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United
States
Environmental
Protection
Agency
Office
of
Water
Regulations
and
Standards
Criteria
and
Standards
Division
Washington
DC
20460
EPA
440/
5­
80­
057
October
1980
EPA
Ambient
Water
Quality
Criteria
for
Lead
AMBIENT
WATER
QUALITY
CRITERIA
FOR
LEAD
Prepared
By
U.
S.
ENVIRONMENTAL
PROTECTION
AGENCY
Office
of
Water
Regulations
and
Standards
Criteria
and
Standards
Division
Washington,
D.
C.

Office
of
Research
and
Development
Environmental
Criteria
and
Assessment
Office
Cincinnati,
Ohio
Carcinogen
Assessment
Group
Washington,
D.
C.

Environmental
Research
Laboratories
Corvalis,
Oregon
Duluth,
Minnesota
Gulf
Breeze,
Florida
Narragansett,
Rhode
Island
i
DISCLAIMER
This
report
has
been
reviewed
by
the
Environmental
Criteria
and
Assessment
Office,
U.
S.
Environmental
Protection
Agency,
and
approved
for
publication.
Mention
of
trade
names
or
commercial
products
does
not
constitute
endorsement
or
recommendation
for
use.

AVAILABILITY
NOTICE
This
document
is
available
to
the
public
through
the
National
Technical
Information
Service,
(
NTIS),
Springfield,
Virginia
22161.

ii
FOREWORD
Section
304
(
a)(
l)
of
the
Clean
Water
Act
of
1977
(
P.
L.
95­
217),
requires
the
Administrator
of
the
Environmental
Protection
Agency
to
publish
criteria
for
water
quality
accurately
reflecting
the
latest
scientific
knowledge
on
the
kind
and
extent
of
all
identifiable
effects
on
health
and
welfare
which
may
be
expected
from
the
presence
of
pollutants
in
any
body
of
water,
including
ground
water.
Proposed
water
quality
criteria
for
the
65
toxic
pollutants
listed
under
section
307
(
a)(
l)
of
the
Clean
Water
Act
were
developed
and
a
notice
of
their
availability
was
published
for
public
comment
on
March
15,
1979
(
44
FR
15926),
July
25,
1979
(
44
FR
43660),
and
October
1,
1979
(
44
FR
56628).
This
document
is
a
revision
of
those
proposed
criteria
based
upon
a
consideration
of
comments
received
from
other
federal
Agencies,
State
agencies,
special
interest
groups,
and
individual
scientists.
The
criteria
contained
in
this
document
replace
any
previously
published
EPA
criteria
for
the
65
pollutants.
This
criterion
document
is
also
published
in
satisfaction
of
paragraph
11
of
the
Settlement
Agreement
in
Natural
Resources
Defense
Council,
et.
al.
vs.
Train,
8
ERC
2120
(
D.
D.
C.
1976),
modified,
12
ERC
1833
(
D.
D.
C.
1979)
.

The
term
"
water
quality
criteria"
is
used
in
two
sections
of
the
Clean
Water
Act,
section
304
(
a)(
l)
and
section
303
(
c)(
2).
The
term
has
a
different
program
impact
in
each
section.
In
section
304,
the
term
represents
a
non­
regulatory,
scientific
assessment
of
ecological
ef­
fects.
The
criteria
presented
in
this
publication
are
such
scientific
assessments.
Such
water
quality
criteria
associated
with
specific
stream
uses
when
adopted
as
State
water
quality
standards
under
section
303
become
enforceable
maximum
acceptable
levels
of
a
pollutant
in
ambient
waters.
The
water
quality
criteria
adopted
in
the
State
water
quality
standards
could
have
the
same
numerical
limits
as
the
criteria
developed
under
section
304.
However,
in
many
situations
States
may
want
to
adjust
water
quality
criteria
developed
under
section.
304
to
reflect
local
environmental
conditions
and
human
exposure
patterns
before
incorporation
into
water
quality
standards.
It
is
not
until
their
adoption
as
part
of
the
State
water
quality
standards
that
the
criteria
become
regulatory.

Guidelines
to
assist
the
States
in
the
modification
of
criteria
presented
in
this
document,
in
the
development
of
water
quality
standards,
and
in
other
water­
related
programs
of
this
Agency,
are
being
developed
by
EPA.

STEVEN
SCHATZOW
Deputy
Assistant
Administrator
Office
of
Water
Regulations
and
Standards
iii
ACKNOWLEDGEMENTS
Aquatic
Life
Toxicology:

Charles
I.
Stephan,
ERL­
Duluth
U.
S.
Environmental
Protection
Agency
Mammalian
Toxicology
and
Human
Health
Effects:

Paul
B.
Hammond
(
author)
University
of
Cincinnati
Michael
L.
Dourson
(
doc.
mgr.)
ECAO­
Cin
U.
S.
Environmental
Protection
Agency
Jerry
F.
Stara
(
doc.
mgr.)
ECAO­
Cin
U.
S.
Environmental
Protection
Agency
Patrick
Durkin
Syracuse
Research
Corporation
W.
Galke,
ECAO­
RTP
U.
S.
Environmental
Protection
Agency
Terri
Laird,
ECAO­
Cin
U
.
S.
Environmental
Protection
Agency
K.
Mahaffey
U.
S.
Food
and
Drug
Administration
John
H.
Gentile,
ERL­
Narragansett
U.
S.
Environmental
Protection
Agency
Roy
E.
Albert*
Carcinogen
Assessment
Group
U.
S.
Environmental
Protection
Agency
R.
J.
Bull,
HERL
U.
S.
Environmental
Protection
Agency
Thomas
Clarkson
University
of
Rochester
Robert
A.
Ewing
Battelle
­
Columbus
Laboratory
T.
J.
Haley
National
Center
for
Toxicological
Research
P.
Landrigan
Center
of
Disease
Control
H.
Needleman
Children's
Hospital
Medical
Center
Technical
Support
Services
Staff:
D.
J.
Reisman,
M.
A.
Garlough,
B.
L.
Zwayer,
P.
A.
Daunt,
K.
S.
Edwards,
T.
A.
Scandura,
A.
T.
Pressley,
C.
A.
Cooper,
M.
M.
Denessen.

Clerical
Staff:
C.
A.
Haynes,
S.
J.
Faehr,
L.
A.
Wade,
D.
Jones,
B.
J.
Bordicks,
B.
J.
Duesnell,
T.
Highland,
B.
Gardiner.

*
CAG
Participating
Members:
Elizabeth
L.
Anderson,
Larry
Anderson,
Dolph
Arnicar,
Steven
Bayard,
David
L.
Bayliss,
Chao
W.
Chen,
John
R.
Fowle
III,
Bernard
Haberman,
Charalingayya
Hiremath,
Chang
S.
Lao,
Robert
McGaughy,
Jeffrey
Rosenblatt,
Dharm
V.
Singh,
and
Todd
W.
Thorslund.

iv
TABLE
OF
CONTENTS
Page
Criteria
Summary
Introduction
A­
1
Aquatic
Life
Toxicology
Introduction
Effects
Acute
Toxicity
Chronic
Toxicity
Plant
Effects
Residues
Miscellaneous
Summary
Criteria
References
Mammalian
Toxicology
and
Human
Health
Effects
Introduction
Exposure
Natural
Background
Levels
Man­
generated
Sources
of
Lead
Ingestion
from
Water
Ingestion
from
Food
Inhalation
Dermal
Miscellaneous
Sources
Pharmacokinetics
Absorption
Dermal
Distribution
Metabolism
Excretion
Contributions
of
Lead
from
Diet
versus
Air
to
PbB
Effects
Carcinogenicity
Teratogenicity
Mutagenicity
Reproductive
Effects
Renal
Effects
Cardiovascular
Effects
Miscellaneous
Effects
Criterion
Formulation
Existing
Guidelines
and
Standards
Current
Levels
of
Exposure
Special
Groups
at
Risk
Basis
and
Derivation
of
Criterion
References
Appendix
B­
1
B­
1
B­
2
B­
2
B­
5
B­
8
B­
9
B­
9
B­
11
B­
11
B­
31
C­
1
C­
1
C­
2
C­
2
C­
3
C­
3
C­
4
C­
9
C­
9
C­
9
C­
15
C­
16
C­
19
C­
19
C­
21
C­
21
C­
22
C­
35
C­
48
C­
63
C­
66
C­
66
C­
68
C­
70
C­
71
C­
72
C­
72
C­
72
C­
73
C­
73
C­
81
C­
104
V
CRITERIA
DOCUMENT
LEAD
CRITERIA
Aquatic
Life
For
total
recoverable
lead,
the
criterion
(
in
µ
g/
l)
to
protect
fresh­

water
aquatic
life
as
derived
using
the
Guidelines,
is
the
numerical
value
given
by
e(
2.35[
1n(
hardness)]­
9.48)
as
a
24­
hour
average
and
the
concen­

tration
(
in
µ
g/
l)
should
not
exceed
the
numerical
value
given
by
e(
1.22[
1n(
hardness)]­
0.47)
at
any
time.
For
example,
at
hardnesses
of
50,

100,
and
200
mg/
1
as
CaCO3
the
criteria
are
0.75,
3.8,
and
20
µ
g/
l,
re­

spectively,
as
24­
hour
averages,
and
the
concentrations
should
not
exceed
74,
170,
and
400
µ
g/
l,
respectively,
at
any
time.

The
available
data
for
total
recoverable
lead
indicate
that
acute
and
chronic
toxicity
to
saltwater
aquatic
life
occur
at
concentrations
as
low
as
668
and
25
µ
g/
l,
respectively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.

Human
Health
The
ambient
water
quality
criterion
for
lead
is
recommended
to
be
iden­

tical
to
the
existing
water
standard
which
is
50
µ
g/
l.
Analysis
of
the
toxic
effects
data
resulted
in
a
calculated
level
which
is
protective
of
human
health
against
the
ingestion
of
contaminated
water
and
contaminated
aquatic
organisms.
The
calculated
value
is
comparable
to
the
present
stan­

dard.
For
this
reason
a
selective
criterion
based
on
exposure
solely
from
consumption
of
6.5
grams
of
aquatic
organisms
was
not
derived.

vi
INTRODUCTION
Lead
(
atomic
weight
207.2)
is
a
soft
gray,
acid­
soluble
metal
(
Windholz,

1976)
and
exists
in
three
oxidation
states,
0,
+
2,
and
+
4.
Lead
is
a
major
constituent
of
more
than
200
identified
minerals.
Most
of
these
are
rare,

and
only
three
are
found
in
sufficient
abundance
to
form
mineral
deposits:

galena
(
PbS)
the
simple
sulfide,
angelesite
(
PbSO4)
the
sulfate,
and
cer­

rusite
(
PbCO3)
the
carbonate
(
U.
S.
EPA,
1979).
Lead
is
used
in
electro­

plating,
metallurgy,
and
the
manufacture
of
construction
materials,
radia­

tion
protective
devices,
plastics,
and
electronics
equipment.

Although
neither
metallic
lead
nor
the
common
lead
minerals
are
classi­

fied
as
soluble
in
water,
they
can
both
be
solubilized
by
some
acids;
in
contrast,
some
of
the
lead
compounds
produced
industrially
are
considered
water
soluble.
Natural
lead
compounds
are
not
usually
mobile
in
normal
ground
or
surface
water
because
the
lead
leached
from
ores
becomes
adsorbed
by
ferric
hydroxide
or
tends
to
combine
with
carbonate
or
sulfate
ions
to
form
insoluble
compounds
(
Hem,
1976).
The
solubility
of
lead
compounds
in
water
depends
heavily
on
pH
and
ranges
from
about
10,000,000
µ
g/
l
of
lead
at
pH
5.5
to
1
µ
g/
l
at
pH
9.0
(
Hem
and
Durum,
1973).
Lead
does
reach
the
aquatic
environment
through
precipitation,
fallout
of
lead
dust,
street
run­

off,
and
both
industrial
and
municipal
wastewater
discharges
(
U.
S.
EPA,

1976).
Inorganic
lead
compounds
are
most
stable
in
the
plus
two
valence
state,
while
organolead
compounds
are
more
stable
in
the
plus
four
state
(
Standen,
1967).

A­
1
REFERENCES
Hem,
J.
D.
1976.
Geochemical
controls
on
lead
concentrations
in
stream
water
and
sediments.
Geochim.
Cosmochim.
Acta.
40:
599.

Hem,
J.
D.
and
W.
H.
Durum.
1973.
Solubility
and
occurrence
of
lead
in
sur­

face
water.
Jour.
Am.
Water
Works.
65:
562.

Standen,
A.
(
ed.)
1967.
Kirk­
Othmer
Encyclopedia
of
Chemical
Technology.

Interscience
Publishers,
New
York.

U.
S.
EPA.
1976.
Quality
criteria
for
water.
Off.
Water
Plan.
Stand.,
U.
S.

Environ.
Prot.
Agency,
Washington,
D.
C.

U.
S.
EPA.
1979.
Water­
related
environmental
fate
of
129
priority
pollut­

ants.
Off.
Water
Plan.
Stand.,
U.
S.
Environ.
Prot.
Agency,
Washington,
D.
C.

Windholz,
M.
(
ed.)
1976.
The
Merck
Index.
9th
ed.
Merck
and
Co.,
Inc.,

Rahway,
New
Jersey.

A­
2
Aquatic
Life
Toxicology*

INTRODUCTION
The
acute
and
chronic
adverse
effects
of
lead
have
been
studied
with
a
variety
of
freshwater
organisms.
Representative
test
animals
listed
in
Ta­

bles
1
through
6
include
fish
from
six
different
families
(
Salmonidae,
Cy­

prinidae,
Catostomidas,
Ictaluridae,
Poeciliidae,
and
Centrarchidae),
and
invertebrate
species
from
the
nine
groups
(
rotifers,
annelids,
snails,
clad­

ocerans,
copepods,
isopods,
mayflies,
stoneflies,
and
caddisflies).
Tox­

icity
tests
have
also
been
conducted
with
freshwater
plants
from
the
algal,

desmid
and
diatom
groups,
and
both
fish
and
invertebrate
species
have
been
used
in
bioconcentration
tests.

Acute
toxicity
tests
have
been
conducted
with
lead
and
a
variety
of
saltwater
invertebrates,
but
no
tests
with
fish
are
available.
Results
in­

dicate
a
range
of
acute
values
from
668
µ
g/
l
for
a
copepod
to
27,000
µ
g/
l
for
the
adult
soft
shell
clam.
A
chronic
test
has
been
conducted
with
one
invertebrate
species,
the
mysid
shrimp,
and
the
chronic
value
was
25
µ
g/
l.

Select
invertebrate
and
algal
species
are
good
accumulators
of
lead.
Bio­

concentration
factors
calculated
on
a
wet
weight
basis
ranged
from
17.5
for
the
hard
clam
to
2,570
for
the
mussel.

Of
the
analytical
measurements
currently
available,
a
water
quality
cri­

terion
for
lead
is
probably
best
stated
in
terms
of
total
recoverable
lead,

because
of
the
variety
Of
forms
of
lead
that
can
exist
in
bodies
of
water
and
the
various
chemical
and
toxicological
properties
of
these
forms.
The
*
The
reader
is
referred
to
the
Guidelines
for
Deriving
Water
Quality
Cri­
teria
for
the
Protection
of
Aquatic
Life
and
Its
Uses
in
order
to
better
un­
derstand
the
following
discussion
and
recommendation.
The
following
tables
contain
the
appropriate
data
that
were
found
in
the
literature,
and
at
the
bottom
of
each
table
are
calculations
for
deriving
various
measures
of
tox­
icity
as
described
in
the
Guidelines.

B­
1
forms
of
lead
that
are
commonly
found
in
bodies
of
water
and
are
not
meas­

ured
by
the
total
recoverable
procedure,
such
as
the
lead
that
is
a
part
of
minerals,
clays
and
sand,
probably
are
forms
that
are
less
toxic
to
aquatic
life
and
probably
will
not
be
converted
to
the
more
toxic
forms
very
readily
under
natural
conditions.
On
the
other
hand,
forms
of
lead
that
are
common­

ly
found
in
bodies
of
water
and
are
measured
by
the
total
recoverable
proce­

dure,
such
as
the
free
ion,
and
the
hydroxide,
carbonate,
and
sulfate
salts,

probably
are
forms
that
are
more
toxic
to
aquatic
life
or
can
be
converted
to
the
more
toxic
forms
under
natural
conditions.

Because
the
criterion
is
derived
on
the
basis
of
tests
conducted
on
sol­

uble
inorganic
salts
of
lead,
the
total
and
total
recoverable
lead
concen­

trations
in
the
tests
will
probably
be
about
the
same,
and
a
variety
of
an­

alytical
procedures
will
produce
about
the
same
results.
Except
as
noted,

all
concentrations
reported
herein
are
expected
to
be
essentially
equivalent
to
total
recoverable
lead
concentrations.
All
concentrations
are
expressed
as
lead,
not
as
the
compound
tested.

Acute
Toxicity
EFFECTS
Table
1
contains
six
acute
values
for
three
freshwater
invertebrate
spe­

cies.
Only
one
of
the
tests
was
flow­
through
(
Spehar,
et
al.
1978)
but
in
two,
the
toxicant
concentrations
were
measured
(
Spehar,
et
al.
1978;

Chapman,
et
al.
Manuscript).
Acute
tests
were
conducted
at
three
different
levels
of
water
hardness
with
Daphnia
maqna
(
Chapman,
et
al.
Manuscript),

demonstrating
that
daphnids
were
three
times
more
sensitive
to
lead
in
soft
water
than
in
hard
water.
This
acute
value
for
Daphnia
magna
in
soft
water
agrees
closely
with
the
value
reported
earlier
for
the
same
species
in
soft
water
by
Biesinger
and
Christensen
(
1972).
Rotifers
tested
for
96
hours
in
B­
2
soft
water
by
Buikema,
et
al.
(
1974)
were
very
resistant
to
lead;
however,

scuds
were
reported
by
Spehar,
et
al.
(
1978)
to
be
more
sensitive
to
lead
than
any
other
invertebrate
thus
far
tested.
Interestingly,
this
same
rela­

tionship
existed
in
longer
exposures
lasting
up
to
28
days
in
which
the
scud
was
far
more
sensitive
to
lead
than
a
snail,
cladoceran,
chironomid,
mayfly,

stonefly,
and
caddisfly
(
Table
6)
(
Spehar,
et
al.
1978;
Biesinger
and
Chris­

tensen,
1972;
Anderson,
et
al.
1980;
and
Nehring,
1976).

Thirteen
acute
toxicity
tests
have
been
conducted
on
lead
with
six
spe­

cies
of
fish
(
Table
1).
Of
the
13
only
three
were
reported
to
be
flow­
through,
and
measured
toxicant
concentrations
were
reported
for
only
one
(
Holcombe,
et
al.
1976).
The
results
of
acute
tests
conducted
by
Davies,
et
al.
(
1976)
with
rainbow
trout
in
hard
water
are
reported
as
unmeasured
values
in
Table
1,
because
total
lead
concentrations
were
not
measured,
even
though
the
dissolved
lead
concentrations
were.

The
data
in
Table
1
indicate
a
relationship
between
water
hardness
and
the
acute
toxicity
of
lead
to
rainbow
trout
(
Davies,
et
al.
1976),
fathead
minnows
and
bluegills
(
Pickering
and
Henderson,
1966),
because
lead
was
gen­

erally
much
more
toxic
in
soft
water.
Another
example
of
the
effect
of
hardness
was
reported
by
Tarzwell
and
Henderson
(
1960)
who
conducted
96­
hour
exposures
of
fathead
minnows
to
lead
in
soft
and
hard
water
(
20
and
400
mg/
l
as
CaC03,
respectively).
Results
from
the
soft
water
test
are
shown
in
Table
1.
The
hard
water
exposure
is
included
in
Table
6
because
an
LC50
value
was
not
obtained
within
96
hours;
however,
this
test
did
show
that
the
hard
water
LC50
value
was
greater
than
75,000
ug/
l
which
meant
that
the
Lc50
in
hard
water
was
at
least
31
times
that
in
soft
water.
Hale
(
1977)

conducted
an
acute
exposure
of
rainbow
trout
to
lead
and
obtained
an
LCso
value
of
8,000
ug/
l.
This
value
is
six
times
greater
than
the
LC50
value
B­
3
obtained
for
rainbow
trout
in
soft
water
by
Davies,
et
al.
(
1976).
Hale
did
not
report
water
hardness;
however,
alkalinity
and
pH
were
reported
to
be
105
mg/
l
and
7.3,
respectively,
which
suggests
that
this
water
was
probably
harder
than
the
soft
test
water
used
by
Davies,
et
al.
(
1976).
Wallen,
et
al.
(
1957)
also
reported
high
acute
lead
values
for
the
mosquitofish;
how­

ever,
these
authors
also
did
not
report
water
hardness
and
the
test
was
con­

ducted
in
turbid
water
contining
suspended
clay
particles
at
apporoximately
300,000
ug/
l
(
Table
6).
Pickering
and
Henderson
(
1966)
found
that
lead
ace­

tate
was
about
as
toxic
as
lead
chloride
to
the
fathead
minnow
in
soft
water
(
Tables
1
and
6).

An
exponential
equation
was
used
to
describe
the
observed
relationship
of
the
acute
toxicity
of
lead
to
hardness
in
fresh
water.
A
least
squares
regression
of
the
natural
logarithms
of
the
acute
values
on
the
natural
logarithms
of
hardness
produced
slopes
of
1.05,
2,48,
1.60,
and
1.01,
re­

spectively,
for
Daphnia
magna,
rainbow
trout,
fathead
minnow,
and
bluegill
(
Table
1).
The
slope
for
Daphnia
magna
was
significant,
but
that
for
rain­

bow
trout
was
not.
The
slopes
for
the
bluegill
and
fathead
minnow
were
based
on
data
for
two
hardnesses
each,
although
four
tests
are
available
with
the
minnow.
An
arithmetic
mean
slope
of
1.22
was
calculated
for
the
three
species
other
than
the
rainbow
trout.
This
mean
slope
was
used
with
the
geometric
mean
toxicity
value
and
hardness
for
each
species
to
obtain
a
logarithmic
intercept
for
each
of
the
nine
freshwater
species
for
which
acute
values
are
available
for
lead.

The
species
mean
intercept,
calculated
as
the
exponential
of
the
loga­

rithmic
intercept,
was
used
to
compare
the
relative
acute
sensitivities
(
Table
3).
The
Guidelines
specify
that
in
order
to
derive
a
criterion
the
minimum
data
base
should
include
at
least
one
acute
value
for
a
benthic
in­

sect.
No
such
value
is
available
for
lead.
However,
7­
to
28­
day
soft
B­
4
water
exposures
of
the
mayfly,
stonefly,
and
caddisfly
to
lead
have
been
re­

ported
by
Nehring
(
1976),
Warnick
and
Bell
(
1969),
and
Spehar,
et
al.
(
1978)

(
Table
6).
Their
results
indicate
that
benthic
insects
are
rather
insensi­

tive
to
lead.
Although
the
data
are
not
really
comparable,
it
appears
that
the
caddisfly
may
be
the
least
sensitive
of
the
three
and
may
be
slightly
less
sensitive
than
the
goldfish.
In
an
attempt
to
account
in
some
way
for
these
insensitive
species
in
the
derivation
of
the
Final
Acute
Intercept,
a
caddisfly
was
entered
as
the
least
sensitive
species
in
the
list
of
fresh­

water
intercepts
in
Table
3.

A
freshwater
Final
Acute
Intercept
of
0.623
kg/
l
was
obtained
for
lead
using
the
species
mean
intercepts
listed
in
Table
3
and
the
calculation
pro­

cedures
described
in
the
Guidelines.
Thus
the
Final
Acute
Equation
is
e(
l.
22[
ln(
hardness)]­
0.47).

No
standard
acute
toxicity
values
for
saltwater
fish
species
are
availa­

ble
but
several
are
available
for
invertebrate
species.
The
most
sensitive
invertebrate
species
was
a
copepod
Acartia
clausi
with
an
LC50
of
668
,.,
9/
l
and
the
least
sensitive
was
the
soft
shell
clam
Mya
arenaria
with
an
LC50
of
27,000.
A
value
of
2,450
was
obtained
with
oyster
larvae
Crassostrea
virginica
in
a
static
test
and
a
LC50
of
2,960
was
recorded
for
mysid
shrimp
Mysidopsis
bahia
in
a
flow­
through
test
in
which
concentrations
were
measured
(
Table
1).
Acute
values
are
not
available
for
enough
appropriate
kinds
of
species
to
allow
calculation
of
a
Saltwater
Final
Acute
Value,

Chronic
Toxicity
Four
tests
of
the
chronic
toxicity
of
lead
to
freshwater
invertebrate
species
have
been
conducted
(
Table
2).
Chapman,
et
al.
(
Manuscript)
studied
the
chronic
toxicity
of
lead
to
Daphnia
magna
at
three
different
hardnes­

ses.
Results
shown
in
Table
2
demonstrate
that
daphnids
were
nearly
11
B­
5
times
more
sensitive
to
lead
in
the
soft
water.
For
the
same
species
in
a
different
soft
water,
a
chronic
value
over
four
times
higher
(
Table
6)
was
obtained
by
Biesinger
and
Christensen
(
1972)
in
a
test
in
which
the
concen­

trations
of
lead
were
not
measured.
Use
of
the
comparable
acute
value
of
450
Vg/
l
(
Table
1)
results
in
an
acute­
chronic
ratio
of
8.2.

A
life
cycle
test
on
lead
in
hard
water
was
conducted
by
Borgmann,
et
al.
(
1978)
with
snails.
These
authors
used
biomass
as
their
endpoint
and
reported
that
lead
concentrations
as
low
as
19
vg/
l
significantly
decreased
survival
but
not
growth
or
reproduction.
After
a
thorough
review
of
this
work,
however,
it
was
not
at
all
clear
how
these
investigators
arrived
at
such
a
low
effect
concentration.
This
publication
did,
however,
contain
suitable
information
for
determining
a
chronic
value.
Chronic
limits
were
taken
directly
from
the
cumulative
percent
survival
figure
which
showed
no
observed
effect
on
survival
at
12
pg/
l
and
almost
complete
mortality
at
54
ug/
l
l
The
chronic
value
for
snails
shown
in
Table
2
was
therefore
estab­

lished
at
25
ug/
l,
which
is
somewhat
lower
than
the
chronic
value
reported
for
daphnids
in
hard
water.

Seven
chronic
tests
on
lead
have
been
conducted
with
six
species
of
freshwater
fish
(
Table
2),
all
of
which
were
in
soft
water.
In
addition,

Davies,
et
al.
(
1976)
described
the
long­
term
effects
on
rainbow
trout
fry
and
finger
lings
exposed
to
various
concentrations
of
lead
for
19
months
in
hard
and
soft
water
(
Table
6).
Although
these
experiments
were
neither
life
cycle
(
no
natural
reproduction)
nor
early
life
stage
(
no
embryos
exposed),

they
do
provide
valuable
information
concerning
the
relationship
between
water
hardness
and
chronic
lead
toxicity
to
fish.
During
these
19­
month
ex­

posures,
most
of
the
trout
(
60
to
100
percent)
developed
spinal
deformities
in
hard
water
at
measured
lead
concentrations
of
850
ug/
l
and
above.

B­
6
However,
during
the
soft
water
exposure
most
trout
(
44
to
97
percent)

developed
spinal
deformities
in
measured
lead
concentrations
as
low
as
31
pg/
l
(
Table
6).
These
results
strongly
demonstrate
that
lead
is
more
chronically
toxic
in
soft
water
than
in
hard
water,

Davies,
et
al.
(
1976)
also
published
results
of
an
early
life
stage
test
with
rainbow
trout
in
soft
water
(
Table
2).
Even
through
this
test
was
started
with
embryos
and
continued
for
19
months
after
hatch,
it
could
not
be
considered
a
life
cycle
test
because
no
reproduction
occurred,
The
chronic
limits
that
these
authors
chose
were
somewhat
lower
than
those
shown
in
Table
2,
because
they
based
their
results
on
a
very
low
incidence
of
black
colored
tails
and
spinal
deformities
(
0.7
and
4.7
percent,
respective­

ly)
l
Because
this
test
was
not
conducted
with
duplicate
exposures,
statis­

tically
significant
differences
could
not
be
determined.
After
careful
ex­

amination
of
their
results
it
was
decided
that
the
chronic
limits
(
Table
2)

should
be
established
on
the
occurrence
of
spinal
curvatures
only
and
at
lead
concentrations
which
caused
a
substantial
increase
in
these
deformi­

ties.
Even
though
the
incidence
of
black
tail
was
apparently
related
to
the
concentration
of
lead,
it
could
not
by
itself
be
considered
an
important
adverse
effect.

Spinal
deformities
have
also
been
cause
by
lead
in
a
life
cycle
test
with
brook
trout
(
Holcombe,
et
al.
1976)
and
in
early
life
stage
tests
with
rainbow
trout,
northern
pike
and
walleye
(
Sauter,
et
al.
1976).
On
the
other
hand,
Sauter,
et
al.
(
1976)
did
not
observe
deformities
during
early
life
stage
tests
with
lake
trout,
channel
catfish,
white
sucker,
and
blue­

gill.
Results
of
tests
by
Sauter,
et
al.
(
1976)
with
northern
pike
and
walleye,
however,
were
not
included
in
Tables
2
and
6
because
of
excessive
mortality
due
to
cannibalism
and
feeding
problems.
The
chronic
value
ob­

tained
for
rainbow
trout
by
Sauter,
et
al.
(
1976)
is
somewhat
higher
than
B­
7
that
chronic
value
derived
from
Davies,
et
al.
(
1976).
Even
though
the
hardnesses
were
about
the
same,
differences
could
be
due
to
differences
in
the
length
of
exposure
(
2
months
vs.
19
months).

As
was
done
with
the
freshwater
acute
values,
the
freshwater
chronic
values
of
Chapman,
et
al.
(
Manuscript)
were
regressed
against
hardness
to
account
for
the
apparent
effect
of
hardness
on
the
chronic
toxicity
of
lead
and
a
slope
of
2.35
was
obtained.
Even
though
this
slope
is
not
significant
because
it
is
based
on
only
three
values,
it
relects
the
obvious
effect
of
hardness
on
chronic
toxicity.
In
the
same
manner
as
for
acute
toxicity,
the
chronic
slope
was
used
with
the
geometric
mean
chronic
toxicity
value
and
hardness
for
each
species
to
obtain
a
logarithmic
intercept
and
a
species
mean
chronic
intercept
for
each
species
for
which
a
chronic
value
is
availa­

ble
(
Table
2).
A
Freshwater
Final
Chronic
Intercept
of
0.000076
ug/
l
was
then
obtained
using
the
calculation
procedures
described
in
the
Guidelines.

Thus,
the
Final
Chronic
Equation
is
e
(
2,35[
ln(
hardness)]­
9.48).

The
mysid
shrimp
Mysidopsis
bahia
is
the
only
saltwater
species
with
which
a
chronic
test
has
been
conducted
on
lead
(
Table
2).
The
most
sens­

itive
observed
adverse
effect
was
reduced
spawning
(
U.
S.
EPA,
1980)
and
the
resulting
chronic
value
was
25
ug/
l.
The
96­
hour
LC50
for
this
same
spe­

cies
in
the
same
study
was
2,960
ug/
l,
producing
an
acute­
chronic
ratio
of
119.

Plant
Effects
Four
static
tests
on
three
species
of
algae
have
been
reported
by
Mona­

han
11976)
(
Table
4).
These
exposures
were
conducted
for
7
days
and
concen­

trations
of
lead
were
not
measured.
Results
of
short
exposures
of
algae
and
diatoms
to
unmeasured
lead
concentrations
have
also
been
published
by
Malan­

chuk
and
Gruendling
(
1973)
(
Table
6).
The
adverse
effect
concentrations
from
these
tests
ranged
from
500
to
28,000
ug/
l.
It
would
appear
therefore
B­
8
that
any
adverse
effects
of
lead
on
plants
are
unlikely
at
concentrations
protective
of
chronic
effects
on
freshwater
animals.

No
saltwater
plant
species
have
been
exposed
to
inorganic
lead,
but
one
saltwater
algal
species
Dunaliella
tertiolecta
has
been
exposed
to
both
tetramethyl
and
tetraethyl
lead.
The
results
(
Table
6)
demonstrate
that
this
species
is
more
sensitive
to
tetraethyl
lead
by
a
factor
greater
than
10.
No
data
are
available
concerning
the
relative
toxicities
of
inorganic
lead
and
these
organolead
compounds.

Residues
Four
freshwater
invertebrate
species
have
been
exposed
to
lead
(
Borg­

mann,
et
al.
1978;
Spehar,
et
al.
1978)
and
the
bioconcentration
factors
ranged
from
499
to
1,700
(
Table
5).
Brook
trout
and
bluegills
were
also
ex­

posed
to
lead
(
Holcombe,
et
al.
1976,
and
Atchison,
et
al.
1977)
and
calcu­

lated
bioconcentration
factors
were
42
and
45,
respectively
(
Table
5).

Some
species
of
saltwater
bivalve
molluscs,
diatoms
and
phytoplankton
are
capable
of
accumulating
lead
(
Table
5).
The
bioconcentration
factors
range
from
17.5
with
the
hard
clam
to
2,570
with
the
mussel.
Because
the
duration
of
the
study
may
be
an
important
consideration
in
bioconcentration
studies,
this
comparison
is
not
entirely
valid
since
the
mussel
was
exposed
for
130
days
and
the
hard
clam
for
only
56
days.

Neither
a
freshwater
nor
a
saltwater
Final
Residue
Value
can
be
calcu­

lated
because
no
maximum
permissible
tissue
concentration
is
available
for
lead.

Miscellaneous
Many
of
the
values
in
Table
6
have
already
been
discussed.
Spehar
(
1978)
found
no
adverse
effects
on
a
freshwater
snail,
scud,
stonefly,
and
caddisfty
in
28
days
at
565
ug/
l.
Pickering
and
Henderson
(
1966)
found
that
lead
chloride
and
lead
acetate
are
about
equally
toxic
to
fathead
minnows
in
B­
9
static
tests
in
soft
water
(
Table
1
and
6),
but
Wallen,
et
al.
(
1957)
found
that
lead
oxide
is
much
less
acutely
toxic
than
lead
nitrate
to
the
mosqui­

tofish
in
turbid
water.

The
lOday
test
conducted
by
Anderson,
et
al.
(
1980)
(
Table
6)
showed
that
the
midge,
Tanytarsus
dissimilis,
is
rather
insensitive
to
lead
with
a
chronic
value
of
258
ug/
l.
This
test
included
exposure
of
the
species
during
most
of
its
life
cycle
and
several
of
the
presumably
sensitive
molts,

and
so
should
probably
be
considered
as
useful
as
the
early
life
stage
test
with
fish.

A
variety
of
other
effects
on
saltwater
organisms
have
been
observed.

Gray
and
Ventilla
(
1973)
observed
a
reduction
in
growth
rate
in
a
ciliate
protozoan
after
a
12
hour
exposure
to
a
lead
concentration
of
150
pg/
l.

Woolery
and
Lewin
(
1976)
observed
a
reduction
in
photosynthesis
and
respira­

tion
in
the
diatom
Pheodactylum
tricornutum
at
concentrations
of
lead
ranging
from
100
to
10,
OCJO
pg/
l.
However,
Hannan
and
Patouillet
(
1972)

obtained
no
growth
inhibition
with
P.
tricornutum
at
a
concentration
of
1,000
ug/
l
after
72
hours.
Rivkin
(
1979)
using
growth
rate
to
determine
toxicity
to
the
diatom,
Skeletonema
costatum,
reported
a
12
day
ECSO
of
5.1
Llg/
l.
Hessler
(
1974)
observed
delayed
cell
division
in
the
phytoplank­

ton,
Platymonas
subcordiformus,
after
treatment
with
2,500
pg/
l
for
72
hours.
At
60,000
ug/
l,
Hessler
(
1974)
reported
not
only
growth
retardation
but
also
death.
Benijts­
Claus
and
Benijts
(
1975)
observed
delayed
larval
development
in
the
mud
crab,
Rhithropanopeus
harrisii,
after
treatment
with
lead
concentrations
of
50
vg/
l.
In
Fundulus
heteroclitus,
Weis
and
Weis
(
1977)
observed
depressed
axis
formation
in
developing
embryos
with
lead
concentrations
of
100
ug/
l.
Reish
and
Cat­
r
(
1978),
found
that
1,000
ug/
l
suppressed
reproduction
of
two
polychaete
species,
Ctenodriluis
serratus
and
Ophryotrocha
disdema,
in
a
2lday
test.

B­
10
Summary
Standard
acute
data
for
lead
are
available
for
nine
freshwater
fish
and
invertebrate
species
with
a
range
from
124
to
542,000
vg/
l.
Chronic
tests
have
been
conducted
with
two
invertebrate
species
and
six
fish
species
with
the
chronic
values
ranging
from
12
to
174
ug/
l.
Both
the
acute
and
chronic
toxicities
of
lead
to
freshwater
animals
decrease
as
hardness
increases.

Freshwater
algae
are
affected
by
concentrations
of
lead
above
500
ugll,

based
on
data
for
three
species.
Bioconcentration
factors
ranging
from
42
to
1,700
are
available
for
four
invertebrate
and
two
fish
species.

Acute
values
for
five
saltwater
species
ranged
from
668
pg/
l
for
a
cope­

pod
to
27,000
ug/
l
for
the
soft
shell
clam.
A
chronic
toxicity
test
was
conducted
for
the
mysid
shrimp
and
adverse
effects
were
observed
at
37
,.,
g/
l
but
not
at
17
ug/
l.
The
acute­
chronic
ratio
for
this
species
is
118.

Delayed
embryonic
development,
suppressed
reproduction
and
inhibition
of
growth
rate
among
fish,
crab,
polychaete
worm,
and
plankton
were
also
caused
by
lead.

CRITERIA
For
total
recoverable
lead
the
criterion
(
in
pg/
l)
to
protect
freshwater
aquatic
life
as
derived
using
the
Guidelines
is
the
numerical
value
given
by
,(
2.35[
ln(
hardness)]­
9.48)
as
a
24­
hour
average
and
the
concentration
(
in
ug/
1)
should
not
exceed
the
numerical
value
given
by
e
(
1.22[
ln(
hardness)]

­
0.47)
at
any
time.
For
example,
at
hardnesses
of
50,
100,
and
200
mg/
l
as
CaC03
the
criteria
are
0.75,
3.8,
and
20
pg/
l,
respectively,
as
24­
hour
averages,
and
the
concentrations
should
not
exceed
74,
170,
and
400
ug/
l,

respectively,
at
any
time.

B­
11
The
available
data
for
total
recoverable
lead
indicate
that
acute
and
chronic
toxicity
to
saltwater
aquatic
life
occur
at
concentrations
as
low
as
668
and
25
ug/
l,
respectively,
and
would
occur
at
lower
concentrations
among
species
that
are
more
sensitive
than
those
tested.

B­
12
.
G
t
i
ul
%
.
F
f
i
SI
8
I
I
I
I
I
I
I
I
I
I
Q
.
r
In
P
,'

2
n
f
R
R
8
I
8
I
4
to
t
t
1
`
E
a
3
s
a
s
t
ac'
a
u;
a
v;
a
t`
­
a
If
a
UT
a
v;
a
u;
8
u­
l
4
*

0
c­
4
c
v
c
n
a
c
.
Ln
w­
l
2
v\
t
n
I
4
m
.
z?
.
Y
3
I
v\
.
R
N
­

K
0
N
A
.
;
ij
t
s
t
2
L
­
8
I:
I
El
I
I
a
4
a
3
I
I
0
:
a
­
I
2
I
2
t
pl
+
+
m
m
8
m
f
s­
4
f
x
s
I
d
3
f
I
I
I
I
I
I
I
I
Q
P
f
0
LO
"
f
a
j
t­
2%
­­
I
I
I
I
I
I
I
I
I
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B­
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Bull.
73:
401.

B­
37
Weis,
J.
S.
1976.
Effects
of
mercury,
cadmium,
and
lead
salts
on
regenera­

tion
and
ecdysis
in
the
fiddler
crab,
Uca
pugilator.
U.
S.
Dep.
Comm.
Natl.

Ocean.
Atmos.
Admin.
Fish.
Bull.
74:
464.

Weis,
J.
S.
and
P.
Weis.
1977.
Effect
of
heavy
metals
on
development
of
the
killifish,
Fundulus
heteroclitus.
Jour.
Fish.
Biol.
11:
49.

Whitley,
L.
S.
1968.
The
resistance
of
tubificid
worms
to
three
comnon
pol­

lutants.
Hydrobiol.
32:
193.

Woolery,
M.
L.
and
R.
A
Lewin.
1976.
The
effects
of
lead
on
algae.
IV.
Ef­

fects
of
lead
on
respiration
and
photocynthesis
of
Phaeodactylum
tricornutum
(
Bacillariophyceae).
Water
Air
Soil
Pollut.
6:
25.

Zaroogian,
G.
E.,
et
al.
1979.
Crassostrea
virginica
as
an
indicator
of
lead
pollution.
Mar.
Biol.
52:
189.

8
­
38
Mammalian
Toxicology
and
Human
Health
Effects
INTRODUCTION
The
hazards
of
lead
exposure
have
been
under
intensive
inves­

tigation
for
many
years.
Research
activities
continue
for
several
reasons.
First,
industrial
production
and
commercial
use
continues
at
a
fairly
steady
rate.
Second,
hazardous
sources
persist
in
the
environment
long
after
the
hazard­
generating
practice
has
been
cur­

tailed.
A
good
example
is
the
persistence
of
lead­
base
paint
in
houses
long
after
the
elimination
of
lead­
containing
pigments
from
new
household
paints.
Finally,
as
biomedical
science
in
general
and
toxicology
in
particular
continue
to
push
back
the
frontiers
of
knowledge,
indices
of
toxicity
change,
generally
with
a
consequent
downward
revision
of
what
is
considered
an
acceptable
level
of
human
exposure
to
environmental
pollutants.

Reassessment
of
acceptable
levels
of
lead
exposure
have
been
fairly
numerous
in
recent
years.
These
have
taken
the
form
of
cri­

teria
documents
and
of
more
academically­
oriented
reviews.
Some
have
been
highly
comprehensive,
covering
effects
on
the
ecosystem
in
general,
as
well
as
on
man
[
National
Academy
of
Sciences
(
NAS),

1972;
Boggess,
1978].
Others
have
been
mainly
concerned
with
ef­

fects
of
lead
on
man
[
World
Health
Organization
(
WHO),
1977;
U.
S.

EPA,
1977;
Hammond,
1977].

The
purpose
of
this
review
is
to
summarize
the
literature
which
is
most
relevant
to
the
question
of
what
is
an
acceptable
level
of
human
exposure
to
lead
via
water.
In
doing
so,
it
is
necessary
to
consider
the
consequences
to
human
health
of
one
or
another
level
of
intake
assignable
to
water
and
to
the
numerous
other
sources.

C­
l
EXPOSURE
Natural
Rackground
Levels
Lead
is
ubiquitous
in
nature,
being
a
natural
constituent
of
the
earth's
crust.
The
usual
concentration
in
rocks
and
in
soils
from
natural
sources
ranges
from
10
to
30
mg/
kg.
Yost
natural
groundwaters
have
concentrations
ranging
from
1
to
10
µ
g/
l.
This
is
well
below
the
United
States'
drinking
water
standard
of
50
µ
g/
l.
It
is
much
easier
to
specify
natural
levels
of
lead
in
rocks
and
soil
than
in
vegetation
since
long­
range
transport
of
lead
from
man­
made
sources
via
the
air
inevitably
contaminates
both
surface
soil
and
plants
growing
thereon.
The
normal
concentration
of
lead
in
rural
vegetation,
however,
ranges
from
0.1
to
1.0
mg/
kg
dry
weight,
or
2
to
20
mg/
kg
ash
weight.
Thus,
nutrient
movement
from
soil
to
the
organic
matter
in
plants
via
water
does
not
result
in
any
noticeable
degree
of
biomagnification.
Again,
because
of
the
impact
of
long­
range
transport
of
lead
via
air
from
man­
generated
sources,
it
is
only
possible
to
specify
lowest
concentrations
found
over
areas
of
the
globe
most
remote
from
human
activity.
These
are
of
the
order
of
0.0001
to
0.001
µ
g/
m3,
mostly
measured
over
Green­

land
and
over
remote
oceans.

Areas
of
abnormally
high
concentrations
of
lead
occur
in
natu­

ral
ores,
usually
in
conjunction
with
high
concentrations
of
cad­

mium
and
zinc.
There
is
essentially
no
transfer
from
natural
ore
beds
into
overlying
streams:
and
there
is
none
if
the
soil
is
even
slightly
alkaline
(
Jennett,
et
al.
1977).

C­
2
Man­
generated
Sources
of
Lead
Lead
consumption
in
the
United
States
has
been
fairly
stable
from
year
to
year
at
about
1.3
x
106
metric
tons.
Approximately
half
of
that
consumption
has
been
for
the
manufacture
of
storage
batteries
and
one­
fifth
has
been
for
the
manufacture
of
gasoline
antiknock
additives,
notably
tetraethyl
lead
and
tetramethyl
lead.

Pigments
and
ceramics
account
for
about
6
percent
of
annual
pro­

duction.
All
other
major
uses
are
for
metallic
lead
products
or
for
lead­
containing
alloys.
The
consumption
of
tetraethyl
lead
and
tetramethyl
lead
is
declining.
Other
uses
that
have
significant
potential
for
input
into
man
are
for
paint
pigment
and
solder.

Paints
applied
to
surfaces
will
eventually
crack,
flake
or
peel.

Children
are
known
to
ingest
this
type
of
deteriorating
paint.

Solder
also
is
a
potential
source
of
lead
exposure
either
when
used
to
seal
water
pipe
joints
or
for
joining
seams
in
metal
food
and
beverage
containers.

Ingestion
from
Water
Lead
does
not
move
readily
through
stream
beds
because
it
easily
forms
insoluble
lead
sulfate
and
carbonate.
Moreover,
it
binds
avidly
to
organic
ligands
of
the
dead
and
living
flora
and
fauna
of
stream
beds.
Nonetheless,
under
special
circumstances,

lead
does
have
considerable
potential
for
hazardous
exposure
to
man
via
drinking
water.
In
areas
where
the
home
water
supply
is
stored
in
lead­
lined
tanks
or
where
it
is
conveyed
to
the
water
tap
by
lead
pipes,
the
concentration
may
reach
several
hundred
micrograms
per
liter
or
even
in
excess
of
1,000
µ
g/
l
(
Beattie,
et
al.
1972).

There
is
a
definite
positive
correlation
between
the
concentration
C­
3
of
lead
in
the
domestic
water
supply
and
the
concentration
of
lead
in
the
blood.
The
concentration
of
lead
in
the
water
conveyed
through
lead
pipes
is
dependent
on
a
number
of
factors.
The
longer
the
water
has
stood
in
the
pipes,
the
higher
the
lead
concentration
(
wong
and
Rerrang,
1976).
The
lower
the
pH
of
the
water
and
the
lower
the
concentration
of
dissolved
salts
in
the
water,
the
great­

er
is
the
solubility
of
lead
in
the
water.
Leaching
of
lead
from
plastic
pipes
has
also
been
documented
(
Heusgem
and
De
craeve,

1973).
The
source
of
lead
was
probably
lead
stearate,
which
is
used
as
a
stabilizer
in
the
manufacture
of
polyvinyl
plastics.
The
magnitude
of
the
problem
of
excessive
lead
in
tap
water
is
not
ade­

quately
known.
In
one
recent
survey
of
969
U.
S.
water
systems,
1.4
percent
of
all
tap
water
exceeded
the
50
ug/
l
standard
(
WCabe,

1970).
Special
attention
should
be
given
in
water
quality
surveil­

lance
to
soft
water
supplies,
es?
eciallv
those
with
a
DH
<
6.5.

Future
survey
work
should
also
indicate
whether
or
not
the
water
was
filtered
before
analysis.
This
appears
to
be
a
common
nractice
among
water
analysts.
Since
a
substantial
fraction
of
the
lead
in
drinking
water
probably
is
in
particulate
form,
filtration
prior
to
analysis
could
give
deceivingly
low
analytical
values
especially
if
a
substantial
fraction
of
the
particulate
lead
in
water
is
avail­

able
for
absorption.
However,
"
drinking
water"
analyses
are
usual­

ly
performed
in
unfiltered
water
and
hence
represent
total
lead.

Ingestion
from
Food
It
is
generally
held
that
food
constitutes
the
major
source
of
lead
ingested
by
people.
Raw
fruits
and
vegetables
acquire
lead
by
surface
denosition
from
rainfall,
dust
and
soil,
as
well
as
from
c­
4
uptake
via
the
root
system.
The
relative
contribution
of
these
two
sources
varies
greatly
depending
upon
whether
the
edible
portion
is
leafy
or
not.
Furthermore,
the
nature
of
food
processing
may
either
lower
or
raise
the
concentration
in
the
raw
product
­
e.
g.,

washing
as
compared
to
packing
in
metal.
cans
with
lead
solder
seams.
There
is
no
evidence
of
biomagnification
in
the
food
chain,

e.
g.,
from
aquatic
vegetation
to
the
edible
portions
of
fish
and
shellfish.
Therefore,
fish
do
not
constitute
an
unusually
signifi­

cant
source
of
lead
in
man's
diet.

Schroeder,
et
al.
(
1961)
reported
0
to
1.5
mg/
kg
of
lead
for
condiments,
0.2
to
2.5
mg/
kg
for
fish
and
seafood,
0
to
0.37
mg/
kg
for
meat
and
eggs,
and
0
to
1.3
mg/
kg
for
vegetables.
Other
more
recent
studies
have
confirmed
this
observation.
Yany
foods
and
beverages
are
packed
in
metal
cans
which
have
a
lead­
soldered
side
seam
and
caps.
The
concentration
of
lead
in
the
contents
is
sub­

stantially
higher
after
packing
than
before,
and
is
also
higher
than
the
same
product
packed
in
glass
[
Mitchell
and
Aldous,
1974;

U.
S.
Food
and
Drug
Administration
(
U.
S.
FDA)
,
19751.
In
some
instances,
the
lead
probably
leaches
from
the
solder
through
cracks
or
pores
in
the
protective
shellac
coating
applied
to
the
inside
of
the
can.
In
many
other
instances,
however,
microscopic
pellets
of
lead
splatter
inside
the
can
during
the
soldering
process.
Their
availability
for
absorption
may
differ
substantially
from
that
of
lead
leached
into
solution.

Yilk
has
been
studied
extensively
as
to
lead
content
because
it
constitutes
a
substantial
fraction
of
the
diet
of
infants
and
young
children.
Whole
raw
cow
milk
has
a
concentration
of
about
9
c­
5
Lq/
l
!
Hammond
and
Aronson,
1964)
whereas
market
milk
has
an
average
of
40
~
q/
l
(
Yitchell
and
Aldous,
1974).
Evaporated
milk
has
`
been
variously
reported
to
contain
an
average
of
202
ug/
l.
("
qitchell
and
Aldous,
1974),
110
f
11
ug/
l
(
Lamm
and
Rosen,
1974),
and
330
to
87C
;
q
:
I
i'!
urthy
and
Rhea,
1971).

7­
e
daily
dietary
intake
of
lead
has
been
estimated
by
numer­

OUS
inl:
e;
tiqators,
using
either
the
duplicate
portions
approach
or
t.:
le
c.::?
osites
technique
wherein
theoretical
diets
are
derived
us
1
n
:
j
'­.
'
2
t
_
witi
on
tables.
The
results
are
qenerally
consistent,
con­

siderinq
variations
in
body
size
and
metabolic
rates.
Thus,
'
lord­

man
(
1975)
reported
an
average
daily
intake
of
231
ug
Pb
for
Fin­

nish
adult
males
and
178
ug
Pb
for
adult
females.
This
is
consis­

tent
with
a
British
study
reporting
274
ug
Pb/
day
for
young
adults
(
Thompson,
1971)
and
with
a
Japanese
study
reporting
293
ug
Pb/
day
for
adult
males
doing
medium
work
(
Horiuchi,
et
al.
1956).
The
first
two
studies
(
Nordman,
1975;
Thompson,
1971)
described
the
duplicate
portions
technique
whereas
the
third
(
Voriuchi,
et
al.

1956)
used
the
composites
approach.
Kolbye,
et
al.
(
1974)
analyzed
the
difficulties
inherent
in
applying
this
approach.
Kehoe
(
1961)

reported
an
average
intake
of
218
ug
Pb/
dav
for
sedentarv
men.

This
is
not
consistent,
however,
with
two
other
American
studies
of
daily
fecal
lead
excretion
(
Griffin,
et
al.
1975;
Tepper
and
ievin,

1972).
From
the
lead
balance
studies
of
Kehoe
(
1961),
it
can
be
estimated
that
gastrointes,
tinal
absorption
of
lead
approxim,
ates
8
percent.
Making
this
adjustment,
daily
lead
intake
from
the
diet
based
on
Eecal
lead
excretion
would
be
113
s­
19
in
sedentarv
adult
males
(
Griffin,
et
al.
1975)
and
119
ug
in
women
(
Tepoer
and
Levin,

1972).

C­
6
Many
studies
of
dietary
lead
intake
are
somewhat
vague
as
to
whether
water
consumption
was
included
in
the
estimates.
Others
specify
"
food
and
beverages."

The
dietary
intake
of
lead
in
infants
and
young
children
has
not
been
studied
as
extensively
as
it
has
in
adults.
Using
the
duplicate
diet
approach,
Alexander,
et
al.
(
1973)
estimated
a
range
of
40
to
210
pg/
day
of
lead
for
children
ranging
in
age
from
three
months
to
8.5
years.
Horiuchi,
et
al.
(
1956)
estimated
126
ug/
day
of
lead
for
youngsters
10
months
old.
These
seemingly
high
values
compared
to
adults
are
not
too
surprising
considering
the
high
caloric
and
fluid
requirements
of
children
in
proportion
to
their
weight.

A
bioconcentration
factor
(
BCF)
relates
the
concentration
of
a
chemical
in
aquatic
animals
to
the
concentration
in
the
water
in
which
they
live.
An
appropriate
BCF
can
be
used
with
data
concern­

ing
food
intake
to
calculate
the
amount
of
lead
which
might
be
ingested
from
the
consumption
of
fish
and
shellfish.
Residue
data
for
a
variety
of
inorganic
compounds
indicate
that
bioconcentration
factors
for
the
edible
portion
of
most
aquatic
animals
are
similar,

except
that
for
some
compounds
bivalve
molluscs
(
clams,
oysters,

scallops,
and
mussels)
should
be
considered
a
separate
group.
An
analysis
(
U.
S.
EPA,
1980)
of
data
from
a
food
survey
was
used
to
estimate
that
the
per
capita
consumption
of
freshwater
and
estua­

rine
fish
and
shellfish
is
6.5
g/
day
(
Stephan,
1980).
The
per
capita
consumption
of
bivalve
molluscs
is
0.8
g/
day
and
that
of
all
other
freshwater
and
estuarine
fish
and
shellfish
is
5.7
g/
day.

c­
7
Several
bioconcentration
factors
are
available
for
the
edible
portions
of
bivalve
molluscs:

Soecies
Oyster,
Crassostrea
virqinica
Oyster,
Crassostrea
virqinica
BCF
536
68
Reference
Zarooqian,
et
al.
1979
Pringle,
et
al.
1968
Oyster,
Crassostrea
virqinica
Auahauq,
hard
clam,
Mercenaria
mercenaria
1,400
17.5
Shuster
and
Prinqle,
1969
Prinqle,
et
al.
1968
'
Soft
shell
clam,
Mya
arenaria
112
Prinqle,
et
al.
1968
Yussel,
Mytilus
edulis
650
Schulz­
Baldes,
1974
Yussel,
Mytilus
edulis
200
Talbot,
et
al.
1976
Yussel,
Yytilus
edulis
2,570
Schulz­
Raldes,
1972
Yussel,
Mytilus
edulis
Yussel,
Mytilus
edulis
2,080
Schulz­
Raldes,
1972
796
Schulz­
Raldes,
1972
The
geometric
mean
bioconcentration
factor
for
lead
in
bivalve
molluscs
is
375,
but
no
data
are
available
for
appropriate
tissues
in
other
aquatic
animals.
Based
on
the
available
data
for
cooper
and
cadmium,
the
mean
BCF
value
for
other
species
is
probably
about
one
percent
of
that
for
bivalve
molluscs.
If
the
values
of
375
and
3.8
are
used
with
the
consumption
data,
the
weighted
average
BCF
for
lead
and
the
edible
portion
of
all
freshwater
and
estuarine
aquatic
organisms
consumed
by
Americans
is
calculated
to
be
49.

C­
8
Inhalation
The
third
major
obligatory
source
of
lead
in
the
general
popu­

lation
is
ambient
air.
A
great
deal
of
controversy
has
been
qener­

ated
regarding
the
contribution
of
air
to
total
daily
lead
absorp­

tion.
Unlike
the
situation
with
food
and
water,
general
ambient
air
lead
concentrations
vary
greatly.
In
metronolitan
areas
aver­

age
air
lead
concentrations
of
2
ug/
m3
with
excursions
of
10
"
q/
m3
in
areas
of
heavy
traffic
or
industrial
point
sources
are
not
un­

common,
whereas
in
nonurban
areas,
average
air
lead
concentrations
usually
are
of
the
order
of
0.1
uq/
m3.
In
addition,
people
are
so
mobile
that
static
air
sampling
devices
are
not
very
useful
for
estimating
the
integrated
air
lead
exposure
of
urban
populations.

Dermal
Exposure
of
the
skin
to
lead
probably
is
significant
only
under
special
circumstances
such
as
among
workers
in
contact
with
lead­
based
gear
compounds
or
greases,
or
blenders
of
alkyl
lead
fuel
additives.
It
is
very
unlikely
that
the
concentrations
of
lead
in
water
or
air
are
sufficient
to
make
dermal
contact
a
siq­

nificant
route
of
exposure.

Miscellaneous
SOUrCeS
Among
adults
not
occupationally
exposed
to
lead,
there
are
sev­

eral
sources
of
lead
which
may
assume
clinically
significant
pro­

portions.
Perhaps
the
most
serious
widespread
problem
is
the
con­

sumption
of
illicitly
distilled
whiskey
(
moonshine)
which
is
often
heavily
contaminated
with
lead.
Many
cases
of
frank
lead
poisoning
have
been
documented.
The
concentration
of
lead
in
moonshine
whis­

key
commonly
exceeds
10
mg/
l,
or
2,000
times
the
drinking
water
c­
9
standard.
Storage
of
acidic
beverages
in
impronerly
qlazed
earth­

enware
has
caused
severe,
sometimes
fatal
poisoninq
in
the
consumer
(
Klein,
et
al.
1970;
Harris
and
Elsea,
1967).

Occupational
exposure
to
lead
may
be
quite
excessive.
Thus,

in
primary
lead
smelters,
the
air
lead
concentration
may
exceed
1,000
pg/
m3.
A
similar
situation
exists
in
some
storage
battery
manufacturing
plants.
Other
hazardous
occupations
include
welding
and
cutting
of
lead­
painted
metal
structures,
automobile
radiator
repair,
and
production
of
lead­
base
paints.
In
these
occupations,

the
principal
hazard
is
generally
considered
to
be
from
inhalation
of
lead
fumes
and
dusts.
Hand­
to­
mouth
transfer
is
probably
siq­

nificant.

The
hazard
of
lead
to
children
is
of
considerable
concern.

The
number
of
children
excessively
exposed
to
lead
from
miscella­

neous
sources
is
impressive.
Thus,
federally
assisted
lead
screen­

ing
programs
reveal
that
excess
lead
absorption
was
found
in
11.1
percent
of
277,347
children
screened
in
1973.
Blood
lead
levels
(
PbB)
were
reported
to
be
in
excess
of
40
pq/
dl.
The
percentage
has
fallen
since
then,
being
6.4
percent
in
1974
and
6.5
percent
in
1975
(
Hopkins
and
Houk,
1976).
By
1976
the
problem
had
not
changed
appreciably
since
1974
and
1975.
In
that
year,
9.7
nercent
of
500,463
children
screened
had
PbBs
>
30
pq/
dl
and
2.7
percent
or
­

13,604
children
had
PbBs
)
50
uq/
dl
(
Center
for
Disease
Control­,
­

1977).

It
has
long
been
held
that
the
major
source
of
elevated
lead
exposure
in
infants
and
young
children
is
lead­
base
paint
in
the
interior
of
home
and
in
the
soil
surroundinq
the
homes.
Yore
re­

c­
10
cently,
the
high
lead
content
of
soil
and
street
dust
attributable
to
the
fallout
of
lead
from
automobile
exhaust
has
become
suspect.

Thus,
in
the
1972
publication
Airborne
Lead
in
Perspective
(
NAS,

1972),
it
is
pointed
out
that
the
daily
inqestion
of
43
mg
of
street
dust
at
2,000
uq
Pb/
g
would
suffice
to
elevate
the
PbB
of
a
young
child
from
20
ug/
dl
to
40
vg/
dl.
In
a
survey
of
77
midwestern
United
States
cities,
it
was
found
that
the
average
lead
concentra­

tion
in
the
street
dust
of
residential
areas
was
1,636
uq/
q
and
that
in
commercial
and
industrial
areas
the
average
concentrations
were,
respectively,
2,413
uq/
q
and
1,512
ug/
q
(
Hunt,
et
al.
1971).

Soil
along
the
shoulder
of
heavily­
traveled
roadways
also
is
heavi­

ly
contaminated,
although
most
values
found
have
been
in
the
range
of
hundreds
of
micrograms
per
gram
rather
than
thousands
(
for
exam­

pie,
Laqerwerff
and
Specht,
1970).

The
relative
contribution
of
soil,
automotive
exhaust
fallout,

and
paint
to
lead
exposure
in
children
remains
uncertain.
There
is
no
question
that
children
in
the
age
range
of
1
to
5
years,
in
which
the
problem
of
elevated
PbBs
exists,
do
indeed
exhibit
pica,
the
habit
of
mouthing
or
ingesting
nonedible
objects,
e.
g.,
pieces
of
Plastic,
gravel,
cigarette
butts,
etc.
(
Rarltrop,
1966).
T'Zle
habit
also
appears
to
be
more
prevalent
among
children
who
have
elevated
PbBs
than
among
those
who
do
not
(
Mooty,
et
al.
1975).
There
is
strong
evidence
that
paint
is
a
major
source
of
lead
in
chiidren
with
pica.
Thus,
Sachs
(
1974)
reported
that
80
percent
of
patients
seen
because
of
evidence
of
excessive
lead
absorption
had
a
histor:
y
of
eating
paint
or
plaster.
Hammond,
et
al.
(
1977)
reported
that
amc\
ng
29
children
with
elevated
PbBs
(
>
40
pq/
dl)
selected
at
ran­
­

C­
11
dom
from
a
lead
screening
program,
all
but
one
came
from
14
homes
classified
as
having
high
hazard
for
lead­
base
paint,
either
exter­

ior
or
interior
(
Table
1).
Yiqh
hazard
consisted
of
there
beinq
at
least
one
accessible
painted
surface
with
>
0.5
percent
Pb,
peelinq
­

or
otherwise
loose.
The
medium
classification
consisted
of
>
O.
S
­

percent
Pb,
but
the
painted
surface
was
generally
tight.
In
this
study
there
was
f,
ound
to
be
a
highly
significant
correlation
(
D
=

0.007)
between
paint
hazard
classification
(
low,
medium,
hiqh)
and
fecal
lead
excretion,
but
no
correlation
between
fecal
lead
excre­

tion
and
traffic
density
(
vehicles
per
day)
in
the
vicinity
of
the
home
(
p
=
0.41).
Unfortunately,
the
correlation
between
traffic
density
and
the
lead
content
of
soil
and
dust
was
not
determined.

Thus,
the
data
are
merely
susqestive.

Ter
Haar
and
Aronow
(
1974)
reported
that
elevated
lead
expo­

sure
in
eight
children,
hospitalized
for
excessive
lead
absorDtion,

could
not
be
caused
by
lead
from
fallout
of
airborne
cornbusted
automobile
exhaust.
Six
of
the
eight
children
had
distinctly
ele­

vated
fecal
lead
excretion
as
compared
to
nine
control
children,

yet
their
excretion
of
210
Pb,
a
marker
for
aerosol
fallout,
was
no
different
from
that
of
the
controls.
However,
the
children
in
this
study
were
supposed
to
have
ingested
oaint.
The
criteria
were
one
or
all
of
the
following:
(
1)
x­
ray
showed
radio
opaque
materials
in
the
gut,
(
2)
history
of
pica,
(
3)
elevated
PbB,
and
(
4)
x­
ray
showed
Pb
lines
on
the
long
bones.

There
is
other
evidence,
however,
which
suqaests
that
dust
and
soil
are,
under
some
circumstances
at
least,
siqnif
icant
sources
of
lead
for
infants
and
children
and
that
their
effect
is
additive
to
c­
12
TABLE
1
Classification
of
Home
Environments
as
to
Lead
Hazarda
Family
Paintb
Hazard
Lead
Concentration,
%
d.
w.
C
Interior
Exterior
Dust
Dust
Soil
Vehicles
per
cl.
x
lo3
A
J3
C
D
F
G
II
J
L
M
N
P
R
s
H
H
H
H
17
M
H
H
JJ
L(
I);
H(
E)

FJ
M(
I);
H(
E)
H
L(
I),
II(
E)
0.45(
2)

20
O.
ll(
2)

0.3(
l)

0.3(
l)
0.7(
l)

0.1(
l)

4.0(
l)

1.9(
l)

0.6(
l)
0.12(
3)

0.06(
2)

0.07(
l)

0.3(
2)

0.1(
l)

0.2(
l)

0.9(
2)

0.05(
l)

0.1(
3)

0
2.5
­
5
30
10
­
15
2.5
­
5
=
0.5
=
0.5
4­
6
l­
2
2.5
­
5
0.5
­
1
l­
2
2.5
­
5
4­
6
5
­
7.5
'
Source:
Hammond,
et
al.
1977
bII
=
high;
?
II
=
medium;
L
=
low;
(
I)
=
interior;
(
I?)
=
exterior.
Absence
of
(
I)
or
(
E)
designation
means
that
both
conformed
to
the
designated
classification
of
H,
M
or
L.

CNumhers
in
parentheses
indicate
number
of
environmental
samples.

c­
13
that
produced­
by
inhalation.
The
best
evidence
is
orovided
in
a
study
of
a
population
of
children
residinq
in
the
immediate
vicini­

ty
of
a
large
secondarv
lead
smelter
near
El
Paso,
Texas
(
Landri­

gan,
et
al.
1975).
Sixty­
nine
percent
of
one­
to
four­
year­
old
children
living
within
one
mile
of
the
El
Base
smelter
had
blood
lead
levels
greater
than
or
equal
to
40
pg/
dl,
the
level
then
con­

sidered
indicative
of
increased
lead
absorption.
By
contrast,
the
prevalence
of
blood
lead
levels
greater
than
or
equal
to
40
uq/
dl
among
98
adults
living
in
the
same
area
was
16
percent.
The
qeo­

metric
mean
lead
concentration
of
soil
in
that
location
was
1,791
porn
and
that
of
house
dust
was
4,022
ppm.
Lead
based
paint
was
not
­

a
problem.
Therefore
it
seems
likely
that
a
proportion
of
the
lead
intake
in
the
children
living
in
El
Paso
was
oral
rather
than
by
inhalation
and
that
the
net
effect
of
the
two
routes
of
exposure
was
to
place
children
at
a
considerably
increased
risk
of
lead
up­

take
than
adults.
The
mere
presence
of
high
concentrations
of
lead
in
soil
accessible
to
children
is
not
enough
to
create
a
hazard.

Thus,
children
living
in
British
homes
built
on
soils
containinq
8,000
1­
14
Pb/
q
showed
a
considerably
smaller
elevation
of
PbB
than
was
found
in
the
El
Paso
study
(
Barltrop,
et
al.
1974).
This
may
be
exDlained
by
other
factors,
e.
g.
rainfall
and
soil
composition.
El
Paso,
Texas
is
a
hot,
dry,
windy
town,
whereas
Britain
has
consid­

erable
rainfall,
probably
resulting
in
a
heavv
protective
cover
of
vegetation.

Certain
miscellaneous
sources
of
lead
are
unique
to
children
by
virtue
of
the
pica
habit.
These
include
colored
newsprint
items
to
which
lead­
base
oiq­
(
Joselow
and
Bogden,
1974)
and
other
c­
14
ment
is
applied.
In
a3dition,
pica
is
known
to
occur
in
some
women,

particularly
during
pregnancy.

P~
AQ!!
ACOYIVlFICS
In
characterizinq
the
accumulation
of
lead
in
the
body
under
various
circumstances
of
exposure,
experimental
animal
data
are
useful
for
establishing
relevant
principles.
The
snecific
rates
of
transfer
into,
within,
and
outside
of
the
animal
system
cannot
be
relied
upon
to
reflect,
with
any
reliabilitv,
the
situation
in
man.

Only
human
data
will
serve
to
indicate
how
much
lead,
in
what
form,

and
by
what
route
the
accumulation
of
lead
in
specific
orqans
and
systems
would
occur.
This
restriction
has
imposed
severe
limita­

tions
on
knowledqe
concerning
lead
metabolism
in
man.
Only
certain
human
biological
fluids
and
tissues
are
accessible
for
sa;­
p,
linq,

except
after
death.
The
human
cadaver,
in
turn,
has
its
own
limi­

tations,
chiefly
that
the
precise
history
of
lead
exposure
prior
to
death
is
not
known.
Ante
mortem
studies
of
lead
metabolism
in
human
volunteers,
on
the
other
hand,
have
their
own
limitation.

They
provide
a
substantial
amount
of
knowledoe
concerning
the
sub­

ject,
but
extrapolation
of
the
data
to
the
general
population
is
tenuous.
Population
studies
materially
overcome
this
restriction,

but
at
the
expense
of
precision
and
detail
of
knowledge.
Ry
com­

bining
data
from
all
sources,
a
reasonable
understandina
of
lead
metabolism
does
emerge,
however.
The
ultimate
objective
of
this
section
is
to
relate
contribution
of
source
(
water)
to
total
expo­

sure.
AS
will
be
seen,
this
can
onlv
be
achieved
by
using
incre­

mental
PSS
as
an
index
of
water
exposure
­
the
approach
also
used
by
the
U.
S.
EPA
in
assessing
air
as
a
s@
urce
of
lead
exposure.

c­
15
In
reviewing
the
metabolism
of
lead
in
man,
it
is
generally
assumed
that
all
inorganic
forms
once
absorbed
behave
in
the
same
manner.
There
is
no
evidence
to
suggest
that
this
assumption
is
erroneous.

Absorption
The
classic
studies
of
lead
metabolism
in
man,
conducted
by
Kehoe
(
1961)
indicate
that,
on
the
average
and
with
considerable
day
to
day
excursions,
approximately
8
percent
of
the
normal
diet­

ary
lead
(
including
beverages)
is
absorbed.
This
conclusion
was
reached
as
a
result
of
long­
term
balance
studies
in
volunteers.

Recent
studies
using
the
nonradioactive
tracer
204
Pb
have
con­

firmed
this
conclusion
(
Rabinowitz,
et
al.
1974).
It
is
of
special
significance
that
these
same
workers
found
that
absorption
of
doses
of
lead
nitrate,
lead
cysteine,
and
lead
sulfide
eaten
after
a
6­

hour
fast
and
followed
by
another
6­
hour
fast
was
up
to
8­
fold
higher
than
when
the
lead
was
taken
with
meals
(
Netherill,
et
al.

1974).
This
finding
has
been
confirmed
in
mice
using
small
doses
of
lead
(
3
ug/
kg)
but
not
when
using
large
doses
(
2,000
pg/
kg)

(
Garber
and
Wei,
1974).
Thus,
lead
in
water
and
other
beverages
taken
between
meals
may
have
a
far
greater
impact
on
total
lead
absorption
than
lead
taken
with
meals.

The
gastrointestinal
absorption
of
lead
in
young
children
is
considerably
greater
than
in
adults.
Alexander,
et
al.
(
1973)

found
that
dietary
lead
absorption
was
approximately
50
percent
in
eight
healthy
children
three
months
to
8.5
years
of
age.
This
finding
has
been
confirmed
using
a
larger
number
of
subjects
less
than
2
years
of
age
(
Ziegler,
et
al.
1978).
It
is
worth
noting
too
C­
16
that
the
same
observation
has
been
made
using
infant
rats,
thus
suggesting
a
similarity
in
lead
absorption
characteristics
(
Forbes
and
Reina,
1974;
Kostial,
et
al.
1971).

Numerous
factors
influence
the
absorption
of
lead
from
the
gastrointestinal
tract.
Low
dietary
Ca
and
Fe
and
high
dietary
fat
enhance
lead
absorption
in
experimental
animals
(
Sobel,
et
al,

1938;
Six
and
Goyer,
1970,
1972).
Lead
absorption
has
also
been
shown
to
be
enhanced
in
experimental
animals
by
high
fat,
low
pro­

tein,
and
high
protein
diets,
and
to
be
decreased
by
high
mineral
diets
(
Barltrop
and
Khoo,
1975).
There
also
has
been
shown
to
be
an
inverse
relationship
between
dietary
lead
absorption
and
the
cal­

cium
content
of
the
diet
of
infants
(
Ziegler,
et
al.
1978).
The
chemical
nature
of
the
lead
also
has
an
influence
on
the
degree
of
absorption.
Thus,
Barltrop
and
Meek
(
1975)
reported
that,
in
mature
rats
in
an
acute
experiment,
lead
naphthenate,
lead
octoate,

and
lead
sulfide
were
absorbed
only
two­
thirds
as
well
as
lead
ace­

tate
and
that
elemental
lead
particles,
180
to
250
urn,
were
ab­

sorbed
only
about
14
percent
as
well.
Lead
phthalate
and
lead
car­

bonate
were
absorbed
somewhat
better
than
lead
acetate.
Some
attention
has
also
been
given
to
the
availability
for
absorption
of
lead
in
dried
paint.
The
absorption
of
lead
naphthenate
is
reduced
50
percent
(
in
rats)
as
a
result
of
incorporation
in
paint
films
(
Gage
and
Litchfield,
1969).
Similarly,
it
has
been
found
in
mon­

keys
that
lead
octoate
in
dried
ground
paint
is
not
absorbed
to
the
same
extent
as
lead
octoate
not
incorporated
into
paint
(
Kneip,
et
al.
1974).

c­
17
There
are
serious
problems
in
reaard
to
assessins
the
absorp­

tion
of
lead
via
the
resniratory
tract.
The
fractional
deposition
of
inhaled
aerosols
is
relatively
easy
to
measure,
even
in
man.

The
problem
lies
in
determining
the
fate
of
the
aerosol
oarticles.

TO
varvinq
degrees,
deoendinq
on
their
solubility
and
narticle
size,
these
particles
will
be
absorbed
from
the
resDiratory
tract
into
the
svstemic
circulation,
or
they
will
be
transferred
to
the
gastrointestinal
tract
by
swallowing
following
either
retrograde
movement
up
the
pulmonary
bed
or
by
drainaqe
into
the
bharynx
from
the
nasal
passages.
Unfortunately,
the
particle
size
distribution
and
solubilitv
of
lead
aerosols
varies
tremendously,
depending
on
their
origin
and
residence
time
in
the
air.
All
of
these
diff
i­

culties
have
frustrated
previous
attempts
to
assess
the
impact
of
lead
inhalation
on
the
body
burden
of
lead.
It
has
always
proved
necessary
to
fall
back
on
a
more
indirect
approach
to
the
broblem,

whereby
the
impact
of
air
lead
concentration
on
the
blood
lead
con­

centration
is
measured.
In
order
for
this
aporoach
to
be
meaning­

ful,
certain
conditions
and
restrictions
must
apply.
First,
a
fairly
large
population
of
subjects
is
needed
in
order
to
overcome
the
background
noise
resulting
from
the
variable
imbact
of
dietary
lead
on
the
subject's
PbBs.
Second,
it
is
necessary
to
monitor
the
air
breathed
by
the
subjects
continuously
and
for
a
substantial
period
of
time.
Third,
the
subiects
must
have
been
in
the
air
envi­

ronment
being
evaluated
for
at
least
three
months
in
order
to
assure
reasonable
equilibration
of
air
lead
versus
PbR.
If
all
these
conditions
are
achieved,
the
results
are
only
anplicable
for
the
barticular
tyoe
of
lead
aerosol
under
study.
Thus,
it
would
C­
18
not
be
reasonable
to
extrapolate
data
obtained
in
a
nobulation
breathing
city
air
to
a
population
of
industrial
workers
for
whom
the
greatest
source
of
input
might
be
lead
oxide
fumes.
h?
eedless
to
say,
these
restrictions
are
so
severe
that
very
few
studies
have
been
performed
which
would
allow
one
to
make
a
reasonable
judgment
concerning
the
relative
importance
of
diet
versus
air
as
sources
of
lead
absorption.
An
assessment
of
available
inFormation
is
de­

ferred
to
the
end
of
this
section
on
lead
metabolism.

Dermal
Very
few
studies
concerning
the
dermal
absorotion
of
lead
in
man
or
experimental
animals
are
available.
Once
again,
the
problem
of
the
chemical
form
of
lead
comes
into
play.
In
an
early
study
of
dermal
absorption
of
lead
in
rats,
it
was
found
that
tetraethyl
lead
was
absorbed
to
a
substantially
greater
degree
than
lead
arse­

nate,
lead
oleate,
or
lead
acetate
(
Laug
and
Kunze,
1948).
Differ­

ences
in
the
degree
of
absorption
among
the
oleate,
arsenate,
and
acetate
were
not
signif
icant.
In
a
more
recent
study,
absorption
of
lead
acetate
and
lead
naphthenate
through
the
intact
skin
was
demonstrated,
based
on
concentrations
of
lead
attained
in
various
organs
as
compared
to
controls
(
Rastoqi
and
Clausen,
1976).
Were
seems
to
be
little
question
that
lead
can
be
absorbed
through
the
intact
skin,
at
least
when
applied
in
high
concentrations
such
as
were
used
in
the
Rastogi
study
(
0.24&
l)
.

Distribution
The
general
features
of
lead
distribution
in
the
body
are
well­
known,
both
from
animal
studies
and
from
human
autopsy
data.

Under
circumstances
of
long­
term
exposure,
anproxim.
ate1.
y
95
percent
c­
19
of
the
total
amount
of
lead
in
the
body
(
body
burden)
is
loca.
lized
in
the
skeleton
after
attainment
of
maturity.
By
contrast,
in
children,
only
72
bercent
is
in
bone
(
Rarrv,
1975).
From
animal
studies
it
also
appears
that
the
very
young
retain
lead
to
a
great­

er
extent
than
adults
(
Jugo,
1977).
The
amount
in
bone
increases
with
old
age
but
the
amount
in
most
soft
tissues,
including
the
blood,
attains
a
steady
state
early
in
adulthood
(
Barry,
1975;

Horiuchi
and
Takada,
1954).
Special
note
should
be
made
regardinq
the
kinetics
of
lead
distribution
with
reference
to
the
blood.

When
human
volunteers
are
introduced
into
a
new
air
environment
containing
substantially
higher
concentration
of
lead
than
the
bre­

vious
one,
the
concentration
of
lead
in
the
blood
rises
rapidly
and
attains
a
new
apparent
steady
state
in
about
60
to
100
days
(
Tola,

et
al.
1973;
Rabinowitz,
et
al.
1974;
Griffin,
et
al.
1975).
This
is
probably
only
an
apparent
steady
state
rather
than
a
true
one
because
the
kinetics
of
disappearance
of
lead
from
the
blood
differ
depending
upon
whether
the
high
level.
was
maintained
for
months
or
for
years.
When
men
were
placed
in
a
high
lead
environment
for
100
days
and
then
returned
to
a
low
lead
environment,
the
PbB
concen­

tration
returned
to
the
pre­
exposure
level
with
a
disapbearance
half­
time
of
only
about
six
weeks.
9y
contrast,
the
rate
of
PbB
decrement
in
workers
who
retire
from
the
lead
trades
is
much
longer
(
Haeqer­
Aronsen,
et
al.
1974;
prerovska
and
Teisinger,
1970).
This
suggests
that
true
equilibrium
between
the
blood
compartment
and
bone
compartment
is
only
slowly
attained
under
constant
state
exDo­

sure
conditions.

C­
20
The
distribution
of
lead
at
the
organ
and
cellular
1evel.
s
has
been
studied
extensively.
In
blood,
lead
is
primarily
localized
in
the
erythrocytes.
The
ratio
of
the
concentration
of
lead
in
the
cell
to
lead
in
the
plasma
is
approximately
16:
l.
Lead
crosses
the
placenta
readily.
The
concentration
of
lead
in
the
blood
of
the
newborn
is
quite
similar
to
the
maternal
blood
concentration.
The
approximate
ratio
of
fetal
to
maternal
Pb'
3
is
somewhat
greater
than
one
(
Clark,
1977;
Schaller,
et
al.
1976).
Studies
of
the
subcellu­

lar
distribution
of
lead
indicate
that
distribution
occurs
to
all
organelles,
suggesting
that
all
cellular
functions
at
least
have
the
opportunity
to
interact
with
lead.

Yetabolism
Upon
entry
into
the
body,
lead
compounds
occurring
in
the
environment
dissociate.
Therefore,
no
question
of
metabolism
of
the
pollutant
is
involved.
The
one
exception
is
the
family
OS
alkyl
lead
compounds,
principally
tetramethyl
lead
and
tetraethyl
lead.
These
are
dealkylated
to
form
trialkyl
and
dialkvl
metabo­

lites,
which
are
more
toxic
than
the
tetraalkyl
forms
(
qolanowska,

et
al.
1967).

Excretion
The
numerous
studies
reported
in
the
literature
concerninq
routes
of
excretion
in
experimental
animals
indicate
wide
interspe­

ties
differences.
In
most
species,
except
the
baboon,
biliary
excretion
predominates
over
urinary
excretion
(
Cohen,
1970).
It
also
appears
that
urinary
excretion
predominates
in
man
(
Pabino­

wits,
et
al.
1973).
This
conclusion,
however,
is
based
on
data
from
one
volunteer.

c­
21
Contributions
of
Lead
from
Diet
versus
Air
to
PbR
Great
concern
has
developed
in
recent
years
regarding
the
impact
of
air
lead
exposure
on
human
health
in
the
general
Dooula­

tion.
Analysis
of
the
contribution
of
ambient
air
to
lead
intake
by
man
has
taken
the
form
of
an
analysis
of
air
lead
versus
PbR
for
reasons
explained
in
the
section
on
lead
absorotion.
An
analysis
of
all
available
data
bearing
on
this
question
first
appeared
in
the
Environmental
Health
Criteria
3
Lead
oublished
by
WHO
(
1977).

A
more
rigorous
and
detailed
analysis
was
published
subsequently
in
Air
Quality
Criteria
for
Lead
(
U.
S.
EPA,
1977).

Most
of
the
data
bearing
on
the
question
of
air
lead
versus
PbB
are
deficient
in
one
of
two
major
respects.
The
most
serious
and
frequent
deficiency
is
the
lack
of
continuous
air
sampling
in
the
breathing
zone
of
the
subjects.
An
almost
equally
serious
but
less
frequent
deficiency
is
the
lack
of
variation
in
the
air
lead
concentration
over
the
range
of
interest.
This
is,
unfortunately,

a
problem
seen
mainly
in
the
clinical
studies
(
as
onposed
to
ooou­

lation
studies)
where
the
number
of
subjects
is
quite
limited.

Another
problem,
also
limited
to
the
clinical
studies,
is
the
arti­

ficial
nature
of
the
lead
aerosol
utilized.
In
spite
of
all
these
apparent
limitations,
calculations
from
the
epidemiologic
and
labo­

ratory
data
sources
indicate
a
fairly
narrow
range
of
blood
Db
to
air
Pb
ratios,
namely
1
to
4
ug/
dl
for
every
microgram
of
air
lead
per
cubic
meter
(
ug/
m3).
This
blood
Pb
to
air
Pb
ratio
appears
to
be
higher
for
children
than
adults
(
Table
2).

Among
all
the
studies,
the
only
one
that
satisfied
all
cri­

teria
for
design
was
the
one
by
Azar,
et
al.
(
1975).
It
should
be
c­
22
5
A­
­­­
e
­
A
mc~­~~
cooc~
cc
ac­+
wawar­
r­
N­
m
.
.
.
.
.
.
.
.
.
.
.
.
.
C'OONOCON4­
4N
­­­­­­­­
­­

Fe
IncD
.
.
NO
2
.
4
3
noted
that
the
regression
eauation
develooed
to
describe
the
data
(
log
PbB
=
1.2557
+
0.153
(
log
ug
Pb/
m3))
has
a
slope
of
less
than
one.
Thus,
the
incremental
rise
in
?
bS
for
each
1
ug
oh/
m3
in
air
becomes
progressively
smaller.
This
relationship
is
consonant
with
experimental
animal
data
showing
that
over
a
wide
ranqe
of
dietary
lead
levels
the
incremental
rise
in
PbB
decreases
progressively
proportional
to
the
rise
in
dietary
lead
levels
(
Prpic­
Yaiic,
et
al.
1973;
Azar,
et
al.
1973).
It
also
is
consonant
with
the
World
Yealth
Organization
analysis
of
data
on
air
lead
exposure
in
a
bat­

tery
plant
(
WHO,
1977).

The
Azar
data
have
been
analyzed
as
to
dose
response
by
the
U.
S.
EPA
(
1977)
and
are
presented
in
Table
3.

So
far
as
the
contribution
of
other
sources
of
lead
to
PbB
is
concerned,
a
quantitive
analysis
such
as
has
been
done
for
air
lead
is
simply
not
possible
using
the
data
currently
available.
An
estimate
of
the
total
dietary
contribution
to
PbB
was
attempted
by
WY0
(
1977)
recently
(
Table
4).

So
far
as
the
soecific
contribution
of
water
is
concerned,

information
is
even
more
scarce
than
for
total
diet.
Estimates
of
the
contribution
of
lead
in
water
to
PbB
have
been
reported
in
four
separate
studies.
The
first
of
these
was
bublished
in
1976
(
El­

wood,
et
al.
1976).
A
linear
regression
was
calculated
for
PbB
and
water
lead
using
"
first
run"
morning
tap
water
in
129
houses
in
northwest
Wales.
Blood
lead
concentrations
were
determined
for
an
adult
female
resident
in
each
house.
The
regression
drawn
was
as
follows:

PbB
(
pg/
dl)
=
19.6
+
7.2
(
mg
Pb/
l
water)

C­
24
TABLE
3
Estimated
Percentage
of
Powlation
Exceeding
a
Specific
Blood
Lead
Level
in
Relation
to
Ambient
Air
Lead
Exposure
Percent
Exceeding
Blood
Lead
Level
of:

Air
Lead,

w/
m3
20.0
30.0
w/
d1
ug/
dl
40.0
w/
d1
0.5
15.22
0.59
0.02
1.0
26.20
1.67
0.07
1.5
34.12
2.88
0.16
2.0
40.23
4.12
0.26
2.5
45.15
5.35
0.38
3.0
49.23
6.57
0.51
3.5
52.69
7.75
0.66
4.0
55.67
8.90
0.81
4.5
58.27
10.01
0.97
5.0
60.57
11.09
1.14
6.0
64.45
13.16
1.48
7.0
67.63
15.10
1.83
8.0
70.28
16.92
2.20
C­
25
TABLE
4
Comparison
of
Daily
Oral
Lead
Intake
With
PbB
Levelsa
Study
Design
Oral
Intake
PbBb
1
w/
day)
h9/
100
ml)
PhB
per
100
u9
oral
Pb
Reference
Fecal
excretion
llgc
(
women)
15.3
13.0
Tepper
and
Levin
(
1972)

Duplicate
portion
230
(
men)
12.3
5.4
Nordman
(
1975)

Duplicate
portion
180
(
women)
7.9
4.4
Nordman
(
1975)

Composites
technique
505
(
men)
34.6
6.8
2ur3.
o
and
Griffini,
1973
d
aSource:
WHO,
1977
b
Contributions
of
air
to
PbB
levels
are
not
reported
in
most
of
these
studies
and
could
not
be
subtracted
from
total
PbB
levels.

'
Calculated
from
daily
faecal
excretion
of
108
pg
of
lead
assuming
gastrointestinal
absorption
10
percent
a
Pb­
B
levels
from
Secchi,
et
al.
1971
C­
26
The
regression
selection
seems
inaoprobriate
from
examination
of
the
scattergram
(
Figure
1).
A
curvilinear
model
would
have
been
more
appropriate
or
at
least
should
have
been
tested,
particularly
since
the
authors'
linear
model
extrapolates
to
PbB
19.6
ug/
dl,
a
rather
high
baseline
value
for
non­
occupationallv
exoosed
women.

Moore,
et
al.
(
1977a)
reported
a
very
similar
study
in
which
the
interaction
of
PbB
with
lead
in
both
"
first
flush"
water
and
running
water
was
determined
(
Moore,
et
al.
1977a).
The
study
was
conducted
in
Glasgow,
Scotland,
where
the
water
is
extremely
soft,

As
in
the
Elwood
study,
blood
was
drawn
from
adult
females
of
the
household.

The
Moore,
et
al.
(
1977a)
study
demonstrated
that
there
is
a
curvilinear
relationship
between
PbB
and
the
concentration
of
lead
in
"
first
flush"
water
(
Figure
2).
The
equation
for
the
regression
line
was
x
=
0.533
+
0.675
y,
with
both
values
being
expressed
as
umol/
l.
Blood
lead
rose
as
the
cube
root
of
"
first
flush"
water.

Actually,
there
is
an
error
in
the
equation.
The
term
x
really
is
PbB
and
y
is
the
cube
root
of
the
"
first
flush"
water.
The
authors
point
out
that
the
lead
concentration
in
running
water
probably
reflects
the
impact
of
drinking
water
on
?
bB
better
than
"
first
flush"
water.
They
found
that
the
same
relationshio
held,
wherein
mean
blood
lead
rose
in
DroDortion
to
the
cube
root
of
runninq
water
lead.
The
correlation
of
running
water
lead
to
PbB
was
even
somewhat
better
than
that
of
"
first
flush"
water
to
PbP
(
I:
=
0.57
vs.
0.52).
According
to
the
authors,
running
water
lead
concentra­

tions
were
approximately
one­
third
the
"
first
flush"
lead
concen­

trations.
These
data
are
useful
in
that
they
provide
an
estimate
C­
27
.
l
*
l
.

L
I
I
_
.
.
..­
­
0
0
IO
020
030
..___.___
010
050
080
010
­
0
80
080
'
00
WbW
MO
0)`)

RI9rrrrion
01
blood­
lstd
on
mornin
watmr
loud
in
CmOrnmrlon~
hir@.

FIGURE
1
Regression
of
Blood­
lead
on
Morning
Water
Lead
in
Caernarfonshire
Source:
Elwood,
et
al.
1976
C­
28
BLOOD
l
po
LEAD
l/
l)

I
Intetv~
ls
of
Yafrt
Lead
­
up
ro
9
72
No.
of
Srmplcs
in
101
104
101
Interval
4
_
_
I
1
1
t
I
I
I
1
.24
.40
1
I.
44
2
3
4
5
6
First
Draw
Water
Lead
i+
mol/
fl
FIGURE
2
Mean
Blood­
lead
Values
for
Nine
Grouus
at
Intervals
of
First­
flush
Water
Lead
Source:
Moore,
et
al.
1977a
c­
29
of
the
consequences
of
changing
the
concentration
of
lead
in
water
from
one
value
to
another.
The
example
provided
is
the
PbB
conse­

quence
of
going
from
a
"
first
flush"
concentration
of
0.24
pmol/
l
(
50
pg/
l)
to
0.48
wok'
1
(
100
w/
l).
Such
a
change
results
in
an
incremental
rise
in
PbB
of
0.11
umol/
l,
or
of
2.3
ug/
dl.
(
3n
a
run­

ning
water
basis,
the
PbB
change
would
occur
going
from
24/
3
or
8
vg/
l
to
48/
3
or
16
ug/
l.
Using
the
authors'
equation,
the
effect
on
PbB
of
lead
in
running
water
can
be
estimated
(
Table
5).
If
this
relationship
is
correct,
the
impact
of
water
lead
on
PbB
is
ex­

tremely
great
in
the
lower
ranges
of
water
lead
but
diminishes
rap­

idly
in
the
higher
range
of
water
lead
(
SO
to
100
ug/
l).

Hubermont,
et
al.
(
1978)
also
reports
the
interaction
of
morn­

ing
tap
water
lead
to
PbB
in
pregnant
women
of
the
household.

Again,
as
in
the
study
of
Moore,
et
al.
(
1977a)
a
curvilinear
rela­

tionship
is
described
for
the
interaction
of
PbB
with
water
lead:

PbB
=
9.62
+
1.74
log
morning
water
Pb,
fug/
l).

The
correlation
was
good
(
K
=
+
0.37:
p
=
0.001).
The
calculated
impact
of
water
Pb
on
PbB
using
this
equation
is
considerably
less
in
the
lower
range
of
water
lead
than
in
the
Moore,
et
al.
(
1977a)

study.
The
data
may
not
be
strictly
comoarable
concerning
water
sampling
procedure.

One
additional
set
of
data
is
available
which
bears
on
the
question
of
the
impact
of
the
concentration
of
lead
in
water
on
PbB.
A
study
was
conducted
by
the
U.
S.
EPA
concerning
the
rela­

tionship
of
lead
in
drinking
water
to
PbB
(
Greathouse
and
Craun,

1976).
Both
early
morning
and
running
water
samples
were
analyzed
for
lead
in
a
soft
water
area
(
Boston,
Yassachusetts),
In
addi­

c­
30
TABLE
5
Effect
of
Running
Water
Lead
on
PbB*

Pb
in
Levels
(
y3)
Pb
Levels
An
(
umol/
l)
(
y3)
Running
Water
(
ucl/
l)
Total
PbB
PbB
due
to
Water
0
0
11.03
0
0.0145
1
14.44
3.41
0.0725
5
16.86
5.83
0.1449
10
18.37
7.34
0.3623
25
20.99
9.96
0.7246
50
23.58
12.55
1.4493
100
26.84
15.81
ay
=
(
J­
ICI
of
Pb/
l
of
running
water)
x
3
207
*
Source:
Voore,
et
al.
1977a
c­
31
tion,
blood
samples
for
members
of
the
househoLd
were
analyzed
for
lead.
These
subjects
included
both
children
and
adults.
Numerous
variables
that
might
have
influenced
?
bR
were
measured,
including
age,
sex,
traffic
density,
lead
in
dust,
and
socio­
economic
status.

The
data
for
interaction
of
PhS
and
water
Pb
were
re­
evaluated
bv
Dr.
Greathouse
specifically
for
the
purpose
of
comparison
to
the
analyses
of
Moore,
et
al.
(
1977a)
and
Hubermont,
et
al.
(
1978).

This
was
done
subsequent
to
publication
of
the
1976
Greathouse
and
Craun
report.
Statistical
analyses
were
performed
using
both
the
Hubermont
model
(
PbB
=
a
+
b
log
Pb
in
water)
and
the
Voore
model
(
PbB
=
a
+
b3
Pb
water).
These
models
were
tested
using
(
1)
all
subjects
aged
20
or
more,
and
(
2)
women
20
to
50.
The
models
were
also
tested
using
running
water
data
and
early
morning
water
data.

Interestingly,
the
relationship
of
early
morning
water
Pb
to
run­

ning
water
Pb
was
almost
identical
to
the
3~
1
relationshio
reported
by
Moore,
et
al.
(
19773).
More
precisely,
the
relationship
was:

Early
morning
water
Pb
=
­
0.028
+
3.081
running
water
Pb
r*
=
0.235;
?
=
0.0001
The
cube
root
model
of
Yoore,
et
al.
(
1977a)
was
more
aopro­

priate
than
the
log
water
Pb
model
of
Hubermont,
et
al.
(
1978),
and
the
correlation
of
PbB
with
running
water
Pb
was
better
than
with
morning
water
Pb.
The
correspondence
between
data
from
all
sub­

jects
20
years
of
age
and
over
and
for
women
aqe
20
to
5r)
was
striking:

Females
20
to
50,
n
=
249
PbB
=
13.38
+
2.487
3Jrunning
water,
Pb,
ug/
l
0
=
0.020
C­
32
All
subjects
20
yrs
+,
n
=
390
PbB
=
14.33
+
2.541
P
=
0.0065
At
this
point
it
is
useful
to
compare
the
data
from
the
three
studies
discussed
above.
These
data
constitute
the
sole
firm
foun­

dation
for
assessing
the
impact
of
lead
in
water
on
the
internal
dose
of
lead
as
reflected
in
PbB.
The
comparison
is
presented
in
Table
6.
Calculations
are
made
as
to
the
PbB
due
to
water
over
a
range
of
1
to
100
pg
Pb/
l.
The
comparison
is
made
on
the
basis
of
running
water
Pb
in
spite
of
the
fact
that
the
equations
for
the
two
European
studies
were
developed
on
the
basis
of
"
first
flush"
or
"
early
morning"
water.
This
adjustment
seems
justified
since
the
ratio
of
these
values
to
running
water
values
has
been
affirmed
to
be
3:
l
in
two
of
the
three
studies
and
therefore
probably
is
ap­

proximately
correct
for
the
third
study,
the
one
by
Yubermont,
et
al.
(
1978).
It
is
seen
that
the
impact
of
lead
in
water
on
PbB
is
quite
different
among
the
three
studies.
Since
there
is
no
basis
for
rejecting
any
of
the
three
studies,
an
estimate
of
the
averaqe
situation
is
made
from
an
average
of
the
three
sets
of
data.
The
reasons
for
any
variation
in
the
relationships
can
only
be
left
to
speculation.
Certainly
the
calcium,
phosohate,
and
iron
concentra­

tions
of
the
waters
in
the
three
studies
were
different
and
mav,
to
some
extent
at
least,
account
for
the
differences
in
the
impact
of
lead
in
water
on
PbB.

It
is
known
that
calcium
profoundly
depresses
lead
absorDtion,

even
over
a
relatively
narrow
range.
For
example,
Ziegler,
et
al.

(
1978)
demonstrated
that
a
mere
doubling
of
the
dietary
calcium
C­
33
level
profoundly
depressed
lead
absorotion
in
infants.
Also,
ani­

mal
studies
have
shown
that
nutritional
iron
deficiency
enhances
lead
absorption.
Attention
should
be
qiven
to
the
siqnificance
of
the
variations
in
calcium
and
iron
content
of
water
aqainst
the
background
variations
of
calcium
and
iron
in
nonaqueous
portions
of
the
diet.
As
with
calcium,
high
phosphate
levels
also
tend
to
de­

press
lead
absorption.

EFFECTS
The
effects
of
lead
on
man
will
be
reviewed
in
a
selective
fashion.
Greatest
emphasis
will
be
Placed
on
those
effects
which
occur
at
the
lower
levels
of
exDosure
and
those
which
are
qrooerly
viewed
with
the
most
concern,
namely
neurobehavioral
effects,
car­

cinogenesis,
mutagenesis,
and
teratoqenesis.
Because
of
the
Dauci­

ty
of
data
in
man
and
the
seriousness
of
the
effect,
some
sections
will
be
specifically
subdivided
into
sections
dealing
with
human
data
and
animal
data.
In
other
cases,
that
does
not
seem
necessary
because
of
the
wealth
of
human
data
available.

There
is
vast
literature
concerning
the
effects
of
lead
on
the
formation
of
hemoglobin
and
more
limited
literature
on
the
related
effects
on
other
hemo­
proteins.
From
the
standpoint
of
standard
setting,
the
effects
of
lead
on
this
system
are
particularly
imcor­

tant
since
current
knowledge
suggests
that
the
hematopoietic
system
is
the
"
critical
organ."
That
is
to
say
that
effects
are
detect­

able
at
lower
levels
of
lead
exposure
than
is
the
case
with
any
other
organ
or
System.
The
mechanism
whereby
lead
reduces
the
cir­

culating
concentration
of
hemoglobin
is
not
thoroughlv
understood.

Many
specific
abnormalities
exist,
some
occurring
at
lower
PbBs
c­
35
than
others.
The
life
span
of
ervthrocytes
is
shortened
in
heavy
lead
exposure
(
PbB
=
59
to
162)
(
Hernberq,
et
al.
1967).
The
mech­

anism
is
not
well
understood,
but
damage
to
the
erythrocyte
mem­

brane
is
likely.
Dose­
response
and
dose­
effect
relationships
have
not
been
established.
It
seems
unlikely,
however,
that
shortened
cell
life
results
in
lead­
induced
reduction
in
circulating
hemo­

globin.
Rather,
it
is
more
likely
that
the
synthesis
of
hemoglobin
is
the
critical
mechanism.

Although
there
is
evidence
that
lead
interferes
with
globin
synthesis
as
well
as
heme
synthesis,
this
effect
seems
to
occur
only
secondarily
to
a
deficit
in
heme
production
(
Piddjngton
and
White,
1974).
Thus,
it
is
the
action
of
lead
on
heme
synthesis
that
appears
most
critical.
This
action
is
complex
and
involves
several
enzymes
in
the
synthesis
of
heme
(
Figure
3).

Clear
evidence
exists
that
lead
inhibits
both
d­
aminolevulinic
acid
dehydrase
(
ALAD)
and
heme
synthetase
both
in
vitro
and
in
vivo
­­
­­

at
relatively
low
levels
of
lead
exnosure.
Elevation
of
the
con­

centration
of
the
substrates
for
these
two
enzymes
in
plasma
and
urine
(
ALA)
and
in
erythrocytes
(
PROTO)
increases
as
PbR
increases.

As
a
matter
of
fact,
rise
in
PROTO
and
ALA
occur
at
PbBs
somewhat
below
those
associated
with
a
decrement
of
hemoqlobin.
Thus,
in
adults,
a
decrement
in
hemoglobin
first
appears
at
PbR
=
50
(
Tola,

et
al.
1973)
and
at
PbB
=
40
in
children
(
Betts,
et
al.
1973;

Pueschel,
et
al.
1972),
whereas
a
distinct
elevation
in
ALA
in
the
urine
(
ALAU)
first
appears
at
PbB
=
40
in
men
(
Selander
and
Cramer,

1970;
Haeger­
Aronsen,
et
al.
1974)
and
children
(
WAS,
1972)
and
somewhat
lower
in
women
(
Roels,
et
al.
1975).
Rises
in
PROW
first
C­
36
(
Mitochondrion)

Succmyl
­
CoA
7
Heme
+
A­&

Glyane
Ferrocdelatose
I
I
ALA
Syntherase
(
ALAS)
Fe%

l<
,
+

8
Amir!
olevulinic
Protoporphyrin
TX
Acid
(
ALA)

1
ALA
Dehyd!
ose
(
ALAI21
~
Cytoplosm~
I
Coprophyrinogen
i3
Pbe
1
Porphobliinogen
I
s
U;
oporphyrinogen
Iii
Pb
/
­
Fe
t
\
FfZ"

FIGURE
3
Effects
of
Lead
on
Heme
Yetabolism
c­
37
appear
at
PbR
=
15
to
30
in
women
and
children
and
at
?
bS
=
25
in
men
(
Sassa,
et
al.
1973;
Roels,
et
al.
1975).
The
most
reasonable
explanation
for
the
rise
in
PRO?
9
at
levels
of
lead
exoosure
below
the
threshold
for
hemoglobin
decrement
is
that
the
primary
event
is
inhibition
of
the
insertion
of
iron
into
PROTO
IX,
whether
it
is
caused
by
inhibition
of
heme
synthetase
or
by
inhibited
entry
of
Fe
into
the
mitochondrion
(
Jandl,
et
al.
1959).
Reqardless
of
that
uncertainty,
the
effect
is
the
same,
a
potential
decrement
in
hemo­

globin,
which
leads
to
feedback
depression
of
ALAS
resultinq
in
a
compensatory
increase
in
the
production
of
ALA
and
other
heme
pre­

cursors.
The
evidence
for
this
compensatory
adjustment
is
to
be
found
both
in
laboratory
animal
studies
(
Strand,
et
al.
1972;

Suketa,
et
al.
1975)
and
in
studies
of
people
with
elevated
lead
exposure
(
Berk,
et
al.
1970;
Meredith,
et
al.
1977).
The
approxi­

mate
threshold
for
ALAD
inhibition
is
Pb9
=
10
to
20
for
adults
(
Tola,
1973)
and
PbB
=
15
in
children
(
Granick,
et
al.
1973).

Roughly
equivalent
inhibition
occurs
concurrently
in
the
liver
of
man
(
Secchi,
et
al.
1974)
and
in
the
liver
and
brain
of
rats
(
Mil­

lar,
et
al.
1970).
The
toxicological
implications
of
ALAD
inhibi­

tion
have
not
been
studied
extensively.
However,
substantial
lead­

induced
depression
of
blood
ALAT)
activity
in
dogs
does
not
reduce
the
blood­
regenerating
response
to
acute
hemorrhaging
in
dogs
(
Yax­

field,
et.
al.
1972).

A
few
studies
have
been
reported
concerning
effects
of
lead
on
hemoproteins
other
than
hemoglobin.
Thus,
the
rate
of
cytochrome
P450­
mediated
drug
metabolism
has
been
found
to
be
deoressed
in
two
cases
of
lead
poisoning
(
PbB
=
60
and
72)
but
not
in
10
cases
where
c3­
38
lead
exposure
ranged
from
PbB
=
20
to
60
(
Alvares,
et
al..
1975).

Cytochrome
content
of
kidney
mitochondria
has
also
been
reported
to
be
depressed
in
rats
(
Rhyne
and
Gayer,
1971).

The
question
arises
as
to
whether
certain
populations
may
be
predisposed
to
the
toxic
effects
of
lead
as
a
result
of
G­
6­
PD
deficiency
or
iron
deficiency.
G­
6­
PD
deficiency
is
known
to
be
associated
with
increased
susceptibility
of
erythrocytes
to
hemoly­

sis.
The
possibility
of
increased
susceptibility
of
G­
6­
PD­
defi­

cient
children
to
the
hematopoietic
toxicity
of
lead
has
not
been
reported.
In
regard
to
possible
enhancement
of
hemoglobin
defi­

ciency
by
coexistent
iron
deficiency,
the
one
study
reported
to
date
was
negative.
There
was
no
significant
difference
in
the
blood
hemoglobin
or
hematocrit
among
29
iron­
deficient
children
with
PbB
20
pg/
dl
as
compared
to
17
iron­
deficient
children
with
PbB
=
20
to
40
ug/
dl
(
Angle,
et
al.
1975).

Dose­
response
relationships
for
the
effect
of
lead
on
various
parameters
of
hematological
indices
have
been
developed
recently
(
Zielhuis,
1975).
These
are
reproduced
in
tabular
form
in
Table
7.

In
considering
these
data,
it
is
obvious
that
FEP
(
essentially
PROTO)
elevation
is
a
more
sensitive
correlate
of
lead
exposure
than
ALAU.
It
should
also
be
noted,
however,
that
an
increase
in
FEP
above
normal
also
occurs
in
iron
deficiency
anemia.
Thus,
the
data
must
be
considered
in
that
light.
In
a
recent
study
of
FEF
in
lead­
exposed
and
non­
lead­
exposed
children,
Roels,
et
al.
(
19781
were
able
to
study
the
interaction
of
FE!?
and
PbB
in
the
absence
of
anemia
as
indicated
by
serum
iron
concentration.
They
proposed
a
maximum
acceptable
limit
for
FEP
at
PbR
=
25
uq/
dl.
The
maximum
c­
39
00000
rnWLnWP
I
I
I
I
I
Odt­
lrlrlrl
Tvr.
lm­=
Fmw
­
rmo
0
mm
0
r+
OOrn.­
lPO
l­
lmln
oo~
m~
w
­
lmr­
co
000000
amv;
PVIwr+
acceptable
point
was
the
mean
FFP
plus
two
standard
deviations
for
rural
children,
which
equalled
79.2
pg
FEP/
dl
erythrocytes.
The
PbB
of
these
children
was
9.1
ug/
dl
+
0.5
with
serum
iron
)
50
­

~
g/
lOOml.
This
maximum
is
very
similar
to
the
maximum
acceotable
FEP
which
would
be
calculated
at
mean
FEP
~
1~
s
two
standard
devia­

tions
(
PbB
=
26
pg/
dl)
cited
in
the
recent
"
Air
Quality
for
Lead"

(
U.
S.
EPA,
1977).
As
was
indicated
earlier,
the
cooperative
effect
of
iron
deficiency
and
lead
exposure
on
FEP
has
not
as
yet
been
ade­

quately
defined.
There
is
just
the
one
study
by
Angle,
et
al.

(
1975)
I
suggesting
no
interaction
at
PbB
=
20
to
40.

The
syndrome
of
lead
encephalopathy
has
been
recoqnized
since
the
time
of
Hippocrates
as
occurring
in
workers
in
the
lead
trades.

The
major
features
were
dullness,
irritabilitv,
ataxia,
headaches,

loss
of
memory
and
restlessness.
These
symptoms
often
progressed
to
delirium,
mania,
coma,
convulsions,
and
even
death.
The
same
general
effects
were
also
described
in
infants
and
young
children.

Encephalopathy
due
to
lead
was
probably
more
frequently
fatal.
in
children
than
in
adults
because
lead
exposure
was
usually
not
sus­

pected
and
because
children
do
not
communicate
siqns
and
symDtoms
as
readily
as
adults.
The
mortality
rate
among
children
has
been
variously
reported
as
being
from
5
to
40
percent.

The
literature
concerning
the
neurological
features
and
the
probable
dose
of
lead
involved
is
far
more
specific
for
children
than
for
adults.
This
is
probably
because
the
problem
persisted
longer
and
hence
benefited
more
from
the
accumulated
sophistication
of
disease
investigation.
Apart
from
the
mortality
statistics,

there
was
a
considerable
toll
recorded
among
survivors
in
the
form
c­
41
of
'
Lonq­
term
neurological
sequel.
ae.
Cortical
atrophy,
convulsive
seizures,
and
mental
retardation
were
commonly
reported
(?
erlstein
and
Attala,
1966;
Ryers
and
Lord,
1943).

The
minimal
level
of
lead
exposure
resulting
in
lead
encepha­

lopathy
is
not
clearly
known
and
oerhaos
never
will
be
in
liqht
of
the
dramatic
decrease
in
the
incidence
of
the
disease,
particularly
durinq
the
last
10
to
15
years.
Drawinq
mainly
from
his
own
experi­

ences,
Chisolm
(
1968)
has
estimated
the
minimal
PbB
associated
with
encephalopathy
as
being
80
uq/
dl.
There
are
occasional
reports
however
of
occurrence
of
encephalopathy
at
PbBs
below
80
ug/
dl
(
Smith,
et
al.
1938;
Gant,
1938).
Although
80
uq/
dl
may
be
a
rea­

sonable
estimate
of
threshold
for
encephalopathy
in
children,
the
usual
values
are
much
higher,
with
a
mean
of
appr0ximatel.
y
328
according
to
one
source
(
NAS,
1972).

It
has
been
reasoned
that
if
lead
exposure
as
specified
above
can
have
such
severe
deleterious
effects
on
the
central
nervous
system,
lower
levels
of
exposure
might
well
result
in
more
subtle
effects.
Specifically,
the
concern
has
been
over
whether
such
effects
occur
in
children
whose
PbRs
are
in
the
40
to
80
us/
da
range.
Given
the
difficulties
of
study
design,
it
is
hardly
sur­

prisln,
q
that
all
of
the
relevant
studies
are
ooen
to
criticism.

The
most
common
deficiencies
encountered
are
overlap
of
lead
expo­

sure
in
the
study
groups
(
Pb
versus
control),
inadequate
matching
for
socio­
economic
status
and
other
variable,
insensitivity
of
the
behavioral
tests,
and
uoor
knowledge
of
the
deqree
of
lead
exoo­

sure.
Zn
regard
to
this
last­
named
problem,
the
index
of
exposure
has
Iusually
been
PbBs
determined
at
the
time
of
behavioral
testinq.

c­
42
In
some
instances
record
of
one
earlier
PbS
determination
w&
s
available.
In
spite
of
these
problems,
when
the
various
studies
are
taken
together,
subtle
neurobehavioral
effects
I?
o
aooear
to
occur
as
a
result
of
exposure
in
the
range
of
PbB
=
40
to
80
ug/
dl.

Two
general
approaches
have
been
used
in
attackin
the
prob­

lem.
The
most
common
approach
has
been
to
evaluate
two
populations
of
children
closely
matched
as
to
age,
sex,
and
socio­
economic
status,
but
differing
as
to
lead
exposure.
These
studies
are
retrospective
and
usually
strictlv
cross­
sectional.
In
onlv
3ne
instance
was
a
follow­
up
repeat
study
of
the
population
oerformed
(
de
la
Burde
and
Choate,
1972,
1975).
The
other
qenera?
approach
has
been
to
identify
children
with
neurobehavioral
deficits
of
un­

known
etiology
and
to
establish
whether
their
lead
exposure
was
excessive
in
comparison
to
appropriate
control
children.
Aside
from
the
usual
specific
flaws
in
experimental
design,
tAere
has
been
the
additional
question
as
to
which
came
first,
the
excessive
lead
exposure
or
the
neurobehavioral
deficit.
Amonq
mentally
sui­

normal
children
whose
problems
were
clearly
attributable
to
etiolo­

gies
other
than
lead,
pica
incidence
and
PbBs
were
both
el.
eva­,
er?

(
Bicknell,
et
al.
1968).

Among
studies
of
the
first
type,
those
of
de
la
Rurde
and
Choate
(
1975)
are
illustrative
of
the
problems
that
exist
in
this
area
of
toxicology.
Fine
motor
dysfunction,
impaired
concept
for­

mation,
and
altered
behavior
profile
were
observed
in
70
preschool
children
exhibiting
pica
and
elevated
PbBs,
all
of
which
were
>
3r!

ug/
dl
l
The
mean
level
was
59
ug/
dl.
The
children
were
examined
at
fuclr
years
and
again
at
seven
years
of
aqe.
Roth
t+
e
lead­
exposed
c­
43
group
and
the
control
group
had
been
followed
from
infancy
through
eight
years
of
age
as
part
of
a
Collaborative
Study
of
Cerebral
Palsy,
Yenta1
Retardation,
and
Neurologic
Disorders
of
Infancy
and
Childhood.
Unfortunately,
the
control
group
did
not
have
blood
lead
analyses
performed.
Yowever,
tooth
lead
and
urinary
copropor­

phyrin
determinations
were
ultimately
performed.
Another
problem
was
the
inference
that
positive
radiographic
findinqs
of
lead
in
long
bones
and/
or
intestines
were
found
in
subjects
with
PbBs
in
the
range
of
30
to
40
ug/
dl.
Lead
lines
in
bones
at
this
level
of
exposure
are
extremely
unlikely
(
Betts,
et
al.
1973),
suggesting
either
that
the
blood
lead
determinations
were
spuriously
low
or
that
they
had
actually
been
higher
at
times
which
did
not
coincide
with
the
time
of
sampling.
Thus,
it
would
seem
that
the
minimal
?
bB
associated
with
neurobehavioral
effects
may
well
have
been
more
on
the
order
of
50
to
60
ug/
dl
rather
than
30
to
40
vg/
dl.
Overall,

the
experimental
design
was
otherwise
generally
sound.

Another
often­
cited
study
by
Perino
and
Ernhart
(
1974)
was
basically
of
the
same
general
design
as
the
one
reported
by
de
la
Burde
and
Choate
(
1972,
1975).
It
concluded
that
neurobehavioral
deficits
occurred
at
PbRs
as
low
as
40
ug/
dl.
The
flaw
in
this
study
was
that
the
parents
in
the
control
qrouo
were
better
educat­

ed
than
those
of
the
lead­
exposed
children.
Differences
found
may
have
been
due
to
the
fact
that
more
hiqhlv
educated
parents
train
their
children
more
on
tasks
related
to
the
behavioral
measures
used.
Low
lead
parent­
child
intelligence
was
correlated
at
0.52
and
high
lead
at
only
0.1.
The
low
correlation
in
high
lead
groups
luence
was
operat
inq
suggests
that
a
factor
other
than
parental
inf
and
probably
was
lead
exposure.

c­
44
Albert,
et
al.
(
1974)
studied
school­
age
children
with
a
his­

tory
of
PbBs
)
60
pg/
dl
early
in
childhood.
[
Jnfortunately,
PbBs
­

for
about
one
half
of
the
control
oooulation
were
not
available
and
some
of
the
control
children
previously
had
PbBs
440
pg/
dl.
­

The
same
types
of
flaws
existed
in
studies
which
came
up
with
negative
results.
Thus,
Kotok's
study
(
1972)
had
a
rather
wide
overlap
between
PbBs
of
control
subjects
and
lead­
exposed
subjects,

and
in
another
negative
study
fewer
than
half
of
the
"
lead­
exposed"

group
had
PbBs
>
40
ug/
dl
(
Lansdown,
et
al.
1974).
Another
problem
­

among
negative
studies
has
been
the
study
of
perhaps
inappropriate
populations.
Lansdown's
pooulation
consisted
of
British
children
living
in
the
vicinity
of
a
smelter.
In
another
negative
study,

the
children
were
Mexican­
Americans
also
living
in
the
vicinitv
of
a
smelter
(
McNeil,
et
al.
1975).
The
problem
population
we
are
dealing
with
in
this
country
is
of
an
entirely
different
socio­

economic
character:
inner
city
children
who
are
predominantly
socially
and
economically
deprived.
The
difference
in
background
may
be
significant
as
a
determinant
of
behavioral
ability.

In
summary,
there
is
sufficient
evidence
to
indicate
that
subtle
neurcbehavioral
effects
of
lead
exposure
occur
in
children
exposed
to
lead
at
levels
which
do
not
result
in
clinical
encepha­

lopathy
.
The
minimal
level
of
lead
exposure,
the
duration
of
expo­

sure
required,
alld
the
period
of
greatest
sensitivity
cannot
be
specified
with
any
degree
of
certainty.
However,
the
conclusions
of
two
recent.
expert
groups
who
have
evaluated
the
literature
in
World
Yealth
Organization
.
ion
great
depth
are
remarkably
similar.
The
concluded
that
the
probabil
itv
of
not
iceable
brain
dvsfunct
c­
4
5
increases
in
children
from
PbB
levels
of
approximately
50
ua/
dl
(
WHO,
1977),
and
the
U.
S.
EPA
Science
Advisory
Board
concurred
in
the
U.
S.
EPA
conclusion
that
"
the
blood
lead
levels
associated
with
neurobehavioral
deficits
in
asymptomatic
children
appear
to
be
in
excess
of
50
to
60
ug/
dl."
Future
research
may
reveal
that
this
cut­
off
point
is
actually
lower.
Effects
of
lead
exposure
on
the
peripheral
nervous
system
of
both
adults
and
children
are
also
documented.
A
number
of
studies
have
documented
the
occurrence
of
slowed
nerve
conduction
with
an
approximate
PbB
maximum
of
50
ug/
dl
(
Hernberg,
et
al.
1967;
Lilis,
et
al.
1977;
Landrigan
and
Baker,

1976).
This
effect
has
been
noted
to
occur
at
this
exposure
level
without
any
overt
signs
of
neuromuscular
impairment.

Although
generally
considered
not
to
be
a
maior
public
health
problem
today,
the
potential
damage
to
the
brain
of
the
fetus
from
lead
exposure
has
received
some
attention.
Reattie,
et
al.
(
1975)

identified
77
retarded
children
and
77
normal
children
matched
for
age,
sex,
and
geography.
Of
64
matched
Dairs,
11
of
the
retarded
children
came
from
homes
in
which
the
concentration
of
lead
in
the
"
first
flush"
water
exceeded
800
ug/
l.
By
contrast,
none
of
the
control
children
came
from
such
homes.
In
a
follow­
up
study,
PbBs
from
the
mental
retardates,
taken
durinq
the
second
week
of
life,

were
found
to
be
significantly
higher
than
those
of
control
sub­

jects
(
25.5
ug/
dl
versus
20.9
ug/
dl)
(
Moore,
et
al.
1977b).
Taken
at
face
value,
those
studies
are
extremely
provocative.
They
sug­

gest
that
the
brain
of
the
fetus
is
considerably
more
sensitive
to
the
toxic
effects
of
lead
than
the
brain
of
the
infant
or
younq
child.
Lambs
exposed
to
low
levels
of
lead
in
utero
(
PbB
=
35)
­­

C­
46
developed
impaired
visual
discrimination
learning
behavior
(
Car­

son,
et
al.
1974).
In
spite
of
this
seemingly
low
level
of
expo­

sure,
control
animals
were
exposed
in
utero
to
lower
levels
of
lead
­­

(
PbB
=
5)
than
are
generally
considered
normal
for
most
species.

Bull
and
coworkers
have
exposed
female
rats
to
Pb
from
14
days
prior
to
breeding
through
weaning
of
pups.
The
normal
postnatal
increase
in
cerebral
cytochromes
(
Bull,
et
al.
1978)
and
synapto­

genesis
in
the
cerebral
cortex
(
McCauley,
et
al.
1979)
were
delayed
by
this
treatment.
These
delays
were
associated
with
delays
in
the
development
of
exploratory
and
locomotor
behavior
during
the
same
development
period
(
Crofton,
et
al.
1978).
The
latter
effect
was
shown
to
be
entirely
due
to
exposure
to
Pb
in
utero.
Blood
lead
­­

concentrations
on
the
18th
day
of
gestation
were
reported
to
be
31.9
pg/
dl.
Further
work
is
urgently
needed
concerning
the
neuro­

behavioral
effects
of
low­
level
lead
exposure
in
utero.
­­

Final.
ly,
a
few
comments
are
in
order
regarding
neurobehavioral
effects
of
low­
level
exposure
in
adults.
A
battery
of
performance
tests
were
administered
to
190
lead­
exposed
workers,
along
with
a
questionnaire
(
Morgan
and
Repko,
1974).
PbBs
were
below
80
vg/
dl
in
many
of
the
workers.
Unfortunately,
there
were
many
methodo­

logical
problems
and
eguipment
failures
which
rendered
the
results
difficult
to
interpret.
Further,
results
of
a
similar
study
by
other
investigators
were
essentially
negative
(
Milburn,
et
al.

1976).
Thus,
although
it
seems
reasonable
to
suppose
that
neuro­

behavioral
effects
do
occur
at
some
level.
of
exposure
in
workers,

it
is
extremely
difficult
to
specify
the
exposure
level
at
which
these
effects
may
occur.

c­
47
Carcinoqenicity
Three
groups
of
investigators
have
reported
epidemiological
studies
of
causes
of
death
among
people
overly
exposed
to
lead.

The
first
such
study
was
of
causes
of
death
among
154
pensioners
who
died
between
1926
and
1961
and
of
183
men
who
died
between
1946
and
1961
while
still
emDloyed
(
Dingwall­
Fordyce
and
Lane,
1963).

The
men
were
categorized
as
to
lead
exposure
based
on
the
nature
of
their
work
and,
in
the
case
of
highly
exposed
men,
on
the
basis
of
urinary
lead
excretion
(
100
to
250
vg/
dl
during
the
past
20
years
and
probably
higher
than
that
earlier
in
the
work
historv).
There
is
a
correlation
between
urinary
lead
and
blood
lead,
wherein
100
ug
Pb/
l
in
urine
corresponds
roughly
to
50
ug/
dl
in
blood
(
Selander
and
Cramer,
1970).

There
were
179
men
in
the
hiqh
exposure
category
for
which
causes
of
death
were
registered,
67
men
in
the
category
of
negligi­

ble
exposure
and
91
men
with
no
exposure.
Althouqh
there
was
a
sig­

nificant
excess
number
of
deaths
among
the
men
who
had
been
exposed
to
the
qreatest
lead
hazard,
this
excess
could
not
be
attributed
to
malignant
neoplasms,
as
the
mortality
rate
from
this
cause
was
actually
somewhat
less
than
expected.
Furthermore,
the
incidence
of
death
from
malignant
neoplasms
in
this
group
has
actually
in­

creased
in
the
more
recent
years
as
working
conditions
have
im­

proved.
It
seems,
rather,
that
the
excess
deaths
in
the
heavily­

exposed
group
was
due
mainly
to
vascular
lesions
of
the
central
nervous
system
among
men
employed
in
the
lead
industries
during
the
first
quarter
of
this
century.

C­
48
The
second
relevant
study
was
of
orchardists
who
at
one
time
sprayed
fruit
trees
with
lead
arsenate.
A
cross­
sectional
study
of
this
population
was
conducted
in
1938
by
the
U.
S.
Public
Yealth
Service
(
Yelson,
et
al.
1973).
The
population
was
classified
as
to
exposure
on
the
basis
of
whether
they
were
adult
orchard
workers,

(
orchardists
and
lesser­
exposed
"
intermediates"
as
separate
cate­

gor
ies)
,
non­
exposed
adults
of
the
area,
and
children
in
the
area.

For
all
categories
blood
lead,
urine
lead,
and
arsenic
concentra­

tions
were
determined.
In
addition,
the
number
of
years
of
spray
exposure
was
recorded
for
the
orchardists
and
"
intermediates."

There
was
a
definite
gradation
in
blood
and
urine
lead
concentra­

tion
corresponding
to
the
degree
of
exposure
as
classified
by
nature
of
orchard­
related
work
or
lack
thereof.
The
orchardists
had
the
highest
PbB
(
x
=
44
for
males
and
43
for
females).
Children
of
the
area
were
intermediate
(
PbB
=
37
in
boys
and
36
in
girls)
and
adult
consumers
and
"
intermediates"
had
PbBs
of
22
to
30.

In
1968
a
follow­
up
study
of
this
population
was
begun.
Re­

sults
were
reported
in
1973
(
Nelson,
et
al.
1973).
Of
the
original
1,229
study
members,
the
status
of
1,175
could
be
determined.
Four
hundred
and
fifty­
two
had
died
and
death
certificates
were
avail­

able
for
442.
No
consistent
differences
in
Standard
hfortality
Ratios
(
SMR)
were
observed
on
the
basis
of
either
exposure
classi­

fication
or
duration
of
exposure.
The
only
deviations
in
SYR
from
expected
were
in
the
direction
of
fewer­
than­
expected
deaths.
The
mortality
records
for
heart
disease,
cancer,
and
stroke
were
exam­

ined
separately.
Again,
there
was
no
suggestion
of
a
relationship
between
lead
exposure
and
death
from
any
of
these
three
maior
causes
of
death.

C­
4
9
The
most
recent
study
of
causes
of
death
amonq
lead­
exposed
workers
was
reported
by
Cooper
and
Caffey
(
1975)
and
Cooper
(
1976).

Since
the
results
were
published,
the
study
population
has
been
re­

examined
(
Cooper,
1978).
Results
from
the
updated
study
will
be
discussed,
although
details
as
to
lead
exposure
history
appear
mainly
in
the
earlier
publication.
The
objective
of
the
study
was
to
determine
what
happened
to
lead
workers
whose
levels
of
lead
absorption
were
below
those
associated
with
clinically­
recogniz­

able
illness
but
above
that
of
the
general
population.
The
oopula­

tion
studied
consisted
of
2,352
smelter
workers
and
4,580
battery
workers.
Death
certificates
were
available
for
1,703
of
these
men.

A
good
record
of
lead
exposure
history
was
considered
important.

Unfortunately
biological
monitoring
programs
(
lead
in
urine
or
blood)
were
not
in
effect
in
many
of
the
plants
during
the
period
of
employment,
particularly
so
for
the
deceased.
Nevertheless,
enouqh
data
were
available
to
indicate
that
exposure
was
heavy.
Thus,
67
percent
of
1,863
workers
had
PbSs
>
40
ug/
dl
and
20
percent
had
­

PbBs
)
70
ug/
dl.
­
Twenty­
six
percent
of
the
battery
workers
and
21.1
percent
of
the
smelter
workers
had
been
employed
for
more
than
20
years.

The
only
causes
of
death
that
showed
a
statistically
signifi­

cant
elevation
were
"
all
malignant
neoplasms"
in
the
battery
work­

ers,
cancers
of
"
other
sites"
in
battery
workers
and
"
symptoms,

senility,
and
ill­
defined
conditions"
in
battery
workers.
In
only
one
of
all
the
cancer
deaths
was
a
renal
tumor
SDecified.
Only
two
tumors
of
the
brain
were
identified
in
the
follow­
up
study.
(
No
specification
is
made
in
the
original
1975
report
as
to
brain
c­
50
tumors.)
The
author
of
the
1978
report
concludes
that
the
excess
deaths
due
to
neoplasms
cannot
be
attributed
to
lead
"
because
there
was
no
consistent
association
between
the
incidence
of
cancer
deaths
and
either
length
of
employment
or
estimated
exposures
to
lead."
It
is
not
clear
from
reading
either
of
the
two
reports
con­

cerning
this
population
as
to
just
how
exposure
categories
were
established.

In
a
letter
to
Science,
Kang,
et
al.
(
1980)
questioned
the
appropriateness
of
basing
the
decision
of
statistical
significance
of
the
results
on
confidence
limits
rather
than
on
calcullations
of
a
more
rigorous
statistical
test.
In
their
reanalysis
of
the
re­

sults
of
the
1975
report
by
Cooper
and
Gaffey,
Kang,
et
al.
(
1980)

used
the
test
statistic
z
=
SMR
­
100
100
dl/
expected
and
calculated
a
sta­

tistically
significant
increase
in
deaths
due
to
all
malignant
neo­

plasms,
cancer
of
the
digestive
organs,
and
cancer
of
the
respira­

tory
system
for
lead
smelter
workers.
For
battery
plant
workers
they
calculated
a
statistically
significant
increase
in
cancer
of
the
digestive
organs
and
cancer
of
the
respiratory
system.
They
did
not
calculate
an
increased
incidence
of
all
malignant
neoplasms
for
these
workers.
Based
on
their
calculations,
the
authors
state
"
observation
of
a
significant
excess
of
cancer
in
two
independent
populations
exposed
to
lead
in
two
different
industrial
settings
lends
credibility
to
the
suggestion
that
lead
is
an
etiological
factor.
"

In
their
responses
to
Kang,
et
al.
(
1980),
Cooper
(
1980)
and
Gaffey
(
1980)
support
the
methods
and
conclusions
of
their
previous
work.

c­
51
In
1953
a
study
was
published
indicating
that
lead
causes
renal
tumors
in
rats
(
Zollinger,
1953).
Since
that
time,
five
other
studies
have
confirmed
this
finding
(
Boyland,
et
al.
1962;

Van
Esch,
et
al.
1962;
Roe,
et
al.
1965;
Mao
and
Molnar,
1967;

Oyasu,
et
al.
1970).
The
same
observation
has
also
been
reported
in
mice
but
could
not
be
elicited
in
hamsters
(
Van
Esch
and
Kroes,

1969).
Other
studies
indicate
that
lead
also
causes
lung
tumors
in
hamsters
(
Kobayshi
and
Okamoto,
1974)
and
cerebral
gliomas
in
rats
(
Oyasu,
et
al.
1970).
All
of
these
studies
were
conducted
using
levels
of
lead
exposure
far
in
excess
of
tolerable
human
doses,
but
most
were
designed
to
study
the
mechanism
of
lead­
induced
carcino­

genesis.

The
first
report
of
lead­
induced
renal
tumors
(
Zollinqer,

1953)
was
essentially
a
lifetime
study
in
rats,
with
administration
of
lead
beginning
at
150
to
180
grams
body
weight
and
continuing
for
up
to
9.5
months.
Single
weekly
doses
of
20
mg
lead
phosphate
were
administered
subcutaneously.
Of
the
112
animals
on
lead
that
were
examined,
many
died
early
in
the
study.
Twenty­
one
had
tumors.
Of
the
29
animals
remaining
after
10
months,
19
had
tumors.
The
last
animals
were
killed
16.5
months
after
initiation
of
the
lead
injections.
All
the
tumors
were
renal
and
were
classi­

fied
as
adenomas,
cystadenomas,
or
Dapillary
adenomas.
Metastases
were
evident
in
only
one
case.
According
to
the
histological
cri­

teria
for
renal
toxicity,
all
the
animals
receiving
lead
had
severe
lead
intoxication.
Among
50
control
animals,
none
develoDed
tumors.

C­
52
The
next
study
reported
(
Royland,
et
al.
1962)
tested
the
hypothesis
that
renal
cancer
due
to
lead
was
actually
caused
by
the
well­
known
accumulation
of
porphyrins
associated
with
lead
toxici­

ty.
To
test
the
hypothesis,
elevated
porphyrin
excretion
was
stim­

ulated
by
administration
of
allyl­
isopropylacetamide
(
AIA)
in
the
diet
of
20
rats
for
one
year.
A
like
number
of
rats
were
fed
1
per­

cent
lead
acetate
in
their
diet
for
one
year.
Roth
groups
of
ani­

mals
were
observed
until
they
became
ill
or
had
palpable
tumors.

During
the
period
of
lead
administration
the
mortality
rate
in
the
two
groups
was
quite
similar.
Subsequently
the
lead­
fed
rats
died
earlier
than
the
AIA
rats.
Subsequent
to
the
l­
year
administration
of
test
compounds
all
but
one
of
the
lead­
fed
rats
had
renal
tumors
whereas
none
of
the
AIA
group
had
tumors
of
any
kind.
It
is
not
clear
whether
the
accelerated
mortality
among
the
lead­
fed
rats
was
due
to
the
tumors
or
to
other
toxic
effects
of
lead.

Van
Esch,
et
al.
(
1962)
presented
the
first
study
in
which
tumor
mortality
was
determined
at
more
than
one
dosage
level
of
lead.
In
this
case
lead
was
administered
in
the
diet
as
basic
lead
acetate,
0.1
percent
in
one
group
and
1.0
percent
in
the
other.

Approximately
equal
numbers
of
males
and
females
were
used.
Each
lead­
fed
group
was
compared
to
its
own
set
of
controls,
not
receiv­

ing
lead.
Prior
to
the
termination
of
the
experiment,
only
mori­

bund
animals
were
killed
and
examined
morphologically.
At
equiva­

lent
durations
of
lead
administration,
using
these
guidelines
for
tumor
assessment,
the
higher
dose
of
lead
was
more
carcinogenic
than
the
lower
dose.
Thus,
at
the
end
of
600
days
of
lead
adminis­

tration,
31
percent
of
the
animals
which
survived
to
400
days
died
c­
53
from
renal
tumors
in
the
1.0
percent
lead
acetate
group,
whereas
only
14
percent
of
the
animals
alive
at
400
days
in
the
0.1
percent
lead
acetate
group
died
of
renal
tumors
(
Figure
4).
Mortalities
with
tumors
in
the
subsequent
200­
day
period
(
600
to
800)
were
not
comparable
because
in
the
case
of
the
1.0
percent
lead
group
all
the
animals
were
killed
at
730
days,
whereas
in
the
case
of
the
0.1
percent
lead
group
the
animals
were
allowed
to
survive
until
985
days
unless
they
became
moribund.
It
should
also
be
noted
(
Table
8)
that
during
the
first
600
days
of
the
0.1
percent
basic
lead
ace­

tate
regimen,
10
of
the
original
26
rats
(
38
percent)
died
without
renal
tumors
as
compared
to
one
of
the
original
26
in
the
control
group
(
4
percent),
indicating
that
at
this
level
the
lead
regimen
was
lethal
in
some
manner
unrelated
to
its
carcinogenicity.
As
a
matter
of
fact,
both
levels
of
lead
administration
caused
reduced
body
weight
gains,
suggesting
toxicity
unrelated
to
carcinogenesis.

The
next
study
of
lead­
induced
tumors
in
rats
was
also
de­

signed
to
shed
light
on
the
mechanism
of
lead
carcinogenesis
rather
than
to
define
dose­
response
relationships.
Roe,
et
al.
(
1965)

sought
to
establish
whether
testosterone
or
xanthopterin
would
influence
the
induction
of
renal
neoplasms
by
lead
in
rats.
In
this
study,
the
forms
of
lead,
lead
orthophosphate,
and
the
mode
of
administration
were
unique.
The
lead
salt
was
administered
subcu­

taneously
once
weekly
for
four
weeks,
then
intraperitoneally
for
nine
weeks;
then
after
a
rest
period
of
four
or
nine
weeks,
depend­

ing
on
the
particular
group
of
rats,
lead
administration
was
re­

sumed
for
an
additional
14
weeks.
All
the
animals
were
males.
The
dosage
schedule
of
lead
is
presented
in
Table
9,
assuming
an
aver­

c­
54
100
90
80
Cumulative
%
Mortality
(
0)
or
%
Animals
Z
"
Tumors
at
Time
of
Death
(
e)
B.

50
40
30
20
10
0
1
1
1
4
'
Total
n
=
29
0.18
PbAc
0
2y
401
2;
40
I
Y
400
600
72'
9
1
1
I
I
*

c
i
rota1
n
=
26
n=
1
E
0
/
0
/
6
/
A
,/,
!
t
:
201
.
k
401
4.
601
&
200
400
600
730
TIME
INTERVALS,
DAYS
FIGURE
4
Cumulative
Mortality
and
Tumor
Incidence
in
Rats
Source:
Van
Esch,
et
al.
1962
c­
55
TABLE
I3
IiEtect
of
L,
ead
Exposure
on
the
Incidence
of
Renal
Tumors
in
Ratsa
O­
200
­
Successive
Time
Intervals,
nays
201­
400
401­
600
601­
729
601­
800
noo­
9A5
Cb
O.
IC
c
0.1
c
0.1
c
0.1
c
0.1
c
o­
1
`
n
at
Ibeginning
of
interval­
15
16
13
16
12
15
10
14
10
14
5
6
dead,
no
rend1
tumors
2
0
1
1
2
1
3
1
5
6
5
1
dead,
renal
tumors
0
0
0
0
0
0
0
3
0
3
0
5
n
at
beginning
of
intclval­
dead,
no
renal
tUlllOKS
dead,
renal
tumors
14
16
14
16
12
15
9
9
9
9
3
5
0
0
2
1
3
6
4
6
3
3
1
0
0
0
0
0
1
0
0
0
0
4
C
l.
od
C
1.0
C
1.0
C
1.0
C
1.0
I,
`
St
twqinninq
ob
ir,
tervAl­
dcdd,
no
venal
tumors
dead,
rend
L
tumors
13
11
13
10
12
7
13
5
13.
5
0
1
0
1
I
1
0
1
12
1
0
0
0
2
0
1
0
2
0
4
rk
dt
beqinninq
cjf
in(
ervGl­
13
13
13
9
13
6
13
2
13
2
dcc3d,
no
renal
twwrs
0
4
0
2
0
1
0
0
13
0
dCdd,
renal
lumorr;
0
0
0
1
0
3
0
2
0
2
":;
0,,,
,:
e
:
V`
lll
kxh,
rl
`
31.
1962
1)
('
­
Cnntrol
"
0.1
­
O.
lN
bdsic:
lead
acetatci
in
diet
tl
I
.
O
=
1%
basic
112ird
<
bc:
etdte
in
diet
t!

C­
56
TABLE
9
Dosage
Schedule
used
by
Roe,
et
al.
(
1365)
in
their
study
of
the
Influence
of
Testosterone
and
Xanthopterin
on
the
Induction
of
Renal
Neoplasms
by
Lead
in
Rats
Group
Pb,
mg/
kg/
d
Days
on
Pb
n*

Pb
alone
Pb
alone
Pb
alone
Pb
+
testosterone
Pb
+
xanthopterin
Pb
+
testosterone
Pb
+
xanthopterin
Xanthopterin
Testosterone
No
treatment
2.63
1.25
0.17
1.25
1.25
0.17
0.17
242
238
238
238
238
238
238
238
238
238
24
24
24
16
16
16
16
16
24
24
*
n
=
number
c­
57
age
body
weight
of
400
g,
and
averaging
the
dose
over
the
total
treatment
period.

In
analyzing
the
cancer
data
for
these
groups,
it
seems
rea­

sonable
to
pool
all
the
groups
receiving
the
same
dosage
of
lead
since
neither
testosterone
nor
xanthopterin
influenced
the
tumor
incidence.
However,
xanthopterin
alone
seemed
to
increase
the
mor­

tality
rate
whereas
testosterone
alone
did
not.
Therefore,
only
the
lead
alone,
the
lead
plus
testosterone,
and
the
no
treatment
and
testosterone
alone
groups
are
pooled
here
at
equivalent
lead
dosages.
The
results
are
summarized
in
Table
10.

It
is
not
possible
to
establish
the
slope
of
the
interaction
between
dosage
of
lead
and
tumor
incidence.
The
highest
dose
was
so
toxic
that
there
were
only
two
survivors
by
the
time
the
first
tumor
appeared
in
that
group
(
Table
10).
The
remaining
two
dosage
levels,
by
contrast,
did
not
cause
death
unrelated
to
tumorigenesis
(
Figure
5).
However,
since
only
one
of
these
two
remaining
dosage
levels
was
tumorigenic,
no
dose­
response
relationshio
in
regard
to
tumorigenesis
is
calculable.

Interstitial
nephritis
occurred
in
all
grouns,
including
con­

trols.
Unfortunately,
other
manifestations
of
toxicity,
­
9.
r
anemia,
reduced
body
weight
gains,
and
food
consumption
were
not
reported.
In
keeping
with
the
observations
of
Van
Esch,
et
al.

(
19621,
Boyland,
et
al.
(
1962)
,
Mao
and
Molnar
(
1967),
and
Zol­

linger
(
1953),
very
few
of
the
affected
animals
exhibited
metasta­

sis
and
no
elevated
incidence
of
other
types
of
tumors
was
noted.

Neither
of
the
two
remaining
reports
concerning
the
carcino­

genic
effects
of
lead
in
rats
(
Yao
and
Yolnar,
1967;
Oyasu,
et
al.

C­
58
TARLE
IO
SurllfnarY
of
Mortality
Data
Resulting
from
Lead
Phsplratc
Adlninistration
to
Ratsa
Successive
Time
Intervals,
Days
­­
___
­
_.
__
­­­­
­­.
­­­­­­
­­__­
o­
100
101­
200
201­
300
301­
400
401­
500
50
l­
600
601­
700
­­__
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Cb
2.6'
1.3'
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2.6
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.
I7
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2.6
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2.6
I.
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2.6
1.3
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I7
C
2.6
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.
I7
C
2.6
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.
I7
___­
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n
ill
hc);
inniiig
01
ihltet
val
48
24
40
40
48
6
37
40
48
3
37
38
46
2
37
34
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I
35
25
26
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ll.
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0
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58
llllllW5
_
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.__­
_
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I_~
_­­­­­­­~.
~­­
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­­­
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._­­­­
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­
.­

d
hurt­
c::
I<
oc,
et
al.
1965
b
c:
4
l,
lllruls
c­
59
8
DEAD
100
90
80
70
60
50
40
30
20
10
0
D
I
m
/
,
y,
w
I
I
t
I
I
:
10.1
201
301
401
100
2o"
O
3iO
4:
o
5;
fb
Y1
600'
6$
1
700
TIME
INTERVALS,
DAYS
FIGURE
5
Cumulative
Mortality
Among
Rats
not
having
Renal
Tumors
Source:
Roe,
et
al.
1965
C­
60
1970)
involved
more
than
one
level
of
lead
administration.
The
results
obtained
by
Mao
and
Yolnar
(
1967)
serve
to
confirm
the
results
of
Van
Esch,
et
al.
(
1962)
in
that
both
grouus
used
the
same
regimen
of
lead
in
the
diet
(
1
percent
lead
acetate)
and
got
simi­

lar
incidences
of
renal
tumors
[
SO
percent
by
Van
Esch
(
1962)
vs.

77.5
percent
by
Mao
and
Yolnar
(
1967)
I.
Both
also
noted
that
the
first
appearance
of
tumors
was
at
about
300
days
following
initia­

tion
of
lead
administration.
Yao
and
Molnar
(
1967)
are
the
only
authors
who
conducted
any
lead
analyses.
They
reported
19.3
to
54.2
ug
Pb/
g
kidney
cortex
as
compared
to
3.1
pg
Pb/
g
in
a
single
normal
specimen.
By
way
of
comparison
to
man,
Barry
(
1975)
report­

ed
a
mean
of
0.66
pg/
g
in
kidney
cortex
of
10
occupationally­

exposed
adult
males,
with
a
standard
deviation
of
+
0.56
uq/
g.

Oyasu,
et
al.
(
1970)
used
a
dietary
regimen
of
lead
subacetate
for
326
to
432
days,
either
alone
or
combined
with
indole
in
one
case
and
acetylaminofluorene
(
AAF)
in
the
other.
Neither
of
these
substances
alone
caused
renal
tumors.
Therefore,
the
data
for
lead
with
and
without
these
additional
substances
could
be
combined.

Fifty­
nine
percent
of
130
animals
receiving
1
percent
lead
sub­
ace­

tate
in
the
diet
eventually
developed
renal
tumors.
This
report,

incidentally,
is
the
only
one
in
which
oral
feeding
of
lead
was
to
cause
tumors
other
than
renal.
Eight
percent
of
the
130
lead­
fed
rats
developed
gliomas.
All
but
one
of
these
were
cerebral.
One
was
cerebellar.
The
incidence
of
gliomas
in
animals
receiving
AAF
alone
was
2.5
percent,
compared
to
0.3
percent
in
controls.
There
did
not
seem
to
be
any
synergistic
effect
between
AAF
and
lead.

Lead
did
not
cause
any
other
types
of
tumors.
The
toxic
effects
of
lead
in
this
study,
apart
from
carcinogenesis,
were
not
reported.

C­
61
Van
Esch
and
Kroes
(
1969)
have
reported
that
basic
lead
ace­

tate
causes
renal
tumors
in
mice,
but
not
in
hamsters.
These
were
lifetime
studies
with
lead
being
incorporated
into
the
diet
beqin­

ning
at
five
weeks
of
aqe
for
the
mice
and
three
to
four
weeks
of
age
for
the
hamsters.
Two
levels
of
lead
were
used,
0.1
Dercent
and
1
percent,
cut
back
to
0.5
percent
early
in
the
study
owing
to
tox­

icity.
Only
one
renal
tumor
was
found
at
the
hiqh
level
of
lead
intake
in
the
mice,
but
this
was
probably
because
most
of
the
mice
died
within
the
first
100
days
of
lead
administration.
Fourteen
percent
of
the
mice
receiving
0.1
percent
basic
lead
acetate
devel­

oped
renal
tumors.
There
were
no
renal
tumors
in
hamsters
at
either
dosage
level
of
lead.
Mortality
was
somewhat
increased
at
both
levels
of
lead
administration.

Another
report
of
experimental
carcinoqenesis
is
a
report
of
induction
of
lung
tumors
in
Syrian
hamsters
using
intratracheal
injection
of
lead
oxide
(
Kobayachi
and
Okamoto,
1974).
Actually,

tumors
were
produced
only
when
benzo(
a)
pyrene
(
BP)
was
injected
simultaneously
with
lead
oxide.
Neither
compound
alone
caused
tumor
formation
under
the
conditions
described.
This
cooperative
effect
was
obtained
using
10
weekly
injections.
The
tumors
were
predominantly
adenomas
of
bronchio­
alveolar
origin.
In
addition
to
this
effect,
both
lead
alone
and
in
combination
with
BP
caused
a
very
high
incidence
of
alveolar
metaplasia,
which
the
authors
spec­

ulate
may
be
a
preneoplastic
change.
RP
alone
caused
a
very
low
incidence
of
alveolar
metaplasia.
All
treatments,
including
the
methylcellulose
injection
vehicle
alone
caused
some
deaths.

C­
62
The
final
study
concerning
the
carcinogenic
effects
of
lead
is
the
most
significant
of
all
(
Azar,
et
al.
1973).
It
confirms
other
studies
showing
that
lead
causes
renal
tumors
in
rats
and
that
male
animals
are
more
susceptible
than
females.
A
dose­
related
effect
is
clearly
evident
(
Table
11)
(
Figure
6).
The
dose
of
lead
re­

quired
to
produce
tumors
did
not
clearly
result
in
increased
mor­

tality
among
the
animals;
however,
at
dietary
lead
intake
above
1,000
ppm,
weight
gains
were
reduced.

In
summary,
there
is
little
doubt
that
certain
compound
of
lead
are
carcinogenic
or
at
least
co­
carcinogenic
in
some
species
of
experimental
animals.

Teratogenicity
There
is
little
information
in
the
literature
to
suggest
that
lead
has
a
teratogenic
effect
in
man.
Although
there
were
numerous
reports
of
a
high
incidence
of
stillbirths
and
miscarriages
among
women
working
in
the
lead
trades,
fetal
anomalies
were
not
de­

scribed.
It
must
also
be
pointed
out
that
these
women
were
Droba­

bly
exposed
to
much
higher
concentrations
of
lead
than
for
occupa­

tionally
exposed
men
today.
Recent
literature
is
devoid
of
any
references
to
teratogenic
effects
of
lead
in
man.

In
experimental
animals,
on
the
other
hand,
lead
has
been
shown
repeatedly
to
have
teratogenic
effects.
Early
studies
demon­

strated
this
in
chick
embryos
by
injection
of
lead
into
the
yolk
sac
(
Catzione
and
Gray,
1941;
Karnofsky
and
Ridgway,
1952).
Tera­

togenesis
has
also
been
observed
in
rodents.
These
studies
were
done
using
high
doses
of
lead
given
intravenously
or
intraperitone­

ally.
For
example,
McClain
and
Becker
(
1975)
used
single
intra­

C­
63
00000
00000
l­
l
PWWWc­
4
mmmmln
00000
OUTU7Ul~
r­
l
ooln
m
000
Ina3
tiOUl
mlnm
000
InInn
000
CVNW
99
90
80
Percent
animals
with
renal
50
tumors
10
1
10.1
0.2
0.5
1
2
5
ppm
Dietary
Pb
x
1'
J3
FIGURE
6
Probit
Plot
of
Incidence
of
Renal
Tulnors
in
Male
Sats
Source:
Azar,
et
al.
1973
90
80
50
10
1
10
C­
65
peritoneal
doses
of
25
to
70
mq/
kg
in
rats.
They
found
that
terato­

logic
effects
occurred
when
administration
was
on
day
9.
Adminis­

tration
later
in
pregnancy
resulted
in
embryotoxicity
(
fetal
re­

sorption)
but
not
in
teratogenic
effects.
Carpenter
and
Ferm
(
1977)
observed
teratologic
effects
in
hamsters
following
the
administration
of
50
mg/
kg
Pb(
N03)
2
intravenously
on
day
8.
Chron­

ic
administration
of
lead
in
the
drinking
water
of
pregnant
rats
at
very
high
concentrations
(
up
to
250
mg/
l)
resulted
in
delayed
fetal
development
and
fetal
resorDtion
without
teratologic
effects
(
Rim­

mel,
et
al.
1976).

In
summary,
it
seems
that,
in
man,
embryotoxicity
precedes
teratogenicity
in
the
lead
sensitivity
scale.
This
is
SupDorted
by
historical
experience
in
occupationally
exposed
women
and
by
animal
studies.

Mutagenicity
Pertinent
data
could
not
be
located
in
the
available
litera­

ture
concerning
the
mutagenicity
of
lead.

Reproductive
Effects
As
was
indicated
in
the
previous
section,
lead
has
been
known
to
cause
miscarriages
and
stillbirths
in
women
working
in
the
lead
trades
during
the
latter
half
of
the
19th
century
and
Drobably
on
into
the
early
Dart
of
the
20th
century,
It
is
very
difficult
to
estimate
minimally
toxic
exposure
for
stillbirth
and
miscarriages
because
exposure
data,
e.
g.,
PbB
are
lacking
for
women
who
experi­

enced
this
Droblem.
The
minimally
toxic
Level
of
exDosure
may
actually
be
quite
low.
Lane
(
1949)
reported
on
the
outcome
of
15
pregnancies
incurred
among
150
women
working
in
an
unspecified
lead
C­
66
trade
during
World
War
II.
Three
of
these
women
had
miscarriages
­

an
incidence
seven
times
normal.
Unfortunately
the
numbers
were
too
small
to
be
assigned
statistical
significance.
Lead
exposure
was
modest,
air
lead
being
75
ug/
m3
and
urinary
lead
excretion
in
men
working
with
these
women
being
75
to
125
ug/
l.
A
more
recent
Japanese
study
also
is
suggestive
of
miscarriages
occurring
among
women
with
only
modest
exposure
(
Nogaki,
1958).
These
women
were
the
wives
of
lead
workers.
Unfortunately,
the
actual
level
of
lead
exposure
was
not
reported.

It
has
recently
been
reported
that
the
incidence
of
premature
fetal
membrane
rupture
in
term
and
preterm
infants
is
much
higher
30
to
50
miles
west
of
a
lead
mining
area
of
Yissouri
(
17
percent)

than
in
a
Missouri
urban
area
remote
from
lead
mining
activities
(
0.41
percent)
(
Fahim,
et
al.
1976).
Maternal
and
fetal
PbSs
at
birth
also
differed
significantly
for
normal
births
vs.
births
with
premature
membrane
rupture.
Maternal
and
fetal
PbBs
for
the
normal
deliveries
were
about
14
and
4
pg/
dl,
respectively,
whereas
they
were
about
26
and
13
respectively
for
mothers
and
infants
with
mem­

brane
rupture.
This
provocative
study
needs
confirmation.
It
is
difficult
to
understand,
for
example,
why
fetal
PbB
should
be
so
much
lower
than
maternal
PbB
in
all
groups.

There
is
a
possibility
that
lead
affects
fertility
as
well
as
the
conception.
Lancranjan,
et
al.
(
1975)
reported
that
signifi­

cant
levels
of
teratospermia
occurred
among
men
working
in
a
lead
storage
battery
factory.
Their
PbBs
were
30
to
80
ug/
dl.
Although
many
studies
have
attempted
t0
correlate
semen
quality
with
fertil­

ity,
the
extent
to
which
abnormally­
shaped
sperms
participate'
in
C­
67
fertilization
is
unclear.
Experimental
animal
studies
have
shown
reduced
fertiiity
of
both
maternal
and
paternal
origin
(
Stowe
and
Goyer,
1971).

There
have
been
numerous
conflicting
reports
concerning
the
occurrence
of
chromosomal
aberrations
in
lymphocytes
of
lead­

exposed
workers
(
O'Riordan
and
Evans,
1974;
Forni,
et
al.
1976).

The
reason
for
these
conflicting
findings
is
not
clear.
DeKnudt,

et
al.
(
1977a)
suggest
that
ancillary
factors
may
be
critical:
for
example,
the
level
of
calcium
intake.
They
base
this
conclusion
on
the
lack
of
correspondence
between
lead
effects
in
two
widely
sepa­

rated
lead­
using
plants,
one
being
a
secondary
lead
smelter
and
the
other
being
a
plant
manufacturing
"
tin"
dishes.
Lead
exposures
were
roughly
comparable:
PbBs
were
on
the
order
of
45­
100
pgidl.

Severe
chromosomal
aberrations
were
found
in
one
plant
whereas
no
such
effects
were
seen
in
the
other.
They
further
point
out
that
no
severe
aberrations
have
been
seen
in
at
least
some
animal
studies
in
which
lead
exposure
was'
heavy
and
nutrition
apparently
adequate
(
Jacquet,
et
al.
1977;
De
Knudt,
et
al.
1977b).
The
implications
of
chromosomal
aberrations
which
have
been
reported
are
not
known.

A
recent
report
by
Wibberley,
et
al.
(
1977),
which
demonstrates
a
striking
increased
incidence
of
high
placental
lead
associated
with
stillbirths
or
congenital
malformations,
further
suggests
that
a
relationship
exists
between
intrauterine
exposure
to
lead
and
re­

productive
casualty.

Renal
Effects
There
is
considerable
information
in
man
concerning
the
renal
effects
of
lead.
Two
distinctive
effects
occur
in
both
adults
and
C­
68
children.
One,
is
reversible
oroximal
tubular
damage,
which
is
seen
mainly
with
short­
term
exposure.
The
other
effect
is
reduced
glo­

merular
function
which
has
qenerally
been
considered
to
be
of
a
slow,
progressive
nature.

Tubular
damage
is
manifested
as
the
Fanconi
triad
of
ql_
vco­

suria,
hypophosphatemia
with
phosphaturia,
and
generalized
amino­

aciduria.
The
last­
named
manifestation
appears
to
occur
more
con­

sistently
than
either
glycosuria
or
phosphaturia.
It
was
first
described
more
than
20
years
ago
in
lead
smelter
workers
(
Clarkson
and
Kench,
1956).
In
adults,
the
condition
probably
is
uncommon
at
PbBs
below
70
ug/
dl.
Thus,
in
a
recent
series
of
seven
workers,
all
of
whom
had
PbBs
70
pg/
dl,
with
a
range
of
71­
109,
none
had
amino­

aciduria
or
qlycosuria.
Significantly,
five
had
hemoglobins
below
12
g/
d1
(
Cramer,
et
al.
1974).
Similarly,
in
a
series
of
15
infants
hospitalized
for
lead
poisoning,
all
havinq
PbSs
)
lOO
ug/
dl
at
­

entry,
only
three
had
aminoaciduria,
with
PbBs
of
246,
299,
and
798
ug/
dl
(
Chisolm,
1968).

Reduced
qlomerular
filtration
with
attendant
rise
in
serus
urea
concentration
is
generally
considered
to
be
a
progressive
dis­

ease,
implying
prolonged
lead
exposure.
It
is
accompanied
by
interstitial
fibrosis,
obliteration
of
glomeruli
and
vascular
lesions
(
Morgan,
et
al.
1966).
It
occurs
at
relatively
low
levels
of
lead
exposure,
at
least
relative
to
the
levels
associated
with
aminoaciduria.
For
example,
in
Cramer's
series
of
seven
workers,

none
of
whom
had
aminoaciduria,
three
had
low
renal
clearance
of
inulin
(
490
ml/
min/
1.73m2).
In
another
study
of
eight
men
with
occupational
lead
exposure
(
PbBs
=
29­
98),
four
had
reduced
glomer­

C­
69
ular
filtration
rates
(
wedeen,
et
al.
1975).
Of
these
four
cases,

one
had
a
PbB
of
48
pg/
dl
at
entry.
The
maximal
PAH
secretion
rate
(
TmPAH)
was
also
reduced,
indicating
coexistent
tubular
damage.

Among
the
other
three
cases,
two
had
only
a
marginal
depression
of
TmPAH'
From
these
and
other
studies,
it
appears
that
the
kidney
is
sensitive
to
glomerular­
vascular
damaqe,
with
an
imprecisely
known
threshold
for
effect
which
may
be
below
PbR
=
50
ug/
dl.

Cardiovascular
Effects
Dingwall­
Fordyce
and
Lane
(
1963)
reported
an
excess
mortality
rate
due
to
cerebrovascular
disease
among
lead
workers.
These
workers
were
employed
during
the
first
quarter
of
the
20th
century
when
lead
exposure
was
considerably
higher
than
it
has
been
more
recently.
There
was
no
similar
elevated
mortality
among
men
em­

ployed
more
recently
however.
Similarly,
in
Cooper's
more
recent
epidemiological
study
there
was
no
excess
mortality
attributable
to
stroke
or
other
diseases
associated
with
hypertension
or
vasculo­

pathy
(
Cooper
and
Gaffey,
1975;
Cooper,
1978).
It
would
appear
from
these
studies
that
the
vascular
effects
of
lead
only
occur
with
heavy
industrial
lead
exposure
­
probably
in
excess
of
what
is
encountered
today.

There
have
been
reports
of
heart
failure
(
Kline,
1960)
and
of
electrocardiographic
abnormalities
(
Kosmider
and
Pentelenz,
1962)

attributable
to
lead
exposure.
Yowever,
these
cases
have
always
involved
clinical
lead
intoxication.
It
does
not
seem
likely,

therefore,
that
the
heart
is
a
critical
target
for
lead
effects.

c­
70
Miscellaneous
Effects
Sporadic
reports
of
other
biological
effects
of
lead
in
man
exist,
but
these
are
difficult
to
evaluate
as
to
associated
lead
exposure.
They
have
frequently
been
reoorted
only
at
high
exposure
levels
and
only
by
one
or
two
investigators.
For
example,
Dodic,

et
al.
(
1971)
reported
signs
of
impaired
liver
function
in
11
of
91
patients
hospitalized
for
lead
poisoning.
No
information
was
pro­

vided
as
to
indices
of
lead
exposure.
Impairment
of
thyroid
func­

tion
has
been
reported
in
moonshine
whiskey
drinkers
hospitalized
for
lead
poisoning
(
Sandstead,
et
al.
1969).
The
degree
of
lead
exposure
was
not
clearly
indicated,
but
it
can
be
assumed
to
have
been
high.
Intestinal
colic
has
long
been
recognized
as
a
sign
of
lead
in
industrially
exposed
people.
It
probably
also
occurs
in
children
with
lead
poisoning.
Beritic
(
1971)
reported
that
it
occurs
with
PbBs
as
low
as
about
40
ug/
dl.
This
seems
unlikely
since
the
cases
he
reported
also
were
anemic,
a
condition
associat­

ed
with
the
considerably
higher
PbBs.
A
number
of
studies
have
suggested
that
a
relationship
exists
between
lead
exnosure
and
amyotrophic
lateral
sclerosis
(
ALS).
The
most
recent
report
on
this
examined
plasma
lead
levels
in
16
cases
of
ALS
and
in
18
con­

trols
and
found
significant
differences
at
the
0.05
level
(
Conradi,

et
al.
1978).

Finally,
animal
studies
indicate
that
relatively
high
levels
of
lead
exposure
interfere
with
resistance
to
infectious
disease
(
Hemphill,
et
al.
1971:
Gainer,
1974).
There
are
no
reports
of
an
abnormal
infectious
disease
incidence
among
people
with
high
lead
exposure,
however.

c­
71
CRITERIOW
FORYULATIOY
Existing
Guidelines
and
Standards
Since
lead
is
ubiquitous
in
the
environment,
several
govern­

ment
agencies
have
become
involved
in
regulating
its
use.
The
most
recent
action
was
taken
by
the
Consumer
Product
Safety
Commission
(
CPSC)
.
In
1977
the
CPSC
lowered
the
maximum
allowable
concentra­

tion
of
lead
in
house
paint
to
0.06
percent.
At
present
the
Occupa­

tional
Safety
and
Health
Administration
(
OSHA)
is
preparing
a
set
of
regulations
regarding
occupational
lead
exposure.
Similarly,

the
U.
S.
EPA
has
set
an
ambient
air
lead
standard.
The
U.
S.
FDA
has
provided
new
guidelines
for
the
regulation
of
sources
of
lead
in
foods
and
cosmetics.
Given
the
multi­
media
nature
of
lead
exposure
to
man,
it
is
essential
that
any
action
taken
in
regard
to
one
source,
such
as
water,
be
coordinated
with
similar
actions
being
taken
for
other
media
such
as
air
and
diet.

Current
Levels
of
Exposure
Approximately
1
percent
of
taowater
samples
have
been
found
to
exceed
the
current
standard
of
50
pg/
l.
This
is
generally
a
prob­

lem
in
softwater
areas,
particularly
where
lead
pipes
convey
the
water
supply
to
the
tap
from
the
surface
connection.
The
contri­

bution
of
the
diet
is
approximately
200
pq/
day
for
adults.
For
children
(
ages
three
months
to
nine
years)
the
diet
contributes
40
to
200
ug
of
lead
per
day.
On
the
basis
of
current
information,
it
is
impossible
to
judge
how
much
dietary
lead
is
attributable
to
the
water
used
in
food
preparation.
The
concentration
of
lead
in
ambi­

ent
air
ranges
from
approximately
0.1
ug/
m3
in
rural
areas
to
as
much
as
10
us/
m'
in
areas
of
heavy
automotive
traffic.

C­
72
Special
Groups
at
Risk
In
addition
to
these
usual
levels
of
exposure
from
environ­

mental
media,
there
exist
miscellaneous
sources
which
are
hazard­

ous.
The
level
of
exposure
resulting
from
contact
is
highly
vari­

able.
Children
with
pica
for
paint
chips
or
for
soil
may
experi­

ence
elevation
in
blood
lead
ranging
from
marginal
to
sufficiently
great
to
cause
clinical
illness.
Certain
adults
may
also
be
ex­

posed
to
hazardous
concentrations
of
lead
in
the
workplace,
notably
in
lead
smelters
and
storage
battery
manufacturing
plants.
Again,

the
range
of
exposure
is
highly
variable.
Women
in
the
workplace
are
more
likely
to
experience
adverse
effects
from
lead
exposure
than
men
due
to
the
fact
that
their
hematoDoietic
system
is
more
lead­
sensitive.

Basis
and
Derivation
of
Criterion
The
approach
that
will
be
taken
here
in
assessing
the
impact
of
lead
in
water
on
human
health
is
basically
the
same
as
has
been
taken
by
the
U.
S.
EPA
(
1977)
for
lead
in
air.
The
critical
target
organ
or
system
must
first
be
identified.
Then,
the
highest
inter­

nal
dose
of
lead
that
can
be
tolerated
without
injury
to
the
tarqet
organ
must
be
specified.
Finally,
the
impact
of
lead
in
water
on
the
maximum
tolerated
internal
dose
must
be
estimated,
as
well
as
the
likely
consequences
of
specific
reductions
in
the
maximum
allowable
concentrations
of
lead
in
water.

In
identifying
the
critical
organ
or
system,
great
reliance
is
placed
on
the
concentration
of
lead
in
the
blood
(
PbB)
as
an
index
of
internal
dose.
Such
an
indirect
measurement
is
necessary
be­

cause
of
the
multi­
media
character
of
lead
intake.
Tt
is
virtually
c­
73
impossible
to
measure
total
lead
intake
in
People
in
any
meaningful
way.
Because
intake
and
output
fluctuate
greatly
from
day
to
day,

measurement
of
total
lead
intake
would
reauire
long­
term
balance
studies.
Variables
have
a
substantial
influence
on
the
rate
and
degree
of
lead
uptake
from
the
external
environment.
Some
groups
have
proposed
alternatives
to
PbB
as
a
measure
of
internal
dose,

e.
g.,
FEP
and
tooth
lead.
FEP
is
not
suitable
because
it
is
a
bio­

logical
response
to
lead.
As
such,
it
is
subject
to
influences
other
than
lead,
notably
iron
deficiency.
Tooth
lead
is
a
poten­

tially
useful
index
of
lead
exposure,
but
with
the
present
state
of
the
art
being
what
it
is,
tooth
lead
is
difficult
to
internret.
It
only
provides
an
integrated
profile
of
past
lead
exposure.
One
is
not
able
to
say
when
the
exposure
occurred.
It
has
the
additional
limitation
of
not
being
available
on
demand.
Teeth
are
shed
spon­

taneously
only
in
childhood.
Yoreover,
only
a
very
small
data
base
is
available
for
dose­
effect
and
dose­
response
using
any
measure
of
dose
other
than
PbB.
The
use
of
PbR
as
a
measure
of
internal
dose
is
widely
accepted,
simply
because
nothing
better
is
available.

Having
specified
that
PbB
is
the
best
measure
of
internal
dose
currently
available,
the
next
question
concerns
the
lowest
PbB
lev­

els
at
which
adverse
health
effects
occur.
Two
recent
documents
(
U.
S.
EPA,
1977;
WHO,
1977)
have
been
oublished
in
which
iudgments
were
rendered
in
this
regard
(
Table
12).
It
will
be
noted
that
the
estimates
are
strikingly
similar.
The
estimated
no­
effect
levels
are
based
on
limited
populations
and
probably
are
lower
to
some
undefinable
degree
in
the
total
population
at
risk.
Slightly
more
information
was
available
to
the
U.
S.
E?
A
panel
than
to
the
WI0
L
c­
74
TABLE
12
Summary
of
Lowest
PbBs
Associated
with
Observed
Biological
Effects
in
Various
Ponulation
Crouosa
Lowest
Observed
Effect
Level
(
pg
Pb/
lOO
ml
Blood)
Effect
Population
Grouu
10
15­
20
25­
30
40
40
40
50
50­
60
50­
60
80­
100
100­
120
ALAD
inhibition
Erythrocyte
protoporbhyrin
elevation
Erythrocyte
protoporohyrin
elevation
Increased
urinary
ALA
excretion
Anemia
Coproporphyrin
elevation
Anemia
Cognitive
(
CNS)
deficits
Peripheral
neuropathies
Encephalopathic
symptoms
Encephalopathic
symptoms
Children
and
adults
Women
and
children
Adult
males
Children
and
adults
Children
Adults
and
children
Adults
Children
Adults
and
children
Children
Adults
No
Observed
Effect
Levels
in
Terms
of
PbBb
No
Observed
Effect
Level
(
pg
Pb/
lOO
ml
Blood)
Effect
Population
Group
10
Erythrocyte
ALAD
inhibition
20­
25
FEP
20­
30
FEP
25­
35
FEP
30­
40
Erythrocyte
ATPase
inhibition
40
ALA
excretion
in
urine
40
CP
excretion
in
urine
40
Anaemia
40­
50
Peripheral
neuropathy
50
Anaemia
50­
60
Minimal
brain
dysfunction
60­
70
Minimal
brain
dysfunction
60­
70
Encephalopathy
80
EncePhalopathv
Adults
and
children
Children
Adult
females
Adult
males
General
Adults
and
children
Adults
Children
Adults
Adults
Children
Adults
Children
Adults
aSource:
U.
S.
EPA,
1977
b
Source:
World
Health
Organization,
1977
c­
75
panel
since
it
reviewed
literature
through
mid­
1977,
whereas
the
WHO
expert
groups
reviewed
literature
only
through
1976.
In
addi­

tion,
the
U.
S.
EPA
performed
statistical
calculations
based
on
the
known
distribution
of
blood
lead
levels
in
the
United
States.

Both
sets
of
data
in
Table
12
are
in
error
in
one
regard.
They
use
the
term
"
anemia"
inappropriately
under
the
"
Effect"
column.

What
they
really
mean
is
"
decrement
in
hemoglobin."
Anemia
is
a
clinical
term
used
to
denote
a
degree
of
hemoglobin
decrement
which
is
below
the
normal
range
for
that
class
of
individuals,
e.
g.,
men
or
children.

The
question
that
arises
in
considering
Table
12
is
which
is
the
critical
effect?
Precisely
the
same
issue
confronted
the
U.
S.

EPA
in
its
deliberations
concerning
establishment
of
a
national
ambient
air
quality
standard
for
lead
(
42
FR
630979).
It
focused
on
the
lead
effects
in
children
since
they
are
more
sensitive
than
adults.

It
ruled
that
the
maximum
safe
blood
lead
level
for
any
given
child
should
be
somewhat
lower
than
the
threshold
for
a
decline
in
hemoglobin
level
(
40
ug
Pb/
dl).
In
considering
how
much
lower
this
limit
should
be,
the
U.
S.
EPA
cited
the
opinion
of
the
Center
for
Disease
Control,
as
endorsed
by
the
American
Academy
of
Pediatrics,

that
the
maximum
safe
blood
lead
level
for
any
given
child
should
be
30
pg/
dl.
Based
upon
epidemiological
and
statistical
considera­

tions,
the
U.
S.
EPA
estimated
that
if
the
geometric
mean
PbB
were
kept
at
15
pg/
dl,
99.5
percent
of
children
would
have
PbB
L30
w/
dl.
This
position
provides
a
substantial
margin
of
safety
which
accomodates
minor
excursions
in
lead
exoosure
due
to
adventitious
C­
76
sources.
Controls
on
lead
in
obligatory
media
(
e.
g.,
air
and
water)
do
not
protect
children
from
the
hazards
of
pica
for
lead­

base
paint
chips
or
soil
and
dust
contaminated
with
lead
from
such
sources
as
fallout
from
the
smoke
zone
of
lead
smelters.

In
its
deliberations
concerning
an
ambient
air
lead
standard,

the
U.
S.
EPA
estimated
that
the
contribution
of
sources
other
than
air
to
PbB
is
i0
to
12
ug/
dl.
This
is
bresclmably
composed
over­

whelmingly
of
dietary
sources
which,
in
turn,
is
composed
of
both
food
and
water.

The
next
question
concerns
the
contribution
of
water
to
lead
exposure.
Only
three
useful
studies
of
the
interrelationship
be­

tween
PbB
and
lead
in
drinking
water
are
available.
Overall,
the
Moore,
et
al.
(
1977a)
study,
the
one
by
Hubermont,
et
al.
(
1978),

and
the
calculations
made
from
U.
S.
EPA
data
collected
in
the
Boston
area
(
Greathouse
and
Craun,
1976)
are
credible
because
they
are
consistent
with
other
information
concerning
the
curvilinear
relationship
between
2bB
and
air
Pb.
The
implication
of
the
equa­

tion
describing
the
relationship
between
PbB
and
water
lead
is
that
with
increasing
lead
in
water,
the
incremental
rise
in
PbB
becomes
progressively
smaller,
as
with
air
lead
vs.
?
bB
and
dietary
lead
vs.
PbB
(
see
"
Contributions
of
Lead
from
Diet
vs.
Air
to
PbB"
in
the
Pharmacokinetics
section).
The
water
lead
vs.
PbB
relationship
differs
in
one
significant
respect,
however,
from
the
air
lead
vs.

PbB
relationship
in
that
the
baseline
PbB
(
0
water
PbB)
is
indepen­

dent
of
the
contribution
of
water
lead
to
PbB.
Thus,
regardless
of
whether
one
starts
with
a
baseline
PbB
of
11
ug/
dl,
as
was
indicat­

ed
in
the
Moore,
et
al..
(
1977a)
study
or
whether
one
starts
at
some
c­
77
other
PbB
level,
e.
g.)
20
ug/
dl,
the
add­
on
PbB
from
any
given
level
in
water
will
be
the
same.
Such
is
not
the
case
in
the
Azar
analysis
of
air
Pb
vs.
PbB
(
see
"
Contributions
of
Lead
from
Diet
vs.
Air
to
PbB"
in
the
Pharmacokinetics
section).
Here,
the
higher
the
baseline,
the
less
is
the
contribution
of
air
Pb.
This
is
be­

cause
log
PbB
is
proportional
to
baseline
PbB
+
log
air
concentra­

tion.
Future
research
may
provide
better
insiqht
into
whether
this
discrepancy
is
real
and,
if
so,
why.
The
question
is
of
some
prac­

tical
importance.
For
instance,
if
you
have
a
baseline
PbB
(
no
lead
in
water)
of
30
pg/
dl,
such
as
in
a
child
acquiring
lead
from
paint,
it
would
be
of
some
importance
to
know
whether
an
additional
increment
of
lead
in
water
would
have
the
same
impact
on
PbB
as
it
would
in
a
child
having
a
baseline
of
PbB
of
10
pg/
dl.
An
Azar­
type
model
would
suggest
a
lesser
impact
starting
from
the
higher
base­

line
PbB.

So
far
as
a
specific
recommendation
regarding
a
water
auality
for
Pb
is
concerned,
a
stand
must
be
taken
using
the
available
data.
Beginning
with
the
assumption
that
a
PbB
of
12
ug/
dl
is
essentially
attributable
to
food
and
water
and
that
the
average
lead
content
of
water
consumed
is
10
ug/
l,
aporoximately
5
ug
Pb/
dl
blood
(
from
Table
6)
is
attributable
to
the
water
that
is
used
in
food
and
beverage
preparation
and
in
direct
consumption.
If
the
water
Pb
were
consistently
consumed
at
the
present
Pb
standard
of
50
ug/
l
instead
of
at
10
pg/
l,
an
additional
contribution
of
ap­

proximately
3.4
uq/
dl
to
PbB
would
result
(
3.57
­
5.13
from
Table
6)
l
This
would
yield
a.
total
PbB
of
12
+
3.5
or
15.4
pg/
dl,
the
approximate
maximum
qeometric
mean
PbB
compatible
with
keeping
99.5
C­
78
percent
of
the
population
under
PbB
=
30
ug/
dl.
Thus,
based
on
most
recent
data,
the
present
water
standard
of
50
ug
Pb/
l
may
be
viewed
as
representing
the
upper
limit
of
acceptability.
This
criteria
is
based
on
empirical
observation
of
blood
lead
in
human
population
groups
consuming
their
normal
amount
of
water
and
food
daily.
Spe­

cific
amounts
of
foods
or
drinking
water
consumed
were
not
quanti­

fied,
but
it
can
be
assumed
that
they
reflect
an
average
consump­

tion
of
water,
fish,
shellfish,
and
other
foods.

All
the
assumptions
that
have
been
made
in
arriving
at
an
estimate
of
the
impact
of
lead
in
water
on
PbB
have
been
on
the
conservative
side.
For
instance,
unpublished
data
from
the
Com­

mission
of
the
European
Communities
suggest
that
the
impact
of
lead
in
water
on
PbE
is
appreciably
less
than
has
been
estimated
from
published
data
used
in
this
document
[
personal
communication
from
Alexander
Berlin,
et
al.
(
19781,
Commission
of
the
European
Com­

munities,
1
Luxembourg]
.
Furthermore,
data
from
a
study
(
Morse,
et
al.
1978)
of
the
effect
of
lead
in
water
on
the
PbB
of
a
population
of
children
in
a
relatively
small
town
are
reassuring,
They
indi­

cate
that
among
children
whose
water
supply
contained
50
to
180
ug
Pb/
l,
PbBs
averaged
17.2
ug/
d12.

1
Subsequent
to
the
writing
of
this
reoort,
these
data
were
submit­
ted
to
the
EPA
by
Dr.
Berlin.
They
were
studied
and
judged
not
to
alter
the
conclusions
arrived
at
in
this
document
concerning
PbB
vs.
lead
in
water
(
see
Appendix).
2
It
should
be
pointed
out,
however,
that
the
contribution
from
other
sources
is
not
indicated,
thus,
the
relative
water
lead
con­
tribution
is
unknown.

c­
79
Finally,
there
remains
the
issue
of
the
carcinogenic
effects
of
lead.
Using
data
from
one
species
of
laboratory
animal
(
the
rat)
it
was
possible
to
construct
a
seemingly
valid
dose­
response
curve
and
to
calculate
a
dietary
level
of
lead
which
would
predict
an
incidence
of
cancer
in
l:
lOO,
OOO
peoole.
This
calculated
diet­

ary
level
of
lead
is
29
ug/
kg.
Since
this
estimate
includes
lead
from
all
sources,
its
implications
are
beyond
the
scone
of
this
document.
It
should
be
noted,
however
that
the
International
Agency
for
Research
on
Cancer,
Lyon,
France
considers
the
exneri­

mental
animal
evidence
to
be
of
dubious
significance
with
regard
to
man
(
IARC,
1972).

The
Agency
has
not
yet
resolve
all
of
the
issues
concerning
the
potential
carcinogenicity
of
lead,
but
will
complete
its
review
in
the
near
future.
All
of
the
data
will
be
subjected
to
an
exten­

sive
peer
review
by
outside
experts
and
in­
house
scientists.
De­

pending
upon
the
final
conclusions
of
the
review,
the
water
quality
criteria
for
lead
may
be
re­
evaluated.
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1972.
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Lead
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Commun.
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c­
103
APPENDIX*

Results
of
the
research
examined
in
the
Commission
of
the
European
Communities
(
CEC)
paper
are
summarized
in
Tables
1
and
2.

The
data
presented
in
Table
1
are
equations
developed
by
the
auth­

ors
concerning
the
relationship
of
blood
lead
(
PbB)
to
water
lead
(
PbW).
Table
2
consists
of
calculations
of
the
contribution
of
100
µ
g
Pb/
l
of
water
to
PbB
(
as
µ
g/
dl).
Some
of
these
calculations
were
made
by
Berlin,
et
al.
(
1978),
interpolating
from
data
points
in
the
articles
cited.
Others
were
made
using
the
equations
provided
by
the
authors
of
the
articles
cited.

Three
types
of
equations
are
presented:

(
1)
PbB
=
a
+
b
PbW
(
2)
PbB
=
a
+
log
PbW
(
3)
PbB
=

In
all
cases
"
a"
is
the
baseline
expressing
PbB
at
PbW
=
0.
Of
these
three
mathematical
relationships,
the
third
appears
to
be
the
most
valid
for
two
reasons:
(
1)
the
largest
number
of
subjects
are
involved
in
studies
using
this
equation,
and
(
2)
it
corresponds
to
the
analysis
of
U.
S.
EPA
data
(
Greathouse
and
Craun,
1976)
as
cited
in
the
lead
criterion
document,
which
also
involved
a
very
large
number
of
subjects.
Moreover
calculations
made
of
PbB
vs.
PbW
using
the
U.
S.
EPA
data
were
for
females
aged
20
to
50,
a
sub­
popu­

lation
which
probably
gets
a
larger
proportion
of
its
water
from
*
Summary
of
"
Research
of
PbB
vs.
Lead
in
Drinking
Water
in
Europe"
as
presented
by
A.
Berlin,
et
al.
(
1978),
Commission
of
the
European
Communities.

C­
104
TABLE
1
Relationships
between
PbW
and
PbB*

Relationship
Remarks
Reference
PbB
=
0.018
PbW
+
22.9
=
0.417
Pbw
=
µ
g/
l
PbB
=
µ
g/
l00
ml
First
morning
flush
Addis
and
Moore,
1974
PbW
and
PbB
in
µ
mol/
l
Moore,
1977a
PbB
=
0.76
+
0.15
PbW
r
=
0.58,
first
morning
flush
PbB
=
0.80
+
0.20
Pbw
r
=
0.52,
running
sample
PbW
and
PbB
in
µ
mol/
l
Moore,
et
al.
1977
PbB
=
0.533
+
0.675
3
Pbw
first
morning
flush
PbB
=
0.304
+
1.036
3
Pbw
running
sample
PbW
in
µ
g/
l
PbB
in
µ
g/
100ml
Lauwerys,
et
al.
1977
PbB
=
9.62
+
1.74
log
Pbw
first
morning
flush
PbW
and
PbB
in
µ
mol/
l
PbB
=
0.8
+
0.19
PbW
first
morning
flush
Moore,
1977b
PbB
=
0.8
+
0.53
PbW
full
flush
(
paired
samples)

PbB
=
19.6
+
7.2
PbW
PbW
in
ppm,
PbB
in
µ
g/
l00ml
first
morning
flush
Elwood,
et
al.
1976
PbB
=
20.7
+
12.6
PbW
As
above.
Beattie,
et
al.
1976
Re­
evaluated
data
*
Source:
Berlin,
et
al.
1978
C­
105
TABLE
2
Increment
in
PbB
for
an
Increase
of
100
~
g/
l
in
PbW
(
for
Concentrations
around
100
llq/
l)*

Increment
in
PbR
Remarks
Reference
1.3
~
g/
lOOml
For
running
sample
(
linear
interpolation)
20­
1040
pg/
l
PbW
1.2
pq/
lOOml
First
flush
(
linear
inter­
polation)
lo­
250
pg/
l
PbW
3.4
c1q/
lOOml
For
running
sample
(
linear
interpolation)
lo­
250
pg/
l
PbW
3.3
lig/
lOOml
For
first
flush
(
linear
inter­
polation)
35­
350
ug/
l
PbW
De
Graeve,
et
al.
1975
Beattie,
et
al.
1972
Covell,
1975
Addis,
et
al.
1974
I..
8
~
g/
lOOml
Using
the
linear
equation
derived
by
the
authors
Addis,
et
al.
1974
2.0
pq/
lOOml
Using
the
linear
equation
derived
by
the
author
for
running
water
samples.
Moore,
1977a
6.0
pq/
lOOml
Using
the
non­
linear
equation
derived
by
the
authors
for
running
water
samples.
Moore,
et
al.
1977
3.9
lig/
lOOml
Using
the
non­
linear
equation
de­
rived
by
the
authors
for
first
morning
flush.
Moore,
et
al.
1977
0.83
J~
g/
lOOml
IJsing
the
log
equation
derived
by
the
authors
Lauwerys,
et
al.
1977
In
view
of
the
low
PbW
value,
the
extrapolation
is
uncertain.
vos,
et
al.
1977
C­
106
TABLE
2
(
Continued)

Increment
in
PbB
Remarks
Reference
1.9
pg/
lOOml
Using
the
linear
equation
derived
by
the
authors
for
morning
flush
Moore,
et
al.
1977
5.3
pg/
lOOml
Using
the
linear
equation
derived
by
the
authors
for
full
flush
Moore,
et
al.
1977
0.72
pg/
lOOml
Using
the
linear
equation
derived
by
the
authors
for
morning
flush.
Elwood,
et
al.
1976
1.3
pg/
lOOml
Using
the
re­
evaluated
linear
equation
derived
by
the
authors
for
morning
flush.
Beattie,
et
al.
1976
*
Source:
Berlin,
et
al.
1978
c­
107
the
domestic
supply
than
the
population
at
large.
In
that
regard,

the
only
comparable
population
was
70
pregnant
female
subjects
in
the
study
of
Hubermont,
et
al.
(
1978)
cited
in
the
CEC
document
as
Lauwreys,
et
al.
(
1977).

In
summary,
of
the
studies
cited
in
the
CEC
document,
most
weight
should
probably
be
given
to
the
Moore,
et
al.
(
1977a)
cita­

tion,
on
the
basis
of
large
numbers
of
samples
of
water
and
study
subjects,
and
to
the
Hubermont,
et
al.
(
1978)
study
on
the
basis
of
a
substantial
number
of
subjects
which
were
probably
partaking
of
more
of
the
domestic
water
supply
than
other
sub­
classes
by
virtue
of
pregnancy
and
sex.

So
far
as
the
actual
calculations
in
Table
2
are
concerned,

there
is
one
error.
The
CEC
document
calculates
that
the
equation
of
Hubermont,
et
al.
(
1978)
(
cited
as
Lauwreys,
et
al.
1977)
would
predict
that
PbW
at
100
F.
rg/
l
would
result
in
a
PbB
contribution
of
0.83
pg/
dl.
The
error
is
obvious.
In
the
equation,
the
PbB
contri­

bution
of
water
is
given
by
PbB
=
1.74
log
PbW.
In
fact,
0.83
=

1.74
log
3,
not
1.74
log
100.
The
correct
calculation
is
PbB
=

1.74
x
2
=
3.48,
since
log
100
=
2.

Of
the
13
estimates
of
PbB
vs.
PbW
in
Table
2,
only
5
could
be
verified.
These
were
Addis,
et
al.
(
1974)
(
interpolation),
Addis,

et
al.
(
1974)
using
authors'
equation,
Voore
(
1977a)
using
author's
equation,
Beattie,
et
al.
(
1976)
using
author's
equation,
and
Moore,
et
al.
(
1977),
non­
linear
morning
flush.
Of
the
remaining
nine,
one
was
miscalculated
by
CEC
and
the
remaining
eight
could
not
be
verified
by
this
author
because
the
oaper
was
unavailable
(
Covell,
1975;
Elwood,
1976),
or
because
the
necessary
data
were
C­
108
not
in
the
paper
(
De
Graeve,
et
al.
1975;
Yoore,
et
al.
1977
using
non­
linear
equation
for
running
water;
Moore,
et
al.
1977
using
linear
equation
for
morning
flush
and
running
water
calculations),

or
because
it
was
not
possible
to
see
how
CEC
made
an
interpolation
from
the
data
cited
(
Beattie,
et
al.
1972).

In
summary,
the
two
most
credible
studies
among
the
nine
actu­

ally
scrutinized
in
this
addendum
were
the
very
ones
utilized
in
the
criterion
document
for
lead.
Of
the
two
reviewed
by
the
CEC
but
not
examined
at
the
time
of
this
writing
(
Covell,
1975;
Elwood,

1976),
one
was
reviewed
prior
to
development
of
the
criterion
docu­

ment
and
rejected
on
the
basis
of
the
seemingly
inappropriate
use
of
a
linear
regression
model
(
see
"
Contributions
of
Lead
from
Diet
vs.
Air
to
PbB"
in
the
Pharmacokinetics
section).
It
is
therefore
concluded
that
information
provided
by
CEC
does
not
alter
the
eval­

uations
made
in
the
criterion
document.

c­
109
REFERENCES
Addis,
C.
and
M.
R.
Moore.
1374.
Lead
levels
in
the
water
of
sub­

urban
Glasgow.
Nature.
252:
120.

Beattie,
A.
D.,
et
al.

urban
soft­
water
area.

Beattie,
A.
D.,
et
al.

Lancet.
2:
200.
1972.
Environmental
lead
oollution
in
an
Br.
Med.
Jour.
32:
491.

1976.
Blood
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