Document ID: EPA-HQ-OW-2003-0074-0257
Agency: epa
Document Type: Supporting & Related Material
Title: 
Posted Date: 2003-12-24T05:00Z

Abt
Associates
Inc.
­
1­
memorandum
Environmental
Research
Area
4800
Montgomery
Lane,
Suite
600
#
Bethesda,
MD
20814­
5341
#
(
301)­
913­
0500
Date:
February
24,
2003
To:
Lynn
Zipf,
USEPA/
Office
of
Water
From:
Susan
Keane
and
Kristina
Watts,
Abt
Associates
Subject:
Evaluation
of
the
Appropriateness
of
Representative
Chemicals
for
Influential
Chemical
Groups
(
Task1.2
for
Work
Assignment
3­
11
under
Contract
No.
68­
C99­
239)

Abt
Associates
previously
determined
a
set
of
the
most
influential
chemical
groups
in
the
Risk
Screening
Environmental
Indicator
(
RSEI)
model.
Our
subcontractor,
Great
Lakes
Environmental
Center
(
GLEC),
evaluated
whether
the
chosen
representative
chemical
for
these
groups
is
reasonable,
given
what
is
known
about
typical
conditions
of
wastewater
treatment
and
water
bodies
to
which
effluents
may
be
discharged.
We
have
attached
their
complete
report.
A
summary
of
the
chemical
representatives
that
should,
in
GLEC's
view,
be
used
in
the
RSEI
model
is
as
follows:

Chemical
Group
Appropriate
Representative
Chemical
Chromium
compounds
Cr(
III),
or
if
bioavailability
in
aquatic
environments
is
important,
dissolved
Cr(
III).

Lead
compounds
Pb(
II),
or
if
bioavailability
in
aquatic
environments
is
important,
dissolved
Pb(
II).

Copper
compounds
Cu(
II),
or
if
aquatic
bioavailability
in
aquatic
environments
is
important,
dissolved
Cu(
II).

Cyanide
compounds
Total
HCN
plus
CN­
(
free
cyanide)

Mercury
compounds
Inorganic
Hg(
II)
and
methylmercury,
or
if
bioavailability
in
aquatic
environments
is
important,
dissolved
Hg(
II)
and
methylmercury.
If
toxicity
is
the
important
consideration,
methylmercury.
Abt
Associates
Inc.
These
findings
indicate
that
the
representative
chemicals
for
both
Chromium
compounds
and
Cyanide
compounds
should
change
from
what
is
currently
used
in
the
RSEI
model.
Replacing
chromium
(
IV)
with
chromium
(
III)
has
been
previously
discussed.
The
reference
dose
(
RfD)
for
free
cyanide
is
0.02
mg/
kgday
This
corresponds
to
an
oral
toxicity
weight
of
25.
Currently
we
are
using
an
oral
toxicity
weight
of
100,
based
on
an
RfD
of
0.005
mg/
kg­
day
for
copper
cyanide.

Upon
your
direction,
Abt
Associates
will
update
the
pollutant
group
toxicity
values
in
the
RSEI
model
(
Task
1.3)
for
both
Chromium
and
Cyanide
compounds.
Abt
Associates
Inc.
Evaluation
of
the
Representativeness
of
Designated
Aqueous
Chemical
Species
for
Five
Chemical
Groups
to
be
used
in
the
RSEI
Model
Submitted
to:

Abt
Associates
55
Wheeler
Street
Cambridge,
MA
02138­
1168
Prepared
by:
739
Hastings
Street
Traverse
City,
Michigan
49686
Phone:
(
231)
941­
2230
Facsimile:
(
231)
941­
2240
E­
mail
Address:
mick@
glec­
tc.
com
February
19,
2003
Abt
Associates
Inc.
INTRODUCTION
EPA's
Risk
Screening
Environmental
Indicators
(
RSEI)
model
uses
effluent
data
reported
in
the
Toxics
Release
Inventory
(
TRI)
to
estimate
relative
environmental
impacts
of
discharges
to
both
water
and
air.
The
RSEI
model
makes
the
assumption
that
all
chemical
compounds
reported
in
the
TRI
have
toxicities
equivalent
to
a
specific
representative
chemical
or
chemical
form.
Abt
Associates
contracted
GLEC
to
evaluate
the
representativeness
of
the
chosen
chemical
form
in
aqueous
media
for
five
chemical
groups:

Chemical
Group
Designated
Representative
Chromium
compounds
(
chromium
VI)
Lead
compounds
(
lead)
Copper
compounds
(
copper)
Cyanide
compounds
(
copper
cyanide)
Mercury
compounds
(
methylmercury)

GLEC
was
requested
to
evaluate
the
appropriateness
of
the
designated
chemical
representative
regarding
its
toxicity
and
physicochemical
properties,
considering
the
wastewater
and
drinking
water
treatment
technologies
commonly
employed.
GLEC
also
was
tasked
with
evaluating
the
representativeness
of
the
designated
chemical
with
respect
to
its
behavior
in
natural
surface
waters
receiving
discharges
containing
the
chemical
groups.
If
there
were
any
cases
in
which
the
designated
chemical
species
was
deemed
to
be
inappropriate,
GLEC
was
requested
to
recommend
a
more
appropriate
form.
Abt
Associates
Inc.
RESULTS
AND
RECOMMENDATIONS
The
five
chemical
groups
and
their
selected
chemical
representatives
were
reviewed
and
evaluated;
the
individual
reviews
follow.
In
GLEC's
judgment,
the
following
chemical
representatives
should
be
used
for
each
group:

°
Chromium
compounds:
Cr(
III),
or
if
bioavailability
in
aquatic
environments
is
important,
dissolved
Cr(
III).

°
Lead
compounds:
Pb(
II),
or
if
bioavailability
in
aquatic
environments
is
important,
dissolved
Pb(
II).

°
Copper
compounds:
Cu(
II),
or
if
aquatic
bioavailability
in
aquatic
environments
is
important,
dissolved
Cu(
II).

°
Cyanide
compounds:
Total
HCN
plus
CN­.

°
Mercury
compounds:
Inorganic
Hg(
II)
and
methylmercury,
or
if
bioavailability
in
aquatic
environments
is
important,
dissolved
Hg(
II)
and
methylmercury.
If
toxicity
is
the
important
consideration,
methylmercury.
Abt
Associates
Inc.
CHROMIUM
COMPOUNDS
The
most
important
chromium
species
in
aqueous
systems
are
the
trivalent
and
hexavalent
forms,
Cr(
III)
and
Cr(
VI).
Tripositive
chromium
behaves
chemically
somewhat
like
iron(
III)
and
forms
a
highly
insoluble
hydroxide.
In
natural
waters
Cr(
VI)
is
present
principally
as
the
chromate
anion,
CrO
4
2­.
Most
surface
waters
naturally
contain
chromium
levels
ranging
from
1
to
10

g/
L.
Atlantic
coastal
river
waters
contain
significantly
higher
concentrations
of
chromium
than
the
Gulf
and
Pacific
river
waters,
and
the
chromium
concentration
in
seawater
is
less
than
that
of
river
waters.
The
highest
concentrations
of
chromium
are
usually
found
in
waters
draining
industrial
and
urban
areas.

Chromium
(
III)
is
the
most
stable
and
most
important
oxidation
state
of
the
element.
The
chemistry
of
Cr(
III)
in
aqueous
solutions
is
coordination
chemistry,
and
it
is
dominated
by
the
formation
of
kinetically
inert,
octahedral
complexes.
The
aqueous
Cr(
III)
ion
system
is
extremely
complex.
The
trivalent
chromium
ion
coordinates
with
almost
all
chelating
agents
and
strong
Lewis
bases.
When
sufficient
hydroxide
is
added
to
an
aqueous
solution
of
Cr(
III)
ion,
a
hydrous
chromium(
III)
oxide
(
Cr
2
O3
°
nH
2
O)
of
indefinite
composition
is
precipitated.
Commonly,
this
compound
is
mistakenly
referred
to
as
chromium(
III)
hydroxide.

The
exact
composition
of
Cr(
VI)
anion(
s)
present
in
aqueous
solutions
is
a
function
of
both
pH
and
Cr(
VI)
concentration.
However,
at
pH
values
above
8,
virtually
all
of
the
Cr(
VI)
is
present
as
the
CrO
4
2­

anion.

Only
Cr(
III)
and
Cr(
VI)
compounds
are
produced
in
large
quantities
accessible
to
most
of
the
population.
Therefore,
the
toxicology
of
chromium
compounds
has
been
historically
limited
to
these
two
oxidation
states.
The
EPA
has
set
the
National
Primary
Drinking
Water
Standard
at
100

g/
L
total
chromium.
Industrial
and
municipal
discharges
of
Cr(
III)
and
Cr(
VI)
are
regulated
by
the
National
Pollutant
Discharge
Elimination
System
(
NPDES)
permits.
Current
freshwater
water
quality
criteria
(
WQC)
for
Cr(
III)
are
570

g/
L
(
acute)
and
74

g/
L
(
chronic),
and
for
Cr(
VI)
the
criteria
are16

g/
L
(
acute)
and
11

g/
L
(
chronic).
These
values
constrain
the
levels
of
chromium
that
may
be
discharged
to
the
nation's
surface
waters.
EPA
has
established
exposure
levels
for
both
Cr(
III)
and
Cr(
VI)
for
the
general
population.
For
exposures
of
short
duration
that
constitute
an
insignificant
fraction
of
the
lifespan,
the
acceptable
intake
subchronic
(
AIS)
by
ingestion
is
979
mg/
d
for
trivalent
chromium
and
1.75
mg/
d
for
hexavalent
chromium.
Both
the
established
exposure
levels
and
the
WQC
clearly
illustrate
that
Cr(
VI)
is
much
more
toxic
than
Cr(
III).

To
meet
ambientWQC
for
Cr(
III)
and/
or
Cr(
VI),
industrial
waste
effluents
are
treated
to
remove
chromium.
Hexavelent
chromium
is
reduced
to
trivalent
chromium,
which
is
then
removed
by
precipitation.
The
pH
of
the
aqueous
solution
is
reduced
to
about
2.0
with
hydrochloric
or
sulfuric
acid;
the
aqueous
pH
controls
the
reaction
rate,
which
is
extremely
slow
above
pH
3.
Then,
reducing
agents
such
as
sulfur
dioxide
and
sodium
metabisulfite
are
added.
Lime
or
caustic
soda
is
added
to
raise
the
pH
and
cause
the
precipitation
of
trivalent
chromium.
Precipitation
is
carried
out
at
pH
8.5­
9.5;
in
this
range
chromium
hydroxide
solubility
is
minimal.
Trivalent
chromium
is
treated
simply
by
raising
the
pH
with
lime
or
soda
to
precipitate
the
hydroxy
Cr(
III)
oxide.
Removal
efficiencies
(
which
are
generally
greater
than
90
percent)
may
be
improved
by
adding
additional
flocculating
agents.
About
80
percent
of
chromium
can
also
be
effectively
removed
from
municipal
waste
water
treatment
effluents
by
activated
sludge
treatment.
Abt
Associates
Inc.
Insoluble
inert
chromium(
III)
oxide,
Cr
2
O
3,
is
the
stable
mineral
end
product
into
which
chromium
compounds
are
converted
in
the
environment
as
the
result
of
natural
processes.
When
trivalent
chromium
enters
a
surface
water,
it
eventually
precipitates
as
hydroxychromium(
III)
oxide
in
neutral
pH
regions.
This
compound
ages
and
mineralizes,
becoming
increasingly
insoluble
and
becoming
incorporated
in
the
sediments,
with
only
a
small
proportion
remaining
in
solution.
When
Cr(
VI)
compounds
reach
natural
waters,
they
are
eventually
reduced
to
Cr(
III)
compounds
by
the
natural
organic
matter
present
in
the
waters.

CONCLUSION
In
summary,
the
principal
species
of
chromium
found
in
natural
surface
waters
is
Cr(
III).
Lesser
amounts
of
the
more
toxic
Cr(
VI)
may
be
present,
but
eventually
will
be
converted
to
Cr(
III).
The
bottom
sediments
act
as
a
sink
for
the
eventual
precipitates
of
chromium
in
an
aqueous
system.
The
material
presented
indicates
that
dissolved
( 
0.45
micron)
Cr(
III)
best
represents
the
principal
bioavailable
form
of
chromium
in
natural
surface
waters.
Abt
Associates
Inc.
GENERAL
REFERENCES
Clifford,
D.,
S.
Subrammian,
and
T.
J.
Sorg.
1986.
Removing
dissolved
inorganic
contaminants
from
water.
Environ.
Sci.
Technol.
20(
11):
1072­
1080.

Faust,
S.
D.
and
O.
M.
Aly.
1981.
Chemistry
of
Natural
Waters,
Chpt.
7,
Ann
Arbor
Science
Publisher,
Ann
Arbor,
MI.

Kirk­
Othmer.
1993.
Encyclopedia
of
Chemical
Technology,
4th
Ed.
Vol.
6:
Chromium
Compounds.
John­
Wiley
and
Son,
N.
Y.,
N.
Y.

Manahan,
S.
E.
1972.
Environmental
Chemistry,
Willard
Grant
Press,
Boston,
MA.

Gerhartz
W(
ed).
1986.
Ullmann's
Encyclopedia
of
Industrial
Chemistry.
5th
ed.
Vol
A7,
Chromium
compounds.
Weinheim,
N.
Y.,
N.
Y.

VanderLeeden,
F.,
F.
L.
Troise,
and
D.
K.
Todd,
1990.
The
Water
Encyclopedia,
2nd
Ed.
Lewis
Publishers,
Inc.
Chelsea,
MI.
Abt
Associates
Inc.
COPPER
COMPOUNDS
The
most
important
oxidation
states
for
copper
are
elemental
copper
(
Cu(
0)),
cuprous
copper
(
Cu(
I)),
and
cupric
copper
(
Cu(
II)).
In
aqueous
solutions
Cu(
I)
is
unstable
and
dissociates
to
Cu(
II)
and
Cu(
0).
Some
cuprous
complexes
are
stable,
but
are
rarely
encountered
in
natural
waters.
Therefore,
the
chemistry
of
copper
dissolved
in
natural
surface
waters
is
essentially
the
chemistry
of
the
more
stable
cupric
or
Cu(
II)
species.
Dissolved
copper
is
naturally
found
in
freshwater
surface
waters
at
an
average
level
of
about
15

g/
L,
with
a
range
from
1

g/
L
to
200

g/
L.

In
aquatic
environments
the
chemistry
of
Cu(
II)
is
principally
the
chemistry
of
Cu(
II)
complexes.
The
percentage
of
freely
dissolved
cupric
ion
in
solution
is
usually
much
less
than
1
percent.
Cu(
II)
readily
complexes
with
dissolved
natural
organic
matter
(
NOM)
and
numerous
inorganic
species.
The
copper
complex
speciation,
and
hence
bioavailability,
is
very
much
controlled
by
NOM
concentration,
pH,
alkalinity,
trace
metal
concentrations,
and
major
cation
concentrations.

EPA
has
proposed
a
National
Primary
Drinking
Water
Standard
for
total
copper
of
1.3
mg/
L.
Industrial
and
municipal
discharges
of
copper
are
regulated
by
the
National
Pollutant
Discharge
Elimination
System
(
NPDES)
permits.
Current
Water
Quality
Criteria
(
WQC)
for
copper
are
13

g/
L
(
acute)
and
9

g/
L
(
chronic)
at
a
hardness
of
100
(
the
bioavailability
of
copper
is
hardness
dependent).
The
criteria
are
for
dissolved
copper,
and
were
generated
using
cupric
salts
in
aquatic
toxicity
tests;
thus,
the
criteria
are
for
dissolved
Cu(
II).
Because
permit
limits
are
based
upon
total
copper,
waste
load
allocation
(
WLA)
values
based
on
the
criteria
must
be
translated
from
dissolved
to
total
values
for
the
discharger
permit
limits
using
a
chemical
translator,
either
a
default
value
or
a
site­
specific
value
developed
by
the
discharger.
The
translator
reflects
the
fraction
of
dissolved
Cu(
II)
relative
to
total
copper
in
the
discharger's
effluent,
which
impacts
bioavailability.
For
example,
the
state
of
Michigan
default
translator
for
copper
is
1.5.
If
a
discharger
in
Michigan
has
a
calculated
WLA
of
50

g/
L
based
on
the
copper
criteria,
the
discharger's
permit
limit
would
be
75

g/
L.
Site­
specific
criteria
using
water­
effect
ratio
(
WER)
studies
also
can
be
developed,
potentially
allowing
for
greater
discharges
of
copper
than
permitted
by
EPA's
criteria.
Copper
WER
values
have
been
developed
at
some
sites
that
allow
for
over
a
ten­
fold
increase
in
copper
discharges.
The
WER
value
reflects
the
bioavailable
Cu(
II)
concentration
in
the
combined
effluent
and
receiving
water
matrix,
and
the
intent
of
the
copper
WQC
is
to
regulate
this
form
of
copper,
rather
than
all
copper
species.

Copper
is
toxic
in
very
low
concentrations
to
most
fungi,
algae,
and
certain
bacteria,
and
can
be
lethal
to
higher
life
forms
in
relatively
high
doses.
The
acute
oral
toxicity
in
humans
(
LD
10)
is
100
mg/
kg;
in
contrast,
Cu(
II)
can
be
acutely
toxic
to
aquatic
organisms
at
the
13

g/
L
level.

To
achieve
NPDES
permit
limits,
industrial
effluents
can
be
treated
to
reduce
copper
by
hydroxide
or
sulfide
precipitation,
or
by
a
combination
of
the
two
processes.
Good
to
very
good
removals
(
60­
90
percent)
can
be
achieved
from
industrial
wastes.
Removal
of
copper
from
municipal
effluents
at
activated
sludge
plants
is
reported
to
be
greater
than
80
percent.

The
biogeochemical
cycling
of
Cu(
II)
is
profoundly
influenced
by
its
speciation.
The
interaction
of
Cu(
II)
ions
with
humic
substances
influences
the
transport,
bioavailability,
and
solubility
in
waters
and
sediments.
Hydrolysis
and
precipitation
reactions
dominate
the
Cu(
II)
chemistry
at
pH
values
expected
in
most
natural
water
systems,
especially
in
the
absence
of
significant
levels
of
complexing
compounds
such
as
humic
acids.
A
significant
process
by
which
Cu(
II)
is
removed
from
aqueous
solution
is
by
precipitation
of
malachite
(
CuCO
3,­
Cu(
OH)
2).
Cupric
sulfide
and
cupric
hydroxide
also
precipitate.
Cu(
II)
also
Abt
Associates
Inc.
adsorbs
to
suspended
particulate
matter,
which
eventually
settles
out
of
the
water
column.
Thus,
the
fate
of
most
of
the
aqueous
Cu(
II)
is
to
eventually
migrate
to
the
sediments,
where
further
conversion
to
insoluble
CuS
and
mineral
forms
such
as
malachite
virtually
sequesters
the
copper
in
the
bottom
sediments.
On
a
global
basis,
less
than
1
percent
of
the
copper
transported
by
rivers
to
the
oceans
is
in
the
soluble
form.

CONCLUSION
Given
that
the
chemistry
of
copper
dissolved
in
natural
surface
waters
is
essentially
the
chemistry
of
Cu(
II)
species,
dissolved
( 
0.45

M)
Cu(
II)
best
represents
the
principal
bioavailable
to
be
used
in
the
RSEI
model.
Abt
Associates
Inc.
GENERAL
REFERENCES
Clifford,
D.,
S.
Subrammian,
and
T.
J.
Sorg.
1986.
Removing
dissolved
inorganic
contaminants
from
water.
Environ.
Sci.
Technol.
20(
11):
1072­
1080.

Kirk­
Othmer.
1993.
Encyclopedia
of
Chemical
Technology,
4th
Ed.
Vol.
6:
Copper
Compounds.
John­
Wiley
and
Son,
N.
Y.,
N.
Y.

Gerhartz
W(
ed).
1986.
Ullmann's
Encyclopedia
of
Industrial
Chemistry,
5th
ed.
Vol
A7,
Copper
compounds.
Weinheim,
N.
Y.
NY.

VanderLeeden,
F.,
F.
L.
Troise,
and
D.
K.
Todd.
1990.
The
Water
Encyclopedia,
2nd
Ed.
Lewis
Publishers,
Inc.
Chelsea,
MI.

Demayo,
A.,
M.
C.
Taylor,
and
K.
W.
Taylor.
1982.
Effects
of
copper
on
humans,
animals,
plants,
and
aquatic
life.
CRC
Critical
Reviews
in
Environmental
Control.
12:
83.
Abt
Associates
Inc.
LEAD
COMPOUNDS
Lead
is
widely
distributed
in
the
crust
of
the
earth,
and
is
also
found
in
the
atmosphere
and
hydrosphere.
Atmospheric
fallout
(
dry
deposition)
and
rainout
of
particulate
lead
are
considered
significant
sources
of
lead
input
into
natural
surface
waters,
especially
in
urban
areas.
Storm
runoff
can
carry
significant
amounts
of
lead
in
solution,
in
addition
to
lead­
bearing
sediments,
into
receiving
waters.
Generally,
non­
urban
areas
with
high
storm
runoff
rates
are
less
likely
to
contribute
high
lead
loadings
to
surface
waters
because
they
are
less
likely
to
have
high
lead
fallout
rates;
runoff
originating
in
urban
areas
will
tend
to
have
high
lead
concentrations.

The
chemical
composition
of
runoff
waters
is
an
important
factor
in
determining
the
concentration
of
dissolved
lead.
Runoff
waters
with
a
pH
value
near
6.5
and
an
alkalinity
below
30
mg/
L
(
as
HCO

3
)
are
expected
to
contain
high
concentrations
of
soluble
lead,
ranging
from
40

g/
L
up
to
several
hundred

g/
L.
However,
waters
with
alkalinities
exceeding
60
mg/
L
and
a
pH
value
near
8.0
are
expected
to
contain
low
concentrations
of
dissolved
lead,
generally
below
10

g/
L.

The
mean
global
natural
lead
concentrations
of
lakes
and
rivers
is
estimated
to
be
1­
10

g/
L.
The
mean
concentration
of
dissolved
lead
in
surface
waters
of
the
United
States
has
been
reported
to
be
23

g/
L.
Newer
"
clean
technique"
analytical
methods,
however,
indicate
that
these
older
estimates
are
probably
on
the
high
side.
Most
lead
in
surface
waters
is
present
either
as
particulate
lead
compounds,
or
as
lead
adsorbed
to
suspended
particulate
matter.

The
principal
oxidation
states
for
lead
are
Pb(
0),
Pb(
II)
and
Pb(
IV).
In
natural
surface
waters,
Pb(
II)
predominates,
and
the
following
species
of
lead
can
be
found:
Pb2+,
PbCO
3,
Pb(
CO
3)
2
2­,
Pb(
OH+),
and
Pb(
OH)
2.
The
speciation
of
Pb(
II)
is
dependent
on
pH
and
alkalinity,
as
well
as
the
concentrations
of
any
lead
complexing
agents
in
the
waters.
Industrial
organo­
Pb(
IV)
compounds
exist,
but
since
the
phasing
out
of
lead
antiknock
additives
for
gasoline,
the
remaining
organo­
lead
compounds
pose
little
pollution
threat
to
surface
waters
in
the
U.
S.

EPA
has
proposed
a
National
Primary
Drinking
Water
Standard
for
total
lead
of
0.05
mg/
L.
Industrial
and
municipal
discharges
of
lead
are
regulated
by
the
National
Pollutant
Discharge
Elimination
System
(
NPDES)
permits.
Current
water
quality
(
WQC)
criteria
for
lead
are
65

g/
L
(
acute)
and
2.5

g/
L
(
chronic)
at
a
hardness
of
100
(
lead
criteria
are
hardness
dependent).
The
criteria
are
for
dissolved
Pb(
II),
and
were
generated
using
lead(
II)
salts
in
aquatic
toxicity
tests.
Because
lead
permit
limits
are
for
total
lead,
waste
load
allocation
(
WLA)
values
based
on
the
criteria
must
be
translated
for
the
discharge
permit
from
dissolved
to
total
values
using
a
chemical
translator,
either
a
default
value
or
a
sitespecific
translator
developed
by
the
discharger.
The
translator
reflects
the
fraction
of
dissolved
Pb(
II)
to
total
lead
in
the
discharger's
effluent.
For
example,
the
State
on
Michigan
default
translator
for
lead
is
4.5.
If
a
permit
writer
in
Michigan
has
a
calculated
WLA
of
50

g/
L,
the
discharger's
permit
limit
would
be
225

g/
L.
Site­
specific
criteria
also
can
be
developed
using
water­
effect
ratio
(
WER)
studies,
allowing
for
greater
discharges
of
lead
reflective
of
the
bioavailability
of
lead
in
the
receiving
water/
effluent
matrix.
WER
studies
are
important,
because
the
intent
of
the
WQC
is
to
regulate
the
dissolved
form
of
lead,
rather
than
all
lead
species.

Lead
is
toxic
in
all
forms,
but
to
different
degrees,
depending
upon
the
chemical
nature
and
solubility
of
the
lead
compound.
In
general,
organolead
compounds
are
more
toxic
than
inorganic
lead
salts
if
ingested,
inhaled,
or
absorbed
by
humans.
Children
are
more
sensitive
than
adults
to
the
effects
of
Abt
Associates
Inc.
lead
exposure.
Such
exposure
may
be
acute
(
a
large
single
dose)
or
chronic
(
repeated
low
doses).
Prolonged
absorption
of
lead
should
be
avoided
because
there
are
many
toxic
effects
which
are
cumulative,
and
severe.

To
achieve
permit
limits,
industrial
waste
effluents
can
be
treated
to
reduce
lead
by
hydroxide
or
sulfide
precipitation,
or
by
a
combination
of
the
processes.
Very
good
removals
(
greater
than
90
percent)
can
be
achieved
in
industrial
waste
waters.
Removal
of
lead
from
municipal
effluents
by
activated
sludge
treatment
plants
is
reported
to
be
greater
than
60
percent.

The
environmental
fate
of
Pb(
II)
in
aqueous
systems
in
dominated
by
precipitation
of
Pb(
CO
3),
Pb(
OH)
2,
and
Pb
3(
OH)
2(
CO
3)
2
(
and
in
the
presence
of
sulfur
under
reducing
conditions),
PbS.
Adsorption
of
Pb(
II)
to
suspended
matter,
which
ultimately
settles,
also
removes
lead
from
the
water
column.
The
important
influence
of
pH
on
the
solubility
of
lead
in
aqueous
systems
must
be
recognized.
The
solubility
of
lead
is
below
10

g/
L
above
pH
8.0,
regardless
of
the
alkalinity.
However,
in
slightly
acidic
waters
(
near
pH
6.5)
and
with
a
low
alkalinity,
the
solubility
of
lead
could
increase
to
values
greater
than
100

g/
L.
Complexation
with
organic
matter
or
inorganic
colloids
also
influences
lead
solubility.
Sediments
are
the
ultimate
"
sink"
for
lead,
where
they
are
in
mineralized
forms
(
PbS,
galena).

CONCLUSION
The
material
presented
indicates
that
dissolved
( 
0.45
micron)
Pb(
II)
best
represents
the
principal
bioavailable
form
of
lead
present
in
natural
surface
waters.
Abt
Associates
Inc.
GENERAL
REFERENCES
Clifford,
D.,
S.
Subrammian,
and
T.
J.
Sorg.
1986.
Removing
Dissolved
Inorganic
Contaminants
From
Water.
Environ.
Sci.
Technol,
20(
11):
1072­
1080.

Faust,
S.
D.
and
O.
M.
Aly.
1981.
Chemistry
of
Natural
Waters
Chpt.
7,
Ann
Arbor
Science
Publishers,
Ann
Arbor,
MI.

Kirk­
Othmer.
1993.
Encyclopedia
of
Chemical
Technology,
4th
Ed.
Vol
15:
Lead
Compounds.
John­
Wiley
and
Son,
N.
Y.
,
NY.

Manahan,
S.
E.
1972.
Environmental
Chemistry,
Willard
Grant
Press,
Boston,
MA.

Gerhartz
W(
ed).
1986.
Ullmann's
Encyclopedia
of
Industrial
Chemistry.
5th
ed.
Vol
A15,
Lead
Compounds.
Weinheim,
N.
Y.
NY.

VanderLeeden,
F.,
F.
L.
Troise,
and
D.
K.
Todd,
1990.
The
Water
Encyclopedia,
2nd
Ed.
Lewis
Publishers,
Inc.
Chelsea,
MI.
Abt
Associates
Inc.
CYANIDE
COMPOUNDS
Cyanide
compounds,
which
can
release
the
cyanide
anion,
(
CN­)
are
differentiated
into
simple
and
complex
cyanides.
Simple
cyanides
have
the
formula
A(
CN)
x,
where
A
is
an
alkaline
earth
element
or
a
metal,
and
x
is
the
number
of
cyanide
groups,
such
as
NaCN.
In
solutions
of
simple
metal
cyanides,
the
CN
group
also
can
be
found
in
the
form
of
metal­
cyanide
complexes
(
e.
g.,
Cu(
CN)
3

and
Ni(
CN)
4

2).
The
metal
cyanides
may
dissociate,
depending
on
several
factors.
Some
metal­
cyanide
complexes
such
as
zinc­,
nickel­,
copper­
and
cadmium­
cyanide
are
easily
dissociable
under
acidic
conditions,
and
are
classified
as
weak
acid
dissociable
(
WAD)
species.
Other
metal
cyanides
are
strongly
complexed
and
difficult
to
dissociate,
such
as
cobalt­
and
iron­
cyanide
complexes.

The
different
types
of
cyanide
species,
free
cyanide
(
HCN
and
CN­),
WAD
metal­
cyanide,
strongly
complexed
metal­
cyanide,
as
well
as
other
cyanide­
related
compounds
(
thiocyanate,
cyanogen
chloride
and
cyanate)
differ
considerably
in
their
toxicity,
reactivity,
and
environmental
fate
and
transport.
From
an
environmental
effects
perspective,
the
most
toxicologically
significant
or
ecologically
important
forms
of
cyanide
are
free
and
WAD
cyanide.

Numerous
cyanide
species
are
found
in
water,
but
these
species
are
not
commonly
identified
and
quantified.
Rather,
cyanide
is
typically
measured
as
"
total
cyanide
by
distillation",
which
does
not
differentiate
among
forms
of
cyanide
(
and
excludes
thiocyanate,
cyanate,
organocyanides,
and
some
other
cyanide
forms).
Because
different
forms
of
cyanide
have
different
toxicity
characteristics
and
physicochemical
properties,
total
cyanide
analytical
data
have
limitations
in
risk
assessments,
in
evaluating
cyanide
fate
and
transport
in
the
aquatic
environment
and
in
developing
treatment
processes.
Analytical
methods
that
focus
on
cyanide
species
and
groups
of
species
have
been
developed,
but
are
used
relatively
infrequently,
in
part
because
of
concern
about
the
ability
of
these
methods
to
achieve
low
(
ppb
level)
detection
limits
in
the
complex
matrices
relevant
to
water
quality
management.

In
natural
waters
hydrolysis
of
the
CN­
ion
produces
HCN,
a
weak
acid.
In
most
surface
waters
HCN
predominates;
at
pH
levels
of
7.5
or
less,
essentially
all
of
the
free
cyanide
is
present
as
HCN.
However,
at
pH
levels
of
8.0,
only
about
5
percent
of
the
cyanide
is
present
as
the
CN­
ion.

The
toxicity
of
cyanide
compounds
is
well
known.
In
the
aquatic
environment
HCN
has
been
shown
to
be
the
most
toxic
of
the
cyanide
species.
Industrial
and
municipal
effluent
discharges
of
cyanide
are
regulated
by
the
National
Pollutant
Discharge
Elimination
System
(
NPDES).
Current
water
quality
(
WQC)
criteria
for
cyanide
are
22

g/
L
(
acute)
and
5.2

g/
L
(
chronic),
expressed
as
free
cyanide.
These
limits
constrain
the
levels
of
cyanide
that
may
be
discharged
to
the
nation's
surface
waters.
A
tentative
Drinking
Water
Standard
of
0.2
mg/
L
has
been
established
by
EPA.
Ingestion
of
cyanide,
unless
prompt
first­
aid
or
medical
treatment
is
given,
can
be
rapidly
fatal
at
concentrations
of
approximately
1mg
of
cyanide
per
kilogram
of
body
weight.

Whenever
cyanide
is
manufactured
or
used,
effluents
and
wastes
containing
various
amounts
of
cyanide
are
produced.
Because
of
the
high
toxicity
of
cyanide
to
all
forms
of
life,
the
effluents
and
wastes
must
be
treated
to
reduce
the
cyanide
concentrations
to
achieve
the
required
discharge
level.
Alkaline
chlorination
is
the
process
most
frequently
used
to
treat
effluents
containing
less
than
1
g
CN

per
liter.
In
principle,
this
method
destroys
all
of
the
commercially
used
simple
and
complex
cyanides,
with
the
exception
of
complex
iron
cyanides,
which
are
only
attacked
at
temperatures
above
80

C.
The
treatment
can
be
carried
out
with
chlorine
and
alkali
[
NaOH,
Ca(
OH)
2]
or
with
ready­
made
hypochlorite
solutions
Abt
Associates
Inc.
that
contain
about
12
percent
NaOCl.
At
first,
toxic
cyanogen
chloride
is
formed,
which
hydrolyzes
quickly
to
cyanate
and
chloride
at
pH
>
11.
The
cyanate
may
be
oxidized
further
to
nitrogen
and
carbonate
by
using
excess
hypochlorite.
Hydrogen
peroxide
use
for
cyanide
detoxification
has
increased
in
recent
years
because,
in
this
case,
the
oxidation
of
cyanide
leads
directly
to
cyanate
without
forming
toxic
intermediates
and
byproducts;
in
addition,
hydrogen
peroxide
does
not
cause
additional
salting.

Complexed
cyanides
in
natural
waters
may
revert
to
free
cyanide
through
hydrolysis
reactions,
or
in
certain
instances
photolytic
reactions
(
iron­
cyanide
complex).
However,
the
reactions
are
slow,
and
the
HCN
formed
follows
the
same
fate
paths
available
to
all
free
cyanide
in
surface
waters:

°
Loss
to
the
atmosphere
°
Biological
(
bacterial)
oxidation
°
Complexation
with
iron
to
form
very
stable,
non­
toxic
compounds
°
Polymerization
CONCLUSION
The
material
presented
indicates
that
free
cyanide
(
HCN
and
CN­)
best
represents
the
principal
bioavailable
form(
s)
of
cyanide
present
in
natural
surface
waters.
Abt
Associates
Inc.
GENERAL
REFERENCES
Clesari,
L.
S.,
A.
E.
Greenberg,
and
R.
R.
Trussell
(
eds).
1989.
Standard
Methods
for
the
Examination
of
Water
and
Wastewater,
17th
Ed.
APHA,
Washington,
D.
C.

Kirk­
Othmer.
1993.
Encyclopedia
of
Chemical
Technology,
4th
Ed.
Vol.
7:
Cyanide
Compounds.
John­
Wiley
and
Son,
N.
Y.,
N.
Y.

Gerhartz
W(
ed).
1986.
Ullmann's
Encyclopedia
of
Industrial
Chemistry.
5th
ed.
Vol.
A8,
Cyanide
compounds.
Weinheim,
N.
Y.,
N.
Y.

Zheng,
A
and
D.
A.
Dzombak.
2003.
Evaluation
and
Testing
of
Analytical
Methods
for
Cyanide
Species
in
Municipal
and
Industrial
Contaminated
Waters.
Environ.
Sci.
Technol.
37:
107­
115.
Abt
Associates
Inc.
MERCURY
COMPOUNDS
Mercury
is
a
toxic,
persistent,
bioaccumulative
pollutant.
Most
of
the
mercury
in
natural
surface
waters
is
in
the
form
of
inorganic
mercury
salts
and
organic
forms
of
mercury
(
e.
g.,
methylmercury).
Mercury
accumulates
very
efficiently
in
the
aquatic
food
web,
and
nearly
all
of
the
mercury
that
accumulates
in
fish
tissue
is
methylmercury.

Mercury
can
exist
in
three
oxidation
states:
Hg0
(
elemental),
Hg
2
2+(
mercurous),
and
Hg2+(
mercuric­
Hg(
II)).
The
properties
and
chemical
behavior
of
mercury
strongly
depend
on
the
oxidation
state.
Mercurous
and
mercuric
mercury
can
form
numerous
inorganic
and
organic
compounds;
however,
mercurous
mercury
is
rarely
stable
under
ordinary
environmental
conditions.
Mercury
is
unusual
among
metals,
because
it
tends
to
form
covalent
rather
than
ionic
bonds.
Most
of
the
mercury
encountered
in
all
environmental
media,
except
the
atmosphere,
is
in
the
form
of
inorganic
mercuric
salts
and
organomercury
compounds.
Organomercurics
are
defined
by
the
presence
of
a
covalent
C­
Hg
bond,
which
differentiates
organomercurics
from
inorganic
mercury
compounds
that
merely
associate
with
the
organic
material
in
the
environment
but
do
not
have
the
C­
Hg
bond.
The
compounds
most
likely
to
be
found
under
environmental
conditions
are
the
mercuric
salts,
HgCl
2,
Hg(
OH)
2
and
HgS;
methylmercury
compounds,
methylmercuric
chloride
(
CH
3
HgCl),
and
methylmercuric
hydroxide
(
CH
3
HgOH);
and,
in
small
fractions,
other
organomercurics
(
i.
e.,
dimethylmercury
and
phenylmercury).
Mercuric
and
organomercuric
ions
undergo
rapid
hydrolysis
to
form
the
corresponding
hydroxides.
Mercury
compounds
in
the
aqueous
phase
often
remain
as
undissociated
molecules,
and
the
reported
solubility
values
reflect
this.
The
dissociation
tendencies
of
mercuric
compounds
(
HgX
2
or
RHgX)
depend
on
the
nature
of
the
ligand
X
¯
:

F
¯
>
OCOCH
3
¯
>
HPO
4
2­

Cl
¯
>
Br
¯
>
NH
3>
OH
¯
>
SR
¯
>
S2­

The
thiol
and
sulfide
compounds
are
particularly
stable
and,
therefore,
have
low
tendencies
to
dissociate.
Solubility
values
for
mercury
compounds
which
do
not
dissociate
are
not
based
on
the
ionic
product.
Most
organomercurics
are
not
soluble,
and
do
not
react
with
weak
acids
or
bases
due
to
the
low
affinity
of
mercury
for
oxygen
bonded
to
carbon.
CH
3
HgOH,
however,
is
highly
soluble
due
to
the
strong
hydrogen
bonding
capability
of
the
hydroxide
group.
The
mercuric
salts
vary
widely
in
solubility.
For
example,
HgCl
2
is
readily
soluble
in
water,
and
HgS
is
as
unreactive
as
the
organomercurics
due
to
the
high
affinity
of
mercury
for
sulfur.

There
are
a
number
of
pathways
by
which
mercury
can
enter
the
freshwater
environment.
Hg(
II)
and
methylmercury
from
atmospheric
deposition
(
wet
and
dry)
can
enter
water
bodies
directly;
Hg(
II)
and
methylmercury
can
be
transported
to
water
bodies
in
runoff
(
bound
to
suspended
soil/
humus
or
attached
to
dissolved
organic
carbon);
or
Hg(
II)
and
methylmercury
can
leach
into
the
water
body
from
groundwater
flow
in
the
upper
soil
layers.
Once
in
the
freshwater
system,
mercury
will
be
complexed
and
transformed
as
with
mercury
species
in
soil,
with
additional
processes
involved
due
to
the
aqueous
environment,
such
as
partitioning
to
suspended
matter,
photolysis
reactions,
and
biotransformation
and
biomagnification
in
the
food
web.
Mercury
concentrations
are
typically
reported
for
particular
components
of
the
water
environment,
the
most
common
of
which
are
the
water
column
(
further
partitioned
as
dissolved
or
attached
to
suspended
material),
the
underlying
sediment
(
further
divided
into
surface
sediments
and
deep
sediments);
and
biota
(
particularly
fish).

Partition
coefficients
have
been
calculated
to
describe
the
relative
affinity
of
Hg(
II)
and
Abt
Associates
Inc.
methylmercury
for
sediment
over
water.
Values
of
the
partition
coefficient
K
d
(
concentration
of
mercury
in
dry
sediment
or
suspended
matter
divided
by
the
dissolved
concentration
in
water)
on
the
order
of
100,000
ml/
g
sediment
and
100,000+
ml/
g
suspended
material
are
typically
found
for
Hg(
II)
and
methylmercury,
indicating
a
strong
preference
for
Hg(
II)
and
methylmercury
to
remain
bound
to
bottom
sediment
or
suspended
matter
(
the
affinity
highest
for
methylmercury).
A
river
or
lake
freshwater
system
has
a
larger
volume
of
water
than
sediment,
and
a
significant
amount
of
Hg(
II)
entering
a
system
may
partition
to
the
water
column,
especially
if
there
is
a
high
concentration
of
suspended
material
in
the
water
column.
It
is
often
unclear
whether
the
mercury
in
sediment
will
be
found
as
HgCl
2
or
Hg(
OH)
2
organic
complexes,
which
are
more
susceptible
to
methylation
(
addition
of
a
methyl
group
to
form
methylmercury),
or
will
be
the
more
unreactive
HgS
and
HgO
forms.

Most
of
the
mercury
in
the
water
column
(
Hg(
II)
and
methylmercury)
will
be
bound
to
organic
matter,
either
to
dissolved
organic
carbon
(
DOC;
consisting
of
fulvic
and
humic
acids,
carbohydrates,
carboxylic
acids,
amino
acids
and
hydrocarbons),
or
to
suspended
particulate
matter.
In
most
cases,
studies
that
refer
to
the
dissolved
mercury
in
water
include
mercury
complexes
with
DOC.
Studies
indicate
that
about
25­
60
percent
of
Hg(
II)
and
methylmercury
organic
complexes
are
particle­
bound
in
the
water
column,
with
the
rest
in
the
dissolved
and
DOC­
bound
phase.
Typically,
total
mercury
and
methylmercury
concentrations
are
positively
correlated
with
DOC
concentrations
in
lake
waters.

Hg0
is
produced
in
freshwater
by
humic
acid
reduction
of
Hg(
II),
or
demethylation
of
methylmercury
mediated
by
sunlight.
A
portion
will
remain
in
the
dissolved
gaseous
state,
while
most
will
volatilize.
As
noted
previously,
Hg0
constitutes
very
little
of
the
total
mercury
in
the
water
column,
but
may
provide
a
significant
pathway
for
the
evolution
of
mercury
out
of
the
water
body
via
Hg(
II)
or
methylmercury
­>
Hg0
­>
volatilization.
For
many
lakes,
however,
sedimentation
of
the
Hg(
II)
and
methylmercury
bound
to
particulate
matter
is
the
dominant
process
for
removal
of
mercury
from
the
water
column.

Generally,
no
more
than
25
percent
of
the
total
mercury
in
the
water
column
exists
as
a
methylmercury
complex;
typically,
less
than
10%
is
observed.
The
water
column
methylmercury
concentration
is
a
result
of
methylation
of
Hg(
II)
in
the
bottom
sediments
and
the
water
column,
by
microbial
action
and
abiotic
processes.
Methylmercury
in
the
water
column
which
is
lost
through
demethylation,
exported
downstream,
or
taken
up
by
biota,
is
thought
to
be
replaced
by
additional
methylation
of
Hg(
II)
compounds
to
sustain
equilibrium.

Mercury
can
enter
surface
water
as
Hg0,
Hg2+,
or
methylmercury.
Once
in
aquatic
systems,
mercury
can
exist
in
dissolved
or
particulate
forms,
and
can
undergo
the
following
transformations.

°
Hg
0
in
surface
waters
can
be
oxidized
to
Hg2+,
or
volatilized
to
the
atmosphere.
°
Hg2+
can
be
methylated
in
sediments
and
the
water
column
to
form
methylmercury.
°
Methylmercury
can
be
alkylated
to
form
dimethylmercury.
°
Hg2+
and
methylmercury
can
form
organic
and
inorganic
complexes
with
sediments
and
suspended
particulate
matter.

Each
of
these
reactions
are
also
possible
in
the
reverse
direction.
The
net
rate
of
production
of
each
mercury
species
is
determined
by
the
balance
between
the
forward
and
reverse
reactions.

The
U.
S.
Geological
Survey
National
Pilot
Study
of
Mercury
Contamination
found
the
mean
methylmercury
concentration
in
surface
water
to
be
0.15
ng/
L.
Total
mercury
concentrations
were
ten
to
Abt
Associates
Inc.
one
hundred
times
higher.
In
contrast,
a
methylmercury
concentration
of
0.07
ng/
L
in
the
nation's
surface
waters
has
been
predicted
by
computer
simulation,
and
more
than
80
percent
of
the
total
mercury
in
the
water
column
was
predicted
to
be
inorganic
divalent
species.
The
fraction
of
the
predicted
background
concentration
as
methylmercury
was
7.8
percent.

EPA
has
proposed
a
National
Primary
Drinking
Water
Standard
for
total
mercury
of
0.002
mg/
L.
Industrial
and
municipal
discharges
of
mercury
are
regulated
by
the
National
Pollutant
Discharge
Elimination
System
(
NPDES)
permits.
Current
water
quality
(
WQC)
criteria
(
aquatic
life)
for
mercury
are
1.4

g/
L
(
acute)
and
0.77

g/
L
(
chronic)
at
a
hardness
of
100
(
mercury
criteria
are
hardness
dependent).
The
mercury
aquatic
life
criteria
are
for
total
dissolved
mercury,
and
were
generated
using
mercury
salts
in
aquatic
toxicity
tests;
thus,
the
criteria
are
for
total
dissolved
Hg(
II).
Permit
limits
are
for
total
mercury;
therefore
waste
load
allocation
(
WLA)
values
based
on
the
criteria
must
be
translated
form
dissolved
to
total
values
for
the
discharger's
permit
using
a
chemical
translator,
either
a
default
value
or
a
site­
specific
one
developed
by
the
discharger.
EPA
has
provided
the
following
"
default"
translators
for
use
in
the
permitting
of
mercury:

Summary
of
Mercury
Translators
for
Mercury
in
Water
fd
(
fraction
dissolved)
value
Lentic
(
static
waters)
Lotic
(
flowing
waters)

fd
Hg
dissolved/
Hg
total
0.598
0.370
fd
Me
Hg
dissolved/
Hg
total
0.032
0.014
fd
Me
Hg
dissolved/
Me
Hg
total
0.613
0.490
The
translators
can
also
be
applied
to
derive
default
concentrations
of
dissolved
Hg(
II)
and
methylmercury
from
total
Hg
values.
For
example,
for
a
total
Hg
concentration
of
1
ng/
L,
a
predicted
dissolved
Hg(
II)
concentration
of
0.598
ng/
L
and
dissolved
methylmercury
of
0.032
ng/
L
can
be
calculated.

Methylmercury
is
by
far
the
most
toxic
form
of
mercury,
and
EPA
has
established
a
water
quality
criterion
for
protection
of
human
health
for
methylmercury.
The
primary
source
of
human
exposure
to
methylmercury
is
through
consumption
of
contaminated
fish
and
seafood.
EPA
has
therefore
established
a
fish
tissue
based
criterion
for
methylmercury,
the
value
of
which
is
0.3
mg
methylmercury
/
kg
in
fish
tissue.

To
achieve
permit
limits,
industrial
waste
effluents
can
be
treated
to
reduce
the
concentration
of
mercury.
Mercury
can
be
removed
by
hydroxide
or
sulfide
precipitation,
or
by
a
combination
of
the
two
processes.
Very
good
removals
(>
90
percent)
can
be
achieved.
Removal
of
mercury
from
municipal
influents
at
activated
sludge
plants
is
reported
to
be
~
80
percent.
Ferric
sulfate
coagulation
can
be
used
to
remove
mercury
from
drinking
water
sources.

To
understand
the
fate
of
mercury
requires
recognitation
that
once
entering
a
water
body,
mercury
can
remain
in
the
water
column,
be
lost
through
drainage
water,
revolatilize
into
the
atmosphere,
settle
into
the
sediment
or
be
taken
up
by
aquatic
biota.
After
entering
the
aquatic
environment,
the
movements
of
mercury
through
any
specific
water
body
may
be
unique.
Mercury
in
the
water
column,
in
the
sediment,
and
in
other
aquatic
biota
appears
to
be
available
to
aquatic
organisms
for
uptake.
Sediments
are
the
principal
sink
for
mercury.

Methylation
is
a
key
step
in
the
entrance
of
mercury
into
the
food
chain.
Inorganic
mercury
species
Abt
Associates
Inc.
can
be
biotransformed
to
methylated
organic
species
in
water
bodies
in
both
the
sediment
and
the
water
column.
Abiotic
processes
(
e.
g.,
humic
and
fulvic
acids
in
solution)
also
appear
to
methylate
the
mercuric
ion.
Not
all
mercury
compounds
entering
an
aquatic
ecosystem
are
methylated,
and
demethylation
reactions
as
well
as
volatilization
of
dimethylmercury
decrease
the
amount
of
methylmercury
available
in
the
aquatic
environment.
There
is
a
large
degree
of
uncertainty
and
variability
concerning
the
processes
that
methylate
mercury
in
aquatic
environments.

Methylmercury
is
very
bioavailable,
and
accumulates
in
fish
through
the
aquatic
food
web;
nearly
100%
of
the
mercury
found
in
fish
muscle
tissue
is
methylated.
Methylmercury
appears
to
be
primarily
passed
to
planktivorous
and
piscivorous
fish
via
their
diet.
Larger,
longer­
lived
fish
species
at
the
upper
end
of
the
food
chain
typically
have
the
highest
concentrations
of
methylmercury
in
a
given
water
body.
A
relationship
exists
between
methylmercury
concentrations
in
fish
and
lake
pH,
with
higher
methylmercury
concentrations
in
fish
tissue
typically
found
in
more
acidic
lakes.
The
mechanisms
for
this
behavior
are
unclear.
Most
of
the
total
methylmercury
production
is
incorporated
in
biota,
particularly
fish.
In
fact,
bioconcentration
factors
(
BCFs)
for
accumulation
of
methylmercury
in
fish
are
on
the
order
of
105­
106.
Bioaccumlation
of
methylmercury
in
fish
muscle
tissue
has
been
documented
in
water
bodies
that
are
remote
from
emission
sources
and
seemingly
pristine,
as
well
as
in
water
bodies
that
are
less
isolated.
Methylmercury
appears
to
be
efficiently
passed
through
the
aquatic
food
web
to
the
highest
trophic
level
consumers
in
the
community
(
e.
g.,
piscivorous
fish).
At
this
point
methylmercury
can
be
ingested
by
fish­
consuming
wildlife
and
humans,
where
it
appears
to
pass
from
the
gastrointestinal
tract
into
the
bloodstream
more
efficiently
than
the
divalent
species.

CONCLUSION
The
material
presented
indicates
that
dissolved
( 
0.45
micron)
Hg(
II)
and
dissolved
methylmercury
best
represent
the
principal
bioavailable
forms
of
mercury
present
in
natural
surface
waters.
Abt
Associates
Inc.
GENERAL
REFERENCES
Clifford,
D.,
S.
Subrammian,
and
T.
J.
Sorg.
1986.
Removing
Dissolved
Inorganic
Contaminants
From
Water.
Environ.
Sci.
Technol,
20(
11):
1072­
1080.

Faust,
S.
D.
and
O.
M.
Aly.
1981.
Chemistry
of
Natural
Water,
Chpt.
7,
Ann
Arbor
Science
Publishers,
Ann
Arbor,
MI.

U.
S.
EPA.
1997.
Mercury
study
report
to
Congress.
Vol.
III.
Fate
and
transport
of
mercury
in
the
environment.
U.
S.
Environmental
Protection
Agency.
December
1997.
EPA­
452/
R97­
005.

U.
S.
EPA.
1997.
Mercury
study
report
to
Congress.
Vol.
VI.
Fate
and
transport
of
mercury
in
the
environment.
U.
S.
Environmental
Protection
Agency.
December
1997.
EPA­
452/
R97­
008.

U.
S.
EPA.
2001.
Water
quality
criterion
for
the
protection
of
human
health:
methylmercury.
U.
S.
Environmental
Protection
Agency.
January
2001.
EPA­
823­
R­
01­
001.