
q/
L
+
24
I&
100
100
L
=
29.5alg/
L*

Because
the
eWER
is
40,
if
the
concentration
of
the
metal
in
100
0
effluent
is
41.8
mg/
L,
there
would
be
chronic
toxicity
inside
the
chronic
mixing
zone.
If
the
concentration
in
100
%
effluent
is
greater
than
41.8
mg/
L,
there
would
be
chronic
toxicity
past
the
edge
of
the
chronic
mixing
zone.
Thus
even
if
the
eWER
and
the
uWER
are
taken
into
account
and
they
are
assumed
to
be
completely
additive,
a
mixing
zone
is
necessary
if
the
assimilative
capacity
of
the
upstream
water
is
used
to
allow
discharge
of
more
metal.

If
the
complete­
mix
downstream
water
consists
of
20
%
effluent
and
80
8
upstream
water
at
design
flow,
the
complete­
mix
WER
would
be
12
as
calculated
above.
The
complete­
mix
approach
to
determining
and
using
downstream
WERs
would
allow
a
maximum
concentration
of
12
xng/
L
at
the
edge
of
the
chronic
mixing
zone,
whereas
the
alternative
approach
resulted
in
a
maximum
allowed
concentration
of
29.5
mg/
L.
The
complete­
mix
approach
would
allow
a
maximum
concentration
of
16.8
mg/
L
in
the
effluent
because:

70(
16.8
n&
L)
+
3OtO.
8
a&
L)
I
1176
m&
L
+
24
m&
L
­
100
100
12
n&
L.

In
this
example,
the
complete­
mix
approach
limits
the
concentration
of
the
metal
in
the
effluent
to
16.8
mg/
L,
even
though
it
is
known
that
as
long
as
the
concentration
in
100
%
effluent
is
leas
than
40
mg/
L,
chronic
toxicity
will
not
occur
inside
or
outside
the
mixing
zone.
If
the
WER
of
12
is
used
to
derive
a
site­
specific
CCC
of
12
mg/
L
that
is
applied
to
a
site
142
that
starts
at
the
edge
of
the
chronic
mixing
zone
and
extends
all
the
way
across
the
stream,
there
would
be
overprotection
at
the
edge
of
the
chronic
mixing
zone
(
because
the
maximum
allowed
concentration
is
12
mg/
L,
but
a
concentration
of
29.5
mg/
L
will
not
cause
chronic
toxicity),
whereas
there
would
be
underprotection
on
the
other
side
of­
the
stream
(
because
the
maximum
allowed
concentration
is
12
mg/
L,
but
concentrations
above
5
mg/
L
can
cause
chronic
toxicity.)

The
Rxnerimental
Determination
of
Additivitv
Experimental
variation
makes
it
difficult
to
quantify
additivity
without
determining
a
large
number
of
WERs,
but
the
advantages
of
demonstrating
additivity
might
be
sufficient
to
make
it
worth
the
effort.
It
should
be
possible
to
decide
whether
the
eWER
and
uWER
are
strictly
additive
based
on
determination
of
the
eWER
in
100
%
effluent,
determination
of
the
uWER
in
100
%
upstream
water,
and
determination
of
WERs
in
1:
3,
l:
l,
and
3:
l
mixtures
of
the
effluent
and
upstream
water,
i.
e.,
determination
of
WERs
in
100,
75,
50,
25,
and
0
%
effluent.
Validating
models
of
partial
additivity
and/
or
interactions
will
probably
require
determination
of
more
WERs
and
more
sophisticated
data
analysis
(
see,
for
example,
Broderius
1991).

In
some
cases
chemical
measurements
or
manipulations
might
help
demonstrate
that
at
least
some
portion
of
the
eWER
and/
or
the
uWRR
is
additive:
1.
If
the
difference
between
the
dissolved
WER
and
the
total
recoverable
WER
is
explained
by
the
difference
between
the
dissolved
and
total
recoverable
concentrations,
the
difference
is
probably
due
to
particulate
refractory
metal.
2.
If
the
WERs
in
different
samples
of
the
effluent
correlate
with
the
concentration
of
metal
in
the
effluent,
all,
or
nearly
all,
of
the
metal
in
the
effluent
is
probably
nontoxic.
3.
A
WEX
that
remains
constant
as
the
pH
is
lowered
to
6.5
and
raised
to
9.0
is
probably
additive.
The
concentration
of
refractory
metal
is
likely
to
be
low
in
upstream
water
except
during
events
that
increase
TSS
and/
or
WC;
the
concentration
of
refractory
metal
is
more
likely
to
be
substantial
in
effluents.
Chemical
measurements
might
help
identify
the
percentages
of
the
eWRR
and
the
uWER
that
are
due
to
refractory
metal,
but
again
experimental
variation
will
limit
the
usefulness
of
chemical
measurements
when
concentrations
are
low.

The
distinction
between
the
two
components
of
a
WER
determined
using
downstream
water
has
the
following
implications:
1.
The
magnitude
of
a
WER
determined
using
downstream
water
will
usually
depend
on
the
percent
effluent
in
the
sample.

143
2.

3.

4.

5.

6.
Insofar
as
the
eWRR
and
uWRR
are
additive,.
the
magnitude
of
a
downstream
WER
can
be
calculated
from
the
eWER,
the
UWER,
and
the
ratio
of
effluent
and
upstream
water­
in
the
downstream
water.
The
derivation
and
implementation
of
site­
specific
criteria
should
ensure
that
each
component
is
applied
only
where
it
occurs.
a.
Underprotection
will
occur
if,
for
example,
any
portion
of
the
eWER
is
applied
to
an
area
of
a
stream
where
the
effluent
does
not
occur.
b.
Overprotection
will
occur
if,
for
example,
an
unnecessarily
small
portion
of
the
eWER
is
applied
to
an
area
of
a
stream
where
the
effluent
occurs.
Even
though
the
concentration
of
metal
might
be
higher
than
a
criterion
in
both
a
regulatory
mixing
zone
and
a
plume,
a
reduced
level
of
protection.
is
allowed
in
a
mixing
zone,
whereas
a
reduced
level
of
protection
is
not
allowed
in
the
portion
of
a
plume
that
is
not
inside
a
mixing
zone.
Regulatory
mixing
zones
are
necessary
if,
and
only
if,
a
discharger
wants
to
make
use
of
the
assimilative
capacity
of
the
upstream
water.
It
might
be
cost­
effective
to
quantify
the
eWER
and
uWER,
determine
the
extent
of
additivity,
study
variability
over
time,
and
then
decide
how
to
regulate
the
metal
in
the
effluent.

Peference
Broderius,
S.
J.
1991.
Modeling
the
Joint
Toxicity
of
Xenobiotics
to
Aquatic
Organisms:
Basic
Concepts
and
Approaches.
In:
Aquatic
Toxicology
and
Risk
Assessment:
Fourteenth
Volume.
(
M.
A.
Mayes
and
M.
G.
Barron,
eds.)
ASTM
STP
1124.
American
Society
for
Testing
and
Materials,
Philadelphia,
PA.
pp.
107­
127.

144
Appendix
EI:
Spmcirl
Conddorationm
Concerning
the
Deteminatiozh
of
WHIl
with
Saltwater
Spociem
1.
The
test
organisms
should
be
compatible
with
the
salinity
of
the
site
water,
and
the
salinity
of
the
laboratory
dilution
water
should
match
that
of
the
site
water.
Low­
salinity
stenohaline
organisms
should
not
be
tested
in
high­
salinity
water,
whereas
high­
salinity
stenohaline
organisms
should
not
be
tested
in
low­
salinity
water;
it
is
not
known,
however,
whether
an
incompatibility
will
affect
the
WER.
If
the
community
to
be
protected
principally
consists
of
euryhaline
species,
the
primary
and
secondary
toxicity
tests
should
use
the
euryhaline
species
suggested
in
Appendix
I
(
or
taxonomically
related
species)
whenever
possible,
although
the
range
of
tolerance
of
the
organisms
should
be
checked.
a.
When
Method
1
is
used
to
determine
cmcWRRs
at
saltwater
sites,
the
selection
of
test
organisms
is
complicated
by
the
fact
that
most
effluents
are
freshwater
and
they
are
discharged
into
salt
waters
having
a
wide
range
of
salinities.
Some
state
water
quality
standards
require
a
permittee
to
meet
an
LCSO
or
other
toxicity
limit
at
the
end
of
the
pipe
using
a
freshwater
species.
However,
the
intent
of
the
site­
specific
and
national
water
quality
criteria
program
is
to
protect
the
communities
that
are
at
risk.
Therefore,
freshwater
species
should
not
be
used
when
WRRs
are
determined
for
saltwater
sites
unless
such
freshwater
species
(
or
closely
related
species)
are
in
the
community
at
risk.
The
addition
of
a
small
amount
of
brine
and
the
use
of
salt­
tolerant
freshwater
species
is
inappropriate
for
the
same
reason.
The
addition
of
a
large
amount
of
brine
and
the
use
of
saltwater
species
that
require
high
salinity
should
also
be
avoided
when
salinity
is
likely
to
affect
the
toxicity
of
the
metal.
Salinities
that
are
acceptable
for
testing
euryhaline
species
can
be
produced
by
dilution
of
effluent
with
sea
water
and/
or
addition
of
a
commercial
sea
salt
or
a
brine
that
is
prepared
by
evaporating
site
water;
small
increases
in
salinity
are
acceptable
because
the
effluent
will
be
diluted
with
salt
water
wherever
the
cozununities
at
risk
are
exposed
in
the
real
world.
Only
as
a
last
resort
should
freshwater
species
that
tolerate
low
levels
of
salinity
and
are
sensitive
to
metals,
such
as
Danhnia
mauna
and
Hvalella
azteca,
be
used.
b.
When
Method
2
is
used
to
determine
cccWERs
at
saltwater
sites:
1)
If
the
site
water
is
low­
salinity
but
all
the
sensitive
test
organisms
are
high­
salinity
btenohaline
organisms,
a
commercial
sea
salt
or
a
brine
that
is
prepared
by
evaporating
site
water
may
be
added
in
order
to
increase
the
salinity
to
the
minimum
level
that
is
acceptable
to
the
test
organisms;
it
should
be
determined
whether
the
145
salt
or
brine
reduces
the
toxicity
of
the
metal
and
thereby
increases
the
WER.
2)
If
the
site
water
is
high­
salinity;
selecting
test
organisms
should
not
be
difficult
because
many
of
the
sensitive
test
organisms
are
compatible
with
high­
salinity
water.

2.
It
is
especially
important
to
consider
the
availability
of
test
organisms
when
saltwater
species
are
to
be
used,
because
many
of
the
cwnly
used
saltwater
species
are
not
cultured
and
are
only
available
seasonally.

3.
Many
standard
published
methodologies
for
tests
with
saltwater
species
reconzaen
d
filtration
of
dilution
water,
effluent,
and/
or
teat
solutions
through
a
37­
m
sieve
or
screen
to
remove
predators.
Site
water
should
be
filtered
only
if
predators
are
observed
in
the
sample
of
the
water
because
filtration
might
affect
toxicity.
Although
recommended
in
some
test
methodologies,
ultraviolet
treatment
is
often
not
needed
and
generally
should
be
avoided.

4.
If
a
natural
salt
water
is
to
be
used
as
the
laboratory
dilution
water,
the
samples
should
probably
be
collected
at
slack
high
tide
(&
2
hours).
Unless
there
is
stratification,
samples
should
probably
be
taken
at
mid­
depth;
however,
if
a
water
quality
characteristic,
such
as
salinity
or
TSS,
is
important,
the
vertical
and
horizontal
definition
of
the
point
of
sampling
might
be
important.
A
conductivity
meter,
aalinometer,
and/
or
transmissometer
might
be
useful
for
determining
where
and
at
what
depth
to
collect
the
laboratory
dilution
water;
any
measurement
of
turbidity
will
probably
correlate
with
TSS.

5.
The
salinity
of
the
laboratory
dilution
water
should
be
within
f
10
percent
or
2
mg/
L
(
whichever
is
higher)
of
that
of
the
site
water.

146
appendix
It
Su~
3u;
ftod
Toxicity
Tostn
for
Dotozminhg
WBIU
for
Selecting
primary
and'
secondary
toxicity
tests
for
determining
WERs
for
metals
should
take
into
account
the
following:
1.

2.

3,

4.
5.

6.

7.

8.

9.
WERs
determined
with
more
sensitive
tests
are
likely
to
be
larger
than
WERs
determined
with
less
sensitive
tests
(
see
Appendix
D).
Criteria
are
derived
to
protect
sensitive
species
and
so
WRRs
should
be
derived
to
be
appropriate
for
sensitive
species.
The
appropriate
regulatory
authority
will
probably
accept
WRRs
derived
with
less
sensitive
tests
because
such
WERs
are
likely
to
provide
at
least
as
much
protection
as
WRRs
determined
with
more
sensitive
tests.
The
species
used
in
the
primary
and
secondary
tests
must
be
in
different
orders
and
should
include
a
vertebrate
and
an
invertebrate.
The
test
organism
(
i.
e.,
species
and
life
stage)
should
be
readily
available
throughout
the
testing
period.
The
chances
of
the
test
being
successful
should
be
high.
The
relative
sensitivities
of
test
organisms
vary
substantially
from
metal
to
metal.
The
sensitivity
of
a
species
to
a
metal
usually
depends
on
both
the
life
stage
and
kind
of
test
used.
Water
guality
characteristics
might
affect
chronic
toxicity
differently
than
they
affect
acute
toxicity
(
Spehar
and
Carlson
1984;
Chapman,
unpublished;
Voyer
and
McGovern
1991).
The
endpoint
of
the
primary
test
in
laboratory
dilution
water
should
be
as
close
as
possible
(
but
must
not
be
below)
the
CMC
or
CCC
to
which
the
WER
is
to
be
applied;
the
endpoint
of
the
secondary
test
should
be
as
close
as
possible
(
and
should
not
be
below)
the
CMC
or
CCC.
Designation
of
tests
as
acute
and
chronic
has
no
bearing
on
whether
they
may
be
used
to
determine
a
cmcwER
or
a
cccWRR.
The
suggested
toxicity
tests
should
be
considered,
but
the
actual
selection
should
depend
on
the
specific
circumstances
that
apply
to
a
particular
WER
determination.

Regardless
of
whether
test
solutions
are
renewed
when
tests
are
conducted
for
other
purposes,
if
the
concentrations
of
dissolved
metal
and
dissolved
oxygen
remain
acceptable
when
determining
WRRs,
tests
whose
duration
is
not
longer
than
48
hours
may
be
static
tests,
whereas
tests
whose
duration
is
longer
than
48
hours
must
be
renewal
tests.
If
the
concentration
of
dissolved
metal
and/
or
the
concentration
of
dissolved
oxygen
does
not
remain
acceptable,
the
test
solutions
must
be
renewed
every
24
hours.
If
one
test
in
a
pair
of
side­
by­
side
tests
is
a
renewal
test,
both
of
the
tests
must
be
renewed
on
the
same
schedule.

Appendix
H
should
be
read
if
WRRs
are
to
be
determined
with
saltwater
species.

147
Suggested
Tests
1
for
Determining
cmcWERs
and
cccWERs2
(
Concentrations
are
to
be
measured
in
all
tests.)

JlY&!&
d
CRICWERS'
CCCWERS'
Netal
Aluminum
Arsenic(
II1)

Cadmium
Chrom(
II1)

chmm(
vx)

Capx=
r
Lead
Mercury
Nickel
Selenium
Silver
Zinc
X
CDC
Fw
DA
X
FMC
Em
F­
MC
X
CDC
GM
FM
AR
X
X
Y
Y
FMC
BMC
Y
X
FMC
BMC
FMC
BMC
Fw
SW
DA
BM
GM
CR
CDC
MYC
SLs
or
FM
CR
Fw
SW
DA
MY
CDC
MYC
Fw
GM
SL
or
DA
FMC
Fw
SW
DA
MY
GM
NE
CDC
MYC
Tw
SW
DA
EM
FM
or
GM
AR
CDC
BMC
Fw
SW
GM
MYC
CDC
MYC
DA
BM
Fw
SW
DA
MY
GM
EM
Y
Y
Fw
SW
DA
MY
FX
EM
CDC
MYC
Fw
SW
Y
CR
Y
MYC
Y
MYC
Fw
SW
DA
BM
FMC
CR
CDC
MYC
Fw
SW
DA
BM
FM
MY
CDC
MYC
The
description
of
a
test
specifies
not
only
the
test
species
and
the
duration
of
the
test
but
also
the
life
stage
of
the
species
and
the
adverse
effect(
s)
on
which
the
endpoint
is
to
be
based.

Some
tests
that
are
sensitive
and
are
used
in
criteria
documents
are
not
suggested
here
because
the
chances
of
the
test
organisms
being
available
and
the
test
being
successful
might
be
low.
Such
tests
may
be
used
if
desired.

148
FW=
Fresh
Water;
SW
=
Salt
Water.

Two­
letter
codes
are
used
for
acute
tests,
whereas
codes
for
chronic
tests
contain
three
letters
and
end
in
'
C'.
One­
letter
codes
are
used
for
cosunents.

In
acute
tests
on
cadmium
with
salmonids,
substantial
numbers
of
fish
usually
die
after
72
hours.
Also,
the
fish
are
sensitive
to
disturbance,
and
it
is
sometimes
difficult
to
determine
whether
a
fish
is
dead
or
immobilized.

AR.
A
48­
hr
EC50
based
on
mortality
and
abnormal
development
from
a
static
test
with
embryos
and
larvae
of
sea
urchins
of
a
species
in
the
genus
Arbacia
(
ASTM
1993a)
or
of
the
species
gtronwlocentrotus
ournuratus
(
Chapman
1992).

EM.
A
48­
hr
EC50
based
on
mortality
and
abnormal
larval
development
from
a
static
test
with
embryos
and
larvae
of
a
species
in
one
of
four
genera
(
Crassostrea,
Mulinia,
Mvtilus,
flercenaria)
of
bivalve
molluscs
(
ASTM
1993b).

CR.
A
48­
hr
EC50
(
or
LC50
if
there
is
no
immobilization)
from
a
static
test
with
Acartia
or
larvae
of
a
saltwater
crustacean;
if
molting
does
not
occur
within
the
first
48
hours,
renew
at
48
hours
and
continue
the
test
to
96
hours
(
ASTM
1993a).

DA.
A
48­
hr
EC50
(
or
LC50
if
there
is
no
immobilization)
from
a
static
test
with
a
species
in
one
of
three
genera
(
U.
S.
EPA?;
93
(
CeriodaD
'
g
DaDhnia
SimoceDhalus)
in
the
family
Daphnidae
a;
ASTM
i993a).

FM.
A
48­
hr
LC50
from
a
static
test
at
25OC
with
fathead
minnow
.
gro
elas)
larvae
that
are
1
to
24
hours
old
(
ASTM
EPAm1993a)
The
embryos
must
be
hatched
in
the
laboratory
dilution
water,
except
that
organisms
to
be
used
in
the
site
water
may
be
hatched
in
the
site
water.
The
larvae
mumt
not
be
fed
before
or
during
the
test
and
at
least
90
percent
muat
survive
in
laboratory
dilution
water
for
at
least
six
days
after
hatch.
Note:
The
following
48­
hr
LC5Os
were
obtained
at
a
hardness
of
50
mg/
L
with
fathead
minnow
larvae
that
were
1
to
24
hours
old.
The
metal
was
measured
using
the
total
recoverable
procedure
(
Peltier
1993)
:

CS
LC50
(
us/
L)
13.87
Copper
6.33
Zinc
100.95
149
FX.
A
96­
hr
LC50
from
a
renewal
test
(
renew
at
48
hours)
at
25OC
with
fathead
minnow
(
PimeDhaleS
nro
elas)
larvae
that
are
1
to
24
hours
old
(
ASTM
1993a;
U.
S.
EFA
1993a).
The
embryos
atit
be
hatched..
in
the
laboratory
dilution
water,
except
that
organisms
to
be
used
in
the
site
water
may
be
hatched
in
the
site
water.
The
larvae
m;
ut
rrdt
be
fed
before
or
during
the
test
and
at
least
90
percent
ru8t
survive
in
laboratory
dilution
water
for
at
least
six
days
after
hatch.
Note:
A
96­
hr
LC50
of
188.14
w/
L
was
obtained
at
a
hardness
of
50
mg/
L
in
a
test
on
nickel
with
fathead
minnow
larvae
that
were
1
to
24
hours
old.
The
metal
was
measured
using
the
total
recoverable
procedure
(
Peltier
1993).
A
96­
hr
LC50
is
used
for
nickel
because
substantial
mortality
occurred
after
48
hours
in
the
test
on
nickel,
but
not
in
the
tests
on
cadmium,
copper,
and
zinc.

GM.
A
96­
hr
EC50
(
or
LC50
if
there
is
no
izzaobilization)
from
a
renewal
test
(
renew
at
48
hours)
with
a
species
in
the
genus
Gamnaw
(
ASTM
1993a).

MY.
A
96­
hr
EC50
(
or
LC50
if
there
is
no
inznobilization)
from
a
renewal
test
(
renew
at
48
hours)
with
a
species
in
one
of
two
.
.
genera
(
JWsidoDsis,
Holmeg&
mvsis
[
nee
Acanthomvsisl)
in
the
family
Mysidae
(
U.
S.
EPA
1993a;
ASTM
1993a).
Feeding
is
required
during
all
acute
and
chronic
tests
with
mysids;
for
determining
WEF&,
mysids
should
be
fed
four
hours
before
the
renewal
at
48
hours
and
minimally
on
the
non­
renewal
days.

NE.
A
96­
hr
LCSO
from
a
renewal
test
(
renew
at
48
hours)
using
juvenile
or
adult
polychaetes
in
the
genus
Nereidae
(
ASTM
1993a).

SL.
A
96­
hr
EC50
(
or
LC50
if
there
is
no
immobilization)
from
a
renewal
test
(
renew
at
48
hours)
with
a
species
in
one
of
two
salmo)
in
the
family
Salmonidae
(
ASTM
CHRONIC
m
BMC.
A
7­
day
IC25
from
a
survival
and
development
renewal
test
(
renew
every
48
hours)
with
a
species
of
bivalve
mollusc,
such
as
a
species
in
the
genus
Mulinig.
One
such
test
has
been
described
by
Burgess
et
al.
1992.
[
Note:
When
determining
WEBS,
sediment
must
not
be
in
the
test.
chamber.]
[
Note:
This
test
has
not
been
widely
used.
1
CDC.
A
7­&
y
IC25
based
on
reduction
in
survival
and/
or
reproduction
in
a
renewal
test
with
a
species
in
the
genus
in
the
family
Daphnidae
(
U.
S.
EPA
1993b).
The
150
test
solutions
mum+
be
renewed
every
48
hours.
(
A
21­
day
life­
cycle
test
with
DaDhI'&
manna
is
also
acceptable.,)

FMC.
A
7­
day
IC25
from
a
survival
and
growth
renewal
test
(
renew
every
48
hours)
with
larvae
(
s
48­
hr
old)
of
the
fathead
minnow
(
PimeDhales
promela%)
(
U;
S.
EPA
1993b).
When
determining
WERs,
the
fish
mumt
be
fed
four
hours
before
each
renewal
and
minimally
during
the
non­
renewal
days.

MYC.
A
7­
day
IC25
based
on
reduction
in
survival,
growth,
and/
or
reproduction
in
a
renewal
test
with
a
species
in
one
of
two
genera
(
Mvs'd
DS'S
family
Mysiiag
(
i.
k.
Holmesimvsis
[
nee
Acanthomvsis))
in
the
EPA
1993c).
Mysids
mwt
be
fed
during
all
acute
and
chronic
tests;
when
determining
WERs,
they
must
be
fed
four
hours
before
each
renewal.
The
test
solutions
must
be
renewed
every
24
hours.

NEC.
A
20­
day
IC25
from
a
survival
and
growth
renewal
test
(
renew
every
48
hours)
with
a
species
in
the
genus
Neanthes
(
Johns
et
al.
1991).
[
Note:
When
determining
WERs,
sediment
mwt
not
be
in
the
test
chamber.]
[
Note:
This
test
has
not
been
widely
used.]

COMMENTS
X.
Another
sensitive
test
cannot
be
identified
at
this
time,
and
so
other
tests
used
in
the
criteria
document
should
be
considered.

Y.
Because
neither
the
CCCs
for
mercury
nor
the
freshwater
criterion
for
selenium
is
based
on
laboratory
data
concerning
toxicity
to
aquatic
life,
they
cannot
be
adjusted
using
a
WER.

REFERENCES
Asm.
1993a.
Guide
for
Conducting
Acute
Toxicity
Tests
with
Fishes,
Macroinvertebrates,
and
Amphibians.
Standard
E729.
American
Society
for
Testing
and
Materials,
Philadelphia,
PA.

ASTM.
1993b.
Guide
for
Conducting
Static
Acute
Toxicity
Tests
Starting
with
Embryos
of
Four
Species
of
Saltwater
Bivalve
Molluscs.
Standard
E724.
American
Society
for
Testing
and
Materials,
Philadelphia,
PA.

Burgess,
R.,
G.
Morrison,
and
S.
Rego.
1992.
Standard
Operating
Procedure
for
7­
day
Static
Sublethal
Toxicity
Tests
for
Mulinia
lateralis.
U.
S.
EPA,
Environmental
Research
Laboratory,
Narragansett,
RI.

151
Chapman,
G.
A.
1992.
Sea
Urchin
(
Stronwlocent
otus
Fertilization
Test
Method.
U.
S.
EPA,
Newport,
f;
R.
­
1
Johns,
D.
M.,
R.
A.
Pastorok,
and
T.
C.
Ginn.
1991.
A
Sublethal
Sediment
Toxicity
Test
using
Juvenile
EJeantheg
sp.
(
Polychaeta:
Nereidae)
.
In:
Aquatic­
Toxicology
and
Risk
Assessment:
Fourteenth
Volume.
ASTM
STP
1124.
(
M.
A.
Mayes
and
M.
G.
Barron,
eds.)
American
Society
for
Testing
and
Materials,
Philadelphia,
PA.
pp.
280­
293.

Peltier,
W.
H.
1993.
Memorandum.
to
C.
E.
Stephan.
October
19.

Spehar,
R.
L.,
and
A.
R.
Carlson.
1984.
Derivation
of
Site­
Specific
Water
Quality
Criteria
for
Cadmium
and
the
St.
Louis
River
Basin,
Duluth,
Minnesota.
Environ.
Toxicol.
Chem.
3:
651­
665.

U.
S.
EPA.
1993a.
Methods
for
Measuring
the
Acute
Toxicity
of
Effluents
and
Receiving
Waters
to
Freshwater
and
Marine
Organisms.
Fourth
Edition.
EPA/
600/
4­
90/
027F.
National
Technical
Information
Service,
Springfield,
VA.

U.
S.
EPA.
1993b.
Short­
term
Methods
for
Estimating
the
Chronic
Toxicity
of
Effluents
and
Receiving
Waters
to
Freshwater
Organisms.
Third
Edition.
EPA/
600/
4­
91/
002.
National
Technical
Information
Service,
Springfield,
VA.

U.
S.
EPA.
1993c.
Short­
term
Methods
for
Estimating
the
Chronic
Toxicity
of
Effluents
and
Receiving
Waters
to
Marine
and
Estuarine
Organisms.
Second
Edition.
EPA/
600/
4­
91/
003.
National
Technical
Information
Service,
Springfield,
VA.

Voyer,
R.
A.,
and
D.
G.
McGovern.
1991.
Influence
of
Constant
and
Fluctuating
Salinity
on
Responses
of
WsidoDsis
bahia
Exposed
to
Cadmium
in
a
Life­
Cycle
Test.
Aquatic
Toxicol.
19:
215­
230.

152
nix
3:
Rmc
mod
Saltm
of
Hotala
The
following
salts
are
recommended
for
use
when
determining
a
WEH
for
the
metal
listed.
If
available,
a
salt
that
meets
American
Chemical
Society
(
ACS)
specifications
for
reagent­
grade
should
be
used.

Aluminum
*
Aluminum
chloride
(
j­
hydrate:
AlCl,=
6H2O
Aluminum
sulfate
18­
hydrate:
A12(
S0,)
1*
18H,
0
Aluminum
potassium
sulfate
la­
hydrate:
AlK(
S0,),=
12HI0
ite:
NaAs02
enate
'
I­
hydrate,
dibasic:
Na2HAs0,=
7H20
­
Cadmium
chloride
2.5~
hydrate:
CdC12=
2.5H20
Cadmium
sulfate
hydrate:
3CdS0,=
8H20
.
Sk0
um(
I
*
Chrtkic
%
oride
6­
hydrate
(
Chromium
chloride):
CrCl,=
6HiO
*
Chromic
nitrate
g­
hydrate
(
Chromium
nitrate):
Cr(
NO,),*
9&
0
Chromium
potassium
sulfate
12­
hydrate:
CrK(
S0,12*
12H20
Potassium
chromate:
K2­
04
Potassium
dichromate:
K,
Cr,
O,
*
Sodium
chromate
I­
hydrate:
­
Na2Cr0,­
4H,
O
Sodium
dichromate
2­
hydrate:
Na2Cr20,=
2H20
izQRRss
*
Cupric
Cupric
Cupric
chloride
a­
hydrate
(
Copper
chloride):
CuC12*
2Ha0
nitrate
2.5­
hydrate
(
Copper
nitrate):
CU(
NO,)~~~.~
H~
O
sulfate
5­
hydrate
(
Copper
sulfate):
CuS0,*
5H20
*
Lead
chloride:
PbC12
Lead
nitrate:
Pb(
NOB)
2
Pe
CUT
Mrrcuric
chloride:
HgC12
Mercuric
nitrate
monohydrate:
Hg(
N03)
2*
H20
Mercuric
sulfate:
HgSO,

153
Nickel
l
Nickelous
chloride
6­
hydrate
(
Nickel
chloride)
:
NiC12=
6H20
l
Nickelous
nitrate
6­
krydrate
(
Nickel
nitrate):
Ni(
NO,),*
6H,
O
Nickelous
sulfate
6­
hydrate
(
Nickel
sulfate):
NiS0,=
6H20.

*
Sodium
selenite
5­
hydrate:
Na,
SeO,­
5H,
O
*
Sodium
selenate
lo­
hydrate:
Na2Se0,=
10H20
Silver
Silver
nitrate:
AgNO,
(
Even
if
acidified,
standards
and
samples
containing
silver
rwt
be
in
amber
containers.)

Zinc
Zinc
chloride:
ZnClr
*
Zinc
nitrate
6­
hydrate:
Zn(
N0,12*
6H,
0
Zinc
sulfate
7­
hydrate:
ZnS0,=
7H20
*
Note:
ACS
reagent­
grade
specifications
might
not
be
available
for
this
salt.

No
salt
should
be
used
until
information
concerning
the
safety
and
handling
of
that
salt
has
been
read.

154
interpreting
the
data
and
determining
the
degree
to
which
the
data
correspond
to
the
assumption(
s).

3.
The
details
of
each
sampling
design
should
be
formulated
with
the
aid
of
people
who
understand
the
site
and
people
who
have
a
working
knowledge
of
WERs.
Because
of
the
complexity
of
designing
a
WER
study
for
large
sites,
the
design
team
should
utilize
the
combined
expertise
and
experience
of
individuals
from
the
appropriate
EPA
Region,
states,
municipalities,
dischargers,
environmental
groups,
and
others
who
can
constructively
contribute
to
the
design
of­
the
study.
Building
a
team
of
cooperating
aquatic
toxicologists,
aquatic
chemists,
limnologists,
oceanographers,
water
quality
modelers,
statisticians,
individuals
from
other
key
disciplines,
as
well
as
regulators
and
those
regulated,
who
have
knowledge
of
the
site
and
the
site­
specific
procedures,
is
central
to
success
of
the
derivation
of
a
WER
for
a
large
site.
Rather
than
submitting
the
workplan
to
the
appropriate
regulatory
authority
(
and
possibly
the
Water
Management
Division
of
the
EPA
Regional
Office)
for
comment
at
the
end,
they
should
be
members
of
the
team
from
the
beginning.

4.
Data
from
one
sampling
event
should
always
be
analyzed
prior
to
the
next
sampling
event
with
the
goal
of
improving
the
sampling
design
as
the
study
progresses.
For
example,
if
the
toxicity
of
the
metal
in
surface
water
samples
is
related
to
the
concentration
of
TSS,
a
water
quality
characteristic
such
as
turbidity
might
be
measured
at
the
time
of
collection
of
water
samples
and
used
in
the
selection
of
the
concentrations
to
be
used
in
the
WER
toxicity
tests
in
site
water.
At
a
minimum,
the
team
that
interprets
the
results
of
one
sampling
event
and
plans
the
next
should
include
an
aquatic
toxicologist,
a
metals
chemist,
a
statistician,
and
a
modeler
or
other
user
of
the
data.

5.
The
final
interpretation
of
the
data
and
the
derivation
of
the
FWER(
s)
should
be
performed
by
a
team.
Sufficient
data
are
likely
to
be
available
to
allow
a
quantitative
estimate
of
experimental
variation,
differences
between
species,
and
seasonal
differences.
It
will
be
necessary
to
decide
whether
one
site­
specific
criterion
can
be
applied
to
the
whole
area
or
whether
separate
site­
specific
criteria
need
to
be
derived
for
two
or
more
sites.
The
interpretation
of
the
data
might
produce
two
or
more
alternatives
that
the
appropriate
regulatory
authority
could
subject
to
a
cost­
benefit
analysis.

Other
aspects
of
the
determination
of
a
WER
for
a
large
site
are
likely
to
be
the
same
as
described
for
Method
1.
For
example:
a.
WERs
should
be
determined
using
two
or
more
sensitive
species;
the
suggestions
given
in
Appendix
I
should
be
considered
when
selecting
the
tests
and
species
to
be
used.

69
b.
Chemical
analyses
of
site
water,
laboratory
dilution
water,
and
test
solutions
should
follow
the
requirements
for
the
specific
test
used
and
those
given
in
this
document.
c.
If
tests
in
many
surface
water
samples
are
compared
to.
one
test
in
a
laboratory
dilution
water,
it
is
very
important
that
that
one
test
be
acceptable.
Use
of
(
1)
rangefinding
tests,
(
2)
additional
treatments
beyond
the
standard
five
concentrations
plus
controls,
and
(
3)
dilutions
that
are
functions
of
the
known
concentration­
effect
relationships
obtained
with
the
toxicity
test
and
metal
of
concern
will
help
ensure
that
the
desired
endpoints
and
WEF&
can
be
calculated.
d.
Measurements
of
the.
concentrations
of
both
total
recoverable
and
dissolved
metal
should
be
targeted
to
the
test
concentrations
whose
data
will
be
used
in
the
calculation
of
the
endpoints.
e.
Samples
of
site
water
and/
or
effluent
should
be
collected,
handled,
and
transported
so
that
the
tests
can
begin
as
soon
as
is
feasible.
f.
If
the
large
site
is
a
saltwater
site,
the
considerations
presented
in
Appendix
H
ought
to
be
given
attention.

70
Figure
2:
Calculating
an
Mjuotd
Geometric
Mean
Where
n
=
the
number
of
experimentally
determined
WERs
in
a
set,
the
'
adjusted
geometric
mean.
of
the
set
is
calculated
as
follows:

a.
Take
the
logarithm
of
each
of
the
WERs.
The
logarithms
can
be
to
any
base,
but
natural
logarithms
(
base
e)
are
preferred
for
reporting
purposes.
b.
Calculate
x'
=
the
arithmetic
mean
of
the
logarithms.
c.
Calculate
8
=
the
sample
standard
deviation
of
the
logarithms:

d.
Calculate
SE
=
the
standard
error
of
the
arithmetic
mean:
SE
­
s/
fi
.
e.
Calculate
A
I??­
(&,,
I
(
SE),
where
f+,,
is
the
value
of
Student's
t
statistic
for
a
one­
sided
probability
of
0.70
with
n
­
1
degrees
of
freedom.
The
values
of
to.,
for
some
common
degrees
of
freedom
(
df)
are:

df
to.
1
0.727
0.617
3
0.584
4
0.569
5
0.559
6
0.553
7
0.549
8
0.546
9
0.543
10
0.542
11
0.540
12
0.539
The
values
of
to,,
for
more
degrees
of
freedom
are
available,
for
example,
on
page
T­
5
of
Natrella
(
1966).
f.
Take
the
antilogarithm
of
A.

This
adjustment
of
the
geometric
mean
accounts
for
the
fact
that
the
means
of
fifty
percent
of
the
sets
of
WERs
are
expected
to
be
higher
than
the
actual
mean;
using
the
one­
sided
value
of
t
for
0.70
reduces
the
percentage
to
thirty.

71
Figtar
3:
An
­
10
Dwivation
of
a
FWER
tis
example
assumes.
that
cccWEF&
were
determined
monthly.
using
simulated
downstream
water
that
was
prepared
by
mixing
upstream
water
with
effluent
at
the
ratio
that
existed
when
the
samples
were
obtained.
Also,
the
flow
of
the
effluent
is
always
10
cfs,
and
the
design
flow
of
the
upstream
water
is
40
cfs.
(
Therefore,
the
downstream
flow
at
design­
flow
conditions
is
50
cfs.)
The
concentration
of
metal
in
upstream
water
at
design
flow
is
0.4
ug/
L,
and
the
CCC
is
2
ug/
L.
Each
FWER
is
derived
from
the
WRS
and
hWERs
that
are
available
through
that
month.

Month
March
April
­
Y
June
July
Aug.
Sept.
Oct.
Nov.
Dec.
Jan.
Feb.
10
10
10
ii
10
10
10
10
10
10
10
850
289
300
430
120
05
t:
150
110
180
244
0.8
;­;:
826.4
82.80
0.6
341.5
34.31
0.6
5:
aE
341.6
34.32
0.6
5.7=
475.8
47.74
0.4
7.0=
177.2
17.88
0.4
10.5'
196.1
19.77
0.4
12.0'
118.4
12.00
0.4
11.0'
119.2
12.08
0.4
7.5=
234.0
23.56
0.4
c
z­;
c
79.6
8.12
0.6
6:
lc
251.4
25.30
0.6
295.2
29.68
hWEX
Neither
Type
1
nor
Type
2;
the
downstream
flow
(
i.
e.,
of
the
eFLOW
and
the
UFLOW)
is
>
500
cfs.
the
sum
The
total
number
of
available
Type
1
and
Type
2
WERs
is
less
than
3.
A
Type
2
WER;
the
downstream
flow
is
between
100
and
500
cfs.
No
Type
1
WER
is
available;
the
FWER
is
the
lower
of
the
lowest
Type
2
WER
and
the
lowest
hWER.
A
Type
1
WER;
the
downstream
flow
is
between
50
and
iO0
cfs.
One
Type
1
WEFI
is
available;
the
FWER
is
the
geometric
mean
of
all
Type
1
and
Type
2
WERs.
Two
or
more
Type
1
WERs
are
available
and
the
range
is
less
than
a
factor
of
5;
the
FWER
is
the
adjusted
geometric
mean
(
see
Figure
2)
of
the
Type
1
WERs,
because
all
the
hWERs
are
higher.
TWo
or
more
Type
1
WERs
are
available
and
the
range
is
not
greater
than
a
factor
of
5;
the
FWEFI
is
the
lowest
hWER
because
the
lowest
hWER
is
lower
than
the
adjusted
geometric
mean
of
the
Type
1
WEF&.
l.
ob
l.
ob
l..
Ob
5.7d
5.7d
6.80'
10.69"
10.880
10.880
8.12h
8.12"
8.12'

72
Figure
4:
Roduciag
the
Impact
of
Bacpertital
Variation
When
the
FWER
is
the
lowest
of,
for
example,
three
WERs,
the
impact
of
experimental
variation
can
be
reduced
by
conducting
additional
primary
tests.
Xf
the
endpoint
of
the
secondary
test
is
above
the
CMC
or
CCC
to
which
the
FWER
is
to
be
applied,
the
additional
tests
can
also
be
conducted
with
the
secondary
test.

Month
April
WY
June
Lowest
Month
April
WY
June
Lowest
Case
1
Case
1
(
Primary
Test)

4.801
2.552
9.164
2.552
(
Primary
Test)

4.801
2.552
9.164
Case
3
(
Primary
(
Second.
Geo.
Test)
Test)
Mean
4.801
3.163
3.897
2.552
5.039
3.586
9.164
7.110
8.072
3.586
Case
2
(
Primary
Geometric
Test1
Mean
3.565
4.137
4.190
3.270
6.736
7.857
3.270
Case
4
(
P;
riyry
(
Second.
Geo.
t)
Test)
Mean
4.801
3.163
3.897
2.552
2.944
2.741
9.164
7.110
8.072
2.741
uses
the
individual
WERs
obtained
with
the
primary
test
for
the
three
months,
and
the*
PWER
is
the
lowest
of
the
three
WERS.
In
Case
2,
duplicate
primary
tests
were
conducted
in
each
month,
so
that
a
geometric
mean
could
be
calculated
for
each
month;
the
PWER
is
the
lowest
of
the
three
geometric
means.

In
Cases
3
and
4,
both
a
primary
test
and
a
secondary
test
were
conducted
each
month
and
the
endpoints
for
both
tests
in
laboratory
dilution
water
are
above
the
CMC
or
CCC
to
which
the
PWER
is
to
be
applied.
In
both
of
these
cases,
therefore,
the
FWER
is
the
lowest
of
the
three
geometric
means.

The
availability
of
these
alternatives
does
not
mean
that
they
are
necessarily
cost­
effective.

73
Fiw
5:
Calculating
an
I&
SO
(
or
BCSO)
by
Interpolation
When
fewer
than
two
treatments
kill
some
but
not
all
of
the
exposed
test
organisms,
a
statistically
sound
estimate
of
an
~~
50
cannot
be
calculated.
Some
programs
and
methods
produce
LCSOs
when
there
are
fewer
than
two
'
partial
kills',
but
such
results
are
obtained
using
interpolation,
not
statistics.
If
(
a)
a
test
is
otherwise
acceptable,
(
b)
a
sufficient
number
of
organisms
are
exposed
to
each
treatment,
and
(
c)
the
concentrations
are
sufficiently
close
together,
a
test
with
zero
or
one
partial
kill
can
provide
all
the
information
that
is
needed
concerning
the
LCSO.
in
LCSO
calculated
by
interpolation
should
probably
be
called
an
'
approximate
LCSO'
to
acknowledge
the
lack
of
a
statistical
basis
for
its
calculation,
but
this
does
not
imply
that
such
an
LCSO
provides
no
useful
toxicological
information.
If
desired,
the
binomial
test
can
be
used
to
calculate
a
statistically
sound
probability
that
the
true
LCSO
lies
between
two
tested
concentrations
(
Stephan
1977).

Although
more
complex
interpolation
methods
can
be
used,
they
will
not
produce
a
more
useful
LCSO
than
the
method
described
here.
Inversions
in
the
data
between
two
test
concentrations
should
be
removed
by
pooling
the
mortality
data
for
those
two
concentrations
and
calculating
a
percent
mortality
that
is
then
assigned
to
both
concentrations.
Logarithms
to
a
base
other
than
10
can
be
used
if
desired.
If
Pl
and
P2
are
the
percentages
of
the
test
organisms
that
died
when
exposed
to
concentrations
Cl
and
C2,
respectively,
and
if
Cl
<
c2,
Pl
<
P2,
0
s
Pl
5
50,
and
50
5
P2
s
100,
then:

c
=
Log
Cl
+
Puog
c2
­
Log
Cl)

Lc50
­
lo=

If
Pl
=
0
and
P2
=
100,
tcso­
m
If
Pl
=
P2
If
Pl
=
50,
Lcso=
m
=
50,
LCSO
=
Cl.
If
P2
=
50,
LCSO
=
c2.
If
Cl
=
4
mg/
L,
C2
=
7
rag/
L,
Pl
=
15
%,
and
P2
=
100
%,
then
LCSO
=
5.036565
mg/
L.

Besides
the
mathematical
requirements
given
above,
the
following
toxicological
recommendations
are
given
in
sections
G.
8
and
1.2:
a.
0.65
<
Cl/
C2
<
0.99.
b.
0
5
Pl
<
37.
c.
63
<
P2
S
100.

74
tiguro
6:
Calculating
a
Time­
Weighted
Average
If
a
sampling
plan
(
e.
g.,
for
measuring
metal
in
a
treatment
in
a
toxicity
test)
is
designed
so
that
a
series
of
values
are
obtained
over
time
in
such
a
way
that
each
value
contains
the
same
amount
of
information
(
i.
e.,
represents
the
same
amount
of
time),
then
the
most
meaningful
average
is
the
arithmetic
average.
In
most
cases,
however,
when
a
series
of
values
is
obtained
over
time,
some
values
contain
more
information
than
others;
in
these
cases
the
most
meaningful
average
is
a
time­
weighted
average
(
TWA).
If
each
value
contains
the
same
amount
of
information,
the
arithmetic
average
will
equal
the
TWA.

A
TWA
is
obtained
by
multiplying
each
value
by
a
weight
and
then
dividing
the
sum
of
the
products
by
the
sum
of
the
weights.
The
simplest
approach
is
to
let
each
weight
be
the
duration
of
time
that
the
sample
represents.
Except
for
the
first
and
last
samples,
the
period
of
time
represented
by
a
sample
starts
halfway
to
the
previous
sample
and
ends
halfway
to
the
next
sample.
The
period
of
time
represented
by
the
first
sample
starts
at
the
beginning
of
the
test,
and
the
period
of
time
represented
by
the
last
sample
ends
at
the
end
of
the
test.
Thus
for
a
96­
hr
toxicity
test,
the
sum
of
the
weights
will
be
96
hr.

The
following
are
hypothetical
examples
of
grab
samples
taken
from
96­
hr
flow­
through
tests
for
two
comon
sampling
regimes:

Sampling
Cont.
Weight
Product
Time­
weighted
average
time
(
hr)
(
mu/
L)
(
hr)
(
hr)
(
mu/
L)
(
ma/
L)

0
12
48
576
96
14
48
96
i%

2:
8
12
6
24
1::
48
7
24
168
72
9
24
216
96
8
u
96
1248/
96
=
13.00
720/
96
=
7.500
When
all
the
weights
are
the
same,
the
arithmetic
average
equals
the
TWA.
Similarly,
if
only
one
sample
is
taken,
both
the
arithmetic
average
and
the
TWA
equal
the
value
of
that
sample.

The
rules
are
more
complex
for
composite
samples
and
for
samples
from
renewal
tests.
In
all
cases,
however,
the
sampling
plan
can
be
designed
so
that
the
TWA
equals
the
arithmetic
average.

75
REFERENCES
ASTM.
1993a.
Guide
for
Conducting
Acute
Toxicity
Tests
with
Fishes,
Macroinvertebrates,
and
Amphibians.
Standard
E729.
American
Society
for
Testing
and
Materials,
Philadelphia,
PA.

ASTM.
1993b.
Guide
for
Conducting
Static
Acute
Toxicity
Tests
Starting
with
Embryos
of
Four
Species
of
Saltwater
Bivalve
Molluscs.
Standard
E724.
American
Society
for
Testing
and
Materials,
Philadelphia,
PA.

ASTM.
1993c.
Guide
for
Conducting
Renewal
Life­
Cycle
Toxicity
Tests
with
Daphnia
magna.
Standard
E1193.
American
Society
for
Testing
and
Materials,
Philadelphia,
PA.

ASTM.
1993d.
Guide
for
Conducting
Early
Life­
Stage
Toxicity
Tests
with
Fishes.
Standard
E1241.
American
Society
for
Testing
and
Materials,
Philadelphia,
PA.

ASTM.
1993e.
Guide
for
Conducting
Three­
Brood,
Renewal
Toxicity
Tests
with
Ceriodaphnia
dubia.
Standard
E1295.
American
Society
for
Testing
and
Materials,
Philadelphia,
PA.

ASTM.
1993f.
Guide
for
Conducting
Acute
Toxicity
Tests
on
Aqueous
Effluents
with
Fishes,
Macroinvertebrates,
and
Amphibians.
Standard
E1192.
American
Society
for
Testing
and
Materials,
Philadelphia,
PA.

Barnthouse,
L.
W.,
G.
W.
Suter,
A.
E.
Rosen,
and
J.
J.
Beauchamp.
1987.
Estimating
Responses
of
Fish
Populations
to
Toxic
Contaminants.
Environ.
Toxicol.
Chem.
6:
811­
824.

Bruce,
R.
D.,
and
D.
J.
Versteeg.
1992.
A
Statistical
Procedure
for
Modeling
Continuous
Toxicity
Data.
Environ.
Toxicol.
Chem.
11:
1485­
1494.

Hoekstra,
J.
A.,
and
P.
H.
Van
Ewijk.
1993.
Alternatives
for
the
No­
Observed­
Effect
Level.
Environ.
Toxicol.
Chem.
12:
187­
194.

Kilpatrick,
F.
A.
1992.
Simulation
of
Soluble
Waste
Transport
and
Buildup
in
Surface
Waters
Using
Tracers.
Open­
File
Report
92­
457.
U.
S.
Geological
Survey,
Books
and
Open­
File
Reports,
Box
25425,
Federal
Center,
Denver,
CO
80225.

Natrella,
M.
G.
1966.
Experimental
Statistics.
National
Bureau
of
Standards
Handbook
91.
(
Issued
August
1,
1963;
reprinted
October
1966
with
corrections).
U.
S.
Government
Printing
Office,
Washington,
DC.

76
Prothro,
M.
G.
1993.
Memorandum
titled
"
Office
of
Water
Policy
and
Technical
Guidance
on
Interpretation
and
Implementation
of
Aquatic
Life
Metals
Criteria'.
October
1.

Stephan,
C.
E.
1977.
Methods
for
Calculating
an
LC50.
In:
Aquatic
Toxicology
and
Hazard
Evaluation.
(
F.
L.
Mayer
and
J.
L.
Hamelink,
eds.)
ASTM
STP
634.
American
Society
for
Testing
and
Materials,
Philadelphia,
PA.
pp.
65­
84.

Stephan,
C.
E.,
and
J.
W.
Rogers.
1985.
Advantages
of
Using
Regression
Analysis
to
Calculate
Results
of
Chronic
Toxicity
Tests.
In:
Aquatic
Toxicology
and
Hazard
Assessment:
Eighth
Symposium.
(
R.
C.
Bahner
and
D.
J.
Hansen,
eds.)
ASTM
STP
891.
American
Society
for
Testing
and
Materials,
Philadelphia,
PA.
pp.
328­
338.

Suter,
G.
W.,
A.
E.
Rosen,
E.
Linder,
and
D.
F.
Parkhurst.
1987.
Endpoints
for
Responses
of
Fish
to
Chronic
Toxic
Exposures.
Environ.
Toxicol.
Chem.
6:
793­
809.

U.
S.
EPA.
1983a.
Water
Quality
Standards
Handbook.
Office
of
Water
Regulations
and
Standards,
Washington,
DC.

U.
S.
EPA.
1983b.
Methods
for
Chemical
Analysis
of
Water
and
Wastes.
EPA­
600/
4­
79­
020.
National
Technical
Information
Service,
Springfield,
VA.

U.
S.
EPA.
1984.
Guidelines
for
Deriving
Numerical
Aquatic
Site­
Specific
Water
Quality
Criteria
by
Modifying
National
Criteria.
EPA­
600/
3­
84­
099
or
PB85­
121101.
National
Technical
Information
Service,
Springfield,
VA.

U.
S.
EPA.
1985.
Guidelines
for
Deriving
Numerical
National
Water
Quality
Criteria
for
the
Protection
of
Aquatic
Organisms
and
Their
Uses.
PB85­
227049.
National
Technical
Information
Service,
Springfield,
VA.

U.
S.
EPA.
1991a.
Technical
Support
Document
for
Water
Quality­
based
Toxics
Control.
EPA/
505/
2­
90­
001
or
PB91­
127415.
National
Technical
Information
Service,
Springfield,
VA.

U.
S.
EPA.
1991b.
Manual
for
the
Evaluation
of
Laboratories
Performing
Aquatic
Toxicity
Tests.
EPA/
600/
4­
90/
031.
National
Technical
Information
Service,
Springfield,
VA.

U.
S.
EPA.
1991c.
Methods
for
the
Determination
of
Metals
in
Environmental
Samples.
EPA­
600/
4­
91­
010.
National
Technical
Information
Service,
Springfield,
VA.

77
U.
S.
EPA.
1992.
Interim
Guidance
on
Interpretation
and
Implementation
of
Aquatic
Life
Criteria
for
Metals.
Office
of
Science
and
Technology,
Health
and
Ecological
Criteria
Division,
Washington,
DC.

Uf;
S.
EPA.
1993a.
Methods
for
Measuring
the
Acute
Toxicity
of
Effluents
and
Receiving
Waters
to
Freshwater
and
Marine
Organisms.
Fourth
Edition.
EPA/
600/
4­
90/
027F.
National
Technical
Information
Service,
Springfield,
VA.

U.
S.
EPA.
1993b.
Short­
term
Methods
for
Estimating
the
Chronic
Toxicity
of
Effluents
and
Receiving
Waters
to
Freshwater
Organisms.
Third
Edition.
EPA/
600/
4­
91/
002.
National
Technical
Information
Service,
Springfield,
VA.

U.
S.
EPA.
1993c.
Short­
Term
Methods
for
Estimating
the
Chronic
Toxicity
of
Effluents
and
Receiving
Waters
to
Marine
and
Estuarine
Organisms.
Second
Edition.
EPA/
600/
4­
91/
003.
National
Technical
Information
Service,
Springfield,
VA.

U.
S.
EPA.
19938.
Dilution
Models
for
Effluent
Discharges.
Second
Edition.
EPA/
600/
R­
93/
139.
National
Technical
Information
Service,
Springfield,
VA.

78
mix
A:
comparison
of
WERs
Datormined
Using
Upstram
and
Downstream
Wator
The
'
Interim
Guidance'
concerning
metals
(
U.
S.
EPA
1992)
made
a
fundamental
change
in
the
way
WERs
should
be
experimentally
determined
because
it
changed
the
source
of
the
site
water.
The
earlier
guidance
(
U.
S.
EPA
1983,1984)
required
that
upstream
water
be
used
as
the
site
water,
whereas
the
newer
guidance
(
U.
S.
EPA
1992)
recommended
that
downstream
water
be
used
as
the
site
water.
The
change
in
the
source
of
the
site
water
was
merely
an
acknowledgement
that
the
WER
that
applies
at
a
location
in
a
body
of
water
should,
when
possible,
be
determined
using
the
water
that
occurs
at
that
location.

Because
the
change
in
the
source
of
the
dilution
water
was
expected
to
result
in
an
increase
in
the
magnitude
of
many
WERs,
interest
in
and
concern
about
the
determination
and
use
of
WERs
increased.
When
upstream
water
was
the
required
site
water,
it
was
expected
that
WERs
would
generally
be
low
and
that
the
determination
and
use
of
WERs
could
be
fairly
simple.
After
downstream
water
became
the
recommended
site
water,
the
determination
and
use
of
WERs
was
examined
much
more
closely.
It
was
then
realized
that
the
determination
and
use
of
upstream
WERs
was
more
complex
than
originally
thought.
It
was
also
realized
that
the
use
of
downstream
water
greatly
increased
the
complexity
and
was
likely
to
increase
both
the
magnitude
and
the
variability
of
many
WERs.
Concern
about
the
fate
of
discharged
metal
also
increased
because
use
of
downstream
water
might
allow
the
discharge
of
large
amounts
of
metal
that
has
reduced
or
no
toxicity
at
the
end
of
the
pipe.
The
probable
increases
in
the
complexity,
magnitude,
and
variability
of
WERs
and
the
increased
concern
about
fate,
increased
the
importance
of
understanding
the
relevant
issues
as
they
apply
to
WERs
determined
using
both
upstream
water
and
downstream
water.

A.
Characteristics
of
the
Site
Water
The
idealized
concept
of
an
upstream
water
is
a
pristine
water
that
is
relatively
unaffected
by
people.
In
the
real
world,
however,
many
upstream
waters
contain
naturally
occurring
ligands,
one
or
more
effluents,
and
materials
from
nonpoint
sources;
all
of
these
might
impact
a
WER.
If
the
upstream
water
receives
an
effluent
containing
Tot
and/
or
TSS
that
contributes
to
the
WER,
the
WER
will
probably
change
whenever
the
quality
or
quantity
of
the
TOC
and/
or
TSS
changes.
In
such
a
case,
the
determination
and
use
of
the
WER
in
upstream
water
will
have
some
of
the
increased
complexity
associated
with
use
of
downstream
water
and
some
of
the
concerns
associated
with
multiple­
discharge
situations
(
see
Appendix
F)
l
The
amount
of
complexity
will
depend
greatly
on
the
79
mxnber
and
type
of
upstream
point
and
nonpoint
sources,
the
frequency
and
magnitude
of
fluctuations,
and
whether
the
WER
is
being
determined
above
or
below
the
point
of
complete
mix
of
the
upstream
sources.

Downstream
water
is
a
mixture
of
effluent
and
upstream
water,
each
of
which
can
contribute
to
the
WEFT,
and
so
there
are
two
cosaponents
to
a
WER
determined
in
downstream
water:
the
effluent
corqmnent
and
the
upstream
component.
The
existence
of
1.

2.

3.

4.

5.

6.
these
two
cmponents
has
the
following
implications:
WERS
determined
using
downstream
water
are
likely
to
be
larger
and
more
variable
than
WEFIs
determined
using
upstream
water.
The
effluent
component
should
be
applied
only
where
the
effluent
occurs,
which
has
implications
concerning
implementation.
The
magnitude
of
the
effluent
caflponent
of
a
WER
will
depend
on
the
concentration
of
effluent
in
the
downstream
water.
(
A
consequence
of
this
is
that
the
effluent
component
will
be
zero
where
the
concentration
of
effluent
is
zero,
which
is
the
point
of
item
2
above.)
The
magnitude
of
the
effluent
component
of
a
WER
is
likely
to
vary
as
the
composition
of
the
effluent
varies.
Compared
to
upstream
water,
many
effluents
contain
higher
concentrations
of
a
wider
variety
of
substances
that
can
impact
the
toxicity
of
metals
in
a
wider
variety
of
ways,
and
so
the
effluent
component
of
a
WER
can
be
due
to
a
variety
of
chemical
effects
in
addition
to
such
factors
as
hardness,
alkalinity,
pH,
and
humic
acid.
Because
the
effluent
component
might
be
due,
in
whole
or
in
part,
to
the
discharge
of
refractory
metal
(
see
Appendix
D)
i
the
WER
cannot
be
thought
of
simply
as
being
caused
by
the
effect
of
water
quality
on
the
toxicity
of
the
metal.
Dealing
with
downstream
WERs
is
so
much
simpler
if
the
effluent
WER
(
eWER1
and
the
upstream
WER
(
WER)
are
additive
that
it
is
desirable
to
understand
the
concept
of
additivity
of
WERS,
its
experimental
determination,
and
its
use
(
see
Appendix
G).

B.
The
Implications
of
Mixing
Zones.

When
WERs
are
determined
using
upstream
water,
the
presence
or
absence
of
mixing
zones
has
no
impact;
the
cmcWER
and
the
cccWER
will
both
be
determined
using
site
water
that
contains
zero
percent
of
the
effluent
of
concern,
i.
e.,
the
two
WERs
will
be
determined
using
the
same
site
water.

When
WERs
are
determined
using
downstream
water,
the
magnitude
of
each
WER
will
probably
depend
on
the
concentration
of
effluent
in
the
downstream
water
used
(
see
Appendix
D).
The
concentration
of
effluent
in
the
site
water
will
depend
on
80
where
the
sample
is
taken,
which
will
not
be
the
same
for
the
cmcWER
and
the
cccWJ3R
if
there
are
mixing
zone(
s).
Most,
if
not
all,
discharges
have
a
chronic
(
CCC)
mixing
zone;
many,
but
not
all,
also
have
an
acute
(
CMC)
mixing
zone.
The
CMC
applies
at
all
points
except
those
inside
a
CMC
mixing
zone;
thus
if
there
is
no
CMC
mixing
zone,
the
CMC
applies
at
the
end
of
the
pipe.
The
CCC
applies
at
all
points
outside
the
CCC
mixing
zone.
It
is
generally
assumed
that
if
permit
limits
are
based
on
a
point
in
a
stream
at
which
both
the
CMC
and
the
CCC
apply,
the
CCC
will
control
the
permit
limits,
although
the
CMC
might
control
if
different
averaging
periods
are
appropriately
taken
into
account.
For
this
discussion,
it
will
be
assumed
that
the
same
design
flow
(
e.
g.,
7910)
is
used
for
both
the
CMC
and
the
CCC.

If
the
cmcwER
is
to
be
appropriate
for
use
inside
the
chronic
mixing
zone,
but
the
CCCWER
is
to
be
appropriate
for
use
outside
the
chronic
mixing
zone,
the
concentration
of
effluent
that
is
appropriate
for
use
in
the
determination
of
the
two
WERs
will
not
be
the
same.
Thus
even
if
the
same
toxicity
test
is
used
in
the
determination
of
the
uncWER
and
the
CCCWER,
the
two
WERs
will
probably
be
different
because
the
concentration
of
effluent
will
be
different
in
the
two
site
waters
in
which
the
WERs
are
determined.

If
the
CMC
is
only
of
concern
within
the
CCC
mixing
zone,
the
highest
relevant
concentration
of
metal
will
occur
at
the
edge
of
the
CMC
mixing
zone
if
there
is
a
CMC
mixing
zone;
the
highest
concentration
will
occur
at
the
end
of
the
pipe
if
there
is
no
CMC
mixing
zone.
In
contrast,
within
the
CCC
mixing
zone,
the
lowest
cmcWER
will
probably
occur
at
the
outer
edge
of
the
CCC
mixing
zone.
Thus
the
greatest
level
of
protection
would
be
provided
if
the
cmcWER
is
determined
using
water
at
the
outer
edge
of
the
CCC
mixing
zone,
and
then
the
calculated
site­
specific
CMC
is
applied
at
the
edge
of
the
CMC
mixing
zone
or
at
the
end
of
the
pipe,
depending
on
whether
there
is
an
acute
mixing
zone.
The
cmcWER
is
likely
to
be
lowest
at
the
outer
edge
of
the
CCC
mixing
zone
because
of
dilution
of
the
effluent,
but
this
dilution
will
also
dilute
the
metal.
If
the
cmcWER
is
determined
at
the
outer
edge
of
the
CCC
mixing
zone
but
the
resulting
site­
specific
CMC
is
applied
at
the
end
of
the
pipe
or
at
the
edge
of
the
CMC
mixing
zone,
dilution
is
allowed
to
reduce
the
WER
but
it
is
not
allowed
to
reduce
the
concentration
of
the
metal.
This
approach
is
environmentally
conservative,
but
it
is
probably
necessary
given
current
implementation
procedures.
(
The
situation
might
be
more
complicated
if
the
WER
is
higher
than
the
eWER
or
if
the
two
WERs
are
less­
than­
additive.)

A
comparable
situation
applies
to
the
CCC.
Outside
the
CCC
mixing
zone,
the
CMC
and
the
CCC
both
apply,
but
it
is
assumed
that
the
CMC
can
be
ignored
because
the
CCC
will
be
more
81
restrictive.
The
CCCWER
should
probably
be
determined
for
the
complete­
mix
situation,
but
the
site­
specific
CCC
will
have
to
be
met
at
the
edge
of
the
CCC
mixing
zone.
Thus
dilution
of
the
WER
from
the
edge
of
the
CCC
mixing
zone
to
the
point
of
complete
mix
is
taken
into
account,
but
dilution
of
the
metal
is
not.

If
there
is
neither
an
acute
nor
a
chronic
mixing
zone,
both
the
CMC
and
the
CCC
apply
at
the
end
of
the
pipe,
but
the
CCC
should
still
be
determined
for
the
complete­
mix
situation.

C.
Definition
of
site.

In
the
general
context
of
site­
specific
criteria,
a
.
site.
may
be
a
state,
region,
watershed,
waterbody,
segment
of
a
waterbody,
category
of
water
(
e.
g.,
ephemeral
streams),
etc.,
but
the
site­
specific
criterion
is
to
be
derived
to
provide
adequate
protection
for
the
entire
site,
however
the
site
is
defined.
Thus,
when
a
site­
specific
criterion
is
derived
using
the
Recalculation
Procedure,
all
species
that
'
occur
at
the
site.
need
to
be
taken
into
account
when
deciding
what
species,
if
any,
are
to
be
deleted
from
the
dataset.
Similarly,
when
a
site­
specific
criterion
is
derived
using
a
WER,
the
WER
is
to
be
adequately
protective
of
the
entire
site.
If,
for
example,
a
site­
specific
criterion
is
being
derived
for
an
estuary,
WERs
could
be
determined
using
samples
of
the
surface
water
obtained
from
various
sampling
stations,
which,
to
avoid
confusion,
should
not
be
called
'
sites'.
If
all
the
WERs
were
sufficiently
similar,
one
site­
specific
criterion
could
be
derived
to
apply
to
the
whole
estuary.
If
the
WERs
were
sufficiently
different,
either
the
lowest
WER
could
be
used
to
derive
a
site­
specific
criterion
for
the
whole
estuary,
or
the
data
might
indicate
that
the
estuary
should
be
divided
into
two
or
more
sites,
each
with
its
own
criterion.

The
major
principle
that
should
be
applied
when
defining
the
area
to
be
included
in
the
site
is
very
simplistic:
The
site
should
be
neither
too
small
nor
too
large.
1.
Small
sites
are
probably
appropriate
for
cmcWERs,
but
usually
are
not
appropriate
for
CCCWERS
because
metals
are
persistent,
although
some
oxidation
states
are
not
persistent
and
some
metals
are
not
persistent
in
the
water
column.
For
cccWERs,
the
smaller
the
defined
site,
the
more
likely
it
is
that
the
permit
limits
will
be
controlled
by
a
criterion
for
an
area
that
is
outside
the
site,
but
which
could
have
been
included
in
the
site
without
substantially
changing
the
WER
or
increasing
the
cost
of
determining
the
WER.
2.
Too
large
an
area
might
unnecessarily
increase
the
cost
of
determining
the
WER.
As
the
size
of
the
site
increases,

82
the
spatial
and
temporal
variability
is
likely
to
increase,
which
will
probably
increase
the
number
of
water
samples
in
which
WERs
will
need
to
be
determined
before
a
site­
specific
criterion
can
be
derived.
3.
Events
that
import
or
resuspend
TSS
and/
or
TW
are
likely
to
increase
the
total
recoverable
concentration
of
the
metal
and
the
total
recoverable
WER
while
having
a
much
smaller
effect
on
the
dissolved
concentration
and
the
dissolved
WER.
Where
the
concentration
of
dissolved
metal
is
substantially
more
constant
than
the
concentration
of
total
recoverable
metal,
the
site
can
probably
be
much
larger
for
a
dissolved
criterion
than
for
a
total
recoverable
criterion.
If
one
criterion
is
not
feasible
for
the
whole
area,
it
might
be
possible
to
divide
it
into
two
or
more
sites
with
separate
total
recoverable
or
dissolved
criteria
or
to
make
the
criterion
dependent
on
a
water
quality
characteristic
such
as
TSS
or
salinity.
4.
Unless
the
site
ends
where
one
body
of
water
meets
another,
at
the
outer
edge
of
the
site
there
will
usually
be
an
instantaneous
decrease
in
the
allowed
concentration
of
the
metal
in
the
water
column
due
to
the
change
from
one
criterion
to
another,
but
there
will
not
be
an
instantaneous
decrease
in
the
actual
concentration
of
metal
in
the
water
column.
The
site
has
to
be
large
enough
to
include
the
transition
zone
in
which
the
actual
concentration
decreases
so
that
the
criterion
outside
the
site
is
not
exceeded.
It
is,
of
course,
possible
in
some
situations
that
relevant
distant
conditions
(
e.
g.,
a
lower
downstream
pH)
will
necessitate
a
low
criterion
that
will
control
the
permit
limits
such
that
it
is
pointless
to
determine
a
WER.

When
a
WER
is
determined
in
upstream
water,
it
is
generally
assumed
that
a
downstream
effluent
will
not
decrease
the
WER.
It
is
therefore
assumed
that
the
site
can
usually
cover
a
rather
large
geographic
area.

When
a
site­
specific
criterion
is
derived
based
on
WERs
determined
using
downstream
water,
the
site
should
not
be
defined
in
the
same
way
that
it
would
be
defined
if
the
WER
were
determined
using
upstream
water.
The
eWER
should
be
allowed
to
affect
the
site­
specific
criterion
wherever
the
effluent
occurs,
but
it
should
not
be
allowed
to
affect
the
criterion
in
places
where
the
effluent
does
not
occur.
In
addition,
insofar
as
the
magnitude
of
the
effluent
component
at
a
point
in
the
site
depends
on
the
concentration
of
effluent,
the
magnitude
of
the
WER
at
a
particular
point
will
depend
on
the
concentration
of
effluent
at
that
point.
To
the
extent
that
the
eWER
and
the
WER
are
additive,
the
WER
and
the
concentration
of
metal
in
the
plume
will
decrease
proportionally
(
see
Appendix
G).

83
D.
The
variability
in
the
experimental
determination
of
a
WER.
When
WERs
are
determined
using
downstream
water,
the
following
considerations
should
be
taken
into
account
when
the
site
is
defined:
1.
If
a
site­
specific
criterion
is
derived
using
a
WER
that
applies
to
the
complete­
mix
situation,
the
upstream
edge
of
the
site
to
which
this
criterion
applies
should
be
the
point
at
which
complete
mix
actually
occurs.
If
the
site
to
which
the
complete­
mix
WEZ
is
applied
starts
at
the
end
of
the
pipe
and
extends
all
the
way
across
the
stream,
there
will
be
an
area
beside
the
plume
that
will
not
be
adequately
protected
by
the
site­
specific
criterion.
2.
Upstream
of
the
point
of
complete
mix,
it
will
usually
be
protective
to
apply
a
site­
specific
criterion
that
was
derived
using
a
WER
that
was
determined
using
upstream
water.
3.
The
plums
might
bs
an
area
in
which
the
concentration
of
metal
could
exceed
a
site­
specific
criterion
without
causing
toxicity
because
of
simultaneous
dilution
of
the
metal
and
the
eWER.
The
fact
that
the
plume
is
much
larger
than
the
mixing
zone
might
not
be
important
if
there
is
no
toxicity
within
the
plume.
As
long
as
the
concentration
of
metal
in
100
%
effluent
does
not
exceed
that
allowed
by
the
additive
portion
of
the
eWER,
from
a
toxicologic&
l
standpoint
neither
the
size
nor
the
definition
of
the
plume
needs
to
be
of
concern
because
the
metal
will
not
cause
toxicity
within
the
plume.
If
there
is
no
toxicity
within
the
plume,
the
area
in
the
plums
might
be
like
a
traditional
mixing
zone
in
that
the
concentration
of
metal
exceeds
the
site­
specific
criterion,
but
it
would
be
different
from
a
traditional
mixing
zone
in
that
the
level
of
protection
is
not
reduced.

Special
considerations
are
likely
to
be
necessary
in
order
to
take
into
account
the
ewER
when
defining
a
site
related
to
multiple
discharges
(
see
Appendix
F).

When
a
WER
is
determined
using
upstream
water,
the
two
major
sources
of
variation
in
the
WEB
are
(
a)
variability
in
the
quality
of
the
site
water,
which
might
be
related
to
season
and/
or
flow,
and
(
b)
experimental
variation.
Ordinary
day­
to­
day
variation
will
account
for
some
of
the
variability,
but
seasonal
variation
is
likely
to
be
more
important.

As
explained
in
Appendix
D,
variability
in
the
concentration
of
nontoxic
dissolved
metal
will
contribute
to
the
variability
of
both
total
recoverable
WERs
and
dissolved
WERs;
variability
in
the
concentration
of
nontoxic
particulate
metal
will
contribute
to
the
variability
in
a
total
recoverable
WER,
but
not
to
the
variability
in
a
dissolved
WER.
Thus,
dissolved
84
WERs
are
expected
to
be
less
variable
than
total
recoverable
WERs,
especially
where
events
conunonly
increase
TSS
and/
or
Tot.
In
some
cases,
therefore,
appropriate
use
of
analytical
chemistry
can
greatly
increase
the
usefulness
of
the
experimental
determination
of
WERs.
The
concerns
regarding
variability
are
increased
if
an
upstream
effluent
contributes
to
the
WER.

When
a
WER
is
determined
in
downstream
water,
the
four
major
sources
of
variability
in
the
WER
are
(
a)
variability
in
the
quality
of
the
upstream
water,
which
might
be
related
to
season
and/
or
flow,
(
b)
experimental
variation,
(
c)
variability
in
the
composition
of
the
effluent,
and
(
d)
variability
in
the
ratio
of
the
flows
of
the
upstream
water
and
the
effluent.
The
considerations
regarding
the
first
two
are
the
same
as
for
WERs
determined
using
upstream
water;
because
of
the
additional
sources
of
variability,
WERs
determined
using
downstream
water
are
likely
to
be
more
variable
than
WEXs
determined
using
upstream
water.

It
would
be
desirable
if
a
sufficient
number
of
WEZRs
could
be
determined
to
define
the
variable
factors
in
the
effluent
and
in
the
upstream
water
that
contribute
to
the
variability
in
WERs
that
are
determined
using
downstream
water.
Not
only
is
this
likely
to
be
very
difficult
in
most
cases,
but
it
is
also
possible
that
the
WER
will
be
dependent
on
interactions
between
constituents
of
the
effluent
and
the
upstream
water,
i.
e.,
the
eWER
and
uWER
might
be
additive,
more­
than­
additive,
or
less­
than­
additive
(
see
Appendix
G).
When
interaction
occurs,
in
order
to
completely
understand
the
variability
of
WERs
determined
using
downstream
water,
sufficient
tests
would
have
to
be
conducted
to
determine
the
means
and
variances
of:
the
effluent
component
of
the
WER.
t:
the
upstream
component
of
the
WER.
C.
any
interaction
between
the
two
components.
An
interaction
might
occur,
for
example,
if
the
toxicity
of
a
metal
is
affected
by
pH,
and
the
pH
and/
or
the
buffering
capacity
of
the
effluent
and/
or
the
upstream
water
vary
considerably.

An
increase
in
the
variability
of
WERs
decreases
the
usefulness
of
any
one
WER.
Compensation
for
this
decrease
in
usefulness
can
be
attempted
by
determining
WERs
at
more
times;
although
this
will
provide
more
data,
it
will
not
necessarily
provide
a
proportionate
increase
in
understanding.
Rather
than
determining
WERs
at
more
times,
a
better
use
of
resources
might
be
to
obtain
more
information
concerning
a
smaller
number
of
specially
selected
occasions.

It
is
likely
that
some
cases
will
be
so
complex
that
achieving
even
a
reasonable
understanding
will
require
unreasonable
resources.
In
contrast,
some
WERs
determined
using
the
85
methods
presented
herein
might
be
relatively
easy
to
understand
if
appropriate
chemical
measurements
are
performed
when
WERs
are
determined.
1.
If
the
variation
of
the
total
recoverable
WER
is
substantially
greater
than
the
variation
of
the
comparable
dissolved
WER,
there
is
probably
a
variable
and
substantial
concentration
of
particulate
nontoxic
metal.
It
might
be
advantageous
to
use
a
dissolved
WER
just
because
it
will
have
less
variability
than
a
total
recoverable
WER.
2.
If
the
total
recoverable
and/
or
dissolved
WER
correlates
with
the
total
recoverable
and/
or
dissolved
concentration
of
metal
in
the
site
water,
it
is
likely
that
a
substantial
percentage
of
the
metal
is
nontoxic.
In
this
case
the
WER
will
probably
also
depend
on
the
concentration
of
effluent
in
the
site
water
and
on
the
concentration
of
metal
in
the
effluent.
These
approaches
are
more
likely
to
be
useful
when
WERs
are
determined
using
downstream
water,
rather
than
upstream
water,
unless
both
the
magnitude
of
the
WER
and
the
concentration
of
the
metal
in
the
upstream
water
are
elevated
by
an
upstream
effluent
and/
or
events
that
increase
TSS
and/
or
Tot.

Both
of
these
approaches
can
be
applied
to
WERs
that
are
determined
using
actual
downstream
water,
but
the
second
can
probably
provide
much
better
information
if
it
is
used
with
WERs
determined
using
simulated
downstream
water
that
is
prepared
by
mixing
a
sample
of
the
effluent
with
a
sample
of
the
upstream
water.
In
this
way
the
composition
and
characteristics
of
both
the
effluent
and
the
upstream
water
can
be
determined,
and
the
exact
ratio
in
the
downstream
water
isknown.

Use
of
simulated
downstream
water
is
also
a
way
to
study
the
relation
between
the
WER
and
the
ratio
of
effluent
to
upstream
water
at
one
point
in
time,
which
is
the
most
direct
way
to
test
for
additivity
of
the
ewER
and
the
uWER
(
see
Appendix
G).
This
can
be
viewed
as
a
test
of
the
assumption
that
WERs
determined
using
downstream
water
will
decrease
as
the
concentration
of
effluent
decreases.
If
this
assumption
is
true,
as
the
flow
increases,
the
concentration
of
effluent
in
the
downstream
water
will
decrease
and
the
WER
will
decrease.
Obtaining
such
information
at
one
point
in
tims
is
useful,
but
confirmation
at
one
or
more
other
times
would
be
much
more
useful
.

E.
The
fate
of
metal
that
has
reduced
or
no
toxicity.

Metal
that
has
reduced
or
no
toxicity
at
the
end
of
the
pipe
might
be
more
toxic
at
some
time
in
the
future.
For
example,
metal
that
is
in
the
water
column
and
is
not
toxic
now
might
become
more
toxic
in
the
water
column
later
or
might
move
into
86
the
sediment
and
become
toxic.
If
a
WER
allows
a
surface
water
to
contain
as
much
toxic
metal
as
is
acceptable,
the
WER
would
not
be
adequately
protective
if
metal
that
was
nontoxic
when
the
WER
was
determined
became
toxic
in
the
water.
column,
unless
a
compensating
change
occurred.
Studies
of
the
fate
of
metals
need
to
address
not
only
the
changes
that
take
place,
but
also
the
rates
of
the
changes.

Concern
about
the
fate
of
discharged
metal
justifiably
raises
concern
about
the
possibility
that
metals
might
contaminate
sediments.
The'possibility
of
contamination
of
sediment
by
toxic
and/
or
nontoxic
metal
in
the
water
column
was
one
of
the
concerns
that
led
to
the
establishment
of
EPA's
sediment
quality
criteria
program,
which
is
developing
guidelines
and
criteria
to
protect
sediment.
A
separate
program
was
necessary
because
ambient
water
quality
criteria
are
not
designed
to
protect
sediment.
Insofar
as
technology­
based
controls
and
water
quality
criteria
reduce
the
discharge
of
metals,
they
tend
to
reduce
the
possibility
of
contamination
of
sediment.
Conversely,
insofar
as
WERs
allow
an
increase
in
the
discharge
of
metals,
they
tend
to
increase
the
possibility
of
contamination
of
sediment.

When
WERs
are
determined
in
upstream
water,
the
concern
about
the
fate
of
metal
with
reduced
or
no
toxicity
is
usually
small
because
the
WERs
are
usually
small.
In
addition,
the
factors
that
result
in
upstream
WERs
being
greater
than
1.0
usually
are
(
a)
natural
organic
materials
such
as
humic
acids
and
(
b)
water
quality
characteristics
such
as
hardness,
alkalinity,
and
pH.
It
is
easy
to
assume
that
natural
organic
materials
will
not
degrade
rapidly,
and
it
is
easy
to
monitor
changes
in
hardness,
alkalinity,
and
pH.
Thus
there
is
usually
little
concern
about
the
fate
of
the
metal
when
WERs
are
determined
in
upstream
water,
especially
if
the
WER
is
small.
If
the
WER
is
large
and
possibly
due
at
least
in
part
to
an
upstream
effluent,
there
is
more
concern
about
the
fate
of
metal
that
has
reduced
or
no
toxicity.

When
WERs
are
determined
in
downstream
water,
effluents
are
allowed
to
contain
virtually
unlimited
amounts
of
nontoxic
particulate
metal
and
nontoxic
dissolved
metal.
It
would
seem
prudent
to
obtain
some
data
concerning
whether
the
nontoxic
metal
might
become
toxic
at
some
time
in
the
future
whenever
(
1)
the
concentration
of
nontoxic
metal
is
large,
(
2)
the
concentration
of
dissolved
metal
is
below
the
dissolved
national
criterion
but
the
concentration
of
total
recoverable
metal
is
substantially
above
the
total
recoverable.
national
criterion,
or
(
3)
the
site­
specific
criterion
is
substantially
above
the
national
criterion.
It
would
seem
appropriate
to:
a.
Generate
some
data
concerning
whether
'
fate'
(
i.
e.,
environmental
processes)
will
cause
any
of
the
nontoxic
metal
to
become
toxic
due
to
oxidation
of
organic
matter,

87
oxidation
of
sulfides,
etc.
For
example,
a
WER
could
be
determined
using
a
sample
of
actual
or
simulated
downstream
water,
the
sample
aerated
for
a
period
of
time
(
e.
g.,
two
weeks),
the
pH
adjusted
if
necessary,
and
another
TIER
determined.
If
aeration
reduced
the
WER,
shorter
and
longer
periods
of
aeration
could
be
used
to
study
the
rate
of
change.
b.
Determine
the
effect
of
a
change
in
water
quality
characteristics
on
the
WER;
for
example,
determine
the
effect
of
lowering
the
pH
on
the
WER
if
influent
lowers
the
pH
of
the
downstream
water
within
the
area
to
which
the
site­
specific
criterion
is
to
apply.
c.
Determine
a
WEX
in
actual
downstream
water
to
demonstrate
whether
downstream
conditions
change
sufficiently
(
possibly
due
to
degradation
of
organic
matter,
multiple
dischargers,
etc.)
to
lower
the
WER
more
than
the
concentration
of
the
metal
is
lowered.
If
environmental
processes
cause
nontoxic
metal
to
become
toxic,
it
is
important
to
determine
whether
the
time
scale
involves
days,
weeks,
or
years.

When
WERs
are
determined
using
downstream
water,
the
site
water
contains
effluent
and
the
WER
will
take
into
account
not
only
the
constituents
of
the
upstream
water,
but
also
the
toxic
and
nontoxic
metal
and
other
constituents
of
the
effluent
as
they
exist
after
mixing
with
upstream
water.
The
determination
of
the
WER
automatically
takes
into
account
any
additivity,
synergism,
or
antagonism
between
the
metal
and
components
of
the
effluent
and/
or
the
upstream
water.
The
effect
of
Calcium,
magnesium,
and
various
heavy
metals
on
competitive
binding
by
such
organic
materials
as
humic
acid
is
also
taken
into
account.
Therefore,
a
site­
specific
criterion
derived
using
a
WER
is
likely
to
be
more
appropriate
for
a
site
than
a
national,
state,
or
recalculated
criterion
not
only
because
it
takes
into
account
the
water
quality
characteristics
of
the
site
water
but
also
because
it
takes
into
account
other
constituents
in
the
effluent
and
upstream
water.

Determination
of
WERs
using
downstream
water
causes
a
general
increase
in
the
complexity,
magnitude,
and
variability
of
WEI&,
and
an
increase
in
concern
about
the
fate
of
metal
that
has
reduced
or
no
toxicity
at
the
end
of
the
pipe.
In
addition,
there
are
some
other
drawbacks
with
the
use
of
downstream
water
in
the
determination
of
a
WER:
1.
It
might
serve
as
a
disincentive
for
some
dischargers
to
remove
any
more
organic
carbon
and/
or
particulate
matter
than
required,
although
WEF&
for
some
metals
will
not
be
related
to
the
concentration
of
Tot
or
TSS.

88
2.
If
conditions
change,
a
WER
might
decrease
in
the
future.
This
is
not
a
problem
if
the
decrease
is
due
to
a
reduction
in
nontoxic
metal,
but
it
might
be
a
problem
if
the
decrease
is
due
to
a
decrease
in
TOC
or
TSS
or
an
increase
in
competitive
binding.
3.
If
a
WBR
is
determined
when
the
effluent
contains
refractory
metal
but
a
change
in
operations
results
in
the
discharge
of
toxic
metal
in
place
of
refractory
metal,
the
site­
specific
criterion
and
the
permit
limits
will
not
provide
adequate
protection.
In
most
cases
chemical
monitoring
probably
will
not
detect
such
a
change,
but
toxicological
monitoring
probably
will.

Use
of
WERs
that
are
determined
using
downstream
water
rather
than
upstream
water
increases:
1.
The
importance
of
understanding
the
various
issues
involved
in
the
determination
and
use
of
WERs.
2.
The
importance
of
obtaining
data
that
will
provide
understanding
rather
than
obtaining
data
that
will
result
in
the
highest
or
lowest
WER.
3.
The
appropriateness
of
site­
specific
criteria.
4.
The
resources
needed
to
determine
a
WBR.
5.
The
resources
needed
to
use
a
WER.
6.
The
resources
needed
to
monitor
the
acceptability
of
the
downstream
water.
A
WER
determined
using
upstream
water
will
usually
be
smaller,
less
variable,
and
simpler
to
implement
than
a
WER
determined
using
downstream
water.
Although
in
some
situations
a
downstream
WBR
might
be
smaller
than
an
upstream
WER,
the
important
consideration
is
that
a
WER
should
be
determined
using
the
water
to
which
it
is
to
apply.

Reference%

U.
S.
EPA.
1983.
Water
Quality
Standards
Handbook.
Office
of
Water
Regulations
and
Standards,
Washington,
DC.

U.
S.
EPA.
1984.
Guidelines
for
Deriving
Numerical
Aquatic
Site­
Specific
Water
Quality
Criteria
by
Modifying
National
Criteria.
EPA­
600/
3­
84­
099
or
PB85­
121101.
National
Technical
Information
Service,
Springfield,
VA.

U.
S.
EPA.
1992.
Interim
Guidance
on
Interpretation
and
Implementation
of
Aquatic
Life
Criteria
for
Metals.
Office
of
Science
and
Technology,
Health
and
Ecological
Criteria
Division,
Washington,
DC.

89
Appmdix
B:
Tha
Roorlenrlrtion
Procodurm
NOTE:
The
National
Toxics
Rule
(
NTR)
does
not
allow
use
of
the
Recalculation
Procedure
in
the
derivation
of
a
site­
specific
criterion.
Thus
nothing
in
this
appendix
applies
to
jurisdictions
that
are
subject
to
the
NTR.

The
Recalculation
Procedure
is
intended
to
cause
a
site­
specific
criterion
to
appropriately
differ
from
a
national
aquatic
life
criterion
if
justified
by
demonstrated
pertinent
toxicological
differences
between
the
aquatic
species
that
occur
at
the
site
and
those
that
were
used
in
the
derivation
of
the
national
criterion.
There
are
at
least
three
reasons
why
such
differences
might
exist
between
the
two
sets
of
species.
First,
the
national
dataset
contains
aquatic
species
that
are
sensitive
to
many
pollutants,
but
these
and
comparably
sensitive
species
might
not
occur
at
the
site.
Second,
a
species
that
is
critical
at
the
site
might
be
sensitive
to
the
pollutant
and
require
a
lower
criterion.
(
A
critical
species
is
a
species
that
is
comnercially
or
recreationally
important
at
the
site,
a
species
that
exists
at
the
site
and
is
listed
as
threatened
or
endangered
under
section
4
of
the
Endangered
Species
Act,
or
a
species
for
which
there
is
evidence
that
the
loss
of
the
species
from
the
site
is
likely
to
cause
an
unacceptable
impact
on
a
commercially
or
recreationally
important
species,
a
threatened
or
endangered
species,
the
abundances
of
a
variety
of
other
species,
or
the
structure
or
function
of
the
coxmuunity.)
Third,
the
species
that
occur
at
the
site
might
represent
a
narrower
mix
of
species
than
those
in
the
national
dataset
due
to
a
limited
range
of
natural
environmental
conditions.
The
procedure
presented
here
is
structured
so
that
corrections
and
additions
can
be
made
to
the
national
dataset
without
the
deletion
process
being
used
to
take
into
account
taxa
that
do
and
do
not
occur
at
the
site;
in
effect,
this
procedure
makes
it
possible
to
update
the
national
aquatic
life
criterion.

The
phrase
'
occur
at
the
site.
includes
the
species,
genera,
families,
orders,
classes,
and
phyla
that:
are
usually
present
at
the
site.
k:
are
present
at
the
site
only
seasonally
due
to
migration.
C.
are
present
intermittently
because
they
periodically
return
to
or
extend
their
ranges
into
the
site.
d.
were
present
at'the
site
in
the
past,
are
not
currently
present
at
the
site
due
to
degraded
conditions,
and
are
expected
to
return
to
the
site
when
conditions
improve.
e.
are
present
in
nearby
bodies
of
water,
are
not
currently
present
at
the
site
due
to
degraded
conditions,
and
are
expected
to
be
present
at
the
site
when
conditions
improve.
The
taxa
that
'
occur
at
the
site"
cannot
be
determined
merely
by
sampling
downstream
and/
or
upstream
of
the
site
at
one
point
in
time.
'
Occur
at
the
site'
does
not
include
taxa
that
were
once
90
present
at
the
site
but
cannot
exist
at
the
site
now
due
to
permanent
physical
alteration
of
the
habitat
at
the
site
resulting
from
dams,
etc.

The
definition
of
the
.
site'
can
be
extremely
important
when
using
the
Recalculation
Procedure.
For
example,
the
number
of
taxa
that
occur
at
the
site
will
generally
decrease
as
the
size
of
the
site
decreases.
Also,
if
the
site
is
defined
to
be
very
small,
the
permit
limit
might
be
controlled
by
a
criterion
that
applies
outside
(
e.
g.,
downstream
of)
the
site.

Note:
If
the
variety
of
aquatic
invertebrates,
amphibians,
and
fishes
is
so
limited
that
species
in
fewer
than
eicrht
ies
occur
at
the
site,
the
general
Recalculation
Procedure
is
not
applicable
and
the
following
special
version
of
the
Recalculation
Procedure
must
be
used:
1.
Data
muat
be
available
for
at
least
one
species
in
each
of
the
families
that
occur
at
the
site.
2.
The
lowest
Species
Mean
Acute
Value
that
is
available
for
a
species
that
occurs
at
the
site
must
be
used
as
the
FAV.
3.
The
site­
specific
CMC
and
CCC
must
be
calculated
as
described
below
in
part
2
of
step
E,
which
is
titled
.
Determination
of
the
CMC
and/
or
CCC'.

The
concept
of
the
Recalculation
Procedure
is
to
create
a
dataset
that
is
appropriate
for
deriving
a
site­
specific
criterion
by
modifying
the
national
dataset
in
some
or
all
of
three
ways:
a.
Correction
of
data
that
are
in
the
national
dataset.
b.
Addition
of
data
to
the
national
dataset.
c.
Deletion
of
data
that
are
in
the
national
dataset.
All
corrections
and
additions
that
have
been
approved
by
U.
S.
EPA
are
required,
whereas
use
of
the
deletion
process
is
optional.
The
Recalculation
Procedure
is
more
likely
to
result
in
lowering
a
criterion
if
the
net
result
of
addition
and
deletion
is
to
decrease
the
number
of
genera
in
the
dataset,
whereas
the
procedure
is
more
likely
to
result
in
raising
a
criterion
if
the
net
result
of
addition
and
deletion
is
to
increase
the
number
of
genera
in
the
dataset.

The
Recalculation
Procedure
consists
of
the
following
steps:
A.
Corrections
are
made
in
the
national
dataset.
B.
Additions
are
made
to
the
national
dataset.
C;
The
deletion
process
may
be
applied
if
desired.
D.
If
the
new
dataset
does
not
satisfy
the
applicable
Minimum
Data
Requirements
(
MDRs),
additional
pertinent
data
mumt
be
generated;
if
the
new
data
are
approved
by
the
U.
S.
EPA,
the
Recalculation
Procedure
mwt
be
started
again
at
step
B
with
the
addition
of
the
new
data.
E.
The
new
CMC
or
CCC
or
both
are
determined.
F.
A
report
is
written.
Each
step
is
discussed
in
more
detail
below.

91
A.
Corrections
1.
Only
corrections
approved
by
the
U.
S.
EPA
may
be
made.
2.
The
concept
of
.
correction.
includes
removal
of
data
that
should
not
have
been
in
the
national
dataset
in
the
first
place.
The
concept
of
.
correction"
does
not
include
removal
of
a
datum
from
the
national
dataset
just
because
the
quality
of
the
datum
is
claimed
to
be
suspect.
If
additional
data
are
available
for
the
same
species,
the
U.
S.
EPA
will
decide
which
data
should
be
used,
based
on
the
available
guidance
(
U.
S.
EPA
1985);
also,
data
based
on
measured
concentrations
are
usually
preferable
to
those
based
on
nominal
concentrations.
3.
Two
kinds
of
corrections
are
possible:
a.
The
first
includes
those
corrections
that
are
known
to
and
have
been
approved
by
the
U.
S.
EPA;
a
list
of
these
will
be
available
from
the
U.
S.
EPA.
b.
The
second
includes
those
corrections
that
are
submitted
to
the
U.
S.
EPA
for
approval.
If
approved,
these
will
be
added
to
EPA's
list
of
approved
corrections.
4.
Selective
corrections
are
not
allowed.
All
corrections
on
EPA's
newest
list
mrmt
be
made.

B.
Additioag
1.
2.

3.

P
Only
additions
approved
by
the
U.
S.
EPA
may
be
made.
Wo
kinds
of
additions
are
possible:
a.
The
first
includes
those
additions
that
are
known
to
and
have
been
approved
by
the
U.
S.
EPA;
a
list
of
these
will
be
available
from
the
U.
S.
EPA.
b.
The
second
includes
those
additions
that
are
submitted
to
the
U.
S.
EPA
for
approval.
If
approved,
these
will
be
added
to
EPA's
list
of
approved
additions.
Selective
additions
are
not
allowed.
All
additions
on
EPA's
newest
list
mast
be
made.

.
&.
me
Deletion
Process
The
basic
principles
are:
1.
Additions
and
corrections
muat
be
made
as
per
steps
A
and
B
above,
before
the
deletion
process
is
performed.
2.
Selective
deletions
are
not
allowed.
If
any
species
is
to
be
deleted,
the
deletion
process
described
below
mamt
be
applied
to
all
species
in
the
national
dataset,
after
any
necessary
corrections
and
additions
have
been
made
to
the
national
dataset.
The
deletion
process
specifies
which
species
muat
be
deleted
and
which
species
aunt
not
be
deleted.
Use
of
the
deletion
process
is
optional,
but
no
deletions
are
optional
when
the
deletion
process
is
used.
3.
Comprehensive
information
must
be
available
concerning
what
species
occur
at
the
site;
a
species
cannot
be
deleted
based
92
on
incomplete
information
concerning
the
species
that
do
and
do
not
satisfy
the
definition
of
'
occur
at
the
site'.
4.
Data
might
have
to
be
generated
before
the
deletion
process
is
begun:
a.
Acceptable
pertinent
toxicological
data
mwt
be
available
for
at
least
one
species
in
each
class
of
aquatic
plants,
invertebrates,
amphibians,
and
fish
that
contains
a
species
that
is
a
critical
species
at
the
site.
b.
For
each
aquatic
plant,
invertebrate,
amphibian,
and
fish
species
that
occurs
at
the
site
and
is
listed
as
threatened
or
endangered
under
section
4
of
the
Endangered
Species
Act,
data
munt
be
available
or
be
generated
for
an
acceptable
Surrogate
Species.
Data
for
each
surrogate
species
must
be
used
as
if
they
are
data
for
species
that
occur
at
the
site.
If
additional
data
are
generated
using
acceptable
procedures
(
U.
S.
EPA
1985)
and
they
are
approved
by
the
U.
S.
EPA,
the
Recalculation
Procedure
must
be
started
again
at
step
B
with
the
addition
of
the
new
data.
5.
Data
might
have
to
be
generated
gfter
the
deletion
process
is
completed.
Even
if
one
or
more
species
are
deleted,
there
still
are
MDRs
(
see
step
D
below)
that
muet
be
satisfied.
If
the
data
remaining
after
deletion
do
not
satisfy
the
applicable
MDRs,
additional
toxicity
tests
must
be
conducted
using
acceptable
procedures
(
U.
S.
EPA
1985)
so
that
all
MDRs
are
satisfied.
If
the
new
data
are
approved
by
the
U.
S.
EPA,
the
Recalculation
Procedure
mumt
be
started
again
at
step
B
with
the
addition
of
new
data.
6.
Chronic
tests
do
not
have
to
be
conducted
because
the
national
Final
Acute­
Chronic
Ratio
(
FACR)
may
be
used
in
the
derivation
of
the
site­
specific
Final
Chronic
Value
(
FCV).
If
acute­
chronic
ratios
(
ACRs)
are
available
or
are
generated
so
that
the
chronic
MDRs
are
satisfied
using
only
species
that
occur
at
the
site,
a
site­
specific
FACR
may
be
derived
and
used
in
place
of
the
national
FACR.
Because
a
FACR
was
not
used
in
the
derivation
of
the
freshwater
CCC
for
cadmium,
this
CCC
can
only
be
modified
the
same
way
as
a
FAV;
what
is
acceptable
will
depend
on
which
species
are
deleted.

If
any
species
are
to
be
deleted,
the
following
deletion
process
at
be
applied:
a.
Obtain
a
copy
of
the
national
dataset,
i.
e.,
tables
1,
2,
and
3
in
the
national
criteria
document
(
see
Appendix
E).
b.
Make
corrections
in
and/
or
additions
to
the
national
dataset
as
described
in
steps
A
and
B
above.
c.
Group
all
the
species
in
the
dataset
taxonomically
by
phylum,
class,
order,
family,
genus,
and
species.
d.
Circle
each
species
that
satisfies
the
definition
of
.
occur
at
the
site'
as
presented
on
the
first
page
of
this
appendix,
and
including
any
data
for
species
that
are
surrogates
of
threatened
or
endangered
species
that
occur
at
the
site.

93
e.
Use
the
following
step­
wise
process
to
determine
which
of
the
uncircled
species
mumt
be
deleted
and
which
muat
not
be
deleted:

1.
Does
the
genus
occur
at
the
site?
If
'
No',
gotostep2.
If
.
Yes.,
are
there
one
or
more
species
in
the
genus
that
occur
at
the
site
but
are
not
in
the
dataset?
If"
No',
go
to
step
2.
If
'
Yes',
retain
the
uncircled
species.*

2.
Does
the
family
occur
at
the
site?
If
'
No',
go
to
step
3.
If
'
Yes.,
are
there
one
or
more
genera
in
the
family
that
occur
at
the
site
but
are
not
in
the
dataset?
If
'
No',
go
to
step
3.
If
'
Yes',
retain
the
uncircled
species.*

3.
Does
the
order
occur
at
the
site?
If
'
No',
go
to
step
4.
If
'
Yes',
does
the
dataset
contain
a
circled
species
that
is
in
the
same
order?
If
'
No',
retain
the
uncircled
species.*
If
'
Yes',
delete
the
uncircled
species.*

4.
Does
the
class
occur
at
the
site?
If
'
No',
go
to
step
5.
If
.
Yesm,
does
the
dataset
contain
a
circled
species
that
is
in
the
same
class?
If
'
No',
retain
the
uncircled
species.*
If
'
Yes',
delete
the
uncircled
species.+

5.
Does
the
phylum
occur
at
the
site?
If
'
No',
delete
the
uncircled
species.*
If
WYesm,
does
the
dataset
contain
a
circled
species
that
is
in
the
same
phylum?
If
.
No',
retain
the
uncircled
species.+
If
.
Yes.,
delete
the
uncircled
species.+

l
=
Continue
the
deletion
process
by
starting
at
step
1
for
another
uncircled
species
unless
all
uncircled
species
in
the
dataset
have
been
considered.

The
species
that
are
circled
and
those
that
are
retained
constitute
the
site­
specific
dataset.
(
An
example
of
the
deletion
process
is
given
in
Figure
Bl.)

This
deletion
process
is
designed
to
ensure
that:
a.
Each
species
that
occurs
both
in
the
national
dataset
and
at
the
site
also
occurs
in
the
site­
specific
dataset.

94
b.
Each
species
that
occurs
at
the
site
but
does
not
occur
in
the
national
dataset
is
represented
in
the
site­
specific
dataset
by
u
species
in
the
national
dataset
that
are
in
the
same
genus.
C.
Each
genus
that
occurs
at
the
site
but
does
not
occur
in
the
national
dataset
is
represented
in
the
site­
specific
dataset
by
all
genera
in
the
national
dataset
that
are
in
the
same
fay.
d.
Each
order,
class,
and
phylum
that
occurs
both
in
the
national
dataset
and
at
the
site
is
represented
in
the
site­
specific
dataset
by
the
one
or
more
species
in
the
national
dataset
that
are
most
closely
related
to
a
species
that
occurs
at
the
site.

D.
Checkina
the
Minimum
Data
Reuuirements
The
initial
MDRs
for
the
Recalculation
Procedure
are
the
same
as
those
for
the
derivation
of
a
national
criterion.
If
a
specific
requirement
cannot
be
satisfied
after
deletion
because
that
kind
of
species
does
not
occur
at
the
site,
a
taxonomically
similar
species
muat
be
substituted
in
order
to
meet
the
eight
MDRs:

If
no
species
of
the
kind
required
occurs
at
the
site,
but
a
species
in
the
same
order
does,
the
MDR
can
only
be
satisfied
by
data
for
a
species
that
occurs
at
the
site
and
is
in
that
order;
if
no
species
in
the
order
occurs
at
the
site,
but
a
species
in
the
class
does,
the
MDR
can
only
be
satisfied
by
data
for
a
species
that
occurs
at
the
site
and
is
in
that
class.
If
no
species
in
the
same
class
occurs
at
the
site,
but
a
species
in
the
phylum
does,
the
MDR
can
only
be
satisfied
by
data
for
a
species
that
occurs
at
the
site
and
is
in
that
phylum.
If
no
species
in
the
same
phylum
occurs
at
the
site,
any
species
that
occurs
at
the
site
and
is
not
used
to
satisfy
a
different
MDR
can
be
used
to
satisfy
the
MDR.
If
additional
data
are
generated
using
acceptable
procedures
(
U.
S.
EPA
1985)
and
they
are
approved
by
the
U.
S.
EPA,
Recalculation
Procedure
must
be
started
again
at
step
the
addition
of
the
new
data.

If
fewer
than
eight
families
of
aquatic
invertebrates,
amphibians,
and
fishes
occur
at
the
site,
a
Species
Mean
Value
murt
be
available
for
at
least
one
species
in
each
the
B
with
Acute
of
the
families
and
the
special
version
of
the
Recalculation
Procedure
described
on
the
second
page
of
this
appendix
must
be
used.

E.
Determinina
the
CMC
and/
or
CCC
1.
Determining
the
FAV:
a.
If
the
eight
family
MDRs
are
satisfied,
the
site­
specific
FAV
muat
be
calculated
from
Genus
Mean
Acute
Values
using
95
the
procedure
described
in
the
national
aquatic
life
guidelines
(
U.
S.
EPA
1985).
b.
If
fewer
than
eight
families
of
aquatic
invertebrates,
amphibians,
and
fishes
occur
at
the
site,
the
lowest
Species
Mean
Acute
Value
that
is
available
for
a
species
that
occurs
at
the
site
mumt
be
used
as
the
FAV,
as
per
the
special
version
of
the
Recalculation
Procedure
described
on
the
second
page
of
this
appendix.
2.
The
site­
specific
CMC
nut
be
calculated
by
dividing
the
site­
specific
FAV
by
2.
The
site­
specific
FCV
rout
be
calculated
by
dividing
the
site­
specific
FAV
by
the
national
FACR
(
or
by
a
site­
specific
FACR
if
one
is
derived).
(
Because
a
FACR
was
not
used
to
derive
the
national
CCC
for
cadmium
in
fresh
water,
the
site­
specific
CCC
equals
the
site­
specific
FCV.)
3.
The
calculated
FAV,
CMC,
and/
or
CCC
mrmt
be
lowered,
if
necessary,
to
(
1)
protect
an
aquatic
plant,
invertebrate,
amphibian,
or
fish
species
that
is
a
critical
species
at
the
site,
and
(
2)
ensure
that
the
criterion
is
not
likely
to
jeopardize
the
continued
existence
of
any
endangered
or
threatened
species
listed
under
section
4
of
the
Endangered
Species
Act
or
result
in
the
destruction
or
adverse
modification
of
such
species'
critical
habitat.

.
.
Fe
Wrltlna
the
ReDort
The
report
of
the
results
of
use
of
the
Recalculation
Procedure
mm+
include:
1.
A
list
of
all
species
of
aquatic
invertebrates,
amphibians,
and
fishes
that
are
known
to
'
occur
at
the
site',
along
with
the
source
of
the
information.
2.
A
list
of
all
aquatic
plant,
invertebrate,
amphibian,
and
fish
species
that
are
critical
species
at
the
site,
including
all
species
that
occur
at
the
site
and
are
listed
as
threatened
or
endangered
under
section
4
of
the
Endangered
Species
Act.
3.
A
site­
specific
version
of
Table
1
from
a
criteria
document
produced
by
the
U.
S.
EPA
after
1984.
4.
A
site­
specific
version
of
Table
3
from
a
criteria
document
produced
by
the
U.
S.
EPA
after
1984.
5.
A
list
of
all
species
that
were
deleted.
6.
me
new
calculated
FAV,
CMC,,
and/
or
CCC.
7.
The
lowered
FAV,
CMC,
and/
or
CCC,
if
one
or
more
were
lowered
to
protect
a
specific
species.

U.
S.
EPA.
1985.
Guidelines
for
Deriving
Numerical
National
Water
Quality
Criteria
for
the
Protection
of
Aquatic
Organisms
and
Their
Uses.
PB85­
227049.
National
Technical
Information
Service,
Springfield,
VA.

96
Figure
81:
An
Xxapqple
of
thm
Dolotion
Procwm
IWing
Three
Phyla
SPECIES
THAT
ABE
IN
THE
THREE
phvlum
Class
Order
Annelida
Hirudin.
Bhynchob.
Bryozoa
(
No
species
in
this
Chordata
Osteich.
Cyprinif.
Chordata
Osteich.
Cyprinif.
Chordata
Osteich.
Cyprinif.
Chordata
Osteich.
Cyprinif.
Chordata
Osteich.
Salmonif.
Chordata
Osteich.
Percifor.
Chordata
Osteich.
Percifor.
Chordata
Amphibia
Caudata
PHYLAANDOCCURATTHE
SITE
Familv
SDecies
Glossiph.
Gloss+.
complanata
phylum
occur
at
the
site.)
Cyprinid.
Carassius
auratus
Cyprinid.
Notropis
anogenus
Cyprinid.
Phoxinus
eos
Catostom.
Carpiodes
carpio
Osmerida.
Osmerus
mordax
Centrarc.
Lepomis
cyanellus
Centrarc.
Legomis
humilis
Ambystom.
Ambystoma
gracile
SPECIES
THAT
ABE
IN
THE
THREE
phvlum
Class
Order
PHYLA
AND
IN
THE
NATIONAL
DATASET
Familv
Annelida
Oligoch.
HaplOtaX.
Bryozoa
Phylact.
­­­
Chordata
Cephala.
Petromyz.
Chordata
Osteich.
Cyprinif.
Chordata
Osteich.
Cyprinif.
Chordata
Osteich.
Cyprinif.
Chordata
Osteich.
Cyprinif.
Chordata
Osteich.
Cyprinif.
Chordata
Osteich.
Cyprinif.
Chordata
Osteich.
Cyprinif.
Chordata
Osteich.
Salmonif.
Chordata
Osteich.
Percifor.
Chordata
Osteich.
Percifor.
Chordata
Osteich.
Percifor.
Chordata
Amphibia
Anura
Tubifici.
LoDhODod.
Petromyz.
Cyprinid.
Cyprinid.
Cyprinid.
Cyprinid.
Cyprinid.
Cyprinid.
Catostom.
Salmonid.
Centrarc.
Centrarc.
Percidae
Pigidae
mecies
Coda
Tubifextubifex
P
Loghopod.
carteri
D
Petromyzon
marinus
D
Carassius
auratus
S
Notropis
hudsonius
G
Notropis
stramineus
G
Phoxinus
eos
S
Phoxinus
oreas
D
Tinca
tinca
D
Ictiobus
bubalus
F
Oncorhynchus
mykiss
0
Lepomis
cyanellus
S
Legomis
macrochirus
G
Perca
flavescens
D
Xenopus
laevis
C
Explanations
of
Codes:
s=
retained
because
this
Species
occurs
at
the
site.
G
=
retained
because
there
is
a
species
in
this
Genus
that
occurs
at
the
site
but
not
in
the
national
dataset.
F
=
retained
because
there
is
a
genus
in
this
Family
that
occurs
at
the
site
but
not
in
the
national
dataset.
o=
retained
because
this
Order
occurs
at
the
site
and
is
not
represented
by
a
lower
taxon.
c=
retained
because
this
Class
occurs
at
the
site
and
is
not
represented
by
a
lower
taxon.
P
=
retained
because
this
Phylum
occurs
at
the
site
and
is
not
represented
by
a
lower
taxon.
D
=
deleted
because
this
species
does
not
satisfy
any
of
the
requirements
for
retaining
species.

97
AppMdix
C:
alaidaaco
ConcorPing
the
Us.
of
'
Clam
Toohniquo8=
Ipd
QA/
QC
when
Meamuing
Tram
Dfotalm
Note:
This
version
of
this
appendix
contains
more
information
than
the
version
that
was
Appendix
B
of
Prothro
(
1993).

Recent
information
(
Shiller
and
Boyle
1987;
Windom
et
al.
1991)
has
raised
questions
concerning
the
quality
of
reported
concentrations
of
trace
metals
in
both
fresh
and
salt
(
estuarine
and
marine)
surface
waters.
A
lack
of
awareness
of
true
ambient
concentrations
of
metals
in
fresh
and
salt
surface
waters
can
be
both
a
cause
and
a
result
of
the
problem.
The
ranges
of
dissolved
metals
that
are
typical
in
surface
waters
of
the
United
States
away
from
the
immediate
influence
of
discharges
(
Bruland
1983;
Shiller
and
Boyle
1985,1987;
Trefry
et
al.
1986;
Windom
et
al.
1991)
are:

Metal
Salt
water
Fresh
water
(
uu/
L)
(
w/
L)

Cadmium
0.01
to
0.2
0.002
to
0.08
Copper
0.1
to
3.
0.4
to
4.
Lead
0.01
to
1.
0.01
to
0.19
Nickel
0.3
to
5.
1.
to
2.
Silver
0.005
to
0.2
­­­­­­­­­­­­­
Zinc
0.1
to
15.
0.03
to
5.

The
U.
S.
EPA
(
1983,1991)
has
published
analytical
methods
for
monitoring
metals
in
waters
and
wastewaters,
but
these
methods
are
inadequate
for
determination
of
ambient
concentrations
of
some
metals
in
some
surface
waters.
Accurate
and
precise
measurement
of
these
low
concentrations
requires
appropriate
attention
to
seven
areas:
1.
Use
of
.
clean
techniques'
during
collecting,
handling,
storing,
preparing,
and
analyzing
samples
to
avoid
contamination.
2.
Use
of
analytical
methods
that
have
sufficiently
low
detection
limits.
3.
Avoidance
of
interference
in
the
quantification
(
instrumental
analysis)
step.
4.
Use
of
blanks
to
assess
contamination.
5.
Use
of
matrix
spikes
(
sample
spikes)
and
certified
reference
materials
(
CRMs)
to
assess
interference
and
contamination.
6.
Use
of
replicates
to
assess
precision.
7.
Use
of
certified
standards.
In
a
strict
sense,
the
term
#
clean
techniques'
refers
to
techniques
that
reduce
contamination
and
enable
the
accurate
and
precise
measurement
of
trace
metals
in
fresh
and
salt
surface
waters.
In
a
broader
sense,
the
term
also
refers
to
related
issues
concerning
detection
limits,
quality
control,
and
quality
98
assurance.
Documenting
data
quality
demonstrates
the
amount
of
confidence
that
can
be
placed
in
the
data,
whereas
increasing
the
sensitivity
of
methods
reduces
the
problem
of
deciding
how
to
interpret
results
that
are
reported
to
be
below
detection.
limits.

This
aDDendiX
is
written
for
those
analvtical
laboratories
that
.
want
aldance
concernina
wavs
to
lower
detection
limits.
increase
accu
acv.
a
d/
o
.
c
e
s
Drecision.
The
ways
to
achieve
these
goal:
are
ti
in&:
isz
:
hz
sensitivity
of
the
analytical
methods,
decrease
contamination,
and
decrease
interference.
Ideally,
validation
of
a
procedure
for
measuring
concentrations
of
metals
in
surface
water
requires
demonstration
that
agreement
can
be
obtained
using
completely
different
procedures
beginning
with
the
sampling
step
and
continuing
through
the
quantification
step
(
Bruland
et
al.
1979),
but
few
laboratories
have
the
resources
to
compare
two
different
procedures.
Laboratories
can,
however,
(
a)
use
techniques
that
others
have
found
useful
for
improving
detection
limits,
accuracy,
and
precision,
and
(
b)
document
data
quality
through
use
of
blanks,
spikes,
CRMs,
replicates,
and
standards.
.
.
NJoth=
contained
or
not
contained
in
this
aDDendiX
adds
to
o
subtracts
from
anv
reaulatorv
reouirement
set
forth
in
other
EPA
docume
ts
conce
'
a
analyses
of
metals.
A
WER
can
be
acceptably
determ?
ned
withE?
the
use
of
clean
techniques
as
long
as
the
detection
limits,
accuracy,
and
precision
are
acceptable.
No
QA/
QC
requirements
beyond
those
that
apply
to
measuring
metals
in
effluents
are
necessary
for
the
determination
of
WEI&.
The
word
'
must'
is
not
used
in
this
appendix.
Some
items,
however,
are
considered
so
important
by
analytical
chemists
who
have
worked
to
increase
accuracy
and
precision
and
lower
detection
limits
in
trace­
metal
analysis
that
'
mhouldn
is
in
bold
print
to
draw
attention
to
the
item.
Most
such
items
are
emphasized
because
they
have
been
found
to
have
received
inadequate
attention
in
some
laboratories
performing
trace­
metal
analyses.

In
general,
in
order
to
achieve
accurate
and
precise
measurement
of
a
particular
concentration,
both
the
detection
limit
and
the
blanks
should
be
less
than
one­
tenth
of
that
concentration.
Therefore,
the
term
.
metal­
free'
can
be
interpreted
to
mean
that
the
total
amount
of
contamination
that
occurs
during
sample
collection
and
processing
(
e.
g.,
from
gloves,
sample
containers,
labware,
sampling
apparatus,
cleaning
solutions,
air,
reagents,
etc.)
is
sufficiently
low
that
blanks
are
less
than
one­
tenth
of
the
lowest
concentration
that
needs
to
be
measured.

Atmospheric
particulates
can
be
a
major
source
of
contamination
(
Moody
1982;
Adeloju
and
Bond
1985).
The
term
'
class­
100"
refers
to
a
specification
concerning
the
amount
of
particulates
in
air
(
Moody
1982)
;
although
the
specification
says
nothing
about
the
composition
of
the
particulates,
generic
control
of
particulates
can
greatly
reduce
trace­
metal
blanks.
Except
during
collection
99
of
samples,
initial
cleaning
of
equipment,
and
handling
of
samples
containing
high
concentrations
of
metals,
all
handling
of
samples,
sample
containers,
labware,
and
sampling
apparatus
should
be
performed
in
a
class­
100
bench,
room,
or
glove
box.

Neither
the
'
ultraclean
techniques'
that
might
be
necessary
when
trace
analyses
of
mercury
are
performed
nor
safety
in
analytical
laboratories
is
addressed
herein.
Other
documents
should
be
consulted
if
one
or
both
of
these
topics
are
of
concern.

v*
l
.
.
*

Measurement
of
trace
metals
in
surface
waters
should
take
into
account
the
potential
for
contamination
during
each
step
in
the
process.
Regardless
of
the
specific
procedures
used
for
collection,
handling,
storage,
preparation
(
digestion,
filtration,
and/
or
extraction),
and
quantification
(
instrumental
analysis),
the
general
principles
of
contamination
control
should
be
a.

b.

C.

d.

e.

f.
applied.
Some
specific
reconnnendations
are:
Powder­
free
(
non­
talc,
class­
1001
latex,
polyethylene,
or
polyvinyl
chloride
(
PVC,
vinyl)
gloves
should
be
worn
during
all
steps
from
sample
collection
to
analysis.
(
Talc
seems
to
be
a
particular
problem
with
zinc;
gloves
made
with
talc
cannot
be
decontaminated
sufficiently.)
Gloves
should
only
contact
surfaces
that
are
metal­
free;
gloves
should
be
changed
if
even
suspected
of
contamination.
The
acid
used
to
acidify
samples
for
preservation
and
digestion
and
to
acidify
water
for
final
cleaning
of
labware,
sampling
apparatus,
and
sample
containers
should
be
metal­
free.
The
quality
of
the
acid
used
should
be
better
than
reagent­
grade.
Each
lot
of
acid
should
be
analyzed
for
the
metal(
s)
of
interest
before
use.
The
water
used
to
prepare
acidic
cleaning
solutions
and
to
rinse
la&
are,
sample
containers,
and
sampling
apparatus
may
be
prepared
by
distillation,
deionization,
or
reverse
osmosis,
and
&
aould
be
demonstrated
to
be
metal­
free.
The
work
area,
including
bench
tops
and
hoods,
should
be
cleaned
(
e.
g.,
washed
and
wiped
dry
with
lint­
free,
class­
100
wipes)
frequently
to
remove
contamination.
All
handling
of
samples
in
the
laboratory,
including
filtering
and
analysis,
should
be
performed
in
a
class­
100
clean
bench
or
a
glove
box
fed
by
particle­
free
air
or
nitrogen;
ideally
the
clean
bench
or
glove
box
should
be
located
within
a
class­
100
clean
room.
Labware,
reagents,
sampling
apparatus,
and
sample
containers
should
never
be
left
open
to
the
atmosphere;
they
should
be
stored
in
a
class­
100
bench,
covered
with
plastic
wrap,
stored
in
a
plastic
box,
or
turned
upside
down
on
a
clean
surface.
Minimizing
the
time
between
cleaning
and
using
will
help
minimize
contamination.

100
g.
Separate
sets
of
sample
containers,
labware,
and
sampling
apparatus
should
be
dedicated
for
different
kinds
of
samples,
.
surface
water
samples
effluent
samples,
etc.
h.
Gog&
oid
contamination
of
clean
rooms,
samples
that
contain
very
high
concentrations
of
metals
and
do
not
require
use
of
@
clean
techniques'
should
not
be
brought
into
clean
rooms.
i.
Acid­
cleaned
plastic,
such
as
high­
density
polyethylene
(
HDPE)
,
low­
density
polyethylene
(
LDPE),
or
a
fluoroplastic,
should
be
the
only
material
that
ever
contacts
a
sample,
except
possibly
during
digestion
for
the
total
recoverable
measurement.
1.
Total
recoverable
samples
can
be
digested
in
some
plastic
containers.
2.
HDPE
and
LDPE
might
not
be
acceptable
for
mercury.
3.
Even
if
acidified,
samples
and
standards
containing
silver
should
be
in
amber
containers.
j.
All
labware,
sample
containers,
and
sampling
apparatus
8hould
be
acid­
cleaned
before
use
or
reuse.
1.
Sample
containers,
sampling
apparatus,
tubing,
membrane
filters,
filter
assemblies,
and
other
labware
rhould
be
soaked
in
acid
until
metal­
free.
The
amount
of
cleaning
necessary
might
depend
on
the
amount
of
contamination
and
the
length
of
time
the
item
will
be
in
contact
with
samples.
For
example,
if
an
acidified
sample
will
be
stored
in
a
sample
container
for
three
weeks,
ideally
the
container
should
have
been
soaked
in
an
acidified
metal­
free
solution
for
at
least
three
weeks.
2.
It
might
be
desirable
to
perform
initial
cleaning,
for
which
reagent­
grade
acid
may
be
used,
before
the
items
are
taken
into
a
clean
room.
For
most
metals,
items
should
be
either
(
a)
soaked
in
10
percent
concentrated
nitric
acid
at
50
°
C
for
at
least
one
hour,
or
(
b)
soaked
in
50
percent
concentrated
nitric
acid
at
room
temperature
for
at
least
two
days;
for
arsenic
and
mercury,
soaking
for
up
to
two
weeks
at
50
°
C
in
10
percent
concentrated
nitric
acid
might
be
required.
For
plastics
that
might
be
damaged
by
strong
nitric
acid,
such
as
polycarbonate
and
possibly
HDPE
and
LDPE,
soaking
in
10
percent
concentrated
hydrochloric
acid,
either
in
place
of
or
before
soaking
in
a
nitric
acid
solution,
might
be
desirable.
3.
Chromic
acid
should
not
be
used
to
clean
items
that
will
be
used
in
analysis
of
metals.
4.
Final
soaking
and
cleaning
of
sample
containers,
labware,
and
sampling
apparatus
should
be
performed
in
a
class­
100
clean
room
using
metal­
free
acid
and
water.
The
solution
in
an
acid
bath
8hould
be
analyzed
periodically
to
demonstrate
that
it
is
metal­
free.
k.
Labware,
sampling
apparatus,
and
sample
containers
should
be
stored
appropriately
after
cleaning:
1.
After
the
labware
and
sampling'apparatus
are
cleaned,
they
may
be
stored
in
a
clean
room
in
a
weak
acid
bath
prepared
using
metal­
free
acid
and
water.
Before
use,
the
items
101
should
be
rinsed
at
least
three
times
with
metal­
free
water.
After
the
final
rinse,
the
items
should
be
moved
ixmnediately,
with
the
open
end
pointed
down,
to
a
class­
100
clean
bench.
Items
may
be
dried
on
a
class­
100
clean
bench;
items
nhould
not
be
dried
in
an
oven
or
with
laboratory
towels.
The
sampling
apparatus
should
be
assembled
in
a
class­
100
clean
room
or
bench
and
double­
bagged
in
metal­
free
polyethylene
zip­
type
bags
for
transport
to
the
field;
new
bags
are
usually
metal­
free.
2.
After
sample
containers
are
cleaned,
they
should
be
filled
with
metal­
free
water
that
has
been
acidified
to
a
pH
of
2
with
metal­
free
nitric
acid
(
about
0.5
mL
per
liter)
for
storage
until
use.
1.
Labware,
sampling
apparatus,
and
sample
containers
Should
be
rinsed
and
not
rinsed
with
sample
as
necessary
to
prevent
high
and
low
bias
of
analytical
results
because
acid­
cleaned
plastic
will
sorb
some
metals
from
unacidified
solutions.
1.
Because
samples
for
the
dissolved
measurement
are
not
acidified
until
after
filtration,
all
sampling
apparatus,
sample
containers,
labware,
filter
holders,
membrane
filters,
etc.,
that
contact
the
sample
before
or
during
filtration
should
be
rinsed
with
a
portion
of
the
solution
and
then
that
portion
discarded.
2.
For
the
total
recoverable
measurement,
labware,
etc.,
that
contact
the
sample
onlv
before
it
is
acidified
should
be
rinsed
with
sample,
whereas
items
that
contact
the
sample
after
it
is
acidified
should
not
be
rinsed.
For
example,
the
sampling
apparatus
should
be
rinsed
because
the
sample
will
not
be
acidified
until
it
is
in
a
sample
container,
but
the
sample
container
should
not
be
rinsed
if
the
sample
will
be
acidified
in
the
sample
container.
3.
If
the
total
recoverable
and
dissolved
measurements
are
to
be
performed
on
the
same
sample
(
rather
than
on
two
samples
obtained
at
the
same
time
and
place),
all
the
apparatus
and
labware,
including
the
sample
container,
should
be
rinsed
before
the
sample
is
placed
in
the
sample
container;
then
an
unacidified
aliguot
should
be
removed
for
the
total
recoverable
measurement
(
and
acidified,
digested,
etc.)
and
an
unacidified
aliguot
should
be
removed
for
the
dissolved
measurement
(
and
filtered,
acidified,
etc.)
(
If
a
container
is
rinsed
and
filled
with
sample
and
an
unacidified
aliguot
is
removed
for
the
dissolved
measurement
and
then
the
solution
in
the
container
is
acidified
before
removal
of
an
aliguot
for
the
total
recoverable
measurement,
the
resulting
measured
total
recoverable
concentration
might
be
biased
high
because
the
acidification
might
desorb
metal
that
had
been
sorbed
onto
the
walls
of
the
sample
container;
the
amount
of
bias
will
depend
on
the
relative
volumes
jnmlved
and
on
the
amount
of
sorption
and
desorption.)
m.
Field
samples
Should
be
collected
in
a
manner
that
eliminates
the
potential
for
contamination
from
sampling
platforms,

102
n.

0.

P*

Q*

r.

S.

t.
probes,
etc.
Exhaust
from
boats
and
the
direction
of
wind
and
water
currents
should
be
taken
into
account.
The
people
who
collect
the
samples
&
ould
be
specifically
trained
on
how
to
collect
field
samples.
After
collection,
all
handling
of
samples
in
the
field
that
will
expose
the
sample
to
air
ahould
be
performed
in
a
portable
class­
100
clean
bench
or
glove
box.
Samples
should
be
acidified
(
after
filtration
if
dissolved
metal
is
to
be
measured)
to
a
pH
of
less
than
2,
except
that
the
pH
should
be
less
than
1
for
mercury.
Acidification
should
be
done
in
a
clean
room
or
bench,
and
so
it
might
be
desirable
to
wait
and
acidify
samples
in
a
laboratory
rather
than
in
the
field.
If
samples
are
acidified
in
the
field,
metal­
free
acid
can
be
transported
in
plastic
bottles
and
poured
into
a
plastic
container
from
which
acid
can
be
removed
and
added
to
samples
using
plastic
pipettes.
Alternatively,
plastic
automatic
dispensers
can
be
used.
Such
things
as
probes
and
thermometers
l
hould
xaot
be
put
in
samples
that
are
to
be
analyzed
for
metals.
In
particular,
pH
electrodes
and
mercury­
in­
glass
thermometers
should
not
be'
used
if
mercury
is
to
be
measured.
If
pH
is
measured,
it
should
be
done
on
a
separate
aliguot.
Sample
handling
should
be
minimized.
For
example,
instead
of
pouring
a
sample
into
a
graduated
cylinder
to
measure
the
volume,
the
sample
can
be
weighed
after
being
poured
into
a
tared
container,
which
is
less
likely
to
be
subject
to
error
than
weighing
the
container
from
which
the
sample
is
poured.
(
For
saltwater
samples,
the
salinity
or
density
should
be
taken
into
account
if
weight
is
converted
to
volume.)
Each
reagent
used
mhould
be
verified
to
be
metal­
free.
If
metal­
free
reagents
are
not
commercially
available,
removal
of
metals
will
probably
be
necessary.
For
the
total
recoverable
measurement,
samples
should
be
digested
in
a
class­
100
bench,
not
in
a
metallic
hood.
If
feasible,
digestion
should
be
done
in
the
sample
container
by
acidification
and
heating.
The
longer
the
time
between
collection
and
analysis
of
samples,
the
greater
the
chance
of
contamination,
loss,
etc.
Samples
should
be
stored
in
the
dark,
preferably
between
0
and
4OC
with
no
air
space
in
the
sample
container.

Achievina
low
detection
limits
a.
Extraction
of
the
metal
from
the
sample
can
be
extremely
useful
if
it
simultaneously
concentrates
the
metal
and
eliminates
potential
matrix
interferences.
For
example,
ammonium
I­
pyrrolidinedithiocarbamate
and/
or
diethylammonium
diethyldithiocarbamate
can
extract
cadmium,
copper,
lead,
nickel,
and
zinc
(
Bruland
et
al.
1979;
Nriagu
et
al.
1993).
b.
The
detection
limit
should
be
less
than
ten
percent
of
the
lowest
concentration
that
is
to
be
measured.

103
a.
Potential
interferences
should
be
assessed
for
the
specific
instrumental
analysis
technique
used
and
for
each
metal
to
be
measured.
b.
If
direct
analysis
is
used,
the
salt
present
in
high­
salinity
saltwater
samples
is
likely
to
cause
interference
in
most
instrumental
techniques.
c.
As
stated
above,
extraction
of
the
metal
from
the
sample
is
particularly
useful
because
it
simultaneously
concentrates
the
metal
and
eliminates
potential
matrix
interferences.

a.
A
laboratory
(
procedural,
method)
blank
consists
of
filling
a
sample
container
with
analyzed
metal­
free
water
and
processing
(
filtering,
acidifying,
etc.)
the
water
through
the
laboratory
procedure
in
exactly
the
same
way
as
a
sample.
A
laboratory
blank
should
be
included
in
each
set
of
ten
or
fewer
samples
to
check
for
contamination
in
the
laboratory,
and
should
contain
less
than
ten
percent
of
the
lowest
concentration
that
is
to
be
measured.
Separate
laboratory
blanks
&
ould
be
processed
for
the
total
recoverable
and
dissolved
measurements,
if
both
measurements
are
performed.
b.
A
field
(
trip)
blank
consists
of
filling
a
sample
container
with
analyzed
metal­
free
water
in
the
laboratory,
taking
the
container
to
the
site,
processing
the
water
through
tubing,
filter,
etc.,
collecting
the
water
in
a
sample
container,
and
acidifying
the
water
the
same
as
a
field
sample.
A
field
blank
mhould
be
processed
for
each
sampling
trip.
Separate
field
blanks
should
be
processed
for
the
total
recoverable
measurement
and
for
the
dissolved
measurement,
if
filtrations
are
performed
at
the
site.
Field
blanks
l
hauld
be
processed
in
the
laboratory
the
same
as
laboratory
blanks.

Assessinu
accuracy
a.
A
calibration
curve
should
be
determined
for
each
analytical
run
and
the
calibration
should
be
checked
about
every
tenth
sample.
Calibration
solutions
should
be
traceable
back
to
a
certified
standard
from
the
U.
S.
EPA
or
the
National
Institute
of
Science
and
Technology
(
NIST).
b.
A
blind
standard
or
a
blind
calibration
solution
hould
be
included
in
each
group
of
about
twenty
samples.
c.
At
least
one
of
the
following
should
be
included
in
each
group
of
about
twenty
samples:
1.
A
matrix
spike
(
spiked
sample;
the
method
of
known
additions).

104
2.
A
CPM,
if
one
is
available
in
a
matrix
that
closely
approximates
that
of
the
samples.
Values
obtained
for
the
CPM
should
be
within
the
published
values.
The
concentrations
in/
blind
standards
and
solutions,
spikes,
and
CRMs
should
not
be
more
than
5
times
the
median
concentration
expected
to
be
present
in
the
samples.

a.
A
sampling
replicate
should
be
included
with
each
set
of
samples
collected
at
each
sampling
location.
b.
If
the
volume
of
the
sample
is
large
enough,
replicate
analysis
of
at
least
one
sample
l
hould
be
performed
along
with
each
group
of
about
ten
samples.

special
considerations
concernina
the
dissolved
measurement
Whereas
total
recoverable
measurements
are
especially
subject
to
contamination
during
digestion,
dissolved
measurements
are
subject
to
both
loss
and
contamination
during
filtration.
a.
Because
acid­
cleaned
plastic
sorbs
metal
from
unacidified
solutions
and
because
samples
for
the
dissolved
measurement
are
not
acidified
before
filtration,
all
sampling
apparatus,
sample
containers,
labware,
filter
holders,
and
membrane
filters
that
contact
the
sample
before'or
during
filtration
should
be
conditioned
by
rinsing
with
a
portion
of
the
solution
and
discarding
that
portion.
b.
Filtrations
should
be
performed
using
acid­
cleaned
plastic
filter
holders
and
acid­
cleaned
membrane
filters.
Samples
should
not
be
filtered
through
glass
fiber
filters,
even
if
the
filters
have
been
cleaned
with
acid.
If
positive­
pressure
filtration
is
used,
the
air
or
gas
should
be
passed
through
a
0.2­
v
in­
line
filter;
if
vacuum
filtration
is
used,
it
should
be
performed
on
a
class­
100
bench.
C.
Plastic
filter
holders
should
be
rinsed
and/
or
dipped
between
filtrations,
but
they
do
not
have
to
be
soaked
between
filtrations
if
all
the
samples
contain
about
the
same
concentrations
of
metal.
It
is
best
to
filter
samples
from
low
to
high
concentrations.
A
membrane
filter
&
ould
not
be
used
for
more
than
one
filtration.
After
each
filtration,
the
membrane
filter
should
be
removed
and
discarded,
and
the
filter
holder
should
be
either
rinsed
with
metal­
free
water
or
dilute
acid
and
dipped
in
a
metal­
free
acid
bath
or
rinsed
at
least
twice
with
metal­
free
dilute
acid;
finally,
the
filter
holder
ohould
be
rinsed
at
least
twice
with
metal­
free
water.
d.
For
each
sample
to
be
filtered,
the
filter
holder
and
membrane
filter
should
be
conditioned
with
the
sample,
i.
e.,
an
initial
portion
of
the
sample
ohould
be
filtered
and
discarded.

105
The
accuracy
and
precision
of
the
dissolved
measurement
should
be
assessed
periodically.
A
large
volume
of
a
buffered
solution
(
such
as
aerated
0.05
N
sodium
bicarbonate
for
analyses
in
fresh
water
and
a
combination
of
sodium
bicarbonate
and
sodium
chloride
for
analyses
in
salt
water)
should
be
spiked
so
that
the
concentration
of
the
metal
of
interest
is
in
the
range
of
the
low
concentrations
that
are
to
be
measured.
Sufficient
samples
should
be
taken
alternately
for
(
a)
acidification
in
the
same
way
as
after
filtration
in
the
dissolved
method
and
(
b)
filtration
and
acidification
using
the
procedures
specified
in
the
dissolved
method
until
ten
samples
have
been
processed
in
each
Way.
The
concentration
of
metal
in
each
of
the
twenty
samples
should
then
be
determined
using
the
same
analytical
procedure.
The
means
of
the
two
groups
of
ten
measurements
should
be
within
10
percent,
and
the
coefficient
of
variation
for
each
group
of
ten
should
be
less
than
20
percent.
Any
values
deleted
as
outliers
should
be
acknowledged.

To
indicate
the
quality
of
the
data,
reports
of
results
of
measurements
of
the
concentrations
of
metals
should
include
a
description
of
the
blanks,
.
spikes,
C??
Ms,
replicates,
and
standards
that
were
run,
the
number
run,
and
the
results
obtained.
All
values
deleted
as
outliers
&
aould
be
acknowledged.

The
items
presented
above
are
some
of
the
important
aspects
of
'
clean
techniques';
some
aspects
of
quality
assurance
and
quality
control
are
also
presented.
This
is
not
a
definitive
treatment
of
these
topics;
additional
information
that
might
be
useful
is
available
in
such
publications
as
Patterson
and
Settle
(
1976),
Zief
and
Mitchell
(
1976),
Bruland
et
al.
(
1979),
Moody
and
Beary
(
1982),
Moody
(
1982),
Bruland
(
1983),
Adeloju
and
Bond
(
1985),
Berman
and
Yeats
(
1985),
Byrd
and
Andreae
(
1986),
Taylor
(
1987),
Sakamoto­
Arnold
(
1987),
Tramontano
et
al.
(
1987),
Puls
and
Barcelona
(
1989),
Windom
et
al.
(
1991),
U.
S.
EPA
(
1992),
Horowitz
et
al.
(
1992),
and
Nriagu
et
al.
(
1993).

106
References
Adeloju,
S.
B.,
and
A.
M.
Bond.
1985.
Influence
of
Laboratory
Environment
on
the
Precision
and
Accuracy
of
Trace
Element
Analysis.
Anal.
Chem.
57:
1728­
1733.

Berman,
S.
S.,
and
P.
A.
Yeats.
1985.
Sampling
of
Seawater
for
Trace
Metals.
CRC
Reviews
in
Analytical
Chemistry
16:
1­
14.

Bruland,
K.
W.,
R.
P.
Franks,
G.
A.
Knauer,
and
J.
H.
Martin.
1979.
Sampling
and
Analytical
Methods
for
the
Determination
of
Copper,
Cadmium,
Zinc,
and
Nickel
at
the
Nanogram
per
Liter
Level
in
Sea
Water.
Anal.
Chim.
Acta
105:
233­
245.

Bruland,
K.
W.
1983.
Trace
Elements
in
Sea­
water.
In:
Chemical
Oceanography,
Vol.
8.
(
J.
P.
Riley
and
R.
Chester,
eds.)
Academic
Press,
.
New
York,
NY.
pp.
157­
220.

Byrd,
J.
T.,
and
M.
O.
Andreae.
1986.
Dissolved
and
Particulate
Tin
in
North
Atlantic
Seawater.
Marine
Chem.
19:
193­
200.

Horowitz,
A.
J.,
K.
A.
Elrick,
and
M.
R.
Colberg.
1992.
The
Effect
of
Membrane
Filtration
Artifacts
on
Dissolved
Trace
Element
Concentrations.
Water
Res.
26:
753­
763.

Moody,
J.
R.
1982.
NBS
Clean
Laboratories
for
Trace
Element
Analysis.
Anal.
Chem.
54:
1358A­
1376A.

Moody,
J.
R.,
and
E.
S.
Beary.
1982.
Purified
Reagents
for
Trace
Metal
Analysis.
Talanta
29:
1003­
1010.

Nriagu,
J.
O.,
G.
Lawson,
H.
K.
T.
Wong,
and
J.
M.
Azcue.
1993.
A
Protocol
for
Minimizing
Contamination
in
the
Analysis
of
Trace
Metals
in
Great
Lakes
Waters.
J.
Great
Lakes
Res.
19:
175­
182.

Patterson,
C.
C.,
and
D.
M.
Settle.
1976.
The
Reduction
in
Orders
of
Magnitude
Errors
in
Lead
Analysis
of
Biological
Materials
and
Natural
Waters
by
Evaluating
and
Controlling
the
Extent
and
Sources
of
Industrial
Lead
Contamination
Introduced
during
Sample
Collection
and
Processing.
In:
Accuracy
in
Trace
Analysis:
Sampling,
Sample
Handling,
Analysis.
(
P.
D.
LaFleur,
ed.)
National
Bureau
of
Standards
Spec.
Publ.
422,
U.
S.
Government
Printing
Office,
Washington,
DC.

Prothro,
M.
G.
1993.
Memorandum
titled
'
Office
of
Water
Policy
and
Technical
Guidance
on
Interpretation
and
Implementation
of
Aquatic
Life
Metals
Criteria'.
October
1.

Puls,
R.
W.,
and
M.
J.
Barcelona.
1989.
Ground
Water
Sampling
for
Metals
Analyses.
EPA/
540/
4­
89/
001.
National
Technical
Information
Service,
Springfield,
VA.

107
Sakamoto­
Arnold,
C.
M.,
A.
K.
Hanson,
Jr.,
D.
L.
Huizenga,
and
D.
R.
Kester.
1987.
Spatial
and
Temporal
Variability
of
Cadmium
in
Gulf
Stream
Warm­
core
Rings
and
Associated
Waters.
J.
Mar.
Res.
45:
201­
230.

Shiller,
A.
M.,
and
E.
Boyle.
1985.
Dissolved
Zinc
in
Rivers.
Nature
317:
49­
52.

Shiller,
A.
M.,
and
E.
A.
Boyle.
1987.
Variability
of
Dissolved
Trace
Metals
in
the
Mississippi
River.
Geochim.
Cosmochim.
Acta
51:
3273­
3277.

Taylor,
J.
K.
1987.
Quality
Assurance
of
Chemical
Measurements.
Lewis
Publishers,
Chelsea,
MI.

Tranmntano,
J.
M.,
J.
R.
Scudlark,
and
T.
M.
Church.
1987.
A
Method
for
the
Collection,
Handling,
in
Precipitation.
and
Analysis
of
Trace
Metals
Environ.
Sci.
Technol.
21:
749­
753.

Trefzy,
J.
H.,
T.
A.
Nelsen,
R.
P.
Trocine,
S.
Metz.,
and
T.
W.
Vetter.
1986.
Delta
System.
Trace
Metal
Fluxes
through
the
Mississippi
River
Rapp.
P.­
v.
Reun.
Cons.
int.
Explor.
Mer.
186:
277­
288.

U.
S.
EPA.
1983.
Methods
for
Chemical
Analysis
of
Water
and
Wastes.
EPA­
600/
4­
79­
020.
National
Technical
Information
Service,
Springfield,
VA.
Sections
4.1.1,
4.1.3,
and
4.1.4
U.
S.
EPA.
1991.
Methods
for
the
Determination
of
Metals
in
Environmental
Samples.
EPA­
600/
4­
91­
010.
National
Technical
Information
Service,
Springfield,
VA.

U.
S.
EPA.
1992.
Evaluation
of
Trace­
Metal
Levels
in
Ambient
Waters
and
Tributaries
to
New
York/
New
Jersey
Harbor
for
Waste
Load
Allocation.
Prepared
by
Battelle
Ocean
Sciences
under
Contract
No.
68­
C8­
0105.

Windom,
H.
L.,
J.
T.
Byrd,
R.
G.
Smith,
and
F.
Huan.
1991.
Inadequacy
of
NASQAN
Data
for
Assessing
Metals
Trends
in
the
Nation's
Rivers.
Environ.
Sci.
Technol.
25:
1137­
1142.
(
Also
see
the
c­
t
and
response:
Environ.
Sci.
Technol.
25:
1940­
1941.)

Fief,
M.,
and
J.
W.
Mitchell.
1976.
Contamination
Control
in
Trace
Element
Analysis.
Chemical
Analysis
Series,
Vol.
47.
Wiley,
New
York,
NY.

108
Appendix
D:
Relationships
between
WXRm
and
the
Chamimtry
and
Toxicology
of
M&
ala
The
aquatic
toxicology
of
metals
is
complex
in
part
because
the
chemistry
of
metals
in
water
is
complex.
Metals
usually
exist
in
surface
water
in
various
combinations
of
particulate
and
dissolved
forms,
some
of
which
are
toxic
and
some
of
which
are
nontoxic.
In
addition,
all
toxic
forms
of
a
metal
are
not
necessarily
equally
toxic,
and
various
water
quality
characteristics
can
affect
the
relative
concentrations
and/
or
toxicities
of
some
of
the
forms.

The
toxicity
of
a
metal
has
sometimes
been
reported
to
be
proportional
to
the
concentration
or
activity
of
a
specific
species
of
the
metal.
For
example,
Allen
and
Hansen
(
1993)
surmnarized
reports
by
several
investigators
that
the
toxicity
of
copper
is
related
to
the
free
cupric
ion,
but
other
data
do
not
support
a
correlation
(
Erickson
1993a).
For
example,
Borgmann
(
1983),
Chapman
and
McCrady
(
19771,
and
French,
and
Hunt
(
1986)
found
that
toxicity
expressed
on
the
basis
of
cupric
ion
activity
varied
greatly
with
pH,
and
Cowan
et
al.
(
1986)
concluded
that
at
least
one
of
the
copper
hydroxide
species
is
toxic.
Further,
chloride
and
sulfate
salts
of
calcium,
magnesium,
potassium,
and
sodium
affect
the
toxicity
of
the
cupric
ion
(
Nelson
et
al.
1986).
Similarly
for
aluminum,
Wilkinson
et
al.
(
1993)
concluded
that
'
mortality
was
best
predicted
not
by
the
free
Al'*
activity
but
rather
as
a
function
of
the
sum
I:([
A13']
+
[
AlF'*])"
and
that
.
no
longer
can
the
reduction
of
Al
toxicity
in
the
presence
of
organic
acids
be
interpreted
simply
as
a
consequence
of
the
decrease
in
the
free
Al'*
concentration'.

Until
a
model
has
been
demonstrated
to
explain
the
quantitative
relationship
between
chemical
and
toxicological
measurements,
aquatic
life
criteria
should
be
established
in
an
environmentally
conservative
manner
with
provision
for
site­
specific
adjustment.
Criteria
should
be
expressed
in
terms
of
feasible
analytical
measurements
that
provide
the
necessary
conservatism
without
substantially
increasing
the
cost
of
implementation
and
site­
specific
adjustment.
Thus
current
aquatic
life
criteria
for
metals
are
expressed
in
terms
of
the
total
recoverable
measurement
and/
or
the
dissolved
measurement,
rather
than
a
measurement
that
would
be
more
difficult
to
perform
and
would
still
require
empirical
adjustment.
The
WER
is
operationally
defined
in
terms
of
chemical
and
toxicological
measurements
to
allow
site­
specific
adjustments
that
account
for
differences
between
the
toxicity
of
a
metal
in
laboratory
dilution
water
and
in
site
water.

109
Fo­
of
Me­
u
Even
if
the
relationship
of
toxicity
to
the
forms
of
metals
is
not
understood
well
enough
to
allow
setting
site­
specific
water
quality
criteria
without
using
empirical
adjustments,
appropriate
use
and
interpretation
of
WERs
requires
an
understanding
of
how
changes
in
the
relative
concentrations
of
different
forms
of
a
metal
might
affect
toxicity.
Because
WERs
are
defined
on
the
basis
of
relationships
between
measurements
of
toxicity
and
measurements
of
total
recoverable
and/
or
dissolved
metal,
the
toxicologically
relevant
distinction
is
between
the
forms
of
the
metal
that
are
toxic
and
nontoxic
whereas
the
chemically
relevant
distinction
is
between
the
forms
that
are
dissolved
and
particulate.
'
Dissolved
metal'
is
defined
here
as
'
metal
that
passes
through
either
a
0.45­
p
or
a
0.40­
p
membrane
filter'
and
l
particulate
metalm
is
defined
as
.
total
recoverable
metal
minus
dissolved
metal'.
Metal
that
is
in
or
on
particles
that
pass
through
the
filter
is
operationally
defined
as
'
dissolved'.

In
addition,
some
species
of
metal
can
be
converted
from
one
form
to
another.
Some
conversions
are
the
result
of
reeguilibration
in
response
to
changes
in
water
quality
characteristics
whereas
others
are
due
to
such
fate
processes
as
oxidation
of
sulfides
and/
or
organic
matter.
Reequilibration
usually
occurs
faster
than
fate
processes
and
probably
results
in
any
rapid
changes
that
are
due
to
effluent
mixing
with
receiving
water
or
changes
in
pH
at
a
gill
surface.
To
account
for
rapid
changes
due
to
reeguilibration,
the
terms
'
labile'
and
'
refractory'
will
be
used
herein
to
denote
metal
species
that
do
and
do
not
readily
convert
to
other
species
when
in
a
noneguilibrium
condition,
with
'
readily'
referring
to
substantial
progression
toward
equilibrium
in
less
than
about
an
hour.
Although
the
toxicity
and
lability
of
a
form
of
a
metal
are
not
merely
yes/
no
properties,
but
rather
involve
gradations,
a
simple
classification
scheme
such
as
this
should
be
sufficient
to
establish
the
principles
regarding
how
WEF&
are
related
to
various
operationally
defined
forms
of
metal
and
how
this
affects
the
determination
and
use
of
WERs.

Figure
Dl
presents
the
classification
scheme
that
results
from
distinguishing
forms
of
metal
based
on
analytical
methodology,
toxicity
tests,
and
lability,
as
described
above.
Metal
that
is
not
measured
by
the
total
recoverable
measurement
is
assumed
to
be
sufficiently
nontoxic
and
refractory
that
it
will
not
be
further
considered
here.
Allowance
is
made
for
toxicity
due
to
particulate
metal
because
some
data
indicate
that
particulate
metal
might
contribute
to
toxicity
and
bioaccumulation,
although
other
data
imply
that
little
or
no
toxicity
can
be
ascribed
to
particulate
metal
(
Erickson
1993b).
Even
if
the
toxicity
of
particulate
metal
is
not
negligible
in
a
particular
situation,
a
dissolved
criterion
will
not
be
underprotective
if
the
dissolved
criterion
was
derived
using
a
dissolved
WER
(
see
below)
or
if
there
are
sufficient
compensating
factors.

110
­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­
Figure
01:
A
Schamo
for
ClauifyiPg
Formm
of
Metal
in
Water
Total
recoverable
metal
Dissolved
Nontoxic
Labile
Refractory
Toxic
Labile
Particulate
Nontoxic
Labile
Refractory
Toxic
Labile
Metal
not
measured
by
the
total
recoverable
measurement
­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­­

Not
only
can
some
changes
in
water
quality
characteristics
shift
the
relative
concentrations
of
toxic
and
nontoxic
labile
species
of
a
metal,
some
changes
in
water
quality
can
also
increase
or
decrease
the
toxicities
of
'
the
toxic
species
of
a
metal
and/
or
the
sensitivities
of
aquatic
organisms.
Such
changes
might
be
caused
by
(
a)
a
change
in
ionic
strength
that
affects
the
activity
of
toxic
species
of
the
metal
in
water,
(
b)
a
physiological
Rffect
whereby
an
ion
affects
the
permeability
of
a
membrane
and
thereby
alters
both
uptake
and
apparent
toxicity,
and
(
c)
toxicological
additivity,
synergism,
or
antagonism
due
to
effects
within
the
organism.

Another.
possible
complication
is
that
a
form
of
metal
that
is
toxic
to
one
aquatic
organism
might
not
be
toxic
to
another.
Although
such
differences
between
organisms
have
not
been
demonstrated,
the
possibility
cannot
be
ruled
out.

The
Imoortance
of
Lailitv
The
only
common
metal
measurement
that
can
be
validly
extrapolated
from
the
effluent
and
the
upstream
water
to
the
downstream
water
merely
by
taking
dilution
into
account
is
the
total
recoverable
measurement.
A
major
reason
this
measurement
is
so
useful
is
because
it
is
the
only
measurement
that
obeys
the
law
of
mass
balance
(
i.
e.,
it
is
the
only
measurement
that
is
conservative).
Other
metal
measurements
usually
do
not
obey
the
law
of
mass
balance
because
they
measure
some,
but
not
all,
of
the
labile
species
of
metals.
A
measurement
of
refractory
metal
111
would
be
conservative
in
terms
of
changes
in
water
quality
characteristics,
but
not
necessarily
in
regards
to
fate
processes;
such
a
measurement
has
not
been
developed,
however.

Permit
limits
apply
to
effluents,
whereas
water
quality
criteria
apply
to
surface
waters.
If
permit.'
limits
and
water
quality
criteria
are
both
expressed
in
terms
of
total
recoverable
metal,
extrapolations
from
effluent
to
surface
water
only
need
to
take
dilution
into
account
and
can
be
performed
as
mass
balance
calculations.
If
either
permit
limits
or
water
quality
criteria
or
both
are
expressed
in
terms
of
any
other
metal
measurement,
lability
needs
to
be
taken
into
account,
even
if
both
are
expressed
in
terms
of
the
same
measurement.

Extrapolations
concerning
labile
species
of
metals
from
effluent
to
surface
water
depend
to
a
large
extent
on
the
differences
between
the
water
quality
characteristics
of
the
effluent
and
those
of
the
surface
water.
Although
equilibrium
models
of
the
speciation
of
metals
can
provide
insight,
the
interactions
are
too
complex
to
be
able
to
make
useful
nonempirical
extrapolations
from
a
wide
variety
of
effluents
to
a
wide
variety
of
surface
waters
of
either
(
a)
the
speciation
of
the
metal
or
(
b)
a
metal
measurement
other
than
total
recoverable.

Rnpirical
extrapolations
can
be
performed
fairly
easily
and
the
most
cormon
case
will
probably
occur
when
petit
limits
are
based
on
the
total
recoverable
measurement
but
water
quality
criteria
are
based
on
the
dissolved
measurement.
The
empirical
extrapolation
is
intended
to
answer
the
question
What
percent
of
the
total
recoverable
metal
in
the
effluent
becomes
dissolved
in
the
downstream
water?'
This
question
can
be
answered
by:
a.
Collecting
samples
of
effluent
and
upstream
water.
b.
Measuring
total
recoverable
metal
and
dissolved
metal
in
both
samples.
c.
Combining
aliquots
of
the
two
samples
in
the
ratio
of
the
flows
when
the
sanrples
were
obtained
and
mixing
for
an
appropriate
period
of
time
under
appropriate
conditions.
d.
Measuring
total
recoverable
metal
and
dissolved
metal
in
the
mixture.
An
example
is
presented
in
Figure
D2.
This
percentage
cannot
be
extrapolated
from
one
metal
to
another
or
from
one
effluent
to
another.
The
data
needed
to
calculate
the
percentage
will
be
obtained
each
time
a
WER
is
determined
using
simulated
downstream
water
if
both
dissolved
and
total
recoverable
metal
are
measured
in
the
effluent,
upstream
water,
and
simulated
downstream
water.

The
interpretation
of
the
percentage
is
not
necessarily
as
straightforward
as
might
be
assumed.
For
example,
some
of
the
metal
that
is
dissolved
in
the
upstream
water
might
sorb
onto
particulate
matter
in
the
effluent,
which
can
be
viewed
as
a
detoxification
of
the
upstream
water
by
the
effluent.
Regardless
of
the
interpretation,
the
described
procedure
provides
a
simple
112
way
of
relating
the
total
recoverable
concentration
in
the
effluent
to
the
concentration
of
concern
in
the
downstream
water.
Because
this
empirical
extrapolation
can
be
used
with
any
analytical
measurement
that
is
chosen
as
the
basis
for
expression
of
aquatic
life
criteria,
use
of
the
total
recoverable
measurement
to
express
permit
limits
on
effluents
does
not
place
any
restrictions
on
which
analytical
measurement
can
be
used
to
express
criteria.
Further,
even
if
both
criteria
and
permit
limits
are
expressed
in
terms
of
a
measurement
such
as
dissolved
metal,
an
empirical
extrapolation
would
still
be
necessary
because
dissolved
metal
is
not
likely
to
be
conservative
from
effluent
to
downstream
water.

Merits
of
Total
Recoverable
and
Dissolved
WERs
and
Criteria
A
WER
is
operationally
defined
as
the
value
of
an
endpoint
obtained
with
a
toxicity
test
using
site
water
divided
by
the
value
of
the
same
endpoint
obtained
with
the
same
toxicity
test
using
a
laboratory
dilution
water.
Therefore,
just
as
aquatic
life
criteria
can
be
expressed
in
terms
of
either
the
total
recoverable
measurement
or
the
dissolved
measurement,
so
can
WERs.
A
pair
of
side­
by­
side
toxicity
tests
can
produce
both
a
total
recoverable
WER
and
a
dissolved
WER
if
the
metal
in
the
test
solutions
in
both
of
the
tests
is
measured
using
both
methods.
A
total
recoverable
WER
is
obtained
by
dividing
endpoints
that
were
calculated
on
the
basis
of
total
recoverable
metal,
whereas
a
dissolved
WER
is
obtained
by
dividing
endpoints
that
were
calculated
on
the
basis
of
dissolved
metal.
Because
of
the
way
they
are
determined,
a
total
recoverable
WER
is
used
to
calculate
a
total
recoverable
site­
specific
criterion
from
a
national,
state,
or
recalculated
aquatic
life
criterion
that
is
expressed
using
the
total
recoverable
measurement,
whereas
a
dissolved
WER
is
used
to
calculate
a
dissolved
site­
specific
criterion
from
a
national,
state,
or
recalculated
criterion
that
is
expressed
in
terms
of
the
dissolved
measurement.

In
terms
of
the
classification
scheme
given
in
Figure
Dl,
the
basic
relationship
between
a
total
recoverable
national
water
quality
criterion
and
a
total
recoverable
WER
is:

l
A
total
recoverable
criterion
treats
allthe
toxic
and
nontoxic
metal
in
the
site
water
as
if
its
average
toxicity
were
the
same
as
the
average
toxicity
of
all
the
toxic
and
nontoxic
metal
in
the
toxicity
tests
in
laboratory
dilution
water
on
which
the
criterion
is
based.

l
A
total
recoverable
WER
is
a
measurement
of
the
actual
ratio
of
the
ave=
e
toxicities
of
the
total
recoverable
metal
and
replaces
the
assumption
that
the
ratio
is
1.

113
Similarly,
the
basic
relationship
between
a
dissolved
national
.
crlterloa
and
a
dissolved
m
is:

l
A
dissolved
criterion
treats
all
the
toxic
and
nontoxic
dissolved
metal
in
the
site
water
as
if
its
average
toxicity
were
the
same
as
the
average
toxicity
of
all
the
toxic
and
nontoxic
dissolved
metal
in
the
toxicity
tests
in
laboratory
dilution
water
on
which
the
criterion
is
based.

l
A
dissolved
m
is
a
measurement
of
the
actual
ratio
of
the
average
toxicities
of
the
dissolved
metal
and
replaces
the
assumption
that
the
ratio
is
1.
In
both
cases,
use
of
a
criterion
without
a
WER
involves
measurement
of
toxicity
in
laboratory
dilution
water
but
only
prediction
of
toxicity
in
site
water,
whereas
use
of
a
criterion
with
a
wER
involves
measurement
of
toxicity
in
both
laboratory
dilution
water
and
site
water.

When
WERs
are
used
to
derive
site­
specific
criteria,
the
total
recoverable
and
dissolved
approaches
are
inherently
consistent.
They
are
consistent
because
the
toxic
effects
caused
by
the
metal
in
the
toxicity
tests
do
not
depend
on
what
chemical
measurements
are
performed;
the
same
number
of
organisms
are
killed
in
the
acute
lethality
tests
regardless
of
what,
if
any,
measurements
of
the
concentration
of
the
metal
are
made.
The
only
difference
is
the
chemical
measurement
to
which
the
toxicity
is
referenced.
Dissolved
WERs
can
be
derived
from
the
same
pairs
of
toxicity
tests
from
which
total
recoverable
WERs
are
derived,
if
the
metal
in
the
tests
is
measured
using
both
the
total
recoverable
and
dissolved
measurements.
Both
approaches
start
at
the
same
place
(
i.
e.,
the
amount
of
toxicity
observed
in
laboratory
dilution
water)
and
end
at
the
same
place
(
i.
e.,
the
amount
of
toxicity
observed
in
site
water).
The
combination
of
a
total
recoverable
criterion
and
WER
accomplish
the
same
thing
as
the
combination
of
a
dissolved
criterion
and
WER.
By
extension,
whenever
a
criterion
and
a
WER
based
on
the
same
measurement
of
the
metal
are
used
together,
they
will
end
up
at
the
same
place.
Because
use
of
a
total
recoverable
criterion
with
a
total
recoverable
m
ends
up
at
exactly
the
same
place
as
use
of
a
dissolved
criterion
with
a
dissolved
m,
whenever
one
WER
is
determined,
both
should
be
determined
to
allow
(
a)
a.
check
on
the
analytical
chemistry,
(
b)
use
of
the
inherent
internal
consistency
to
check
that
the
data
are
used
correctly,
and
(
c)
the
option
of
using
either
approach
in
the
derivation
of
permit
limits.
.
An
examination
of
how
the
two
approaches
(
the
total
recoverable
approach
and
the
dissolved
approach)
address
the
four
relevant
forms
of
metal
(
toxic
and
nontoxic
particulate
metal
and
toxic
and
nontoxic
dissolved
metal)
in
laboratory
dilution
water
and
in
site
water
further
explains
why
the
two
approaches
are
inherently
consistent.
Here,
only
the
way
in
which
the
two
approaches
address
each
of
the
four
forms
of
metal
in
site
water
will
be
considered:

114
a.
Toxic
dissolved
metal:
This
form
contributes
to
the
toxicity
of
the
site
water
and
is
measured
by
both
chemical
measurements.
If
this
is
the
only
form
of
metal
present,
the
two
WERs
will
be
the
same.
b.
Nontoxic
dissolved
metal:
This
form
does
not
contribute
to
the
toxicity
of
the
site
water,
but
it
is
measured
by
both
chemical
measurements.
If
this
is
the
only
form
of
metal
present,
the
two
WERs
will
be
the
same.
(
Nontoxic
dissolved
metal
can
be
the
only
form
present,
however,
only
if
all
of
the
nontoxic
dissolved
metal
present
is
refractory.
If
any
labile
nontoxic
dissolved
metal
is
present,
equilibrium
will
require
that
some
toxic
dissolved
metal
also
be
present.)
c.
Toxic
particulate
metal:
This
form
contributes
to
the
toxicological
measurement
in
both
approaches;
it
is
measured
by
the
total
recoverable
measurement,
but
not
by
the
dissolved
measurement.
Even
though
it
is
not
measured
by
the
dissolved
measurement,
its
presence
is
accounted
for
in
the
dissolved
approach
because
it
increases
the
toxicity
of
the
site
water
and
thereby
decreases
the
dissolved
WER.
It
is
accounted
for
because
it
makes
the
dissolved
metal
appear
to
be
more
toxic
than
it
is.
Most
toxic
particulate
metal
is
probably
not
toxic
when
it
is
particulate;
it
becomes
toxic
when
it
is
dissolved
at
the
gill
surface
or
in
the
digestive
system;
in
the
surface
water,
however,
it
is
measured
as
particulate
metal.
d.
Nontoxic
particulate
metal:
This
form
does
not
contribute
to
the
toxicity
of
the
site
water;
it
is
measured
by
the
total
recoverable
measurement,
but
not
by
the
dissolved
measurement.
Because
it
is
measured
by
the
total
recoverable
measurement,
but
not
by
the
dissolved
measurement,
it
causes
the
total
recoverable
WER
to
be
higher
than
the
dissolved
WER.
In
addition
to
dealing
with
the
four
forms
of
metal
similarly,
the
WERs
used
in
the
two
approaches
comparably
take
synergism,
antagonism,
and
additivity
into
account.
Synergism
and
additivity
in
the
site
water
increase
its
toxicity
and
therefore
decrease
the
WER;
in
contrast,
antagonism
in
the
site
water
decreases
toxicity
and
increases
the
WER.

Each
of
the
four
forms
of
metal
is
appropriately
taken
into
account
because
use
of
the
WERs
makes
the
two
approaches
internally
consistent.
In
addition,
although
experimental
variation
will
cause
the
measured
WERs
to
deviate
from
the
actual
WERs,
the
measured
WERs
will
be
internally
consistent
with
the
data
from
which
they
were
generated.
If
the
percent
dissolved
is
the
same
at
the
test
endpoint
in
the
two
waters,
the
two
WERs
will
be
the
same.
If
the
percent
of
the
total
recoverable
metal
that
is
dissolved
in
laboratory
dilution
water
is
less
than
100
percent,
changing
from
the
total
recoverable
measurement
to
the
dissolved
measurement
will
lower
the
criterion
but
it
will
115
comparably
lower
the
denominator
in
the
WER,
thus
increasing
the
WER.
If
the
percent
of
the
total
recoverable
metal
that
is
dissolved
in
the
site
water
is
less
than
100
percent,
changing
from
the
total
recoverable
measurement
to
the
dissolved
.
measurement
will
lower
the
concentration
in
the
site
water
that
is
to
be
compared
with
the
criterion,
but
it
also
lowers
the
numerator
in
the
WER,
thus
lowering
the
WER.
Thus
when
WEI&
are
used
to
adjust
criteria,
the
total
recoverable
approach
and
the
dissolved
approach
result
in
the
same
interpretations
of
concentrations
in
the
site
water
(
see
Figure
D3)
and
in
the
same
maximum
acceptable
concentrations
in
effluents
(
see
Figure
D4).

Thus,
if
WEFU
are
based
on
toxicity
tests
whose
endpoints
equal
the
CMC
or
CCC
and
if
both
approaches
are
used
correctly,
the
two
measurements
will
produce
the
same
results
because
each
WER
is
based
on
measurements
on
the
site
water
and
then
the
WER
is
used
to
calculate
the
site­
specific
criterion
that
applies
to
the
site
water
when
the
same
chemical
measurement
is
used
to
express
the
site­
specific
criterion.
The
eguivalency
of
the
two
approaches
applies
if
they
are
based
on
the
same
sample
of
site
water.
When
they
are
applied
to
multiple
samples,
the
approaches
can
differ
depending
on
how
the
results
from
replicate
samples
are
used:
a.
If
an
appropriate
averaging
process
is
used,
the
two
will
be
equivalent.
b.
If
the
lowest
value
is
used,
the
two
approaches
will
probably
be
equivalent
only
if
the
lowest
dissolved
WER
and
the
lowest
total
recoverable
WER
were
obtained
using
the
same
sample
of
site
water.

There
are
several
advantages
to
using
a
dissolved
criterion
even
when
a
dissolved
WER
is
not
used.
In
some
situations
use
of
a
dissolved
criterion
to
interpret
results
of
measurements
of
the
concentration
of
dissolved
metal
in
site
water
might
demonstrate
that
there
is
no
need
to
determine
either
a
total
recoverable
WER
or
a
dissolved
WER.
This
would
occur
when
so
much
of
the
total
recoverable
metal
was
nontoxic
particulate
metal
that
even
though
the
total
recovdrable
criterion
was
exceeded,
the
corresponding
dissolved
criterion
was
not
exceeded.
The
particulate
metal
might
come
from
an
effluent,
a
resuspension.
event,
or
runoff
that
washed
particulates
into
the
body
of
water.
In
such
a
situation
the
total
recoverable
WER
would
also
show
that
the
site­
specific
criterion
was
not
exceeded,
but
there
would
be
no
need
to
determine
a
WER
if
the
criterion
were
expressed
on
the
basis
of
the
dissolved
measurement.
If
the
variation
over
time
in
the
concentration
of
particulate
metal
is
much
greater
than
the
variation
in
the
concentration
of
dissolved
metal,
both
the
total
recoverable
concentration
and
the
total
recoverable
WER
are
likely
to
vary
so
much
over
time
that
a
dissolved
criterion
would
be
much
more
useful
than
a
total
recoverable
criterion.

116
Use
of
a
dissolved
criterion
without
a
dissolved
WER
has
three
disadvantages,
however:
1.
Nontoxic
dissolved
metal
in
the
site
water
is
treated
as
if
it
is
toxic.
2.
Any
toxicity
due
to
particulate
metal
in
the
site
water
is
ignored.
3.
Synergism,
antagonism,
and
additivity
in
the
site
water
are
not
taken
into
account.
Use
of
a
dissolved
criterion
with
a
dissolved
WER
overcomes
all
three
problems.
For
example,
if
(
a)
the
total
recoverable
concentration
greatly
exceeds
the
total
recoverable
criterion,
(
b)
the
dissolved
concentration
is
below
the
dissolved
criterion,
and
(
c)
there
is
concern
about
the
possibility
of
toxicity
of
particulate
metal,
the
determination
of
a
dissolved
WER
would
demonstrate
whether
toxicity
due
to
particulate
metal
is
measurable.

Similarly,
use
of
a
total
recoverable
criterion
without
a
total
recoverable
WER
has
three
comparable
disadvantages:
1.
Nontoxic
dissolved
metal
in
site
water
is
treated
as
if
it
is
toxic.
2.
Nontoxic
particulate
metal
in
site
water
is
treated
as
if
it
is
toxic.
3.
Synergism,
antagonism,
and
additivity
in
site
water
are
not
taken
into
account.
Use
of
a
total
recoverable
criterion
with
a
total
recoverable
WER
overcomes
all
three
problems.
For
example,
determination
of
a
total
recoverable
WER
would
prevent
nontoxic
particulate
metal
(
as
well
as
nontoxic
dissolved
metal)
in
the
site
water
from
being
treated
as
if
it
is
toxic.

pelationshim
between
WERs
and
the
Forms
of
Metal%

Probably
the
best
way
to
understand
what
WERs
can
and
cannot
do
is
to
understand
the
relationships
between
WERs
and
the
forms
of
metals.
A
WER
is
calculated
by
dividing
the
concentration
of
a
metal
that
corresponds
to
a
toxicity
endpoint
in
a
site
water
by
the
concentration
of
the
same
metal
that
corresponds
to
the
same
toxicity
endpoint
in
a
laboratory
dilution
water.
using
the
classification
scheme
given
in
Figure
Dl:
Therefore,

mR=
R,,
+
N,
+
T,
+
dV'
+
AT,
Rt
+
NL
+
Tt
+
AN,
+
AT,

The
subscripts
IS.
and
.
t'
denote
site
water
and
laboratory
dilution
water,
respectively,
and:

R
=
the
concentration
of
Befractory
metal
in
a
water.
definition,
(
By
all
refractory
metal
is
nontoxic
metal.)

117
N
=
the
concentration
of
Nontoxic
labile
metal
in
a
water.

T
=
the
concentration
of
xoxic
labile
metal
in
a
water.

AN
=
the
concentration
of
metal
added
during
a
WEX
determination
that
is
Nontoxic
labile
metal'after
it
is
added.

AT
=
the
concentration
of
metal
added
during
a
WER
determination
that
is
sxic
labile
metal
after
it
is
added.

For
a
total
recoverable
WER;
each
of
these
five
concentrations
includes
both
particulate
and
dissolved
metal,
if
both
are
present;
for
a
dissolved
WER
only
dissolved
metal
is
included.

Because
the
two
side­
by­
side
tests
use
the
same
endpoint
and
are
conducted
under
identical
conditions
with
comparable
test
organisms,
T,+
AT,
­
Tt
+
AT,
when
the
toxic
species
of
the
metal
are
equally
toxic
in
the
two
waters.
If
a
difference
in
water
quality
causes
one
or
more
of
the
toxic
species
of
the
metal
to
be
mOre
toxic
in
one
water
than
the
other,
or
causes
a
shift
in
the
ratios
of
various
toxic
species,
we
can
define
II=
Tn
+
AT,
T&+
AT,
'

Thus
H
is
a
multiplier
that
accounts
for
a
proportional
increase
or
decrease
in
the
toxicity
of
the
toxic
forms
in
site
water
as
compared
to
their
toxicities
in
laboratory
dilution
water.
Therefore,
the­
general
WE3
equation
is:

WERt
Rs+
N,+
aN,+
H(
TL+~
TL)
R,+
N,+
aN,+
(
T,+
AT,)
l
Several
things
are
obvious
from
this
equation:
1.
A
WER
should
not
be
thought
of
as
a
simple
ratio
such
as
H.
H
is
the
ratio
of
the
toxicities
of
the
toxic
species
of
the
metal,
whereas
the
WER
is
the
ratio
of
the
sum
of
the
toxic
and
the
nontoxic
species
of
the
metal.
Only
under
a
very
specific
set
of
conditions
will
wg~
­
H.
If
these
conditions
are
satisfied
and
if,
in
addition,
H=
1,
then
wg~
­
I.
Although
it
might
seem
that
all
of
these
conditions
will
rarely
be
satisfied,
it
is
not
all
that
rare
to
find
that
an
experimentally
determined
WER
is
close
to
1.
2.
When
the
concentration
of
metal
in
laboratory
dilution
water
is
negligible,
RL
­
NL
=
Tt
­
0
and
wm­
RH
+
NB
+
dV"
+
HAT,)
A&
+
AT,
.

118
Even
though
laboratory
dilution
water
is
low
in
Tot
and
TSS,
when
metals
are
added
to
laboratory
dilution
water
in
toxicity
tests,
ions
such
as
hydroxide,
carbonate1
and
chloride
react
with
some
metals
to
form
some
particulate
species
and.
some
dissolved
species,
both
of
which
might
be
toxic
or
nontoxic.
The
metal
species
that
are
nontoxic
contribute
to
A&,
whereas
those
that
are
toxic
contribute
to
AT,.
Hydroxide,
carbonate,
chloride,
Tot,
and
TSS
can
increase
AN'.
Anything
that
causes
AN',
to
differ
from
AN'
will
cause
the
WER
to
differ
from
1.
3.
Refractory
metal
and
nontoxic
labile
metal
in
the
site
water
above
that
in
the
laboratory
dilution
water
will
increase
the
WER.
Therefore,
if
the
WER
is
determined
in
downstream
water,
rather
than
in
upstream
water,
the
WER
will
be
increased
by
refractory
metal
and
nontoxic
labile
metal
in
the
effluent.
Thus
there
are
three
major
reasons
why
WERs
might
be
larger
or
smaller
than
1:
a.
The
toxic
species
of
the
metal
might
be
more
toxic
in
one
water
than
in
the
other,
i.
e.,
H
l
1.
b.
AN
might
be
higher
in
one
water
than
in
the
other.
C.
R
and/
or
N
might
be
higher
in
one
water
than
in
the
other.

The
last
reason
might
have
great
practical
importance
in
some
situations.
When
a
WER
is
determined
in
downstream
water,
if
most
of
the
metal
in
the
effluent
is
nontoxic,
the
WER
and
the
endpoint
in
site
water
will
correlate
with
the
concentration
of
metal
in
the
site
water.
In
addition,
they
will
depend
on
the
concentration
of
metal
in
the
effluent
and
the
concentration
of
effluent
in
the
site
water.
This
correlation
will
be
best
for
refractory
metal
because
its
toxicity
cannot
be
affected
by
water
quality
characteristics;
even
if
the
effluent
and
upstream
water
are
quite
different
so
that
the
water
quality
characteristics
of
the
site
water
depend
on
the
percent
effluent,
the
toxicity
of
the
refractory
metal
will
remain
constant
at
zero
and
the
portion
of
the
WER
that
is
due
to
refractory
metal
will
be
additive.

The
Denendence
of
WERs
on
the
Sensitivitv
of
Toxicity
Tests
It
would
be
desirable
if
the
magnitude
of,
the
WER
for
a
site
water
were
independent
of
the
toxicity
test
used
in
the
determination
of
the
WER,
so
that
any
convenient
toxicity
test
could
be
used.
It
can
be
seen
from
the
general
WER
equation
that
the
WER
will
be
independent
of
the
toxicity
test
only
if:

which
would
require
that
R,­
N,
­
A&
­
RL
­
NL
­
tiL
­
0.
(
It
would
be
easy
to
assume
that
Tt
­
0,
but
it
can
be
misleading
in
some
situations
to
make
more
simplifications
than
are
necessary.)

119
This
is
the
simplistic
concept
of
a
WER
that
would
be
advantageous
if
it
were
true,
but
which
is
not
likely
to
be
true
very
often.
Any
situation
in
which
one
or
more
of
the
terms
is
greater
than
zero
can
cause
the
WER
to
depend
on
the
sensitivity
of
the
toxicity
test,
although
the
difference
in
the
WERs
might
be
small.

Two
situations
that
might
be
common
can
illustrate
how
the
WER
can
depend
on
the
sensitivity
of
the
toxicity
test.
For
these
illustrations,.
there
is
no
advantage
to
assuming
that
H=
1,
so
a
will
be
retained
for
generality.
1.

2.
The
simplest
situation
is
when
R,>
0,
i.
e.,
when
a
substantial
concentration
of
refractory
metal
occurs
in
the
site
water.
If,
for
simplification,
it
is
assumed
that
N,­
aN,=
RL=
NL­
tiL­
Or
then:

Nmi­
R,
+
H(
T,
+
AT,)
I
R6
(
T,
+
AT,)
(
TL
+
AT,)
+
H.

The
quantity
T,+
AT,
obviously
changes
as
the
sensitivity
of
the
toxicity
test
changes.
When
R,­
0,
then
MER­
H
and
the
WER
is
independent
of
the
sensitivity
of
the
toxicity
test.
when
R#>
0,
then
the
WER
will
decrease
as
the
sensitivity
of
the
test
decreases
because
TL
+
AT,
will
increase.

More
complicated
situations
occur
when
(
N,
+
a~,)
>
0.
If,
for
simplification,
it
is
assumed
that
R6
­
RL
­
N,
­
AN,
­
0,
then:

=
­
(
N6
+
d7,)
+
H(
T,
+
A?`,)
­
(
4
+
fl#)
(
T
+
AT
+
H.
L
L)
(
T,
+
AT,)

a.
If
(
N,
+
AN,)
>
0
because
the
site
water
contains
a
substantial
concentration
of
a
complexing
agent
that
has
an
affinity
for
the
metal
and
if
corqplexation
converts
toxic
metal
into
nontoxic
metal,
the
complexation
reaction
will
control
the
toxicity
of
the
solution
(
Allen
1993).
A
complexation
cume
can
be
graphed
in
several
ways,
but
the
S­
shaped
curve
presented
in
Figure
DS
is
most
convenient
here.
The
vertical
axis
is
'%
uncomplexed',
which
is
assumed
to
correlate
with
'%
toxic'.
then
the
'%
nontoxic'.
The
ratio
of
toxic
metal
is:

%
nontoxic,
%
ccnl@
ad
%
toxic
%
Llncoilplm!
Bd
For
the
complexed
nontoxic
metal:
The
'%
complexed'
is
nontoxic
metal
to
­
v.

vt
concentration
ofnontoxdcmetal
concentration
of
toxicmetal
l
120
In
the
site
water,
the
concentration
of
complexed
nontoxic
metal
is
(
N,
+
ti6)
and
the
concentration
of
toxic
metal
is
(
T,+
AT,),
so
that:.

(
Iy,
+
u#)
(
N4
+
a#)
v6
­
(
T6
+
AT,)
­
H(
T,
+
AT,)
l
and
wBRN
V&(
TL
+
AT,)
+
H(=,
+
AT,)

(
TL
+
AT,)
­
vfi+
H=
H(
V,+
1)
.

If
the
WRR
is
determined
using
a
sensitive
toxicity
test
so
that
the
%
uncomplexed
(
i.
e.,
the
%
toxic)
is
10
%,
then
v,­
(
90
%)/(
I0
%)
­
9,
whereas
if
a
less
sensitive
test
is
used
so
that
the
%
uncomplexed
is
50
%,
then
v,­
(
50
%)/(
50
%)
­
1.
Therefore,
if
a
portion
of
the
WER
is
due
to
a
complexing
agent
in
the
site
water,
the
magnitude
of
the
WER
can
decrease
as
the
sensitivity
of
the
toxicity
test
decreases
because
the
%
uncomplexed
will
decrease.
In
these
situations,
the
largest
WER
will
be
obtained
with
the
most
sensitive
toxicity
test;
progressively
smaller
WERs
will
be
obtained
with
less
sensitive
toxicity
tests.
The
magnitude
of
a
WER
will
depend
not
only
on
the
sensitivity
of
the
toxicity
test
but
also
on
the
concentration
of
the
complexing
agent
and
on
its
binding
constant
(
complexation
constant,
stability
constant).
In
addition,
the
binding
constants
of
most
complexing
agents
depend
on
pH.

If
the
laboratory
dilution
water
contains
a
low
concentration
of
a
complexing
agent,

VL
­
ru,
+
flL
TL
+
AT,

and
mR­
v&
T,
+
AT,)
+
H(
T,
+
A?',)
=
v,$
+
H
I
H(
v,
+
1)
.

VL(
TL
+
AT,)
+
tTL
+
AT,)
v,
+
1
v,
+
1
The
binding
constant
of
the
complexing
agent
in
the
laboratory
dilution
water
is
probably
different
from
that
of
the
complexing
agent
in
the
site
water.
Although
changing
from
a
more
sensitive
test
to
a
less
sensitive
test
will
decrease
both
V,
and
v,,
the
amount
of
effect
is
not
likely
to
be
proportional.

If
the
change
from
a
more
sensitive
test
to
a
less
sensitive
test
were
to
decrease
v,
proportionately
more
than
v,,
the
change
could
result
in
a
larger
WER,
rather
121
than
a
smaller
WER,
as
resulted
in
the
case
above
when
it
was
assumed
that
the
laboratory
dilution
water
did
not
contain
any
complexing
agent.
This
is
probably
most
likely
to
occur
if
H?
1
and
if
v#
<
V,,
which
would
mean
that
lygR<
l.
Although
this
is
likely
to
be
a
rare
situation,
it
does
demonstrate
again
the'importance
of
determining
WERS
using
toxicity
tests
that
have
endpoints
in
laboratory
dilution
water
that
are
close
to
the
CMC
or
CCC
to
which
the
WER
is
to
be
applied.

b.
If
(
N,+
AN,)
>
0
because
the
site
water
contains
a
substantial
concentration
of
an
ion
that
will
precipitate
the
metal
of
concern
and
if
precipitation
converts
toxic
metal
into
nontoxic
metal,
the
precipitation
reaction
will
control
the
toxicity
of
the
solution.
The
'
precipitation
curve.
given
in
Figure
D6
is
analogous
to
the
'
coxnplexation
curve.
given
in
Figure
D5;
in
the
precipitation
curve,
the
vertical
axis
is
'%
dissolved.,
which
is
assumed
to
correlate
with
'
8
toxic'.
If
the
endpoint
for
a
toxicity
test
is
below
the
solubility
limit
of
the
precipitate,
(
N,+
AN,)
­
Or
whereas
if
the
endpoint
for
a
toxicity
test
is
above
the
solubility
limit,
(
N,
+
ti#)
>
0.
If
WERs
are
determined
with
a
series
of
toxicity
tests
that
have
increasing
endpoints
that
are
above
the
solubility
limit,
the
WER
will
reach
a
maximum
value
and
then
decrease.
The
magnitude
of
the
WER
will
depend
not
only
on
the
sensitivity
of
the
toxicity
test
but
also
on
the
concentration
of
the
precipitating
agent,
the
solubility
limit,
and
the
solubility
of
the
precipitate.

Thus,
depending
on
the
composition
of
the
site
water,
a
WER
obtained
with
an
insensitive
test
might
be
larger,
smaller,
or
similar
to
a
WER
obtained
with
a
sensitive
test.
Because
of
the
range
of
possibilities
that
exist,
the
best
toxicity
test
to
use
in
the
experimental
determination
of
a
WEX
is
one
whose
endpoint
in
laboratory
dilution
water
is
close
to
the
CMC
or
CCC
that
is
to
be
adjusted.
This
is
the
rationale
that
was
used
in
the
selection
of
the
toxicity
tests
that
are
suggested
in
Appendix
I.

The
available
data
indicate
that
a
less
sensitive
toxicity
test
usually
gives
a
smaller
WER
than
a
more
sensitive
test
(
Hansen
1993a).
Thus,
use
of
toxicity
tests
whose
endpoints
are
higher
than
the
CMC
or
CCC
probably
will
not
result
in
underprotection;
in
contrast,
use
of
tests
whose
endpoints
are
substantially
below
the
CMC
or
CCC
might
result
in
underprotection.

The
factors
that
cause
~~
and
(
N,
+
AN,)
to
be
greater
than
zero
are
all
external
to
the
test
organisms;
they
are
chemical
effects
that
affect
the
metal
in
the
water.
The
magnitude
of
the
WER
is
therefore
expected
to
depend
on
the
toxicity
test
used
only
in
regard
to
the
sensitivity
of
the
test.
If
the
endpoints
for
two
122
different
tests
occur
at
the
same
concentration
of
the
metal,
the
magnitude
of
the
WERs
obtained
with
the
two
tests­
should
be
the
same;
they
should
not
depend
on
(
a)
the
duration
of
the
test,
(
b)
whether
the
endpoint
is
based
on
a
lethal
or
sublethal
effect,
or
(
c)
whether
the
species
is
a
vertebrate
or
an
invertebrate.

Another
interesting
consequence
of
the
chemistry
of
complexation
is
that
the
%
uncosnplexed
will
increase
if
the
solution
is
diluted
(
Allen
and
Hansen
1993).
The
concentration
of
total
metal
will
decrease
with
dilution
but
the
%
uncomplexed
will
increase.
The
increase
will
not
offset
the
decrease
and
so
the
concentration
of
uncomplexed
metal
will
decrease.
Thus
the
portion
of
a
WER
that
is
due
to
complexation
will
not
be
strictly
additive
(
see
Appendix
G),
but
the
amount
of
nonadditivity
might
be
difficult
to
detect
in
toxicity
studies
of
additivity.
A
similar
effect
of
dilution
will
occur
for
precipitation.

The
illustrations
presented
above
were
simplified
to
make
it
easier
to
understand
the
kinds
of
effects
that
can
occur.
The
illustrations
are
qualitatively
valid
and
demonstrate
the
direction
of
the
effects,
but
real­
world
situations
will
probably
be
so
much
mOre
complicated
that
the
various
effects
cannot
be
dealt
with
separately.

.
Pther
ProDertles
of
WQ&

1.
Because
of
the
variety
of
factors
that
can
affect
WERs,
no
rationale
exists
at
present
for
extrapolating
WERs
from
one
metal
to
another,
from
one
effluent
to
another,
or
from
one
surface
water
to
another.
Thus
WERs
should
be
individually
determined
for
each
metal
at
each
site.

2.
The
most
important
infonaation
that
the
determination
of
a
WER
provides
is
whether
simulated
and/
or
actual
downstream
water
adversely
affects
test
organisms
that
are
sensitive
to
the
metal.
A
WER
cannot
indicate
how
much
metal
needs
to
be
removed
from
or
how
much
metal
can
be
added
to
an
effluent.
a.
If
the
site
water
already
contains
sufficient
metal
that
it
is
toxic
to
the
test
organisms,
a
WER
cannot
be
determined
with
a
sensitive
test
and
so
an
insensitive
test
will
have
to
be
used.
Even
if
a
WER
could
be
determined
with
a
sensitive
test,
the
WER
cannot
indicate
how
much
metal
has
to
be
removed.
For
example,
if
a
WER
indicated
that
there
was
20
percent
too
much
metal
in
an
effluent,
a
30
percent
reduction
by
the
discharger
would
not
reduce
toxicity
if
only
nontoxic
metal
was
removed.
Thenext
WER
determination
would
show
that
the
effluent
still
contained
too
much
metal.
Removing
metal
is
useful
only
if
the
metal
removed
is
toxic
metal.
Reducing
the
total
recoverable
concentration
does
not
necessarily
reduce
toxicity.

123
3.
b.
If
the
simulated
or
actual
downstream
water
is
not
toxic,
a
WER
can
be
determined
and
used
to
calculate
how
much
additional
metal
the
effluent
could
contain
and
still
be
acceptable.
Because
an
unlimited
amount
of
refracmry
metal
can
be
added
to
the
effluent
without
affecting
the
organisms,
what
the
WER
actually
determines
is
how
much
additional
toxic
metal
can.
be
added
to
the
effluent.

The
effluent
component
of
nearly
all
WEFU
is
likely
to
be
due
mostly
to
either
(
a)
a
reduction
in
toxicity
of
the
metal
by
TSS
or
'
pot,
or
(
b)
the
presence
of
refractory
metal.
For
both
of
these,
if
the
percentage
of
effluent
in
the
downstream
water
decreases,
the
magnitude
of
the
WER
will
usually
decrease.
If
the
water
quality
characteristics
of
the
effluent
and
the
upstream
water
are
quite
different,
it
is
possible
that
the
interaction
will
not
be
additive;
this
can
affect
the
portion
of
the
WER
that
is
due
to
reduced
toxicity
caused
by
sorption
and/
or
binding,
but
it
cannot
affect
the
portion
of
the
WER
that
is
due
to
refractory
metal.

4.
Test
organisms
are
fed
during
some
toxicity
tests,
but
not
during
others;
it
is
not
clear
whether
a
WER
determined
in
a
fed
test
will
differ
from
a
WER
determined
in
an
unfed
test.
Whether
there
is
a
difference
is
likely
to
depend
on
the
metal,
the
type
and
amount
of
food,
and
whether
a
total
recoverable
or
dissolved
WER
is
determined.
This
can
be
evaluated
by
determining
two
WERs
using
a
test
in
which
the
organisms
usually
are
not
fed
­
one
WER
with
no
food
added
to
the
tests
and
one
with
food
added
to
the
tests.
Any
effect
of
food
is
probably
due
to
an
increase
in
TOC
and/
or
TSS.
If
food
increases
the
concentration
of
nontoxic
metal
in
both
the
laboratory
dilution
water
and
the
site
water,
the
food
will
probably
decrease
the
WER.
Because
complexes
of
metals
are
usually
soluble,
cqlexation
is
likely
to
lower
both
total
recoverable
and
dissolved
WEF&;
sorption
to
solids
will
probably
reduce
only
total
recoverable
WERs.
The
food
might
also
affect
the
acute­
chronic
ratio.
Any
feeding
during
a
test
should
be
limited
to
the
minimum
necessary.

The
acceptable
WERs
found
by
Brungs
et
al.
(
1992)
were
total
recoverable
WERs
that
were
determined
in
relatively
clean
fresh
water.
These
WERs
ranged
from
about
1
to
15
for
both
copper
and
cadmium,
whereas
they
ranged
from
about
0.7
to
3
for
zinc.
The
few
WEF&
that
were
available
for
chromium,
lead,
and
nickel
ranged
from
about
1
to
6.
Both
the
total
recoverable
and
dissolved
WERs
for
copper
in
New
York
harbor
range
from
about
0.4
to
4
with
most
of
the
WERs
being
between
1
and
2
(
Hansen
1993b).

124
Figure
D2:
An
&
amp10
of
the
Empirical
S%
trapolation
Proces8
Assume
the
following
hypothetical
effluent
and
upstream
water:

Effluent:
T,:
100
ug/
L
D6:
10
ug/
L
U*:
24
cfs
Upstream
water:
T,:
40
ug/
L
Do:
38
ug/
L
Pa:
48
cfs
Downstream
water:
Tn
:
60
ug/
L
4.8:
36
ug/
L
QD:
72
cfs
where:

T
=
concentration
of
D
=
concentration
of
u
=
flow.
(
10
8
dissolved)

(
95
%
dissolved)

(
60
%
dissolved)

total
recoverable
metal.
dissolved
metal.

The
subscripts
E,
U,
and
D
signify
effluent,
upstream
water,
and
downstream
water,
respectively.

By
conservation
of
flow:
&­
Q6+
PO.

By
conservation
of
total
recoverable
metal:
TIpo­
Tp,+
Trpo.

If
P
=
the
percent
of
the
total
recoverable
metal
in
the
effluent.
that
becomes
dissolved
in
the
downstream
water,

pI
loo(~
D­
DI&)

W6
'

For
the
data
given
above,
the
percent
of
the
total
recoverable
metal
in
the
effluent
that
becomes
dissolved
in
the
downstream
water
is:

p
I
100
[
(
36
ug/
t)
(
72
CfB)
­
(
38
W/
t)
(
48
Cfd
1
.
32
Q
,
(
100
u&
L)
(
24
Cf8)

which
is
greater
than
the
10
%
dissolved
in
the
effluent
and
less
than
the
60
%
dissolved
in
the
downstream
water.

125
Figure
03:
The
Intormal
Coa8imtuxcy
of
the
Two
Approachem
The
internal
consistency
of
the
total
recoverable
and
dissolved
approaches
can
be
illustrated
by
considering
the
use
of
WERS
to
interpret
the
total
recoverable
and.
dissolved
concentrations
of
a
metal
in
a
site
water.
For
this
hypothetical
example,
it
will
be
assumed
that
the
national
CCCs
for
the
metal
are:
200
ug/
L
as
total
recoverable
metal.
160
ug/
L
as
dissolved
metal.
It
will
'
also
be
assumed
that
the
concentrations
of
the
metal
in
the
site
water
are:
300
ug/
L
as
total
recoverable
metal.
120
ug/
L
as
dissolved
metal.
The
total
recoverable
concentration
in
the
site
water
exceeds
the
national
CCC,
but
the
dissolved
concentration
does
not.

The
following
results
might
be
obtained
if
WERs
are
determined:

atorv
.
.
Dilution
Water
Total
recoverable
LCSO
=
400
ug/
L.
%
of
the
total
recoverable
metal
that
is
dissolved
=
80.
(
This
is
based
on
the
ratio
of
the
national
CCCs,
which
were
determined
in
laboratory
dilution
water.)
Dissolved
LCSO
=
320
ug/
L.

Total
recoverable
LCSO
=
620
ug/
L.
%
of
the
total
recoverable
metal
that
is
dissolved
=
40.
(
This
is
based
on
the
data
given
above
for
site
water).
Dissolved
LCSO
=
248
ug/
L.

Total
recoverable
WEB
=
(
620
ug/
L)/(
400
ug/
L)
=
1.55
Dissolved
WER
=
(
248
ug/
L)/(
320
ug/
L)
=
0.775
Tmalrocowrabl~
WgR
nissolvedclrgR
91.55.
0.775
lab
water
%
dissolvd
*
80
N
2
site
hater
%
dfssolwd
40
Total
recoverable
ssCCC
=
(
200
ug/
L)(
l.
Sf)
=
310
ug/
L.
Dissolved
ssCCC
=
(
160
ug/
L)(
O.
775)
=
124
ug/
L.

Both
concentrations
in
site
water
are
below
the
respective
sscccs.

126
In
contrast,
the
following
results
might
have
been
obtained
when
the
WERs
were
determined:

In
Laboratorv
Dilution
Wate
Total
recoverable
LCSO
=
r400
ug/
L.
%
of
the
total
recoverable
metal
that
is
dissolved
=
80.
Dissolved
LCSO
=
320
ug/
L.

In
Site
ate
TotalWre&
erable
LC50
=
580
ug/
L.
%
of
the
total
recoverable
metal
that
is
dissolved
=
40.
Dissolved
LC50
=
232
ug/
L.

S
Total
recoverable
WER
=
(
580
ug/
L)/(
QoO
ug/
L)
=
Dissolved
WER
=
(
232
ug/
L)/(
320
ug/
L)
=
0.725
Checkina
the
Calculations
mta1
reBcoverable
WER
1.45
t­
9
lab
water
0
dissolwd
Dissolved
WHZ
0.725
site
water
0
ciissolv8d
1.45
980.2
40
Site­
snecific
CCCs
(
ssCCCs)

Total
recoverable
ssCCC
=
(
200
ug/
L)
(
1.45)
=
290
ug/
L.
Dissolved
ssCCC
=
(
160
ug/
L)
(
0.725)
=
116
ug/
L.

In
this
case,
both
respective
ssCCCs.
concentrations
in
site
water
are
above
the
In
each
case,
both
approaches
resulted
in
the
same
concerning
whether
the
concentration
in
site
water
site­
specific
criterion.
conclusion
exceeds
the
The
two
key
assumptions
are:
1.
The
ratio
of
total
recoverable
metal
to
laboratory
dilution
water
when
the
WERs
.
the
ratio
of
the
national
CCCs.
2.
The
ratio
of
total
recoverable
metal
to
site
water
when
the
WERs
are
determined
the
concentrations
reported
in
the
site
Differences
in
the
ratios
that
are
outside
experimental
variation
will
cause
problems
dissolved
metal
in
are
determined
equals
dissolved
metal
in
equals
the
ratio
of
water.
the
range
of
for
the
derivation
of
site­
specific
criteria
and,
therefore,
with
the
internal
consistency
of
the
two
approaches.

127
Figure
01:
The
AOplicrtion
of
tha
SW0
Approrchm
Hypothetical
upstream
water
and
effluent
will
be
used
to
demonstrate
the
equivalence
of
the
total
recoverable
and
dissolved
approaches.
The
upstream
water
and
the
effluent
will
be
assumed
to
have
specific
properties
in
order
to
allow
calculation
of
the
properties
of
the
downstream
water,
which
will
be
assumed
to
be
a
1:
l
mixture
of
the
upstream
water
and
effluent.
It
will
also
be
assumed
that
the
ratios
of
the
forms
of
the
metal
in
the
upstream
water
and
in
the
effluent
do
not
change
when
the
total
recoverable
concentration
changes.

YEfFEergiF
=
3
cfs)
Refractory
par&
ulate:
400
ug/
L
200
ug/
L
Toxic
dissolved:
200
ug/
L
(
50
%
dissolved)

Bffluea
(
Flow
=
3
cfs)
Total
recoverable:
440
ug/
L
Refractory
particulate:
396
ug/
L
Labile
nontoxic
particulate:
44
ug/
L
Toxic
dissolved:
0
ug/
L
(
0
%
dissolved)
(
The
labile
nontoxic
particulate,
which
is
10
%
of
the
total
recoverable
in
the
effluent,
becomes
toxic
dissolved
in
the
downstream
water.)

(
Flow
=
6
cfs)
Downstream
wateq
Total
recoverable:
420
ug/
L
Refractory
particulate:
298
ug/
L
Toxic
dissolved:
122
ug/
L
(
29
%
dissolved)

The
values
for
the
downstream
water
are
calculated
from
the
values
for
the
upstream
water
and
the
effluent:
Total
recoverable:
[
3(
400)
+
3(
440)]/
6
=
420
ug/
L
Dissolved:
[
3(
200)
+
3(
44+
0)]/
6
=
122
ug/
L
Refractory
particulate:
[
3(
200)
+
3(
396))/
6
=
298
ug/
L
Assumed
National
CCC
(
CCC)
Total
recoverable
tn300
ug/
L
Dissolved
=
240
ug/
L
128
Uostream
site­
snecific
CCC
(
ussCCC)

Assume:
Dissolved
CCCWER
=
1.2
Dissolved
ussCCC
=
(
1.2)(
240
ug/
L)
=
288
ug/
L
By
calculation:
TR
ussCCC
=
(
288
ug/
L)/(
O.
S)
=
576
ug/
L
Total
recoverable
CCCWER
=
(
576
ug/
L)/(
300
ug/
L)
=
1.92
30:
ccc
ECCWER
ussccc
co
.
Total
recoverable:
ug/
L
1.92
576
ug/
L
Dissolved:
400n:
g,
L
240
ug/
L
1.2
288
ug/
L
200
ug/
L
%
dissolved
80
%
­­­­
50
%
50
%
Neither
concentration
exceeds
its
respective
ussCCC.

lmalreowrablemR
91.92.
lab
water
0
dissolwd
Dissolved
HER
Sit8
diseolwd
N
1.2
water
%
80
50
N1
l
6
Downstream
site­
soecific
CCC
(
dssCCC)

Assume:
Dissolved
cccWER
=
1.8
Dissolved
dssCCC
=
(
1.8)
(
240
ug/
L)
=
432
ug/
L
By
calculation:
TR
dssCCC
=
((
432
ug/
L­
I(
200
ug/
L)/
2l)/
O.
ll+((
400
ug/
L)/
2)
=
3520
ug/
L
This
calculation
determines
the
amount
of
dissolved
metal
contributed
by
the
effluent,
accounts
for
the
fact
that
ten
percent
of
the
total
recoverable
metal
in
the
effluent
becomes
dissolved,
and
adds
the
total
recoverable
metal
contributed
by
the
upstream
flow.
Total
recoverable
cccWER
=
(
3520
ug/
L)/(
300
ug/
L)
=
11.73
ccc
Total
recoverable:
30:
ug/
L
cccl+
mR
dssCCC
co
c.
11.73
Dissolved:
3520
ug/
L
420nug,
L
240
ug/
L
1.80
432
ug/
L
%
dissolved
122
ug/
L
80
%
­­­­
12.27
%
29
%
Neither
concentration
exceeds
its
respective
dssCCC.

Total
recowrable
HER
80
Dissolved
MER
N
11.73
N
lab
water
0
disoolwd
1.80
site
­
­
water
0
dissolwd
12.27
­
6.52
Calculatina
the
Maximum
Accentable
Concentration
in
the
Effluent
Because
neither
the
total
recoverable
concentration
nor
the
dissolved
concentration
in
the
downstream
water
exceeds
its
respective
site­
specific
CCC,
the
concentration
of
metal
in
the
effluent
could
be
increased.
Under
the
assumption
that
the
ratios
of
the
two
forms
of
the
metal
in
the
effluent
do
not
change
when
the
total
recoverable
concentration
changes,
the
maximum
acceptable
concentration
of
total
recoverable
metal
in
the
effluent
can
be
calculated
as
follows:

129
Starting
with
the
total
recoverable
dssCCC
of
3520
ug/
L
(
6
cfs)
(
3520
w/
L)
­
(
3
cfs)(
400
up/
=)
3
cfs
­
6640
ug/
L
Starting
with
the
dissolved
dssCCC
of
432
ug/
L
(
6
cfs)
(
432
u&
L)
­
(
3
cf8)
(
400
ug/=)
(
o­
5)
N
6640
ug,~
(
3
cfs)
(
0.10)

Total
recoverable:

(
3
cfs)
(
6640
u#/
L)
+
(
3
cfs)
(
400
ug/
='
6
cfs
­
3520
u&
L.

Dissolved:

(
3
cf8)
(
6640
ug/
L)
(
0.10)
+
(
3
Cf8)
(
400
Ug/
L)(
0*
50)
6
cfs
­
432
ug/
L
.

The
value
of
0.10
is
used
because
this
is
the
percent
of
the
total
recoverable
metal
in
the
effluent
that
becomes
dissolved
in
the
downstream
water.

The
values
of
3520
ug/
L
and
432
ug/
L
equal
the
downstream
site­
specific
CCCs
derived
above.

Another
Wav
to
Calculate
the
Maximum
Accentable
Concentration
The
maximum
acceptable
concentration
of
total
recoverable
metal
in
the
effluent
can
also
be
calculated
from
the
dissolved
dssCCC
of
432
ug/
L
using
a
partition
coefficient
to
convert
from
the
dissolved
dssCCC
of
432
ug/
L
to
the
total
recoverable
dssCCC
of
3520
ug/
L:

16
cfi]
[
4;
21g;
L
­
(
3
cfs)
(
400
I&
L)]
.
3
cfs
­
6640
tag/
L.

Note
that
the
value
used
for
the
partition
coefficient
in
this
calculation
is
0.1227
(
the
one
that
applies
to
the
downstream
water
when
the
total
recoverable
concentration
of
metal
in
the
effluent
is
6640
ug/
L),
not
0.29
(
the
one
that
applies
when
the
concentration
of
metal
in
the
effluent
is
only
420
ug/
L).
The
three
ways
of
calculating
the
maximum
acceptable
concentration
give
the
same
result
if
each
is
used
correctly.

130
Figure
D5:
A
Qenoralimd
Coqplexation
Cufpe
The
curve
is
for
a
constant
concentration
of
the
complexing
ligand
and
an
increasing
concentration
of
the
metal.

100
h
W
ti
.
.
.

.
.
.

.

.

.

.

.

.

.

.

.
.
.
.
I
I
I
LOG
OF
CONCENTRATION
OF
METAL
131
?
iguro
D6t
A
Gumralir~
Precipitation
CURO
The
cume
is
for
a
constant
concentration
of
the
precipitating
ligand
and
an
increasing
concentration
of
the
metal.

loo­*
l
l
.

l
0
I
I
I
I
LOG
OF
CONCENTRATION
OF
METAL
132
Allen,
H.
E.
1993.
Importance
of
Metal
Speciation
to
Toxicity.
Proceedings
of
the
Water
Environment
Federation
Workshop
on
Aguatic
Life
Criteria
for
Metals.
Anaheim,
CA.
pp.
55­
62.

Allen,
H.
E.,
and
D.
J.
Hansen.
1993.
The
Importance
of
Trace
Metal
Speciation
to
Water
Quality
Criteria.
Paper
presented
at
Society
for
Environmental
Toxicology
and
Chemistry.
Houston,
TX.
November
15.

Borgmanxi,
U.
1983.
Metal
Speciation
and
Toxicity
of
Free
Metal
Ions
to
Aquatic
Biota.
IN:
Aquatic
Toxicology.
(
J.
O.
Nriagu,
ed.)
Wiley,
New
York,
NY.

Brungs,
W.
A.,
T.
S.
Holderman,
and
M.
T.
Southerland.
1992.
Synopsis
of
Water­
Effect
Ratios
for
Heavy
Metals
as
Derived
for
Site­
Specific
Water
Quality
Criteria.
U.
S.
EPA
Contract
68­
CO­
0070.

Chapman,
G.
A.,
and
J.
K.
McCrady.
1977.
Copper
Toxicity:
A
Question
of
Form.
In:
Recent
Advances
in
Fish
Toxicology
Tubb,
ed.)
EPA­
600/
3­
77­
085
or
PB­
273
500.
National
T&~~
c~
l
.
Information
Service,
Springfield,
VA.
pp.
132­
151.

Erickson,
R.
1993a.
Memorandum
to
C.
Stephan.
July
14.

Erickson,
R.
1993b.
Memorandum
to
C.
Stephan.
November
12.
.
French,
P.,
and
D.
T.
E.
Hunt.
1986.
The
Effects
of
Inorganic
Complexing
upon
the
Toxicity
of
Copper
to
Aquatic
Organisms
(
Principally
Fish).
IN:
Trace
Metal
Speciation
and
Toxicity
to
Aguatic
Organisms
­
A
Review.
(
D.
T.
E.
Hunt,
ea.
1
Report
TR
247.
Water
Research
Centre,
United
Kingdom.

Hansen,
D.
J.
1993a.
Memorandum
to
C.
E.
Stephan.
April
29.

Hansen,
D.
J.
1993b.
Memorandum
to
C.
E.
Stephan.
October
6.

Nelson,
H.,
D.
Benoit,
R.
Erickson,
V.
Mattson,
and
J.
Lindberg.
1986.
The
Effects
of
Variable
Hardness,
pH,
Alkalinity,
Suspended
Clay,
and
Humics
on
the
Chemical
Speciation
and
Aquatic
Toxicity
of
Copper.
PB86­
171444.
National
Technical
Information
Service,
Springfield,
VA.

Wilkinson,
K.
J.,
P.
M.
Bertsch,
C.
H.
Jagoe,
and
P.
G.
C.
Campbell.
1993.
Surface
Complexation
of
Aluminum
on
Isolated
Fish
Gill
Cells.
Environ.
Sci.
Technol.
27:
1132­
1138.

133
Appmdix
St
U.
S.
SPA
Aquatic
Lifo
Criteria
Dccunntm
for
Hotrlm
Aluminum
Antimony
Arsenic
Beryllium
Cadmium
chromium
Copper
Lead
Mercury
Nickel
Selenium
Silver
Thallium
Zinc
EPA
440/
S­
86­
008
EPA
440/
S­
80­
020
EPA
440/
S­
84­
033
EPA
440/
S­
80­
024
EPA
440/
S­
84­
032
EPA
440/
S­
84­
029
EPA
440/
S­
84­
031
EPA
440/
S­
84­
027
EPA
440/
S­
84­
026
EPA
440/
S­
86­
004
EPA
440/
S­
87­
006
EPA
440/
S­
80­
071
EPA
440/
S­
80­
074
EPA
440/
S­
87­
003
PB88­
245998
PB81­
117319
PB85­
227445
PB81­
117350
PB85­
227031
PB85­
227478
PB85­
227023
PB85­
227437
PB85­
227452
PB87­
105359
PB88­
142237
PB81­
117822
PB81­
117848
PB87­
153581
All
are
available
from:
National
Technical
Information
Service
(
NTIS)
5285
Port
Royal
Road
Springfield,
VA
22161
TEL:
703­
487­
4650
134
­
ix
F:
Coruidorrtionm
Concoming
Multiple­
Metal,
Ibultiplo­
Dischugo,
axad
Special
Flowing­
Water
SitUatiOM
Multinle­
Metal
Situation9
Both
Method
1
and
Method
2
work
well
in
multiple­
metal
situations,
although
the
amount
of
testing
required
increases
as
the
number
of
metals
increases.
The
major
problem
is
the
same
for
both
methods:
even
when
addition
of
two
or
more
metals
individually
is
acceptable,
simultaneous
addition
of
the
two
or
more
metals,
each
at
its
respective
maximum
acceptable
concentration,
might
be
unacceptable
for
at
least
two
reasons:
1.
Additivity
or
synergism
might
occur
between
metals.
2.
More
than
one
of
the
metals
might
be
detoxified
by
the
same
complexing
agent
in
the
site
water.
When
WERs
are
determined
individually,
each
metal
can
utilize
all
of
the
complexing
capacity;
when
the
metals
are
added
together,
however,
they
cannot
simultaneously
utilize
all
of
the
complexing
capacity.
Thus
a
discharger
might
feel
that
it
is
cost­
effective
to
try
to
justify
the
lowest
site­
specific
criterion
that
is
acceptable
to
the
discharger
rather
than
trying
to
justify
the
highest
site­
specific
criterion
that
the
appropriate
regulatory
authority
might
approve.

There
are
two
options
for
dealing
with
the
possibility
of
additivity
and
synergism
between
metals:
a.

b.
WERs
could
be
developed
using
a
mixture
of
the
metals
but
it
might
be
necessary
to
use
several
primary
toxicity
tests
depending
on
the
specific
metals
that
are
of
interest.
Also,
it
might
not
be
clear
what
ratio
of
the
metals
should
be
used
in
the
mixture.
If
a
WER
is
determined
for
each
metal
individually,
one
or
more
additional
toxicity
tests
must
be
conducted
at
the
end
to
show
that
the
combination
of
all
metals
at
their
proposed
new
site­
specific
criteria
is
acceptable.
Acceptability
must
be
demonstrated.
with
each
toxicity
test
that
was
used
as
a
primary
toxicity
test
in
the
determination
of
the
WEXs
for
the
individual
metals.
Thus
if
a
different
primary
test
was
used
for
each
metal,
the
number
of
acceptability
tests
needed
would
equal
the
number
of
metals.
It
is
possible
that
a
toxicity
test
used
as
the
primary
test
for
one
metal
might
be
more
sensitive
than
the
'
CMC
(
or
CCC)
for
another
metal
and
thus
might
not
be
usable
in
the
combination
test
unless
antagonism
occurs.
When
a
primary
test
cannot
be
used,
an
acceptable
alternative
test
must
be
used.
The
second
option
is
nreferred
because
it
is
more
definitive;
it
provides
data
for
each
metal
individually
and
for
the
mixture.
The
first
option
leaves
the
possibility
that
one
of
the
metals
is
antagonistic
towards
another­
so
that
the
toxicity
of
the
mixture
would
increase
if
the
metal
causing
the
antagonism
were
not
present.

135
Because
the
National
Toxics
Rule
WTR)
incorporated
WERs
into
the
aquatic
life
criteria
for
some
metals,
it
might
be
envisioned
that
more
than
one
criterion
could
apply
to
a
metal
at
a
site
if
different
investigators
obtained
different
WERs
for
the
same
metal
at
the
site.
.
.
.
.
.
;
EI1
mzqdictrons
sublect
to
the
NTR.
as
well,
l
.
ens,
EPA
intends
that
Dere
s­
d
be
.
no
­
re
one
w&
szaon
for
a
Dolluta
t
at
a
Dolnt
in
a
bide
Thus
whenever
a
site­
specificncriterion
is
to
be
of
water.
derived
using
a
WER
at
a
site
at
which
more
than
one
discharger
has
permit
limits
for
the
same
metal,
it
is
important
that
all
dischargers
work
together
with
the
appropriate
regulatory
authority
to
develop
a
workplan
that
is
designed
to
derive
a
site­
specific
criterion
that
adequately
protects
the
entire
site.

Method
2
is
ideally
suited
for
taking
into
account
more
than
one
discharger.

Method
1
is
straightforward
if
the
dischargers
are
sufficiently
far
downstream
of
each
other
that
the
stream
can
be
divided
into
a
separate
site
for
each
discharger.
Method
1
can
also
be
fairly
straightforward
if
the
WERs
are
additive,
but
it
will
be
complex
if
the
WERs
are
not
additive.
Deciding
whether
to
use
a
simulated
downstream
water
or
an
actual
downstream
water
can
be
difficult
in
a
flowing­
water
multiple­
discharge
situation.
Use
of
actual
downstream
water
can
be
complicated
by
the
existence
of
multiple
mixing
zones
and
plumes
and
by
the
possibility
of
varying
discharge
schedules;
these
same
problems
exist,
however,
if
effluents
from
two
or
more
discharges
are
used
to
prepare
simulated
downstream
water.
Dealing
with
a
multiple­
discharge
situation
is
much
easier
if
the
WERs
are
additive,
and
use
of
simulated
downstream
water
is
the
best
way
to
determine
whether
the
WERs
are
additive.
Taking
into
account
all
effluents
will
take
into
account
synergism,
antagonism,
and
additivity.
If
one
of
the
discharges
stops
or
is
modified
substantially,
however,
it
will
usually
be
necessary
to
determine
a
new
WER,
except
possibly
if
the
metal
being
discharged
is
refractory.
Situations
concerning
intermittent
and
batch
discharges
need
to
be
handled
on
a
case­
by­
case
basis.

Method
1
is
intended
to
apply
not
only
to
ordinary
rivers
and
streams
but
also
to
streams
that
some
people
might
consider
'
special.,
such
as
streams
whose
design
flows
are
zero
and
streams
that
some
state
and/
or
federal
agencies
might
refer
to
as
'
effluent­
dependent',
'
habitat­
creating',
'
effluent­
dominated',
etc.
(
Due
to
differences
between
agencies,
some
streams
whose
design
flows
are
zero
are
not
considered
'
effluent­
dependent',

136
etc.,
and
some
'
effluent­
dependent'
streams
have
design
flows
that
are
greater
than
zero.)
The
application
of
Method
1
to
these
kinds
of
streams
has
the
following
implications:
1.
If
the
design
flow..
is
zero,
at
least
some
WRRs
ought
ti
be
determined
in
100%
effluent.
2.
If
thunderstorms,
etc.,
occasionally
dilute
the
effluent
substantially,
at
least
one
WER
should
be
determined
in
diluted
effluent
to
assess
whether
dilution
by
rainwater
might
result
in
underprotection
by
decreasing
the
WRR
faster
than
it
decreases
the
concentration
of
the
metal.
This
might
occur,
for
example,
if
rainfall
reduces
hardness,
alkalinity,
and
pH
substantially.
This
might
not
be
a
concern
if
the
WER
demonstrates
a
substantial
margin
of
safety.
3.
If
the
site­
specific
criterion
is
substantially
higher
than
the
national
criterion,
there
should
be
increased
concern
about
the
fate
of
the
metal
that
has
reduced
or
no
toxicity.
Even
if
the
WRR
demonstrates
a
substantial
margin
of
safety
(
e.
g.,
if
the
site­
specific
criterion
is
three
times
the
national
criterion,
but
the
experimentally
determined
WE%
is
111,
it
might
be
desirable
to
study
the
fate
of
the
metal.
4.
If
the
stream
merges
with
another
body
of
water
and
a
site­
specific
criterion
is
desired
for
the
merged
waters,
another
WER
needs
to
be
determined
for
the
mixture
of
the
waters.
5.
Whether
WET
testing
is
reguired
is
not
a
WER
issue,
although
WET
testing
might
be
a
condition
for
determining
and/
or
using
aWER.
6.
A
concern
about
what
species
should
be
present
and/
or
protected
in
a
stream
is
a
beneficial­
use
issue,
not
a
WER
issue,
although
resolution
of
this
issue
might
affect
what
species
should
be
used
if
a
WER
is
determined.
(
If
the
Recalculation
Procedure
is
used,
determining
what
species
should
be
present
and/
or
protected
is
obviously
important.)
7.
Human
health
and
wildlife
criteria
and
other
issues
might
restrict
an
effluent
more
than
an
aquatic
life
criterion.
Although
there
are
no
scientific
reasons
why
'
effluent­
dependent',
etc.,
streams
and
streams
whose
design
flows
are
zero
should
be
subject
to
different
guidance
than
other
streams,
a
regulatory
decision
(
for
example,
see
40
CFR
131)
might
require
or
allow
some
or
all
such
streams
to
be
subject
to
different
guidance.
For
example,
it
might
be
decided
on
the
basis
of
a
use
attainability
analysis
that
one
or
more
constructed
streams
do
not
have
to
comply
with
usual
aquatic
life
criteria
because
it
is
decided
that
the
water
quality
in
such
streams
does
not
need
to
protect
sensitive
aquatic
species.
Such
a
decision
might
eliminate
any
further
concern
for
site­
specific
aquatic
life
criteria
and/
or
for
WET
testing
for
such
streams.
The
water
quality
might
be
unacceptable
for
other
reasons,
however.

In
addition
to
its
use
with
rivers
and
streams,
Method
1
is
also
appropriate
for
determining
cmcWERs
that
are
applicable
to
near­
field
effects
of
discharges
into
large
bodies
of
fresh
or
salt
water,
such
as
an
ocean
or
a
large
lake,
reservoir,
or
estuary:

137
a.

b.
The
near­
field
effects
of
a
pipe
that
extends
far
into
a
large
body
of
fresh
or
salt
water
that
has
a
current,
such
as
an
ocean,
can
probably
best
be
treated
the
same
as
a
single
discharge
into
a
flowing
stream.
For
example,
if
a
mixing
zone
is
defined,
the
concentration
of
effluent
at
the
edge
of
the
mixing
zone
might
be
used
to'define
how
to
prepare
a
simulated
site
water.
A
dye
dispersion
study
(
Kilpatrick
1992)
might
be
useful,
but
a
dilution
model
(
U.
S.
EPA
1993)
is
likely
to
be
a
more
cost­
effective
way
of
obtaining
information
concerning
the
amount
of
dilution
at
the
edge
of
the
mixing
zone.
The
near­
field
effects
of
a
single
discharge
that
is
near
a
shore
of
a
large
body
of
fresh
or
salt
water
can
also
probably
best
be
treated
the
same
as
a
single
discharge
into
a
flowing
stream,
especially
if
there
is
a
definite
plume
and
a
defined
mixing
zone.
The
potential
point
of
impact
of
near­
field
effects
will
often
be
an
embayment,
bayou,
or
estuary
that
is
a
nursery
for
fish
and
invertebrates
and/
or
contains
conmercially
important
shellfish
beds.
Because
of
their
importance,
these
areas
should
receive
special
consideration
in
the
determination
and
use
of
a
WER,
taking
into
account
sources
of
water
and
discharges,
mixing­
patterns,
and
currents
(
and
tides
in
coastal
areas).
The
current
and
flushing
patterns
in
estuaries
can
result
in
increased
pollutant
concentrations
in
confined
embayments
and
at
the
terminal
up­
gradient
portion
of
the
estuary
due
to
poor
tidal
flushing
and
exchange.
Dye
dispersion
studies
(
Kilpatrick
1992)
can
be
used
to
determine
the
spatial
concentration
of
the
effluent
in
the
receiving
water,
but
dilution
models
(
U.
S.
EPA
1993)
might
not
be
sufficiently
accurate
to
be
useful.
Dye
studies
of
discharges
in
near­
shore
tidal
areas
are
especially
complex.
Dye
injection
into
the
discharge
should
occur
over
at
least
one,
and
preferably
two
or
three,
complete
tidal
cycles;
subsequent
dispersion
patterns
should
be
monitored
in
the
ambient
water
on
consecutive
tidal
cycles
using
an
intensive
sampling
regime
over
time,
location,
and
depth.
Information
concerning
dispersion
and
the
comunity
at
risk
can
be
used
to
define
the
appropriate
mixing
zone(
s),
which
might
be
used
to
define
how
to
prepare
simulated
site
water.

Kilpatrick,
F.
A.
19'
92.
Simulation
of
Soluble
Waste
Transport
and
Buildup
in
Surface
Waters
Using
Tracers.
Open­
File
Report
92­
457.
U.
S.
Geological
Survey,
Books
and
Open­
File
Reports,
Box
25425,
Federal
Center,
Denver,
CO
80225.

U.
S.
EPA.
1993.
Dilution
Models
for
Effluent
Discharges.
Second
Edition.
EPA/
600/
R­
93/
139.
National
Technical
Information
Service,
Springfield,
VA.

138
­
ix
0:
Additivity
and
the
Two
Compoxmnta
of
a
WER
Determined
Using
Downstrwm
Water
The
Concent
of
Additivitv
of
WERs
In
theory,
whenever
samples
of
effluent
and
upstream
water
are
taken,
determination
of
a
WRR
in
100
%
effluent
would
quantify
the
effluent
WRR
(
eWER)
and
determination
of
a
WRR
in
100
%
upstream
water
would
quantify
the
upstream
WRR
(
uWRR);
determination
of
WRRs
in
known
mixtures
of
the
two
samples
would
demonstrate
whether
the
eWER
and
the
uWRR
are
additive.
For
example,
if
eWRR
=
40,
uWRR
=
5,
and
the
two
WERs
are
additive,
a
mixture
of
20
%
effluent
and
80
%
upstream
water
would
give
a
WER
of
12,
except
possibly
for
experimental
variation,
because:

2O(
emR)
+
8O(
umR)
t
20(
40)
+
80(
S)
t
800
+
400
100
100
100
dE!
L~
2
100
l
Strict
additivity
of
an
eWRR
and
an
uWER
will
probably
be
rare
because
one
or
both
WERs
will
probably
consist
of
a
portion
that
is
additive
and
a
portion
that
is
not.
The
portions
of
the
ewER
and
uWER
thut
are
due
to
refractory
metal
will
be
strictly
additive,
because
a
change
in
water
quality
will
not
make
the
metal
more
or
less
toxic.
In
contrast,
metal
that
is
nontoxic
because
it
is
complexed
by
a
complexing
agent
such
as
RDTA
will
not
be
strictly
additive
because
the
%
uncomplexed
will
decrease
as
the
solution
is
diluted;
the
amount
of
change
in
the
%
uncomplexed
will
usually
be
small
and
will
depend
on
the
concentration
and
the
binding
constant
of
the
complexing
agent
(
see
Appendix
D).
Whether
the
nonrefractory
portions
of
the
UWER
and
eWRR
are
additive
will
probably
also
depend
on
the
differences
between
the
water
quality
characteristics
of
the
effluent
and
the
upstream
water,
because
these
will
determine
the
water
quality
characteristics
of
the
downstream
water.
If,
for
example,
85
%
of
the
eWRR
and
30
%
of
the
uWER
are
due
to
refractory
metal,
the
WRR
obtained
in
the
mixture
of
20
%
effluent
and
80
%
upstream
water
could
range
from
8
to
12.
The
WRR
of
8
would
be
obtained
if
the
only
portions
of
the
eWER
and
uWRR
that
are
additive
are
those
due
to
refractory
metal,
because:

ZO(
O.
85)
(
eMEN
+
8OtO.
30)
(
WER)
­
20(
0.85)
(
40)
+
SO(
O.
30)
(
5)
=
8
100
100
The
WRR
could
be
as
high
as`
12
depending
on
the
percentages
of
the
other
portions
of
the
WERs
that
are
also
additive.
Even
if
the
eWER
and
uWER
are
not
strictly
additive,
the
concept
of
additivity
of
WRRs
can
be
useful
insofar
as
the
eWER
and
uWER
are
partially
additive,
i.
e.,
insofar
as
a
portion
of
at
least
one
of
the
WERs
is
additive.
In
the
example
given
above,
the
WER
determined
using
downstream
water
that
consisted
of
20
%
effluent
139
and
80
%
upstream
water
would
be
12
if
the
eWER
and
uwER
were
strictly
additive;
the
downstream
WER
would
be
less
than
12
if
the
eWER
and
uWER
were
partially
additive.

The
major
advantage
of
additivity
of
WERs
can
be
dmnstrated
using
the
effluent
and
upstream
water
that
were
used
above.
TO
sir&
ify
this
illustration,
the
acute­
chronic
ratio
will
be
assumed
to
be
large,
and
the
eWEF!
of
40
and
the
uWER
of
5
will
be
assumed
to
be
cccWERs
that
will
be
assumed
to
be
due
to
refractory
metal
and
will
therefore
be
strictly
additive.
In
addition,
the
coap?
lete­
mix
downstream
water
at
design­
flow
conditions
will
be
assumed
to
be
20
8
effluent
and
80
%
upstream
water,
so
that
the
downstream
WER
will
be
12
as
calculated
above
for
strict
additivity.

Because
the
ewER
and
the
USER
are
cccWERs
and
are
strictly
additive,
this
metal
will
cause
neither
acute
nor
chronic
toxicity
in
downstream
water
if
(
a)
the
concentration
of
metal
in
the
effluent
is
less
than
40
times
the
CCC
and
(
b)
the
concentration
of
metal
in
the
upstream
water
is
less
than
5
times
the
CCC.
As
the
effluent
is
diluted
by
mixing
with
upstream
water,
both
the
ewER
and
the
concentration
of
metal
will
be
diluted
simultaneously;
proportional
dilution
of
the
metal
and
the
eWER
will
prevent
the
metal
from
causing
acute
or
chronic
toxicity
at
any
dilution.
When
the
upstream
flow
equals
the
design
flow,
the
WER
in
the
plume
will
decrease
from
40
at
the
end
of
the
pipe
to
12
at
complete
mix
as
the
effluent
is
diluted
by
upstream
water;
because
this
WER
is
due
to
refractory
metal,
neither
fate
processes
nor
changes
in
water
quality
characteristics
will
affect
the
WER.
When
stream
flow
is
higher
or
lower
than
design
flow,
the
complete­
mix
WER
will
be
lower
or
higher,
respectively,
than
12,
but
toxicity
will
not
occur
because
the
concentration
of
metal
will
also
be
lower
or
higher.

If
the
eWER
and
the
uWER
are
strictly
additive
and
if
the
national
CCC
is
1
mg/
L,
the
following
conclusions
are
valid
when
the
concentration
of
the
metal
in
100
%
effluent
is
less
than
40
xng/
L
and
the
concentration
of
the
metal
in
100
%
upstream
water
is
less
than
5
mg/
L:
1.
This
metal
will
not
cause
acute
or
chronic
toxicity
in
the
upstream
water,
in
100
%
effluent,
in
the
plume,
or
in
downstream
water.
2.
There
is
no
need
for
an
acute
or
a
chronic
mixing
zone
where
a
lesser
degree
of
protection
is
provided.
3.
If
no
mixing
zone
exists,
there
is
no
discontinuity
at
the
edge
of
a
mixing
zone
where
the
allowed
concentration
of
metal
decreases
instantaneously.
These
results
also
apply
to
partial
additivity
as
long
as
the
concentration
of
metal
does
not
exceed
that
allowed
by
the
amount
140
of
additivity
that
exists.
It
would
be
more
difficult
to
take
into
account
the
portions
of
the
eWER
and
uWER
that
are
not
additive.

The
concept
of
additivity
becomes
unimportant
when
the
ratios,
concentrations
of
the
metals,
or
WRRs
are
very
different.
For
example,
if
eWER
=
40,
uWER
=
5,
and
they
are
additive,
a
mixture
of
1
%
effluent
and
99
%
upstream
water
would
have
a
WRR
of
5.35.
Given
the
reproducibility
of
toxicity
tests
and
WERs,
it
would
be
extremely
difficult
to
distinguish
a
WER
of
5
from
a
WER
of
5.35.
In
cases
of
extreme
dilution,
rather
than
experimentally
determining
a
WER,
it
is
probably
acceptable
to
use
the
limiting
WER
of
5
or
to
calculate
a
WRR
if
additivity
has
been
demonstrated.

Traditionally
it
has
been
believed
that
it
is
environmentally
conservative
to
use
a
WER
determined
in
upstream
water
(
i.
e.,
the
uWRR)
to
derive
a
site­
specific
criterion
that
applies
downstream
(
i.
e.,
that
applies
to
areas
that
contain
effluent).
This
belief
is
probably
based
on
the
assumption
that
a
larger
WER
would
be
obtained
in
downstream
water
that
contains
effluent,
but
the
belief
could
also
be
based
on
the
assumption
that
the
uWER
is
additive.
It
is
possible
that
in
some
cases
neither
assumption
is
true,
which
means
that
using
a
uWRR
to
derive
a
downstream
site­
specific
criterion
might
result
in
underprotection.
It
seems
likely,
however,
that
WRRs
determined
using
downstream
water
will
usually
be
at
least
as
large
as
the
uWER.

Several
kinds
of
concerns
about
the
use
of
WERs
are
actually
concerns
about
additivity:
1.
Do
WERs
need
to
be
determined
at
higher
flows
in
addition
to
being
determined
at
design
flow?
2,
Do
WERs
need
to
be
determined
when
two
bodies
of
water
mix?
3.
Do
WERs
need
to
be
determined
for
each
additional
effluent
in
a
multiple­
discharge
situation.
In
each
case,
the
best
use
of
resources
might
be
to
test
for
additivity
of
WERs.

Mixinu
Zones
In
the
example
presented
above,
there
would
be
no
need
for
a
regulatory
mixing
zone
with
a
reduced
level
of
protection
if:
1.
The
eWRR
is
always
40
and
the
concentration
of
the
metal
in
100
%
effluent
is
always
less
than
40
mg/
L.
2.
The
uWER
is
always
5
and
the
concentration
of
the
metal
in
100
%
upstream
water
is
always
less
than
5
mg/
L.
3.
The
WERs
are
strictly
additive.
If,
however,
the
concentration
exceeded
40
mg/
L
in
100
%
effluent,
but
there
is
some
assimilative
capacity
in
the
upstream
water,
a
regulatory
mixing
zone
would
be
needed
if
the
discharge
were
to
be
allowed
to
utilize
some
or
all
of
the
assimilative
141
capacity.
The
concept
of
additivity
of
WERs
can
be
used
to
calculate
the
maximum
allowed
concentration
of
the
metal
in
the
effluent
if
the
eWER
and
the
uWER
are
strictly
additive.

If
the
concentration
of
metal
in
the
upstream
water
never
exceeds
0.8
mg/
L,
the
discharger
might
want'to
determine
how
much
above
40
mg/
L
the
concentration
could
be
in
100
%
effluent.
If,
for
example,
the
downstream
water
at
the
edge
of
the
chronic
mixing
zone
under
design­
flow
conditions
consists
of
70
%
effluent
and
30
b
upstream
water,
the
WER
that
would
apply
at
the
edge
of
the
mixing
zone
would
be:

70(
e)
+
3o(
uwim)
­
70(
40)
+
30(
s)
I
2800
+
150
I
100
100
100
2g
.
5
.

Therefore,
the
maximum
concentration
allowed
at
this
point
would
be
29.5
lng/
L.
If
the
concentration
of
the
metal
in
the
upstream
water
was
0.8
m/
L,
the
maximum
concentration
allowed
in
100
%
effluent
would
be
41.8
mg/
L
because:

70(
41.8
a&
L)
l
3OfO.
8
n&
L)
I
2926
m/
q/
L
+
24
I&
100
100
L
=
29.5alg/
L*

Because
the
eWER
is
40,
if
the
concentration
of
the
metal
in
100
0
effluent
is
41.8
mg/
L,
there
would
be
chronic
toxicity
inside
the
chronic
mixing
zone.
If
the
concentration
in
100
%
effluent
is
greater
than
41.8
mg/
L,
there
would
be
chronic
toxicity
past
the
edge
of
the
chronic
mixing
zone.
Thus
even
if
the
eWER
and
the
uWER
are
taken
into
account
and
they
are
assumed
to
be
completely
additive,
a
mixing
zone
is
necessary
if
the
assimilative
capacity
of
the
upstream
water
is
used
to
allow
discharge
of
more
metal.

If
the
complete­
mix
downstream
water
consists
of
20
%
effluent
and
80
8
upstream
water
at
design
flow,
the
complete­
mix
WER
would
be
12
as
calculated
above.
The
complete­
mix
approach
to
determining
and
using
downstream
WERs
would
allow
a
maximum
concentration
of
12
xng/
L
at
the
edge
of
the
chronic
mixing
zone,
whereas
the
alternative
approach
resulted
in
a
maximum
allowed
concentration
of
29.5
mg/
L.
The
complete­
mix
approach
would
allow
a
maximum
concentration
of
16.8
mg/
L
in
the
effluent
because:

70(
16.8
n&
L)
+
3OtO.
8
a&
L)
I
1176
m&
L
+
24
m&
L
­
100
100
12
n&
L.

In
this
example,
the
complete­
mix
approach
limits
the
concentration
of
the
metal
in
the
effluent
to
16.8
mg/
L,
even
though
it
is
known
that
as
long
as
the
concentration
in
100
%
effluent
is
leas
than
40
mg/
L,
chronic
toxicity
will
not
occur
inside
or
outside
the
mixing
zone.
If
the
WER
of
12
is
used
to
derive
a
site­
specific
CCC
of
12
mg/
L
that
is
applied
to
a
site
142
that
starts
at
the
edge
of
the
chronic
mixing
zone
and
extends
all
the
way
across
the
stream,
there
would
be
overprotection
at
the
edge
of
the
chronic
mixing
zone
(
because
the
maximum
allowed
concentration
is
12
mg/
L,
but
a
concentration
of
29.5
mg/
L
will
not
cause
chronic
toxicity),
whereas
there
would
be
underprotection
on
the
other
side
of­
the
stream
(
because
the
maximum
allowed
concentration
is
12
mg/
L,
but
concentrations
above
5
mg/
L
can
cause
chronic
toxicity.)

The
Rxnerimental
Determination
of
Additivitv
Experimental
variation
makes
it
difficult
to
quantify
additivity
without
determining
a
large
number
of
WERs,
but
the
advantages
of
demonstrating
additivity
might
be
sufficient
to
make
it
worth
the
effort.
It
should
be
possible
to
decide
whether
the
eWER
and
uWER
are
strictly
additive
based
on
determination
of
the
eWER
in
100
%
effluent,
determination
of
the
uWER
in
100
%
upstream
water,
and
determination
of
WERs
in
1:
3,
l:
l,
and
3:
l
mixtures
of
the
effluent
and
upstream
water,
i.
e.,
determination
of
WERs
in
100,
75,
50,
25,
and
0
%
effluent.
Validating
models
of
partial
additivity
and/
or
interactions
will
probably
require
determination
of
more
WERs
and
more
sophisticated
data
analysis
(
see,
for
example,
Broderius
1991).

In
some
cases
chemical
measurements
or
manipulations
might
help
demonstrate
that
at
least
some
portion
of
the
eWER
and/
or
the
uWRR
is
additive:
1.
If
the
difference
between
the
dissolved
WER
and
the
total
recoverable
WER
is
explained
by
the
difference
between
the
dissolved
and
total
recoverable
concentrations,
the
difference
is
probably
due
to
particulate
refractory
metal.
2.
If
the
WERs
in
different
samples
of
the
effluent
correlate
with
the
concentration
of
metal
in
the
effluent,
all,
or
nearly
all,
of
the
metal
in
the
effluent
is
probably
nontoxic.
3.
A
WEX
that
remains
constant
as
the
pH
is
lowered
to
6.5
and
raised
to
9.0
is
probably
additive.
The
concentration
of
refractory
metal
is
likely
to
be
low
in
upstream
water
except
during
events
that
increase
TSS
and/
or
WC;
the
concentration
of
refractory
metal
is
more
likely
to
be
substantial
in
effluents.
Chemical
measurements
might
help
identify
the
percentages
of
the
eWRR
and
the
uWER
that
are
due
to
refractory
metal,
but
again
experimental
variation
will
limit
the
usefulness
of
chemical
measurements
when
concentrations
are
low.

The
distinction
between
the
two
components
of
a
WER
determined
using
downstream
water
has
the
following
implications:
1.
The
magnitude
of
a
WER
determined
using
downstream
water
will
usually
depend
on
the
percent
effluent
in
the
sample.

143
2.

3.

4.

5.

6.
Insofar
as
the
eWRR
and
uWRR
are
additive,.
the
magnitude
of
a
downstream
WER
can
be
calculated
from
the
eWER,
the
UWER,
and
the
ratio
of
effluent
and
upstream
water­
in
the
downstream
water.
The
derivation
and
implementation
of
site­
specific
criteria
should
ensure
that
each
component
is
applied
only
where
it
occurs.
a.
Underprotection
will
occur
if,
for
example,
any
portion
of
the
eWER
is
applied
to
an
area
of
a
stream
where
the
effluent
does
not
occur.
b.
Overprotection
will
occur
if,
for
example,
an
unnecessarily
small
portion
of
the
eWER
is
applied
to
an
area
of
a
stream
where
the
effluent
occurs.
Even
though
the
concentration
of
metal
might
be
higher
than
a
criterion
in
both
a
regulatory
mixing
zone
and
a
plume,
a
reduced
level
of
protection.
is
allowed
in
a
mixing
zone,
whereas
a
reduced
level
of
protection
is
not
allowed
in
the
portion
of
a
plume
that
is
not
inside
a
mixing
zone.
Regulatory
mixing
zones
are
necessary
if,
and
only
if,
a
discharger
wants
to
make
use
of
the
assimilative
capacity
of
the
upstream
water.
It
might
be
cost­
effective
to
quantify
the
eWER
and
uWER,
determine
the
extent
of
additivity,
study
variability
over
time,
and
then
decide
how
to
regulate
the
metal
in
the
effluent.

Peference
Broderius,
S.
J.
1991.
Modeling
the
Joint
Toxicity
of
Xenobiotics
to
Aquatic
Organisms:
Basic
Concepts
and
Approaches.
In:
Aquatic
Toxicology
and
Risk
Assessment:
Fourteenth
Volume.
(
M.
A.
Mayes
and
M.
G.
Barron,
eds.)
ASTM
STP
1124.
American
Society
for
Testing
and
Materials,
Philadelphia,
PA.
pp.
107­
127.

144
Appendix
EI:
Spmcirl
Conddorationm
Concerning
the
Deteminatiozh
of
WHIl
with
Saltwater
Spociem
1.
The
test
organisms
should
be
compatible
with
the
salinity
of
the
site
water,
and
the
salinity
of
the
laboratory
dilution
water
should
match
that
of
the
site
water.
Low­
salinity
stenohaline
organisms
should
not
be
tested
in
high­
salinity
water,
whereas
high­
salinity
stenohaline
organisms
should
not
be
tested
in
low­
salinity
water;
it
is
not
known,
however,
whether
an
incompatibility
will
affect
the
WER.
If
the
community
to
be
protected
principally
consists
of
euryhaline
species,
the
primary
and
secondary
toxicity
tests
should
use
the
euryhaline
species
suggested
in
Appendix
I
(
or
taxonomically
related
species)
whenever
possible,
although
the
range
of
tolerance
of
the
organisms
should
be
checked.
a.
When
Method
1
is
used
to
determine
cmcWRRs
at
saltwater
sites,
the
selection
of
test
organisms
is
complicated
by
the
fact
that
most
effluents
are
freshwater
and
they
are
discharged
into
salt
waters
having
a
wide
range
of
salinities.
Some
state
water
quality
standards
require
a
permittee
to
meet
an
LCSO
or
other
toxicity
limit
at
the
end
of
the
pipe
using
a
freshwater
species.
However,
the
intent
of
the
site­
specific
and
national
water
quality
criteria
program
is
to
protect
the
communities
that
are
at
risk.
Therefore,
freshwater
species
should
not
be
used
when
WRRs
are
determined
for
saltwater
sites
unless
such
freshwater
species
(
or
closely
related
species)
are
in
the
community
at
risk.
The
addition
of
a
small
amount
of
brine
and
the
use
of
salt­
tolerant
freshwater
species
is
inappropriate
for
the
same
reason.
The
addition
of
a
large
amount
of
brine
and
the
use
of
saltwater
species
that
require
high
salinity
should
also
be
avoided
when
salinity
is
likely
to
affect
the
toxicity
of
the
metal.
Salinities
that
are
acceptable
for
testing
euryhaline
species
can
be
produced
by
dilution
of
effluent
with
sea
water
and/
or
addition
of
a
commercial
sea
salt
or
a
brine
that
is
prepared
by
evaporating
site
water;
small
increases
in
salinity
are
acceptable
because
the
effluent
will
be
diluted
with
salt
water
wherever
the
cozununities
at
risk
are
exposed
in
the
real
world.
Only
as
a
last
resort
should
freshwater
species
that
tolerate
low
levels
of
salinity
and
are
sensitive
to
metals,
such
as
Danhnia
mauna
and
Hvalella
azteca,
be
used.
b.
When
Method
2
is
used
to
determine
cccWERs
at
saltwater
sites:
1)
If
the
site
water
is
low­
salinity
but
all
the
sensitive
test
organisms
are
high­
salinity
btenohaline
organisms,
a
commercial
sea
salt
or
a
brine
that
is
prepared
by
evaporating
site
water
may
be
added
in
order
to
increase
the
salinity
to
the
minimum
level
that
is
acceptable
to
the
test
organisms;
it
should
be
determined
whether
the
145
salt
or
brine
reduces
the
toxicity
of
the
metal
and
thereby
increases
the
WER.
2)
If
the
site
water
is
high­
salinity;
selecting
test
organisms
should
not
be
difficult
because
many
of
the
sensitive
test
organisms
are
compatible
with
high­
salinity
water.

2.
It
is
especially
important
to
consider
the
availability
of
test
organisms
when
saltwater
species
are
to
be
used,
because
many
of
the
cwnly
used
saltwater
species
are
not
cultured
and
are
only
available
seasonally.

3.
Many
standard
published
methodologies
for
tests
with
saltwater
species
reconzaen
d
filtration
of
dilution
water,
effluent,
and/
or
teat
solutions
through
a
37­
m
sieve
or
screen
to
remove
predators.
Site
water
should
be
filtered
only
if
predators
are
observed
in
the
sample
of
the
water
because
filtration
might
affect
toxicity.
Although
recommended
in
some
test
methodologies,
ultraviolet
treatment
is
often
not
needed
and
generally
should
be
avoided.

4.
If
a
natural
salt
water
is
to
be
used
as
the
laboratory
dilution
water,
the
samples
should
probably
be
collected
at
slack
high
tide
(&
2
hours).
Unless
there
is
stratification,
samples
should
probably
be
taken
at
mid­
depth;
however,
if
a
water
quality
characteristic,
such
as
salinity
or
TSS,
is
important,
the
vertical
and
horizontal
definition
of
the
point
of
sampling
might
be
important.
A
conductivity
meter,
aalinometer,
and/
or
transmissometer
might
be
useful
for
determining
where
and
at
what
depth
to
collect
the
laboratory
dilution
water;
any
measurement
of
turbidity
will
probably
correlate
with
TSS.

5.
The
salinity
of
the
laboratory
dilution
water
should
be
within
f
10
percent
or
2
mg/
L
(
whichever
is
higher)
of
that
of
the
site
water.

146
appendix
It
Su~
3u;
ftod
Toxicity
Tostn
for
Dotozminhg
WBIU
for
Selecting
primary
and'
secondary
toxicity
tests
for
determining
WERs
for
metals
should
take
into
account
the
following:
1.

2.

3,

4.
5.

6.

7.

8.

9.
WERs
determined
with
more
sensitive
tests
are
likely
to
be
larger
than
WERs
determined
with
less
sensitive
tests
(
see
Appendix
D).
Criteria
are
derived
to
protect
sensitive
species
and
so
WRRs
should
be
derived
to
be
appropriate
for
sensitive
species.
The
appropriate
regulatory
authority
will
probably
accept
WRRs
derived
with
less
sensitive
tests
because
such
WERs
are
likely
to
provide
at
least
as
much
protection
as
WRRs
determined
with
more
sensitive
tests.
The
species
used
in
the
primary
and
secondary
tests
must
be
in
different
orders
and
should
include
a
vertebrate
and
an
invertebrate.
The
test
organism
(
i.
e.,
species
and
life
stage)
should
be
readily
available
throughout
the
testing
period.
The
chances
of
the
test
being
successful
should
be
high.
The
relative
sensitivities
of
test
organisms
vary
substantially
from
metal
to
metal.
The
sensitivity
of
a
species
to
a
metal
usually
depends
on
both
the
life
stage
and
kind
of
test
used.
Water
guality
characteristics
might
affect
chronic
toxicity
differently
than
they
affect
acute
toxicity
(
Spehar
and
Carlson
1984;
Chapman,
unpublished;
Voyer
and
McGovern
1991).
The
endpoint
of
the
primary
test
in
laboratory
dilution
water
should
be
as
close
as
possible
(
but
must
not
be
below)
the
CMC
or
CCC
to
which
the
WER
is
to
be
applied;
the
endpoint
of
the
secondary
test
should
be
as
close
as
possible
(
and
should
not
be
below)
the
CMC
or
CCC.
Designation
of
tests
as
acute
and
chronic
has
no
bearing
on
whether
they
may
be
used
to
determine
a
cmcwER
or
a
cccWRR.
The
suggested
toxicity
tests
should
be
considered,
but
the
actual
selection
should
depend
on
the
specific
circumstances
that
apply
to
a
particular
WER
determination.

Regardless
of
whether
test
solutions
are
renewed
when
tests
are
conducted
for
other
purposes,
if
the
concentrations
of
dissolved
metal
and
dissolved
oxygen
remain
acceptable
when
determining
WRRs,
tests
whose
duration
is
not
longer
than
48
hours
may
be
static
tests,
whereas
tests
whose
duration
is
longer
than
48
hours
must
be
renewal
tests.
If
the
concentration
of
dissolved
metal
and/
or
the
concentration
of
dissolved
oxygen
does
not
remain
acceptable,
the
test
solutions
must
be
renewed
every
24
hours.
If
one
test
in
a
pair
of
side­
by­
side
tests
is
a
renewal
test,
both
of
the
tests
must
be
renewed
on
the
same
schedule.

Appendix
H
should
be
read
if
WRRs
are
to
be
determined
with
saltwater
species.

147
Suggested
Tests
1
for
Determining
cmcWERs
and
cccWERs2
(
Concentrations
are
to
be
measured
in
all
tests.)

JlY&!&
d
CRICWERS'
CCCWERS'
Netal
Aluminum
Arsenic(
II1)

Cadmium
Chrom(
II1)

chmm(
vx)

Capx=
r
Lead
Mercury
Nickel
Selenium
Silver
Zinc
X
CDC
Fw
DA
X
FMC
Em
F­
MC
X
CDC
GM
FM
AR
X
X
Y
Y
FMC
BMC
Y
X
FMC
BMC
FMC
BMC
Fw
SW
DA
BM
GM
CR
CDC
MYC
SLs
or
FM
CR
Fw
SW
DA
MY
CDC
MYC
Fw
GM
SL
or
DA
FMC
Fw
SW
DA
MY
GM
NE
CDC
MYC
Tw
SW
DA
EM
FM
or
GM
AR
CDC
BMC
Fw
SW
GM
MYC
CDC
MYC
DA
BM
Fw
SW
DA
MY
GM
EM
Y
Y
Fw
SW
DA
MY
FX
EM
CDC
MYC
Fw
SW
Y
CR
Y
MYC
Y
MYC
Fw
SW
DA
BM
FMC
CR
CDC
MYC
Fw
SW
DA
BM
FM
MY
CDC
MYC
The
description
of
a
test
specifies
not
only
the
test
species
and
the
duration
of
the
test
but
also
the
life
stage
of
the
species
and
the
adverse
effect(
s)
on
which
the
endpoint
is
to
be
based.

Some
tests
that
are
sensitive
and
are
used
in
criteria
documents
are
not
suggested
here
because
the
chances
of
the
test
organisms
being
available
and
the
test
being
successful
might
be
low.
Such
tests
may
be
used
if
desired.

148
FW=
Fresh
Water;
SW
=
Salt
Water.

Two­
letter
codes
are
used
for
acute
tests,
whereas
codes
for
chronic
tests
contain
three
letters
and
end
in
'
C'.
One­
letter
codes
are
used
for
cosunents.

In
acute
tests
on
cadmium
with
salmonids,
substantial
numbers
of
fish
usually
die
after
72
hours.
Also,
the
fish
are
sensitive
to
disturbance,
and
it
is
sometimes
difficult
to
determine
whether
a
fish
is
dead
or
immobilized.

AR.
A
48­
hr
EC50
based
on
mortality
and
abnormal
development
from
a
static
test
with
embryos
and
larvae
of
sea
urchins
of
a
species
in
the
genus
Arbacia
(
ASTM
1993a)
or
of
the
species
gtronwlocentrotus
ournuratus
(
Chapman
1992).

EM.
A
48­
hr
EC50
based
on
mortality
and
abnormal
larval
development
from
a
static
test
with
embryos
and
larvae
of
a
species
in
one
of
four
genera
(
Crassostrea,
Mulinia,
Mvtilus,
flercenaria)
of
bivalve
molluscs
(
ASTM
1993b).

CR.
A
48­
hr
EC50
(
or
LC50
if
there
is
no
immobilization)
from
a
static
test
with
Acartia
or
larvae
of
a
saltwater
crustacean;
if
molting
does
not
occur
within
the
first
48
hours,
renew
at
48
hours
and
continue
the
test
to
96
hours
(
ASTM
1993a).

DA.
A
48­
hr
EC50
(
or
LC50
if
there
is
no
immobilization)
from
a
static
test
with
a
species
in
one
of
three
genera
(
U.
S.
EPA?;
93
(
CeriodaD
'
g
DaDhnia
SimoceDhalus)
in
the
family
Daphnidae
a;
ASTM
i993a).

FM.
A
48­
hr
LC50
from
a
static
test
at
25OC
with
fathead
minnow
.
gro
elas)
larvae
that
are
1
to
24
hours
old
(
ASTM
EPAm1993a)
The
embryos
must
be
hatched
in
the
laboratory
dilution
water,
except
that
organisms
to
be
used
in
the
site
water
may
be
hatched
in
the
site
water.
The
larvae
mumt
not
be
fed
before
or
during
the
test
and
at
least
90
percent
muat
survive
in
laboratory
dilution
water
for
at
least
six
days
after
hatch.
Note:
The
following
48­
hr
LC5Os
were
obtained
at
a
hardness
of
50
mg/
L
with
fathead
minnow
larvae
that
were
1
to
24
hours
old.
The
metal
was
measured
using
the
total
recoverable
procedure
(
Peltier
1993)
:

CS
LC50
(
us/
L)
13.87
Copper
6.33
Zinc
100.95
149
FX.
A
96­
hr
LC50
from
a
renewal
test
(
renew
at
48
hours)
at
25OC
with
fathead
minnow
(
PimeDhaleS
nro
elas)
larvae
that
are
1
to
24
hours
old
(
ASTM
1993a;
U.
S.
EFA
1993a).
The
embryos
atit
be
hatched..
in
the
laboratory
dilution
water,
except
that
organisms
to
be
used
in
the
site
water
may
be
hatched
in
the
site
water.
The
larvae
m;
ut
rrdt
be
fed
before
or
during
the
test
and
at
least
90
percent
ru8t
survive
in
laboratory
dilution
water
for
at
least
six
days
after
hatch.
Note:
A
96­
hr
LC50
of
188.14
w/
L
was
obtained
at
a
hardness
of
50
mg/
L
in
a
test
on
nickel
with
fathead
minnow
larvae
that
were
1
to
24
hours
old.
The
metal
was
measured
using
the
total
recoverable
procedure
(
Peltier
1993).
A
96­
hr
LC50
is
used
for
nickel
because
substantial
mortality
occurred
after
48
hours
in
the
test
on
nickel,
but
not
in
the
tests
on
cadmium,
copper,
and
zinc.

GM.
A
96­
hr
EC50
(
or
LC50
if
there
is
no
izzaobilization)
from
a
renewal
test
(
renew
at
48
hours)
with
a
species
in
the
genus
Gamnaw
(
ASTM
1993a).

MY.
A
96­
hr
EC50
(
or
LC50
if
there
is
no
inznobilization)
from
a
renewal
test
(
renew
at
48
hours)
with
a
species
in
one
of
two
.
.
genera
(
JWsidoDsis,
Holmeg&
mvsis
[
nee
Acanthomvsisl)
in
the
family
Mysidae
(
U.
S.
EPA
1993a;
ASTM
1993a).
Feeding
is
required
during
all
acute
and
chronic
tests
with
mysids;
for
determining
WEF&,
mysids
should
be
fed
four
hours
before
the
renewal
at
48
hours
and
minimally
on
the
non­
renewal
days.

NE.
A
96­
hr
LCSO
from
a
renewal
test
(
renew
at
48
hours)
using
juvenile
or
adult
polychaetes
in
the
genus
Nereidae
(
ASTM
1993a).

SL.
A
96­
hr
EC50
(
or
LC50
if
there
is
no
immobilization)
from
a
renewal
test
(
renew
at
48
hours)
with
a
species
in
one
of
two
salmo)
in
the
family
Salmonidae
(
ASTM
CHRONIC
m
BMC.
A
7­
day
IC25
from
a
survival
and
development
renewal
test
(
renew
every
48
hours)
with
a
species
of
bivalve
mollusc,
such
as
a
species
in
the
genus
Mulinig.
One
such
test
has
been
described
by
Burgess
et
al.
1992.
[
Note:
When
determining
WEBS,
sediment
must
not
be
in
the
test.
chamber.]
[
Note:
This
test
has
not
been
widely
used.
1
CDC.
A
7­&
y
IC25
based
on
reduction
in
survival
and/
or
reproduction
in
a
renewal
test
with
a
species
in
the
genus
in
the
family
Daphnidae
(
U.
S.
EPA
1993b).
The
150
test
solutions
mum+
be
renewed
every
48
hours.
(
A
21­
day
life­
cycle
test
with
DaDhI'&
manna
is
also
acceptable.,)

FMC.
A
7­
day
IC25
from
a
survival
and
growth
renewal
test
(
renew
every
48
hours)
with
larvae
(
s
48­
hr
old)
of
the
fathead
minnow
(
PimeDhales
promela%)
(
U;
S.
EPA
1993b).
When
determining
WERs,
the
fish
mumt
be
fed
four
hours
before
each
renewal
and
minimally
during
the
non­
renewal
days.

MYC.
A
7­
day
IC25
based
on
reduction
in
survival,
growth,
and/
or
reproduction
in
a
renewal
test
with
a
species
in
one
of
two
genera
(
Mvs'd
DS'S
family
Mysiiag
(
i.
k.
Holmesimvsis
[
nee
Acanthomvsis))
in
the
EPA
1993c).
Mysids
mwt
be
fed
during
all
acute
and
chronic
tests;
when
determining
WERs,
they
must
be
fed
four
hours
before
each
renewal.
The
test
solutions
must
be
renewed
every
24
hours.

NEC.
A
20­
day
IC25
from
a
survival
and
growth
renewal
test
(
renew
every
48
hours)
with
a
species
in
the
genus
Neanthes
(
Johns
et
al.
1991).
[
Note:
When
determining
WERs,
sediment
mwt
not
be
in
the
test
chamber.]
[
Note:
This
test
has
not
been
widely
used.]

COMMENTS
X.
Another
sensitive
test
cannot
be
identified
at
this
time,
and
so
other
tests
used
in
the
criteria
document
should
be
considered.

Y.
Because
neither
the
CCCs
for
mercury
nor
the
freshwater
criterion
for
selenium
is
based
on
laboratory
data
concerning
toxicity
to
aquatic
life,
they
cannot
be
adjusted
using
a
WER.

REFERENCES
Asm.
1993a.
Guide
for
Conducting
Acute
Toxicity
Tests
with
Fishes,
Macroinvertebrates,
and
Amphibians.
Standard
E729.
American
Society
for
Testing
and
Materials,
Philadelphia,
PA.

ASTM.
1993b.
Guide
for
Conducting
Static
Acute
Toxicity
Tests
Starting
with
Embryos
of
Four
Species
of
Saltwater
Bivalve
Molluscs.
Standard
E724.
American
Society
for
Testing
and
Materials,
Philadelphia,
PA.

Burgess,
R.,
G.
Morrison,
and
S.
Rego.
1992.
Standard
Operating
Procedure
for
7­
day
Static
Sublethal
Toxicity
Tests
for
Mulinia
lateralis.
U.
S.
EPA,
Environmental
Research
Laboratory,
Narragansett,
RI.

151
Chapman,
G.
A.
1992.
Sea
Urchin
(
Stronwlocent
otus
Fertilization
Test
Method.
U.
S.
EPA,
Newport,
f;
R.
­
1
Johns,
D.
M.,
R.
A.
Pastorok,
and
T.
C.
Ginn.
1991.
A
Sublethal
Sediment
Toxicity
Test
using
Juvenile
EJeantheg
sp.
(
Polychaeta:
Nereidae)
.
In:
Aquatic­
Toxicology
and
Risk
Assessment:
Fourteenth
Volume.
ASTM
STP
1124.
(
M.
A.
Mayes
and
M.
G.
Barron,
eds.)
American
Society
for
Testing
and
Materials,
Philadelphia,
PA.
pp.
280­
293.

Peltier,
W.
H.
1993.
Memorandum.
to
C.
E.
Stephan.
October
19.

Spehar,
R.
L.,
and
A.
R.
Carlson.
1984.
Derivation
of
Site­
Specific
Water
Quality
Criteria
for
Cadmium
and
the
St.
Louis
River
Basin,
Duluth,
Minnesota.
Environ.
Toxicol.
Chem.
3:
651­
665.

U.
S.
EPA.
1993a.
Methods
for
Measuring
the
Acute
Toxicity
of
Effluents
and
Receiving
Waters
to
Freshwater
and
Marine
Organisms.
Fourth
Edition.
EPA/
600/
4­
90/
027F.
National
Technical
Information
Service,
Springfield,
VA.

U.
S.
EPA.
1993b.
Short­
term
Methods
for
Estimating
the
Chronic
Toxicity
of
Effluents
and
Receiving
Waters
to
Freshwater
Organisms.
Third
Edition.
EPA/
600/
4­
91/
002.
National
Technical
Information
Service,
Springfield,
VA.

U.
S.
EPA.
1993c.
Short­
term
Methods
for
Estimating
the
Chronic
Toxicity
of
Effluents
and
Receiving
Waters
to
Marine
and
Estuarine
Organisms.
Second
Edition.
EPA/
600/
4­
91/
003.
National
Technical
Information
Service,
Springfield,
VA.

Voyer,
R.
A.,
and
D.
G.
McGovern.
1991.
Influence
of
Constant
and
Fluctuating
Salinity
on
Responses
of
WsidoDsis
bahia
Exposed
to
Cadmium
in
a
Life­
Cycle
Test.
Aquatic
Toxicol.
19:
215­
230.

152
nix
3:
Rmc
mod
Saltm
of
Hotala
The
following
salts
are
recommended
for
use
when
determining
a
WEH
for
the
metal
listed.
If
available,
a
salt
that
meets
American
Chemical
Society
(
ACS)
specifications
for
reagent­
grade
should
be
used.

Aluminum
*
Aluminum
chloride
(
j­
hydrate:
AlCl,=
6H2O
Aluminum
sulfate
18­
hydrate:
A12(
S0,)
1*
18H,
0
Aluminum
potassium
sulfate
la­
hydrate:
AlK(
S0,),=
12HI0
ite:
NaAs02
enate
'
I­
hydrate,
dibasic:
Na2HAs0,=
7H20
­
Cadmium
chloride
2.5~
hydrate:
CdC12=
2.5H20
Cadmium
sulfate
hydrate:
3CdS0,=
8H20
.
Sk0
um(
I
*
Chrtkic
%
oride
6­
hydrate
(
Chromium
chloride):
CrCl,=
6HiO
*
Chromic
nitrate
g­
hydrate
(
Chromium
nitrate):
Cr(
NO,),*
9&
0
Chromium
potassium
sulfate
12­
hydrate:
CrK(
S0,12*
12H20
Potassium
chromate:
K2­
04
Potassium
dichromate:
K,
Cr,
O,
*
Sodium
chromate
I­
hydrate:
­
Na2Cr0,­
4H,
O
Sodium
dichromate
2­
hydrate:
Na2Cr20,=
2H20
izQRRss
*
Cupric
Cupric
Cupric
chloride
a­
hydrate
(
Copper
chloride):
CuC12*
2Ha0
nitrate
2.5­
hydrate
(
Copper
nitrate):
CU(
NO,)~~~.~
H~
O
sulfate
5­
hydrate
(
Copper
sulfate):
CuS0,*
5H20
*
Lead
chloride:
PbC12
Lead
nitrate:
Pb(
NOB)
2
Pe
CUT
Mrrcuric
chloride:
HgC12
Mercuric
nitrate
monohydrate:
Hg(
N03)
2*
H20
Mercuric
sulfate:
HgSO,

153
Nickel
l
Nickelous
chloride
6­
hydrate
(
Nickel
chloride)
:
NiC12=
6H20
l
Nickelous
nitrate
6­
krydrate
(
Nickel
nitrate):
Ni(
NO,),*
6H,
O
Nickelous
sulfate
6­
hydrate
(
Nickel
sulfate):
NiS0,=
6H20.

*
Sodium
selenite
5­
hydrate:
Na,
SeO,­
5H,
O
*
Sodium
selenate
lo­
hydrate:
Na2Se0,=
10H20
Silver
Silver
nitrate:
AgNO,
(
Even
if
acidified,
standards
and
samples
containing
silver
rwt
be
in
amber
containers.)

Zinc
Zinc
chloride:
ZnClr
*
Zinc
nitrate
6­
hydrate:
Zn(
N0,12*
6H,
0
Zinc
sulfate
7­
hydrate:
ZnS0,=
7H20
*
Note:
ACS
reagent­
grade
specifications
might
not
be
available
for
this
salt.

No
salt
should
be
used
until
information
concerning
the
safety
and
handling
of
that
salt
has
been
read.

154