Ambient Water Quality Criteria
Criteria and Standards Division
Office of Water Planning and Standards
U.S. Environmental Protection Agency
Washington, D.C.

Aquatic Life
j~For lead, the criterion to protect freshwater aquatic
lifeycts derived using the Guidelined^'Ts "e^1,51 ln(hardness)" 3-37)
as a 24-hour average~j(see the figure "24-hour average
lead concentration vs. hardness")fand|the concentration
[should not exceed "e'1"51 Inlhardnesa) -l-39)J(see tl)#
figure "maximum lead concentration vs. hardness") at any
^For saltwater aquatic life, no criterion for lead can
be derived using the Guidelines^ and there are insufficient
data to estimate a criterion using other procedures.
Human Health
tf£c the protection of human health from the toxic properties
of lead ingested through water, the water criterion is 50

Lead is a soft gray, acid-soluble metal. It is used
in electroplating, metallurgy, and the manufacture of con-
struction materials, radiation protective devices, plastics
and electronics equipment. The solubility of lead compounds
in water depends heavily on pH and ranges from about 10,000,000
pg/1 of lead at pH 5.5 to 1 jug/1 at pH 9.0 (Durum, 1973).
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). Lead concentrations
in seawater have been reported at 0.03 pg/1 (Tatsumoto and
Patterson, 1963) and in fresh waters from 2 to 140 pg/1,
with a mean of 23 pg/1 (Kopp and Kroner, 1967).
Lead reaches the aquatic environment through precipitation,
fallout of lead dust, leaching from soil, street runoff,
and both industrial and municipal wastewater discharges
(U.S. EPA, 1976). It can be removed from the water column
by adsorption to solids or chemical precipitation or copreci-
in the aquatic environment, lead has been reported
to be acutely toxic to invertebrates at concentrations as
low as 450 ^ig/1 (Biesinger and Christensen, 1972) and chroni-
cally toxic at less than 100 pg/1 (Biesinger and Christensen,
1972). The comparable figures for vertebrates are 900 pg/1
for acute toxicity (Brown, 1968) and 7.6 pg/1 for chronic
toxicity (Davies, et al. 1976). Toxicity is also affected
by water hardness (Tarzwell and Henderson, 1960; Pickering

and Henderson, 1966). Hard water is protective of organisms
exposed to lead.
Bacterial action has been shown capable of converting
inorganic lead to organic forms (Wong, et al. 1975; Silverberg,
et al. 1976). Algae reportedly can concentrate lead in
their tissues to levels as much as 31,000 times ambient
water concentrations (Trollope and Evans, 1976). Since lead
is an element, it will not be destroyed and may be expected
to persist indefinitely in the environment in some form.
Lead has been shown to be teratogenic in animals (McLain
and Becker, 1975). Lead exposure has been reported to decrease
reproductive ability in men (Lancranjan, et al. 1975) and
women (Lane, 1949). It has also been shown to cause distur-
bances of blood chemistry (Roels, et al. 1978), neurological
disorders (Perlstein and Attala, 1966; Byers and Lord, 1943),
kidney damage (Clarkson and Kench, 1956) and adverse cardio-
vascular effects (Dingwall - Fordyce and Lane, 1963).

Biesinger, K.E., and G.M. Christensen. 1972. Effect of
various metals on survival, growth, reproduction, and meta-
bolism of Daphnia magna. Jour. Fish. Res. Board Can. 29:
Brown, V.M. 1968. Calculation of the acute toxicity of
mixtures of poisons to rainbow trout. Water Res. 2: 723.
Byers, R.K. and E.E. Lord. 1943. Late effects of lead
poisoning on metal development. Am. Jour. Child. 66: 471.
Clarkson, T.W., and J.E. Kench. 1956. Urinary excretion
of amino acids by men absorbing heavy metals. Biochem.
Jour. 62: 361.
Davies, P.H., et al. 1976. Acute and chronic toxicity
of lead to rainbow trout, Salmo gairdneri, in hard and soft
water. Water Res. 10: 199.
Dingwall-Fordyce, J., and R.E. Lane. 1963. A follow-up
study of lead workers. Br. Jour. Ind. Mech. 30: 313.
Kopp, J.F., and R.C. Kroner. 1976. Trace metals in waters
of the United States. Fed. Water Pollut. Control Admin.
U.S. Dep. Inter., Cincinnati, Ohio.

Lancranjan, I. et al. 1975. Reproductive ability of workmen
occupationally exposed to lead. Arch. Environ. Health 30:
Lane, R.E. 1949. The care of the lead worker. Br. Jour.
Ind. Med. 6: 1243.
McLain, R.M., and B.A. Baker. 1975. Teratogenicity, fetal
toxicity and placental transfer of lead nitrate in rats.
Toxicol. Appl. Pharmacol. 31: 72.
Perlstein, M.A., and R. Atlala. 1966. Neurologic sequelae
of plumbism in children. Clin. Pediat. 6: 266.
Pickering, Q.H., and C. Henderson. 1966. The acute toxicity
of some heavy metals to different species of freshwater
fishes. Air Water Pollut. Int. Jour. 10: 453.
Roels, H.A., et al. 1978. Lead and cadmium absorption
among children near a noferrous metal plant. A follow-up
study of a test case. Environ. Res. 15: 290.
Standen, A., ed. 1967. Kirk-Othmer encyclopedia of chemical
technology. Interscience Publishers, New York.
Tarzwell, C.M., and C. Henderson. 1960. Toxicity of less
common metals to fishes. Ind. Wastes. 5: 12.

Tatsumoto, M., and C.C. Patterson. 1963. The concentration
of common lead in seawater. Page 74 in Earth science and
meteorites. North Holland Publishing Co. Amsterdam.
Trollope, D.R., and B. Evans. 1976. Concentration of copper,
iron, lead, nickel, and zinc in freshwater algae blooms.
Environ. Pollut. 11: 109.
U.S. EPA. 1976. Quality criteria for water. Off. Water
Plan. Stand. U.S. Environ. Prot. Agency, Washington, D.C.
Wong, P.T.S., et al. 1975. Methylation of lead in the
environment. Nature 253: 263.

The toxic effects of lead have been extensively tested
on a wide variety of freshwater organisms. Representative
test animals used to determine these effects included fish
from six different families (Salmonidae, Cyprinidae, Catosto-
midas, Ictaluridae, Poeciliidae and Centrarchidae); inverte-
brate species were represented by rotifers, annelids, snails,
cladocerans, copepods, isopods, mayflies, stoneflies and
caddisflies; and plants from the algal, desmid and diatom
groups. Consequently, the available data base is quite
large and clearly demonstrates the relative sensitivity
of freshwater organisms to lead.
Acute Toxicity
As shown in Table 1, 15 LC50 values were available
for eight species of fish. All acute tests except one were
conducted for 96 hours, so only one test involving the red
shiner needed adjustment for test duration. Of the 15 acute
exposures only three were reported to be flow-through experi-
ments and of these three only two (Davies, et al. 1976 and
Holcombe, et al. 1976) reported measured values of total
lead; therefore, most of the LC50 values in Table 1 were
adjusted for static exposures and unmeasured test concentra-
*The reader is referred to the Guidelines for Deriving Water
Quality Criteria for the Protection of Aquatic Life £43
FR 21506 (May 18, 1978) and 43 FR 29028 (July 5, 1978)]
and the Methodology Document in order to better understand
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 the calcu-
lations for deriving various measures of toxicity as described
in the Guidelines.	n <•

The first three soft water fathead minnow acute tests
were conducted with lead chloride, and unadjusted LC50 values
ranged from 2,400 to 7,330 jig/1 (Table I). The close agreement
between these tests demonstrates that lead LC50 values for
fish can be reproduced with reasonable accuracy. The fourth
soft water fathead minnow test (Table 1) was conducted with
lead acetate and the calculated LC50 value agreed closely
with the lead chloride exposures. These results demonstrate
that these different lead salts have similar LC50 values.
All other tests with fish were conducted with either lead
chloride or lead nitrate.
Acute tests have been conducted with lead in both hard
and soft water with rainbow trout, fathead minnows and blue-
gills (Davies, et al. 1976 and Pickering and Henderson,
1966). Results from these tests showed that the adjusted
LC50 values for lead were greatly different in hard and
soft water and varied by a factor of 237 times for rainbow
trout, 65 times for fathead minnows and 19 times for bluegills
(Table 1). Another example of hardness related lead toxicity
to fish was reported by Tarzwell and Henderson (1960).
These authors conducted 96-hour exposures of fathead minnows
to lead in hard (400 mg/1) and soft (20 mg/1) water. Results
from the soft water test are shown in Table 1. The hard
water exposure was not included 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 ^g/1
which meant that the difference between hard and soft water

exposures varied by a factor greater than 31 times. Hale
(1977) conducted an acute exposure of rainbow trout to lead
and obtained an adjusted LC50 value of 6,160 jug/1. This
value is almost 6 times greater than the LC50 value obtained
for rainbow trout in soft water by Davies, et al. (1976).
Hale (1977) did not report water hardness? however, alkali-
nity and pH were reported to be 105 mg/1 and 7.3, respectively,
which suggests that this water was probably harder than
the test water used by Davies, et al. Acute values obtained
by Wallen, et al. (1957) for the red shiner and mosquitofish
were also quite high; however, the authors did not report
hardness and both tests were conducted in turbid water con-
taining suspended clay particles at approximately 300,000
Following the Guidelines, an exponential equation de-
scribing the relationship of toxicity to hardness for each
species was fit by least squares regression of the natural
logarithms of the toxicity values and hardness. For lead,
sufficient acute toxicity data and hardness ranges were
available for rainbow trout, fathead minnows, and bluegills
to fit regression equations. The slopes of these equations
ranged from 1.01 for bluegills to 2.16 for rainbow trout,
with a geometric mean of 1.51.
As a measure of relative species sensitivity to lead,
logarithmic intercepts were calculated for each species
by fitting the mean slope (1.51) through the geometric mean
toxicity value and hardness for each species. These inter-
cepts varied from 2.60 for brook trout to 5.23 for goldfish,

with a mean intercept of 3.86 for all six fish species.
This variation in logarithmic intercepts indicates a range
of species sensitivity to lead of only 14-fold when adjusted
for hardness effects. The adjusted mean intercept
(2.50) is slightly lower than that for the two salmonid
species tested. Thus the Final Fish Acute Value is given
k e(1.51 ln(hardness) + 2.50).
The acute toxicity data for invertebrate species (Table
2) contains 10 values for 9 species. Only one test was
conducted for 96 hours; however, the standard test for clado-
ceran and copepods is 48 hours so these exposures needed
no adjustment for test duration. None of the acute tests
shown were flow-through exposures and only one (Brown, 1976)
involved measured concentrations of total lead.
Whitley (1968) reported 24-hour LC50 values of 49,000
and 27,500 /ig/1 for sludge worms obtained from tests conducted
at pH levels of 6.5 and 8.5, respectively. In a separate
test without lead, these pH values were determined to be
near the lower and upper 72 hour LC50 value limits for sludge
worms. The author also showed that at the optimum pH survival
level of 7.5 all sludge worms lived when exposed for 24
hours to concentrations of lead ranging from 10,000 to 50,000
Because lead toxicity to fish was shown to be significantly
related to water hardness, it was necessary to know the
hardness values for as many tests as possible with invertebrate
species. Hardness values (Table 2) for rotifers, Daphnia
magna and isopods were reported by the authors. Cairns,

et al. (1976) did not report water hardness; therefore,
a value was taken from the authors' previously reported
work conducted at the same laboratory- Baudouin and Scoppa
(1974) also did not report water hardness; however, it was
possible to calculate a hardness value by using the authors
reported test water values for pH, alkalinity, calcium and
conductivity. A comparison of adjusted LC50 values between
fish and invertebrates species shows that except for rotifers
the invertebrate species are generally more sensitive to
lead than fish in either hard or soft water.
Although a wide variety of invertebrate species have
been tested (Table 2), no reports were found in the literature
which tested lead toxicity on the same species in both hard
and soft water. However, it seems logical to assume that
a similar relationship exists between acute lead toxicity
and water hardness for invertebrate species as was demonstrated
for acute exposures of fish. This relationship was therefore
estimated for invertebrate species by using the slope (1.51)
from fish acute values. Calculated logarithmic intercepts,
as a measure of relative species sensitivity, ranged from
-0.10 for Daphnia hyalina to 5.59 for the rotifer (Philodina
acuticornis), with a geometric mean of 1.65. Following
adjustment using the species sensitivity factor (21) , the
intercept is -1.39. This would indicate that the invertebrate
species tested are slightly more sensitive to lead than
fish. Thus the Final Invertebrate Acute Value is given
by: e*1,51 In (hardness) - 1.39) which also becomes the equation
for the Final Acute Value.

Chronic Toxicity
Chronic tests have been conducted with lead and six
species of fish (Table 3). All chronic tests were conducted
in soft water (33-44 mg/1 as CaCO^).
No acceptable hard water chronic tests were found in
the literature to compare with the soft water data. Davies,
et al. (1976) reported the long-term effects of lead on
rainbow trout in hard and soft water (Table 7). Although
these tests were neither life cycle, partial life cycle,
nor embryo-larval tests, they do provide useful information.
During these 19-month exposures a significant number of
trout developed spinal deformities, eroded fins and blacktails
in both hard (353 mg/1 as CaCOj) and soft (28 mg/1 as CaCO^)
water at measured lead concentrations of 380 and 13 ug/1,
respectively (Table 7). These results, therefore, established
a definite relationship between water hardness and chronic
lead toxicity to fish in which the rainbow trout sensitivity
varied by a factor of 29 times.
Since no other appropriate fish data were available
to establish a significant relationship between chronic
toxicity values and hardness, a relationship was estimated
by using the slope (1.51) from fish acute values. Calculated
intercepts for the six species tested ranged from -1.81
for lake trout to -1.03 for white suckers, with a mean of
-1.47. The adjusted mean intercept (-3.37) is below that
for all species. Thus the Pinal Pish Chronic Value is obtained
from	ln(hardness) - 3.37).

Only one invertebrate chronic test result was found
in the literature (Table 4). This test with Daphnia magna
was conducted in soft water and the resulting chronic value
was seven times lower than the acute value (Table 2) in
the same water. Daphnids were among the most sensitive
invertebrate species tested in the acute exposures; therefore,
it would seem reasonable to assume that the chronic lead
value for Daphnia magna would be equal to or lower than
most other invertebrate chronic values. Therefore, it would
appear to be inappropriate to use the Guidelines' species
sensitivity factor of 5.1 with the chronic data for Daphnia
magna, since it is one of the most sensitive invertebrate
species. Consequently, that sensitivity factor will not
be used in the calculations to derive the Final Invertebrate
Chronic Value. It is also interesting to note that the
Daphnia magna chronic value for lead is very close to the
fish chronic values (Table 3). Even though invertebrate
chronic tests have not been conducted in hard water it would
again seem logical to assume that a similar relationship
probably exisits between chronic lead toxicity and water
hardness for invertebrate species as was demonstrated for
acute and chronic exposures of fish (Tables 1 and 7).
Since appropriate invertebrate data were not available
to establish a relationship between chronic toxicity values
and hardness, a relationship was estimated by using the
slope (1.51) from fish acute values and the lead value
and water hardness from the Daphnia magna chronic test.
Thus the Final Invertebrate Chronic Value is obtained
from e*'^* (kar<*ness) ~ 1*75).

Since the Final Fish Chronic Value is lower than that
for invertebrate species, it is used to establish the 24-
hour average lead concentrations for the protection of fresh-
water organisms in waters of various hardness.
Plant Effects
Fifteen tests on eight different species of aquatic
plants were found in the literature and are shown in Table
5. Plant exposures by Malanchuk and Gruendling (1973) and
Monahan (1976) were conducted for 24 hours and 7 days, respec-
tively. All tests were static and all concentrations unmea-
sured. The lowest lead value for these plants (500 jjg/1)
was established as the Final Plant Value and is well above
the 24-hour average lead concentrations.
Table 6 contains equilibrium bioconcentration factors
for lead for two fish species. The bioconcentration factor
for brook trout was calculated from a laboratory exposure
by Holcombe, et al. (1976) which included 20 measurements
of lead concentrations in the water during the 140-day test.
Lead residues reported by Atchison, et al. (1977) were obtained
from a mixed population of bluegills collected from a small
300-acre lake. The average concentration for lead in water
for this contaminated lake was determined from 36 separate
measurements. Since no maximum permissible tissue concen-
tration is available for lead, no Residue Limited Toxicant
Concentration can be calculated.
Table 7 contains no data that would appear to alter
the 24-hour average lead concentrations.

Freshwater - Aquatic Life
Summary of Available Data
Final Fish Acute Value * ed»51 ln(hardness) + 2.50)
Final Invertebrate Acute Value = ed»51 In(hardness)-1.39)
Final Acute Value = e1*51 hardness) - !-39>
Final Fish Chronic Value =	ln(hardness) - 3.37)
Final Invertebrate Chronic Value » e(l«51 In(hardness)-1.75)
Final Chronic Value = e^1*51 ln(hardness) - 3.37)
Final Plant Value = 500 jug/1
Residue Limited Toxicant Concentration = not available
The maximum concentration of lead is the Final Acute
Value of e^1'^"*" ln(hardness) ~ 1»39) an(^ tjie 24-hour average
concentration is the Final Chronic Value of	(hardness) 3.37) .
No important adverse effects on freshwater aquatic organisms
have been reported to be caused by concentrations lower
than the 24-hour average concentration.
CRITERION: For lead the criterion to protect freshwater
aquatic life as derived using the Guidelines is	^-n^~
(hardness) 3.37 as a 24-hour average (see the figure
"24-hour average lead concentration vs. hardness") and the
concentration should not exceed	In (hardness) - 1.39)„
(see the figure "maximum lead concentration vs. hardness")
at any time.

In scale

In scale

Table 1. Freshwater Ciah
Bioaasay Test.
oraanisi	Method * Coiic. *~
Rainbow trout,	S	U
Salmo galrdnerl
Rainbow trout,	FT	M
Salmo galrdnerl
Rainbow trout (2 moa), FT	U
Salmo gairdneri
Brook trout (18 mos),	FT M
Salvellnua fontinalia
Red shiner,	S	U
Notropia lutrenaia
Fathead minnow,	S	U
Pimephalea promelas
Fathead minnow,	S	U
Pimephalea promelas
Fathead minnow,	S	U
Piaephales promelas
Fathead minnow,	S	U
Pimephalea promelas
Fathead minnow,	S	U
Pimephalea promelas
Goldfish.	S	U
Carasalua auratus
Mosquitofish (adult),	S	U
Gambusia affinis
Guppy (6 moa),	S	U
Pr.ecilla reticulata
Bltiagill,	S	U
l.epomls macrochirua
Bluegill.	S	U
Lcpomia macrochirua
:ute values for lead
(mq/X as Time
(uq/l)	 (vetetence
Davies, et al.
Davies, et al.
Hale, 1977
Holcombe, et al.
Wallen, et al.
Tarzwell &
Henderson, 1960
Pickering &
Henderson, 1966
Pickering &
Henderson, 1966
Pickering &
Henderson, 1966
Pickering &
Henderson, 1966
Pickering &
Henderson, 1966
Uallen, et al.
Pickering &
Henderson, 1966
Pickering &
Henderson, 1966
Pickering &
Henderson, 1966

Tatue 1. (CouLiiiuod)
Hdcaneiis Adjusted
IliCitb^dy Teat 1 is I XBt>i LCbO LCMi
..J.ii M	H«iti «oq * Cone. ** CaCOj) JIU.S)	iUlidil	
* S - static, FT *» flow through
** U - unmeasured, H ¦ Measured
Adjusted LC50 vs. hardness;
Rainbow trout: slope - 2.16, Intercept - - 0.12, r « 1.0, Not significant, N - 2
Fathead minnow; slope - 1.57, intercept ¦» 3.26, r - 0.97, p - 0.01, N « 5
Bluegill: slope » 1.01, intercept •» 6,45, r - 1,0, Not significant, N = 2
Geometric mean slope « 1.51
Hean intercept for 6 species * 3.86
Adjusted ncan intercept » 3,86 - ln(3.9) « 2.50
Final Fish Acute Value - (1.51-ln(hardness) + 2.50)

Table 2. Freshwater invertebrate acute values for lead
Table 3. Freshwater fish chronic values for lead
Rainbow trout
Salmo gairdneri
Brook trout,
Salvellnus fontlnalls
Lake trout,
Salvellnus namaycush
Channel catfish,
Ictalurus punctatus
White sucker,
Catostowus coa—eraonl
Lepomis macrochlrus

lnq/l as
Sauter, et al.	1976
liolcombe, et al. 1976
Sauter, et al.	1976
Sauter, et al.	1976
Sauter, et al.	1976
Sauter, et al.	1976
* E-L » embryo-larval, LC » life cycle or partial life cycle
Fish Chronic Value vs. hardness:
No hardness relationship could be derived for any fish species
Slope » 1.51 fro* Fish Acute Value
Ceonetric mean intercept for 6 species ¦ - 1.47
Adjusted mean intercept » -1.47- ln(6,7) ° - 3.37
Final Fish Chronic Value « (1.51•In(hardness) - 3.37)
Brook trout,
Salvellnus fontlnalls
Application Factor Values**
96-hr LC50	MATC
(MR/1)	(MB/1)	AF
liolcombe, et al. 1976
Geometric mean AF » 0.02
The Final Fish Chronic Value is below the chronic value derived using the application i.'factor

Table 4. Freshwater Invertebrate chronic values for lead (Biesinger & Christensen, 1972)
Chronic	Haianes6
Limits Value	(iwj/1 as
Organise	Ttat* 
Table 5, Freshwater plant effects for le
Ef tect	iugAU	
Anabaena sp.
Anabaena sp.
Anabaena sp.
Anklstrodeamua sp.
Chlaaydowonaa sp.
Chlaa>ydoaonas sp.
Chlorella sp.
Cosmarlum sp.
Cosaarlua sp.
Cosnarlii sp.
Mavlcula sp.
Mavlcula sp.
Mavlcula sp.
Scenedeswus sp.
Selenastrms sp.
501 reduction	15,000
of 14C02 fixation
50% reduction	26,000
of JI,C02 fixation
501 reduction	15,000
of 1hC02 fixation
241 growth	1,000
501 reduction	17,000
of '""COj fixation
50X reduction	17,000
of ''•COj fixation
531 growth	500
501 reduction	5,000
of 1%C02 fixation
501 reduction	5,000
of ^OOj fixation
50Z reduction	5,000
of »I,C02 fixation
501 reduction	17,000
of II,C02 fixation
501 reduction	28,000
of ll|002 fixation
50Z reduction	17,000
of ,I»C02 fixation
35Z grov/ch	500
52% growth	500
final Plant Value " 500 yg/1
Malanchuk & Gruendling,	1973
Malanchuk & Gruendling,	1973
Malanchuk & Gruendling,	1973
Monahan, 1976
Malanchuk & Gruendling,	1973
Malanchuk & Gruendling,	1973
Monahan, 1976
Malanchuk & Gruendling,	1973
Malanchuk & Gruendling,	1973
Malanchuk & Gruendling,	1973
Malanchuk & Gruendling,	1973
Malanchuk & Gruendling,	1973
Malanchuk & Gruendling,	1973
Monahan, 1976
Monahan, 1976

Table 6. Freshwater residues for lead
Organiaa	Bioconcentration Factor*
Brook CrouC (embryo-3 nos),	42
Salvelinua fontlnalls
Blueglll,	AS
Lepoals macrochlrus
(days)	Reference
140	Holcombe, et al. 1976
Atchison, et al. 1977
* Bioconcentratlon factors have been converted from dry weight to wet weight.
Geometric mean bioconcentratlon factor for all species * 43.

Table 7. Other freshwater daca for lead
Organ!aw	Duration
Cladoceran,	21 days
Daphnta magna
Isopod,	20 days
Aaellua merldianus
Mayfly,	14 days
Epheaerella grandis
Mayfly (nymph),	14 days
Ephemerella grandla
Mayfly,	7 days
Ephemerella aubraria
Stonefly.	14 days
Pceronarcys callfornlca
Caddisfly,	7 days
Hydropsyche betten!
Frog (adult),	30 days
Rana plplens
Rainbow trout,	28 days
Salmo gairdnerl
Rainbow trout (12 nos), 14 days
Salmo gairdneri
Rainbow trout,	19 mos
Salno gairdneri.
Rainbow trout,	19 mos
Salroo gairdneri
Brook trout,	21 days
Salvelinus fontinalis
Brook trout (12 mos), 14 days
Salvelinus fontinalis
Reduced growth
Bioconcentration factor
Bioconcentration factor
Inhibition of ALA-D
Inhibition of ALA-D
Lordoscoliosis, eroded fins
black tail
Lordoscoliosis, eroded fins
black tail
Inhibition of ALA-D
(mg/1 as	Result
CaCOj)		(mr/1)	Reference
43	300	Biesinger & Christensen, 1972
25	100	Brown, 1976
50	3,500	Nehring, 1976
50	-	Nehring, 1976
44	16,000	Warnick & Bell, 1969
50	-	Nehring, 1976
44	32,000	Uarnick & Bell, 1969
100	Kaplan, et al. 1967
135	13	Hodson, 1976
135	10	Hodson, et al. 1977
353	380	Davies, et al. 1976
28	13	Davies, et al. 1976
14	Adams, 1975
135	90	Hodson, et al. 1977

Table 7. (continued)
Brook trout
(embryo-21 day),
Salvellnus fontinalis
Brook trout (12 nos),
Salvellnus fontlnalla
Goldfish (<12 moa),
Caraflslus auratua
Puapkinseed (>12 nos),
1-epoals gibbosus
Qamicn Sttesk
38 days Elevation of ALP and ACH
36 days Decrease of hemoglobin and
inhibition of GOT activity
14 days Inhibition of ALA-D
1A days Inhibition of ALA-D .
(mg/1 as	Result
CaCO^)	(tig/i >	Reference
44	525	Christensen, 1975
44	58	Christensen, et al. 1977
135	470	Hodson, et al. 1977
135	90	Hodson, et al. 1977

The data base for the effects of lead on saltwater
species is quite limited when compared to that available
for freshwater species. A study on the effect of lead on
cholinesterase inhibition in shiner perch is the only study
conducted with saltwater fish. There are limited data for
shellfish and various algae but no chronic test data are
Acute Toxicity
The acute toxicity data base for saltwater organisms
is limited to static tests with invertebrate species. The
LC50 values ranged from 2,450 pg/1 for oyster larvae (Calabrese,
et al. 1973) to 22,869 yg/1 for adults for soft shell clams
(Eisler, 1977). After adjustment for testing procedures
and species sensitivity according to the Guidelines, these
data result in the Final Invertebrate Acute value of 50
pg/1 which becomes the Pinal Acute Value.
Chronic Toxicity
No life cycle or embryo-larval tests have been conducted
with lead and saltwater organisms.
Traditionally, shellfish have been used in bioconcentra-
tion studies since they are known to be excellent bioconcen-
trators of metals. Schulz-Baldes (1972) reported that mussels
(Mytilus edulis) could bioconcentrate lead 2,568 times that
concentration found in their immediate environment (Table
9). The hard clam (Mercenaria mercenaria) appears to be

the least efficient concentrator of lead (Table 2). It
is evident in Table 2 that different bioconcentration values
are obtained with the same species. This can best be explained
by the fact that bioconcentration of metals by molluscs
is affected seasonally, by differences in the weight of
the individuals, length of exposure to the metal, water
temperature, experimental design and the chemical form of
the metal in salt water. Attempts to normalize the bioconcen-
tration data according to exposure time would be futile
since so many other parameters (as mentioned above) equally
affect the accumulation, therefore each study has to be
accepted on its own merits.
Diatoms and other phytoplankton also bioconcentrate
lead (Table 9). Since these organisms serve as food for
molluscs, studies have been reported whereby lead accumulation
by molluscs is affected by the concentration of lead in
food organisms. Schulz-Baldes (1974) showed that mussels
(Mytilus edulis) took up 23.5 percent of the lead available
in food organisms as compared to 29 percent of that available
in the water. Abalone (Haliotus rufescens) accumulated
21 mg/kg of lead while feeding on a brown alga pretreated
with lead (Stewart, et al. 1976).
According to the Guidelines a maximum permissible tissue
concentration is needed to calculate a Residue Limited Toxicant
Concentration (RLTC). Since none is available for lead,
no RLTC can be calculated.

Studies have been reported on the sublethal effects
of lead to saltwater invertebrate species, fish and plankton
(Table 10). Included in these sublethal effects is reduction
in growth rate and development of adult and larval forms,
inhibition of enzymes and physiological processes (respiration,
photosynthesis) and delayed cell division. These effects
were observed in lead concentrations ranging from 14 jag/1
to 60,000 ug/1. The significance of these sublethal effects
has not been established for saltwater species. While these
are much less subtle than death to assess survival of a
population, they clearly cause stress to the individuals
which increases susceptibility predation and parasitization.
This can have detrimental effects on a population if the
stress is allowed to continue in a natural environment.

Saltwater - Aquatic Life
Summary of Available Data
The concentrations below have been rounded to two signif-
icant figures. All concentrations herein are expressed
in terms of lead.
Final Fish Acute Value » not available
Final Invertebrate Acute Value = 50 jug/1
Final Acute Value * 50 ;ig/l
Final Fish Chronic Value = not available
Final Invertebrate Chronic Value = not available
Final Plant Value = not available
Residue Limited Toxicant Concentration = 8.5 ^g/1
Final Chronic Value = 8.5 pq/1
0.44 x Final Acute Value = 21 pg/1
No saltwater criterion can be derived for lead using
the Guidelines because no Final Chronic Value for either
fish or invertebrate species or a good substitute for either
value is available, and there are insufficient data to estimate
a criterion using other procedures.
CRITERION: For saltwater aquatic life, no criterion
for lead can be derived using the Guidelines, and there
are insufficient data to estimate a criterion using other

Table 8. Marine invertebrate acute values for lead
Capitella capltata
Oyster (larva),
CraaaosCrea virginlca
Hard clam (larva),
Nercenaria nercenaria
Bioassay Test	Tine
Mfetnoo* Cone.** (ntb)
Soft ahell clam (adult), S
Mya arenaria
tun/ii lun/ll i.ettience
1,016 Relsh, et al. 1976
U	48 2,200-3.600 1,860-3,040 Calabrese, et al. 1973
U	48
U	96	27,000 22,869 Eisler, 1977
720-800 609-677 Calabrese 6. Nelson, 1974
* S » static
** U - unmeasured
Geometric nean of adjusted values «* 2,460 fg/1 ^49^ ** Mg/1

Table 9. Marine
Crassostrea virginica
Cras8oatrea vlrginica
Crasaostrea virglnics
Quahaug, hard clan,
Hercenarla mercenarla
Soft shell clan,
Mya arenaria.





Phaeodaccylum trlcomuturo
Platyaonas subcordiforais
jldues for lead
Bioconcei.trataon Factoi	(days)
536	140
68*	49
1,400	70
17.5*	56
112*	70
650*	40
200*	37
2,568*	130
2,077*	130
796*	130
1,050*	1/24
933*	1/24
Zarooglan, Manuscript
Pringle, et al. 1968
Shuster 6, Pringle. 1969
Pringle, et ai. 1968
Pringle, et al. 1968
Schulz-Baldez, 1974
Talbot, et al. 1976
Schulz-Baldes, 1972
Schulz-Baldes, 1972
Schulz-Baldes, 1972
Schulz-Baldes, 1976
Schulz-Baldes, 1976
* Dry weight to wet weight conversion
Geowetric mean bioconcentration factor for all species ~ 293

Table 10. Other marine data for lead
Clliate protozoan,
Cristigera sp.
Clliate protozoan,
Cristigera sp.
Ophryotrocha labronica
Oyster,	1 yr
Crasaoatrea vlrginica
Abalone,	6 raos
Haliotus rufescens
Soft shell clan,	168 hrs
Mya arenaria
Mussel,	40 days
Mytllla edulls
Mussel,	ISO hrs
Hytilus edulis
Mud crab,
Rhithropanopeus harlsii
Fiddler crab,	2 tries
Uca pugilator
Reduced growth rate by 150
Reduced growth rate by 300
11. 71
LC50	1,000
Oysters accumulated
0,88 ug/g wet wt. Sea
U20 sampled monthly
no lead added (BCF =
Accumulated 21 ug/g wet wt
while being fed a brown
alga (Egregia laevigata)
which was precreatea with
1 mg Pb/1.
LC50	8,800
LC50	. 30,000
LT50 for adults	500
Delayed larval develop- 50
Accumulated 2.04 i>g Pb/g 100
wet wt (BCF *» 20)
12 hrs
12 hrs
>600 hrs
Gray & Ventilla, 1973
Gray & Ventilla, 1973
Brown & Ahsanullah, 1971
Kopfler & Mayer, 1973
Steward & Schulz-Baldes, 1976
Eisier, 1977
Talbot, et al. 1976
Schulz-Baldes, 1972
Benjyts-Claus & Benijbs, 1975
Weis, 1976

Table 10. (Continued)
Duration Etiect
Sea urchin,
Arbacia punctulata
Shiner perch,
Cymatogaater ag^regata
Alga,	30-31 days
Lamlnarla digitata
Phaeodac tyluw
subcordi formia
subcordi forraia
24 hra
48-72 hra
Few gastrula developed	14
27% Inhibition of brain	7.8
50-60% reduction in	1,000
Completely inhibited 10,000
Reduced photoaynthesis	100
and respiration by 25-50%
72 hr8 No growth inhibition	1,000
Retarded population grow 2,500
growth by delaying cell
Caused inhibition of 60,000
growth and death
„ occurred
2 days 48% of cells in culture 2,500
6 days 98% of cells in culture 60,000
Refer e/.cfc
Waterman, 1937
Abou-Donia & Menzel, 1967
Bryan, 1976
Woolery & Lewin, 1976
Woolery & Lewin, 1976
Hannan & Patouillet, 1972
Hessler, 1974
Hessler, 1974
Hessler, 1975
Uessler, 1975

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Warnick, S.L., and H.L. Bell. 1969. The acute toxicity
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Mammalian Toxicology and Human Health Effects
The hazards of lead exposure have been under intensive
investigation for many years. Research activities continue
unabated 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 curtailed. 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 environ-
mental pollutants.
Reassessment of acceptable levels of lead exposure
have been fairly numerous in recent years. These have taken
the form of criteria documents and of more academically-
oriented reviews. Some have been highly comprehensive,
covering effects on the ecosystem in general, as well as
on man (Natl. Acad. Sci/Natl. Res. Counc., 1972a; Boggess,
1978). Others have been mainly concerned with effects of
lead on man (World Health Organ., 1977; U.S. EPA, 1977a;
Hammond, 1977a).
The purpose of this review is to summarize the litera-
ture 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.
Natural Background Levels
Lead is ubiquitous in nature, being a natural consti-
tuent of the earth's crust. The usual concentration in
rocks and in soils from natural sources ranges from 10 to
30 mg/kg. Most natural groundwaters have concentrations
ranging from 1 to 10 jlig/1. This is well below the United
States' drinking water standard of 50 jig/1. 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 concen-
tration of lead in rural vegetation, however, range 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 jug/nr*, mostly
measured over Greenland and over remote oceans.
Areas of abnormally high concentrations of lead occur
in natural ores, usually in conjunction with high concentra-
tions of cadmium and zinc. There is essentially no transfer
from natural ore beds into overlying streams. There is
none if the soil is even slightly alkaline (Jennett, et
al. 1978).

Man-generated Sources oC Lead
Lead consumption in the United States has been fairly
stable from year to year at about 1.3 x 10® 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- and tetramethyllead. Pigments and ceramics
account for about 6 percent of annual production. All other
major uses are for metallic lead products or for lead-contain-
ing alloys. The consumption of tetraethyl- and tetramethyl-
lead are declining. Other uses that have significant poten-
tial 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 deteri-
orating 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. More-
over, 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 movement into 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 1000 jug/1 (Beattie, et al. 1972a).
There is a definite positive correlation between the concen-
tration of lead in the domestic water supply and the concen-

tration 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 Ber-
rang, 1976). The lower the pH of the water and the lower
the concentration of dissolved salts in the water, the greater
the solubility of lead in the water. Leaching of lead from
plastic pipes has also been documented (Heusgem and De Graeve,
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 adequately known. In one recent survey
of 969 water systems, 1.4 percent of all tap water exceeded
the 50 jag/1 standard (McCabe, 1970). Special attention
should be given in water quality surveillance to soft water
supplies, especially those on the acid side of pH 6.5.
Future survey work should also indicate whether or not the
water was filtered before analysis. This appears to be
a common practice among water analysts. Since a substan-
tial 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 sub-
stantial fraction of the particulate lead in water is avail-
able for absorption. However, "drinking water" analyses
are usually performed in unfiltered water and hence repre-
sent total lead.
Ingestion from Food
It is generally held that food constitutes the major
source of lead ingested by people. Raw fruits and vege-
tables acquire lead by surface deposition from rainfall,

dust and soil, as well as from uptake via the root system.
The relative contribution of these two sources varies great-
ly 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 significant source of lead in man's
diet. There seems to be about as much variation in the
concentration of lead in specific food items as there is
between food items. Thus, 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. Many 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 substantially higher than in the filler before packing,
or in the same item packed in glass (Mitchell and Aldous,
1974; U.S. Food and Drug Administration, 1975). 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

Milk 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 pg/1 (Hammond and Aronson, 1964)
whereas market milk has an average of 40 pg/1 (Mitchell
and Aldous, 1974). Evaporated milk has been variously reported
to contain an average of 202 jug/1 (Mitchell and Aldous),
110 + 11 jug/1 (Lamm and Rosen, 1974), and 330 to 870 pg/1
(Murthy and Rhea, 1971).
The daily dietary intake of lead has been estimated
by numerous investigators, using either the duplicate por-
tions approach or the composites technique wherein theoret-
ical diets are derived using nutrition tables. The results
are generally consistent, considering variations in body
size and metabolic rates. Thus, Nordman (1975) reported
an average daily intake of 231 jug Pb for Finnish adult males
and 174 pg Pb for adult females. This is consistent with
a British study reporting 274 pq Pb/day for young adults
(Thompson, 1971) and with a Japanese study reporting 299
pg Pb/day for adult males doing medium work (Horiuchi, et
al. 1956). The first two studies described above used the
duplicate portions technique whereas the third used the
composites approach. The one duplicate portions study is
consistent with the studies cited above. Kehoe (1961) reported
an average intake of 218 jug Pb/day for sedentary men. This
is not consistent, however, with two other American studies
of daily fecal lead excretion (Coulston, 1962 and Levin,
1972). From the lead balance studies of Kehoe (1961), it
can be estimated that gastrointestinal absorption of lead
approximates 10 percent. Making this adjustment daily lead

intake from the diet based on fecal lead excretion would
be 113 pq in sedentary adult males (Coulston, et al. 1962)
and 119 jug in women (Tepper and Levin, 1972).
Many studies of dietary lead intake are somewhat vague
as to whether water consumption was included in the esti-
mates. Others specify "food and beverages."
The dietary intake of lead in infants and young chil-
dren 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 jjg/day of lead for children
ranging in age from 3 months to 8.5 years. Horiuchi, et
al. (1956) estimated 126 pg/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.
The third major obligatory source of lead in the general
population is ambient air. A great deal of controversy
has been generated regarding the contribution of air to
total daily lead absorption. Unlike the situation with
food and water, general ambient air lead concentrations
vary greatly. In metropolitan areas average air lead concen-
3	3
trations of 2 jig/m with excursions of 10 jjg/m in areas
of heavy traffic or industrial point sources are not uncom-
mon, whereas in non-urban areas average air lead concentra-
tions usually are of the order of 0.1 pg/m3. in addition,
people are so mobile that static air sampling devices are
not very useful for estimating the integrated air lead expo-
sure of urban populations.

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 blend-
ers of alkyl lead full additives. It is very unlikely that
the concentrations of lead in water or air are sufficient
to make dermal contact a significant route of exposure.
Miscellaneous Sources
Among adults not occupationally exposed to lead there
are several sources of lead which may assume clinically-
significant proportions. Perhaps the most widespread serious
problem is the consumption of illicitly-distilled whiskey
("moonshine") which often is heavily contaminated with lead.
Many cases of frank lead poisoning have been documented.
The concentration of lead in moonshine whiskey commonly
exceeds 10 mg/1, or 2000 times the drinking water standard.
Storage of acidic beverages in improperly glazed earthenware
has caused severe, sometimes fatal poisoning 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 1000 jig/m3 in certain areas of the plant. A
similar situation exists in some storage battery manufac-
turing 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 consi-
dered to be from inhalation of lead fumes and dusts. Hand-
to-mouth transfer probably also is significant.

The hazard of lead to children is of considerable con-
cern. The number of children unduly exposed to lead from
miscellaneous sources is impressive. Thus, federally assisted
lead screening programs reveal that undue lead absorption
(PbB > 40 jjg/dl) was found in 11.1 percent of 277,347 chil-
dren screened in 1973. The percentage has fallen since
then, being 6.4 percent in 1974 and 6.5 percent in 1975
(Hopkins and Houk, 1976). Even in 1976 the problem had
not changed appreciably since 1974 and 1975. In that year
8.7 percent of 500,463 children screened had PbB's 7 30
pg/dl and 2.7 percent or 13,604 children had PbB's > 50
pg/dl (Commun. Dis. Center, 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 surrounding
the homes. More recently, 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 (Natl. Acad. Sci/Natl.
Res. Counc., 1972), it is pointed out that the daily inges-
tion of 44 mg of street dust at 2000 jugPb/g would suffice
to elevate the PbB of a young child from 20 jag/dl to 40
jjg/dl. In a survey of 77 midwestern United States cities
it was found that the average lead concentration in the
street dust of residential areas was 1636 ^ig/g and that
in commercial and industrial areas the average concentra-
tions were, respectively, 2413 jig/g and 1512 /ig/g (Hunt,
et al. 1971). Soil along the shoulder of heavily-traveled
roadways also is heavily contaminated, although most values

found have been in the range of hundreds of micrograms per
gram rather than thousands (see, for example, Langerwerff
and specht, 1970).
The relative contribution of soil, automotive exhaust
fallout and paint to lead exposure in children remains uncer-
tain. There is no question that children in the age range
of 1 to 5 years, in which the problem of elevated PbB's
exists, do indeed exhibit pica, the habit of mouthing or
ingesting non-edible objects, e.g., pieces of plastic, gravel,
cigarette butts, etc. (Barltrop, 1966). The habit also
appears to be more prevalent among children who have ele-
vated PbB's than among those who don't (Mooty, et al. 1975).
There is strong evidence that paint is a major source of
lead in children with pica. Thus, Sachs (1.974) reported
that 80 percent of patients seen because of evidence of
excessive lead absorption had a history of eating paint
or plaster. Hammond, et al. (1977) reported that among
29 children with elevated PbB's (? 40 ug/dl) selected at
random from a lead screening program, all but one came from
14 homes classified as having high hazard for lead-base
paint, either exterior or interior (see Table 1). High
hazard consisted of there being at least one accessible
painted surface with > °*5 percent Pb, peeling or other-
wise loose. The medium classification consisted of _>0.5
percent Pb, but painted surface was generally tight. In
this study there was found to be a highly significant correla-
tion (p ¦ 0.007) between paint hazard classification (low,
medium, high) and fecal lead excretion, but no correlation
between fecal lead excretion and traffic density in the
vicinity of the home (p » 0.41). Unfortunately, the correla-

Classification of Home Environments as to Lead Hazard3
Paint .
Lead Concentration, %
d. w. c
Vehicles 3
per d. x 10
2.5 - 5
10 - 15
2.5 - 5
2.5 - 5
L(I ); H(E)
0.5 - 1
M(I); H(E)
L(I), H(E)
5 - 7.5
aHammond, et al. 1977.
^ » high; M * medium; L ¦ low; (I) » interior; (E) » exterior. Absence
of (I) or (E) designation means that both conformed to the designated clas-
sification of H, M or L.
cNumbers in parentheses indicate number of environmental samples.

tion between traffic density and the lead content of soil
and dust was not determined. Thus, the data are merely
Ter Haar and Aronow (1974) reported that elevated lead
exposure in a series of eight children hospitalized for
excessive lead absorption could not be accounted for by
lead from fallout of airborne combusted auto exhaust. Six
of the eight children had distinctly elevated fecal lead
excretion as compared to nine control children. Yet their
excretion of Pb, a marker for aerosol fallout, was no
different from that of controls. However, the children
admitted to this study were selected to have ingested paint.
The criteria were one of 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 suggests that
dust and soil are, under some circumstances at least, signifi-
cant sources of lead for infants and children and that their
effect is additive to that produced by inhalation. The
best evidence is provided in a study of a population of
children residing in the immediate vicinity of a large secon-
dary lead smelter near El Paso, Texas (Landrigan, et al.
1975). Sixty-nine percent of one-to-four year old children
living within 1 mile of the El Paso smelter had blood lead
levels greater than or equal to 40 jig/dl, the level then
considered indicative of increased lead absorption. By
contrast, the prevalence of blood lead levels greater than
or equal to 40 /jg/dl among 98 adults living in the same
area was 16 percent. The geometric mean lead concentration

of soil there was 1,791 ppm 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 considerably increased risk of lead
uptake 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 containing 8,000 pgPb/g showed a considerably
lesser elevation of PbB than was found in the El Paso study
(Barltrop, el al. 1974). This may be explained by other
factors, e.g. rainfall and soil composition. El Paso, Texas,
is a hot, dry, windy town whereas Britain has considerable
rainfall, probably resulting in a heavy protective cover
of vegetation.
Certain miscellaneous sources of lead are unique to
children by virtue of the pica habit. These include colored
newsprint (Joselow and Bogden, 1974) and other items to
which lead-base pigment is applied.

In characterizing the accumulation of lead in the body
under various circumstances of exposure, experimental animal
data are useful for establishing relevant principles. The
specific rates of transfer into, within, and out of the
animal system cannot be relied upon, to reflect with any
reliability the situation in man. Only human data will
serve to indicate how much lead, in what form, and by what
route will lead to the accumulation of one or another concen-
tration of lead in specific organs and systems. This restric-
tion has imposed severe limitations on knowledge concerning
lead metabolism in man. Only certain human biological fluids
and tissues are accessible for sampling, except after death.
The human cadaver, in turn, has its own limitations, 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 limita-
tion. They provide a substantial amount of knowledge concern-
ing the subject, but extrapolation of the data to the general
population is hazardous. Population studies materially
overcome this restriction, but at the expense of precision
and detail of knowledge. By combining data from all sources,
a reasonable understanding of lead metabolism does emerge,
In reviewing the metabolism of lead in man, it is gener-
ally assumed that all inorganic forms behave in the same
manner once absorbed. There is no evidence to suggest that
this assumption is erroneous.

Gastrointestinal tract
The classic studies of lead metabolism in man, conducted
by R. A. Kehoe and colleagues (Kehoe, 1961) indicate that,
on the average and with considerable day-to-day excursions,
approximately 8 percent of the normal dietary lead (includ-
ing beverages) is absorbed. This conclusion was reached
as a result of long-term balance studies in volunteers.
Recent studies using the non-radioactive tracer Pb have
confirmed 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 (Wetherill, et al. 1974).
This finding has been confirmed in mice using small doses
of lead (3 pg/kg) but not when using large doses (2000 pg/kg)
(Garber and Wei, 1974). Thus, lead in water and other bever-
ages 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 chil-
dren is considerably greater than in adults. Alexander,
et al. (1973) found that dietary lead absorption was approxi-
mately 50 percent in eight healthy children 3 months to
8.5 years of age. This finding has been confirmed using
a larger number of subjects less than 2 years of age (Zieg-
ler, et al. 1978). It is worth noting too 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; Six and Goyer,
1972). Lead absorption has also been shown to be enhanced
in experimental animals by high fat, low protein, 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 calcium 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, relative to the absorption of lead acetate,
lead naphthenate, lead octoate, and lead sulfide were absorbed
only two-thirds as well as lead acetate and that elemental
lead particles, 180 to 250 jum, were absorbed only about
14 percent as well. Lead thallate and lead carbonate 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 monkeys that lead octoate in dried
ground paint is absorbed only one-third as well as lead
octoate not incorporated into paint (Kneip, et al. 1974).

Respiratory tract
There are serious problems in regard to assessing the
absorption of lead via this route. The fractional deposi-
tion of inhaled aerosols is relatively easy to measure,
even in man. The problem lies in determining the fate of
the aerosol particles. To varying degrees, depending on
their solubility and particle size, these particles will
be absorbed from the respiratory tract into the systemic
circulation or they will be transferred to the gastrointes-
tinal tract by swallowing following either retrograde move-
ment up the pulmonary bed or by drainage into the pharynx
from the nasal passages. Unfortunately, the particle size
distribution and solubility of lead aerosols varies tremen-
dously, depending on their origin and residence time in
the air. All of these difficulties 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 problem, where-
by the impact of air lead concentration on the blood lead
concentration is measured. In order for this approach to
be meaningful, 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 impact of dietary lead on the subject's
PbB's. Second, it is necessary to monitor the air breathed
by the subjects continuously and for a substantial period
of time. Third, the subjects must have been in the air

environment being evaluated for at least three months in
order to assure reasonable equilibration of air lead versus
PbB. If all these conditions are achieved, the results
are only applicable for the particular type of lead aerosol
under study. Thus, it would not be reasonable to extrapo-
late data obtained in a population breathing city air to
a population of industrial workers for whom the greatest
source of input might be lead oxide fumes. Needless to
say, these restrictions are so severe that very few good
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 assess-
ment of available information is deferred to the end of
this section on lead metabolism.
Very few studies concerning the dermal absorption 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 tetraethyllead was absorbed to a
substantially greater degree than lead arsenate* lead oleate,
or lead acetate (Laug and Kunze, 1948). Differences in
the degree of absorption among the oleate, arsenate, and
acetate were not significant. In a more recent study, absorp
tion of lead acetate and lead naphthenate through the intact
skin was demonstrated, based on concentrations of lead at-
tained in various organs as compared to controls (Rastogi

and Clausen, 1976). There 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 (0.24M).
Distribution and Retention
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,
approximately 95 percent of the total amount of lead in
the body (body burden) is localized in the skeleton after
attainment of maturity. By contrast, in children only 72
percent is in bone (Barry, 1975). The amount in bone in-
creases with old age but the amount in most soft tissues,
including the blood, attains a steady state early in adult-
hood (Barry, 1975; Horiuchi and Takada, 1954). Special
note should be made regarding the kinetics of lead distribu-
tion with reference to the blood, when human volunteers
are introduced into a new air environment containing substan-
tially higher concentration of lead than the previous 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; Coulston, et
al. 1962). 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 disappear-
ance half-time of only about 6 weeks. By contrast, the
rate of PbB decrement in workers who retire from the lead
trades is much longer (Haeger-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 exposure conditions,
if it ever is within the human life span.
The distribution of lead at the organ and cellular
levels has been studied extensively. In blood, lead is
primarily localized in the erythrocytes. The ratio of the
concentration of Pb in the cell to lead in the plasma is
approximately 16:1. Lead crosses the placenta readily.
The concentration of lead in the blood of the newborn is
quite similar to the maternal blood concentration. Studies
of the subcellular distribution of lead indicate that distri-
bution occurs to all organelles, suggesting that all cel-
lular functions at least have the opportunity to interact
with lead.
Upon entry into the body, lead compounds occurring
in the environment disassociate into lead. Therefore no
question of metabolism of the pollutant is involved. The
one exception is the family of alkyl lead compounds, princi-
pally tetramethyl- and tetraethyl lead. These are dealky-
lated to form trialkyl and dialkyl metabolites, which are
more toxic than the tetralkyl form (Bolinowska, et al. 1967).
The numerous studies reported in the literature con-
cerning routes of excretion in experimental animals indicate

wide interspecies differences. In most species biliary
excretion predominates in comparison to urinary excretion,
except in the baboon (Eisenbud and Wrenn, 1970). It also
appears that urinary excretion predominates in man (Rabin-
owitz, et al. 1973). This conclusion, however, is based
on data from one volunteer.
Contributions of Lead from Diet versus Air to PbB
Great concern has developed in recent years regarding
the impact of air lead exposure on human health in the general
population. Analysis of the contribution of ambient air
to lead input in man has taken the form of an analysis of
air lead versus PbB for reasons explained in the section
on lead absorption. An analysis of all available data bearing
on this question first appeared in the Environmental Health
Criteria 3 Lead published by the World Health Organization
(World Health Organ., 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 or another 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 concentra-
tion over the range of interest. This is, unfortunately,
a problem seen in the clinical studies (as opposed to popula-
tion studies) where the number of subjects is quite limited.
Another problem, also limited to the clinical studies, is
the artificial nature of the lead aerosol utilized. In
spite of all these apparent limitations, calculations from

the epidemiologic and laboratory data sources indicate a
fairly narrow range of blood Pb to air Pb concentrations,
namely 1 to 4 fig/dl for every microgram of air lead per
cubic meter (jug/m ). This blood Pb to air Pb ratio appears
to be higher for children than adults.
Among all the studies, the only one that satisfied
all criteria for good design was the one by Azar, et al.
(1975a). It should be noted that the regression equation
developed to describe the data (log PbB = 1.2557 + 0.153
(log ^igPb/m3)) has a slope of less than one. Thus, the
incremental rise in PbB for each 1 jjgPb/m in air becomes
progressively smaller. This relationship is consonant with
experimental animal data showing that over a wide range
of dietary lead levels the incremental rise in PbB decreases
progressively proportional to the rise in dietary lead levels
(Prpic-Majic, et al. 1973; Azar, et al. 1975b). It also
is consonant with the World Health Organ, analysis of data
on air lead exposure in a battery plant (World Health Organ-
ization, 1977).

Estimated Blood Lead to Air Lead Ratios for Four Air Lead Concentrations3
Ratio at
lead concentrations
2ar	. c
oldsmi th
oldsmith	.
ankel-von Lindern
hambejlain -Williams
riff in
ross	d
iharoberlain -Kehoe
Si ze
3.5 5.0
Adult males
0.89 0.66
Adult females
0. 92
1.00 1.08
Adult males

Adult females


Adult males

Adult females

Adult males

Children males

Children females

1.27 1.37

Black females

Adult males
Adult males
Adult males
11 e 10.9
14 8 3.2
(21,000 person-days)
(1.7, 2.5)
i.S. EPA, 1977.
uthors regression equation evaluated at specific air
.S. EPA calculation	H c air lead-
uthors calculations
8 presented in Parentheses are not calculated from any regression equation.

Dose-effect relationships such as may be derived from
the Azar data are conceptually useful, but are not as satis-
factory as dose-response analyses wherein one can estimate
the proportion of people who would exceed any specific PbB
at any specific air lead level. Only the Azar data could
be used for this purpose. Dose-response relationships as
calculated from the Azar data by the U.S. EPA (1977) are
presented in Table 3.
Estimated Percentage of Population
Exceeding a Specific Blood Lead Level in Relation
to Ambient Air Lead Exposure
Percent exceeding
blood lead level of:
,ir lead,
15. 22
0. 59
26. 20
5. 35
aAzar, et al. 1975.

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 contri-
bution to PbB was attempted by World Health Organization
recently (Table 4).
Comparison of Daily Oral Lead Intake With PbB Levels3
iudy Design	Oral Intake PbB	PbB per Reference
(yjg/day) (^g/100 ml) 100 /jg
oral Pb
^plicate portion
^cal excretion
bplicate portion
aplicate portion
hnposites technique
mo, 1977.
Contributions of air to PbB levels are not reported in most of these studies
(ind could not be subtracted from total PbB levels.
Calculated from daily fecal excretion of 108 ^ig of lead assuming gastro-
intestinal absorption 10%.
NbB levels from Secchi, et al. (1971).
18. 3
Coulston, et al. (1972)
15. 3
] 3. 0
Tepper & Levin (1972)
12. 3
Nordman (1975)
Nordman (1975) .
34. 6
Zurlo & Griffini, 1973

The great discrepancy between American and European
data cannot be explained. It should be noted, however,
that the Coulston subjects were prisoners, whose diet per-
haps is far from typical and that the Tepper and Levin data
were based on fecal lead excretion, adjusted upward to compen-
sate for an assumed 10 percent lead absorption. The Euro-
pean data are impressive in that they are consistent among
studies over a fairly wide range of PbB's and dietary intake
levels, with a range of 4.4 to 6.8 and an average of 5.5.
So far as the contribution of water specifically is
concerned, information is even scarcer 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 published in 1976 (Elwood, et al. 1976). A linear regres-
sion 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 (^jg/dl) = 19.6 + 7.2 (mg Pb/1 water)
The regression selection seems inappropriate from examina-
tion 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 >ig/dl, a rather high baseline
value for non-occupationally exposed women.

4S0 r
40 0 ¦
6 aod liQtf
; vl •/«)
Figure 1. Regression of blood-lead on morning water lead
in Caernarfonshire (Elwood, et al. 1976).
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. study demonstrated that there is
a curvilinear relationship between PbB and the concentration
of lead in "first flush" water (Figure 2).
Figure 2. Mean blood-lead values for nine groups at intervals of
first-flush water lead (Moore, et al. 1977a).

The equation for the regression line was x = 0.533 + 0.675
y, with both values being expressed as jjmol/1. 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
PbB better than "first flush" water. They found that the
same relationship held, wherein mean blood lead rose in
proportion to the cube root of running water lead. The
correlation of running water lead to PbB was even somewhat
better than that of "first flush" water to PbB (r ¦ 0.57
vs. = 0.52). According to the authors, running water lead
concentrations were approximately one-third the "first flush"
lead concentrations. These data are useful in that they
provide an estimate of the consequences of changing the
concentration of lead in water from one value to another.
The example provided is the PbB consequence of going from
a "first flush" concentration of 0.24 pmol/1 (50 ug/1) to
0.48 jumol/1 (100 jig/1). Such a change results in an incre-
mental rise in PbB of 0.11 jumol/1, or of 2.3 pg/61. On
a running water basis, the PbB change would occur going
from 24/3 or 8 jug/1 to 48/3 or 16 jig/1. Using the authors'
equation, the effect of lead in running water on PbB can
be estimated (Table 5).

Effect of Running Water Lead on PbB Based on
Results of the Moore, et al. (1977a) Study
Pb in levels
(jumol/1) (y )
(y^) Pb levels in
running water3
PbB due to
ay = (jug of Pb/1 of running water) X 3
If this relationship is correct, the impact of water lead
on PbB is extremely great in the lower ranges o£ water lead
but diminishes rapidly in the higher range of water lead
= 50 to 100 >ag/l.
Hubermont, et al. (1978) also reports the interaction
of morning tap water lead to PbB in pregnant women of the
household. Again, as in the study of Moore, et al. a curvi-
linear relationship is described for the interaction of
PbB with water lead:
PbB = 9.62 + 1.74 log morning water Pb, (jig/1)
The correlation was good (r » + 0.37; p 0.001). The calcu-
lated impact of water Pb on PbB using this equation is consi-
derably less in the lower range of water lead than in the
Moore, et al. study. The data may not be strictly comparable
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 relationship 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, Massachusetts). In addition, blood samples
for members of the household were analyzed for lead. These
subjects included both children and adults. Numerous vari-
ables that might have influenced PbB were measured, including
age, sex, traffic density, lead in dust, and socio-economic
status. The data for interaction of PbB and water Pb were
reevaluated by 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 publi-
cation 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 Moore model (PbB = a +
b 3>/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 running water Pb was almost identical to the
3:1 relationship reported by Moore, et al. (1977a). More
precisely, the relationship was:
early morning water Pb » -0.028 + 3.081 running water Pb
r2 - 0.235; p » 0.0001

The cube root model of Moore, et al. (1977a) was more
appropriate 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 subjects 20 years of age and over
and for women age 20 to 50 was striking:
females 20 to 50, n = 249
PbB ¦ 13.38 + 24.87 ^/running water, Pb, pg/1
p = 0.020
all subjects 20 yrs +, n = 390
PbB = 14. 33 + 25.41 ^running water, Pb, )ig/l
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 foundation 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/1. 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 3:1 in two of the three studies
and therefore probably is approximately correct for the
third study, the one by Hubermont, et al. (1978). It is
seen that the impact of lead in water on PbB is quite dif-
ferent among the three studies. Since there is no basis

for rejecting any of the three studies, it can only be sur-
mised that an average of the three sets of data is as good
an estimate of the average situation as any. The reasons
for the variation in the relationships can only be left
to speculation.
PbB Levels due to Water Lead as Predicted by the Moore, et al. (1977a)
and the Hubermont, et al. (1978) Studies
PbB due to water
et al/6
et al.
all 3
2. 26
4. 07
6. 88
8. 57
10. 64
^Calculated from the data of Greathouse and Craun, (1976).
These values were all calculated using morning or "first flush"
water values which were taken to be three times the running water
levels in the table.
Certainly the calcium, phosphate, and iron concentrations
of the waters in the three studies were different and may,
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
absorption, even over a relatively narrow range. For example,
Ziegler, et al. (1978) demonstrated that a mere doubling
of the dietary calcium level profoundly depressed lead absorp-
tion in infants. Also, animal studies have shown that nutri-
tional iron deficiency enhances lead absorption. Attention

should be given to the significance of the variations in
calcium and iron content of water against the background
variations of calcium and iron content in non-water elements
of the diet. As with calcium, high phosphate levels also
tend to depress lead absorption.
The effects of lead on man will be reviewed in a selec-
tive fashion. Greatest emphasis will be placed on those
effects which occur at the lower levels of exposure and
those which are properly viewed with the most concern, namely
neurobehavioral effects, carcinogenesis, mutagenesis, and
teratogenesis. Because of the paucity of data in man and
the seriousness of the effect, some sections will be specifi-
cally subdivided into subsections dealing with human data
and animal data. In other cases that does not seem neces-
sary because of the wealth of human data available.
Hematological Effects
There is a vast literature concerning the effects of
lead on the formation of hemoglobin and more limited litera-
ture on the related effects on other hemo-proteins. From
the standpoint of standard setting, the effects of lead
on this system are particularly important 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 circulating concentration of hemoglobin is not thoroughly
understood. Many specific abnormalities exist, some occurring

at lower PbB's than others. The lifespan of erythrocytes
is shortened in heavy lead exposure (PbB = 59 to 162) (Hern-
berg, et al. 1967). The mechanism is not well understood,
but damage to the erythrocyte membrane is likely. Dose-
response and dose-effect relationships have not been estab-
lished. It seems unlikely, however, that shortened cell
life is the most sensitive locus of lead-induced reduction
in circulating hemoglobin. 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 produc-
tion (Piddington 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 syn-
thesis of heme (Fig. 3).
Acid (ALA)
ALA Dehydrase (ALAD)
Uroporphyrinogen HI
Figure 3. Effects of Lead on Heme Metabolism

Clear evidence exists that lead inhibits both d-aminolevulinic
acid dehydrase (ALAD) and heme synthetase both Jjn vitro
and ^n vivo at relatively low levels of lead exposure.
Elevation of the concentration of the substrates for these
two enzymes in plasma and urine (ALA) and in erythrocytes
(PROTO) increases as PbB increases. As a matter of fact,
rises in PROTO and ALA occur at PbB's somewhat below those
associated with a decrement of hemoglobin. Thus, in adults,
a decrement in hemo-globin first appears at PbB = 50 (Tola,
et al. 1973) and at PbB = 40 in children (Betts, et al.
1973; Pueschel, et al. 1973), 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 (Natl. Acad. Sci/Natl. Res. Counc., 1972)
and somewhat lower in women (Roels, et al. 1975). Rises
in PROTO first appear at PbB » 15 to 30 in women and children
and at PbB = 25 in men (Sassa, et al. 1973; Roels, et al.
1975). The most reasonable explanation for the rise in
PROTO at levels of lead exposure 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
Pe into the mitochrondrion (Jandl, et al. 1959). Regardless
of that uncertainty, the effect is the same, a potential
decrement in hemoglobin which leads to feedback depression
of ALAS resulting in a compensatory increase in the production
of ALA and other heme precursors. The evidence for this
conpensatory adjustment is to be found both in laboratory

animal studies (Strand, et al. 1973; Suketa and Yamamoto,
1975) and in studies of people with elevated lead exposure
(Berk, et al. 1970; Meredith, et al. 1977). The approximate
threshold £or ALAD inhibition is PbB = 10 to 20 for adults
(Tola, 1973) and PbB = 15 in children (Granick, et al. 1973).
Inhibition of roughly equivalent degree occurs concurrently
in the liver of man (Secchi, et al. 1974) and in the liver
and brain of rats (Millar, et al. 1970). The toxicological
implications of ALAD inhibition have not been studied exten-
sively. However, substantial lead-induced depression of
blood ALAD activity in dogs does not reduce the blood-regen-
erating response to acute hemorrhage in dogs (Maxfield,
et. al. 1972).
A few studies have been reported concerning effects
of lead on hemoprotiens other than hemoglobin. Thus, the
rate of cytochrome P450-mediated drug metabolism has been
found to be depressed in 2 cases of lead poisoning (PbB
» 60 & 72) but not in 10 cases where 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 Goyer, 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 hemolysis. The possibility of increased
susceptibility of G-6-PD-deficient children to the hemato-
poietic toxicity of lead has not been reported. In regard

to possible enhancement of hemoglobin deficiency by coexis-
tent 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 jjg/dl as compared to 17 iron-deficient children
with PbB = 20 to 40 jjg/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.
Dose Response Relationships for the Effect of Lead
on Various Parameters of Hemotological Indices®
Percentage of adult female subjects
with FEP levels that exceeded those
found in control subjects with
	PbB «20 jjg/100ml
PbB level No.
(jug/100 ml)
% with PEP level
higher than normal
Percentage of children with FEP
levels that exceeded those found in
control subjects with PbB»20
n<3/i00 ml
PbB level No.
100 ml)
% with PEP level
higher than normal


Percentage of adult male subjects
with PEP levels that exceeded those
	with PbB »20 jug/100 ml
PbB level No. % with FEP
(jjg/100 ml) level higher
	than normal
Percentage of male adults with ALA-U
levels « 5 mg/litre and ¦ 10 mg/litre
	according to PbB level
PbB level No. ALA-P level (mg/litre)
-5	»10
(jjg/100 ml)
aZielhuis, (1975).

In considering these data, it is obvious that FEP (es-
sentially PROTO) elevation is a more sensitive correlate
of lead exposure than ALAU. It should also be noted, how-
ever, 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 FEP in lead-exposed
and non-lead-exposed children, Roels, et al. (1978) were
able to study the interaction of FEP and PbB in the absence
of anemia as indicated by serum iron concentration. They
proposed a maximum acceptable limit for FEP at PbB =» 25
^jg/dl. The maximum acceptable point was the mean FEP plus
two standard deviations for rural children, which equalled
79.2 jjg FEP/dl erythroytes. The PbB of these children was
9.1 jjg/dl + 0.5 with serum iron 50 pg/100 ml. This maxi-
mum is very similar to the maximum acceptable FEP which
would be calculated from the data of Piomelli at mean FEP
plus two standard deviations (PbB » 26 jjg/dl) cited in the
recent "Air Quality for Lead" (U.S. EPA, 1977a). As was
indicated earlier, the cooperative effect of iron deficiency
and lead exposure on FEP has not as yet been adequately
defined. There is just the one study by Angle, et al. 1975,
suggesting no interaction at PbB ¦ 20 to 40.
Neurological and Behavioral Effects
The syndrome of lead encephalopathy has been recognized
since the time of Hippocrates as occurring in workers in
the lead trades. The major features were dullness, irritabi-
lity, ataxia, headaches, loss of memory and restlessness.

These signs often progressed to delirium, mania, coma, convul-
sions, and death. The same general effects were also de-
scribed 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 suspected
and because children do not communicate signs and symptoms
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 consider-
able toll recorded among survivors in the form of long-term
neurological sequelae. Cortical atrophy, convulsive sei-
zures, and mental retardation were commonly reported (Perl-
stein and Attala, 1966; Byers and Lord, 1943).
The minimal level of lead exposure resulting in lead
encephalopathy is not clearly known and perhaps never will
be in light of the dramatic decrease in the incidence of
the disease, particularly during the last 10 to 15 years.
Drawing mainly from his own experiences, Chisolm has esti-
mated the minimal PbB associated with encephalopathy as
being 80 ug/dl (NAS/NRC, 1972b). There are occasional reports
however of occurrence of encephalopathy at PbB's below 80
pg/dl (Smith, et al. 1938; Gant, 1938). Although 80 >ig/dl
may be a reasonable estimate of threshold for encephalopathy
in children, the usual values are much higher, with a mean
of approximately 328 according to one source (NAS/NRC, 1972).

It has been reasoned that if lead exposure as specified
above can have such severe deleterious effects on the cen-
tral 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
PbB's are in the 40 to 80 )iq/dl range. Given the difficul-
ties of study design, it is hardly surprising that all of
the relevant studies are open to criticism. The most common
deficiencies encountered are overlap of lead exposure in
the study groups (Pb versus control), inadequate matching
for socio-economic status and other variable, insensitivity
of the behavioral tests, and poor knowledge of the degree
of lead exposure. In regard to this last-named problem,
the index of exposure has usually been PbB's determined
at the time of behavioral testing. In some instances record
of one earlier PbB determination was available. In spite
of these problems, when the various studies are taken to-
gether, subtle neurobehavioral effects do appear to occur
as a result of exposure in the range of PbB ¦ 40 to 80 ug/dl.
Two general approaches have been used in attacking
the problem. 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 expo-
sure. These studies are retrospective and usually strictly
cross-sectional. In only one instance was a follow-up repeat
study of the population performed (de la Burde and Choate,
1972; de la Burde and Choate, 1975). The other general
approach has been to identify children with neurobehavioral

deficits of unknown 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, there has been the additional question
as to which came first, the excessive lead exposure or the
neurobehavioral deficit. Among mentally subnormal children
whose problems were clearly attributable to etiologies other
than lead, pica incidence and PbB's were both elevated (Bick-
nell, et al. 1968).
Among studies of the first type, those of de la Burde
and Choate are illustrative of the problems that exist in
this area of toxicology. Pine motor dysfunction, impaired
concept formation, and altered behavior profile were observed
in 70 preschool children exhibiting pica and elevated PbB's,
all of which were < 30 jug/dl. The mean level was 59 jiq/dl.
The children were examined at 4 years and again at 7 years
of age. Both the lead-exposed group and the control group
had been followed from infancy through 8 years of age as
part of a Collaborative Study of Cerebral Palsy, Mental
Retardation, and Neurologic Disorders of Infancy and Child-
hood. Unfortunately, the control group did not have blood
lead analyses performed. However tooth lead and urinary
coproporphyrin determinations ultimately were performed.
Another problem was that positive radiographic findings
of lead in long bones and/or intestines were inferred to
have been found in subjects with PbB's in the range of 30
to 40 jig/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 PbB associated with neurobehavioral
effects may well have been more on the order of 50 to 60
pg/dl rather than 30 to 40 ^jg/dl. Overall, the experimental
design was otherwise generally sound.
Another oft-quoted study by Perino and Ernhart (1974)
was basically of the same general design as the one reported
by de la Burde and Choate. It concluded that neurobehav-
ioral deficits occurred at PbB's as low as 40 jiq/dl. The
flaw in this study was that the parents in the control group
were better educated than those of the lead-exposed chil-
dren. Differences found may have been due to the fact that
more highly 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
suggests that a factor other than parental influence was
operating and probably was lead exposure.
Albert, et al. (1974) studied school-age children who
had had PbB's >60 jig/dl early in childhood. Unfortunately,
PbB's for about one half of the control population were
not available and some of the control children had had PbB's
< 40 )ig/dl.

The same types of flaws existed in studies which came
up with negative results. Thus, Kotok's study (Kotok, 1972)
had a rather wide overlap between PbB's of control subjects
and lead-exposed subjects, and in another negative study
fewer than half of the "lead-exposed" group had PbB's > 40
jjg/dl (Lansdown, et al. 1974). Another problem among negative
studies has been the study of perhaps inappropriate popula-
tions. Lansdown's population consisted of British children
living in the vicinity of a smelter. In another negative
study, the children were Mexican-Americans also living in
the vicinity 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 neurobehavioral effects of lead exposure occur
in children exposed to lead at levels which do not result
in clinical encephalopathy. The minimal level of lead expo-
sure, the duration of exposure required, and 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 great depth
are remarkably similar. The World Health Organization con-
cluded that the probability of noticeable brain dysfunction
increases in children from PbB levels of approximately 50
jiq/dl (WHO, 1977), and the U.S. Environmental Protection

Agency's Science Advisory Board concurred in the U.S. EPA
conclusion that "the blood lead levels associated with neuro-
behavioral deficits in asymptomatic children appear to be
in excess of 50 to 60 pg/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 pg/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 major public
health problem today, the potential damage to the brain
of the fetus from lead exposure has received some attention.
Beattie, et al. (1975) identified 77 retarded children and
77 normal children matched for age, sex, and geography.
Of 64 matched pairs, 11 of the retarded children came from
homes in which the concentration of lead in the "first flush"
water exceeded 800 jug/1. By contrast, none of the control
children came from such homes. In a follow-up study, PbB's
from the mental retardates, taken during the second week
of life, were found to be significantly higher than those
of control subjects (25.5 jug/dl versus 20.9 jig/dl) (Moore,
et al. 1977b). Taken at face value, those studies are ex-
tremely provocative. They suggest that the brain of the
fetus is considerably more sensitive to the toxic effects
of lead than the brain of the infant or young child. Lambs
exposed to low levels of lead Jjn utero (PbB « 35) developed
impaired visual discrimination learning behavior (Carson,

et al. 1974). In spite of this seemingly low level of expo-
sure, control animals were exposed ni 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 synaptogenesis in the cerebral cortex
(McCauley, et al. 1977) were delayed by this treatment.
These delays were associated with delays in the development
of exploratory and locomotor behavior during the same develop-
ment 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 jjg/dl. Further work is urgently needed concerning
the neurobehavioral effects of low-level lead exposure in
Finally, a few comments are in order regarding neuro-
behavioral effects of low-level exposure in adults. A bat-
tery of performance tests were administered to 190 lead-
exposed workers, along with a questionnaire concerning af-
fected (Morgan and Repko, 1974). PbB's were below 80 jug/dl
in many of the workers. Unfortunately, there were many
methodological problems and equipment 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 neurobehavioral effects do occur
at some level of exposure in workers, it is extremely diffi-
cult to specify the exposure level at which these effects
may occur.

Human Studies
Three groups of investigators have reported epidemio-
logical studies of causes of death among people overly exposed
to lead. The first such study was of causes of death among
184 pensioners who died between 1926 and 1961 and of 183
men who died between 1946 and 1961 while still employed
(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 jig/dl during the past 20 years
and probably higher than that earlier in the work history).
There is a correlation between urinary lead and blood lead,
wherein 100 yq Pb/1 in urine corresponds roughly to 50 jjg/dl
in blood (Selander and Cramer, 1970).
There were 179 men in the high exposure category for
which causes of death were registered, 67 men in the cate-
gory of negligible exposure and 91 men with no exposure.
Although there was a significant excess number of deaths
among the men who had been exposed to the greatest lead
hazard, this excess could not be attributed to malignant
neoplasms, as the mortality rate from this cause was actual-
ly somewhat less than expected. Furthermore, the incidence
of death from malignant neoplasms in this group has actually
increased in the more recent years as working conditions
have improved. 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

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 Health Service. The population was classi-
fied as to exposure on the basis of whether they were adult
orchard workers, (orchardists and lesser-exposed "intermedi-
ates" as separate categories), non-exposed adults of the
area, and children in the area. For all categories blood
lead and urine lead and arsenic concentrations were deter-
mined. 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 "inter-
mediates" had PbB's of 22 to 30.
In 1968 a follow-up study of this population was begun.
Results were reported in 1973 (Nelson, et al. 1973). Of
the original 1229 study members, the status of 1175 could
be determined. Four hundred and fifty-two had died and
death certificates were available for 442. No consistent
differences in Standard Mortality Ratios (SMR) were observed
on the basis of either exposure classification or duration
of exposure. The only deviations in SMR from expected were
in the direction of fewer-than-expected deaths. The morta-
lity record for heart disease, cancer, and stroke were exa-
mined separately. Again, there was no suggestion of a rela-
tionship between lead exposure and death from any of these
three major causes of death.

The most recent study of causes o£ death among lead-
exposed workers was reported by Cooper and Gaffey (1975).
Since the results were published, the study population has
been reexamined (Cooper, 1978). Results from the updated
study will be discussed, although details as to lead expo-
sure history appear mainly in the 1975 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-recognizable illness but
above that of the general population. The population 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 impor-
tant. 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, enough data were available to indi-
cate that exposure was heavy. Thus, 67 percent of 1863
workers had PbB's^ 40 ug/dl and 20 percent had PbB's 70-Z~
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
significant elevation were "all malignant neoplasms" in
the battery workers, 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 specified. 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 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 esti-
mated exposures to lead." It is not clear from reading
either of the two reports concerning this population as
to just how exposure categories were established.
Animal Studies
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 (Boy-
land, et al. 1962; Van Esch, et al. 1962; Roe, et al. 1965;
Mao and Molnar, 1967; Oyasu, et al. 1970). The same observa-
tion has also been reported in mice but could not be eli-
cited in hamsters (Van Esch and Kroes, 1969). Other studies
indicate that lead also causes lung tumors in hamsters (Kobay-
shi 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. In
its assessment of the available literature the International
Agency for Research on Cancer (IARC, 1972) commented as
It must be noted that the level of human exposure
equivalent to the levels of lead acetate producing
renal tumours in rats is 810 mg per day (550 mg
Pb). This level appears to exceed by far the maxi-
mum tolerated dose for man.
As will become apparent in the discussion of the seven
substantive papers dealing with experimentally induced can-
cer, signs of lead poisoning and even high mortalities often

supervened using the cancer-inducing dosage regimens. It
will also become apparent that none of the studies were
designed in anticipation of subsequent use of the data for
the purpose of extrapolating to low incidence doses. Few
of the studies utilized more than one dosage level and only
one utilized more than two dosage levels. An additional
problem is the non-comparability of the modes of administra-
tion from one study to another. This makes it impossible
to pool data or to compare dose-response curves for consistency.
The first report of lead-induced renal tumors (Zollinger,
1953) was essentially a lifetime study in rats, with admini-
stration 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 ani-
mals 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 papillary adenomas.
Metastases were evident in only one case, according to the
text. All the animals receiving lead had severe lead intoxi-
cation, according to the author's histological criteria
as applied to the kidneys. Among 50 control animals, none
developed tumors.
The next study reported (Boyland, et al. 1962) tested
the hypothesis that renal cancer due to lead was actually
caused by the well-known accumulation of porphyrins associ-
ated with lead toxicity. To test the hypothesis, elevated
prophyrin excretion was stimulated by administration of

allyl-isopropylacetamide (AIA) in the diet of 20 rats Cor
one year. A like number of rats were fed 1 percent lead
acetate in their diet for one year. Both groups of animals
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. Subsequently to
the 1-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 com-
pared to its own set of controls, not receiving lead. Prior
to the termination of the experiment, only moribund animals
were killed and examined morphologically. At equivalent
durations of lead administration, using these guidelines
for tumor assessment, the higher dose of lead was more carcin-
ogenic than the lower dose. Thus, at the end of 600 days
of lead administration, 31 percent of the animals which
survived to 400 days died with renal tumors in the 1.0 per-
cent lead acetate group, whereas only 14 percent of the
animals alive at 400 days in the 0.1 percent lead acetate
group died with renal tumors (see Figure 4). Mortalities
with tumors in the subsequent 200-day period (600 to 800)

Cumulative %
Mortality (©)
or % Animals c
Tumors at Time
of Death (&)
i r
Total n - 29
0.1% PbAc
1 I
Total n » 26
1.0* PbAc
401 601 0 201 401
•Ir	4; 4^ 4*
600 729 200 400 600
Figure 4. Cumulative mortality and tumor incidence in rats
(Van Esch, et al. 1962).

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 (see Table 8) that during the first
600 days of the 0.1 percent basic lead acetate regimen,
10 of the original 26 rats (38 percent) died without renal
tumors as compared to 1 of the original 26 in the control
group (4 percent), indicating that at this level the lead
regimen was of itself lethal in some manner unrelated to
its carcinogenicity. As a matter of fact, both levels of
lead administration caused reduced body weight gains, suggest-
ing toxicity unrelated to carcinogenesis.
The next study of lead-induced tumors in rats was also
designed 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 form of lead, lead
orthophosphate, and the mode of administration were unique.
The lead salt was administered subcutaneously once weekly
for 4 weeks, then intraperitoneally for 9 weeks; then after
a rest period of 9 or 4 weeks, depending on the particular
group of rats, lead administration was resumed for an addi-
tional 14 weeks. All the animals were males. The dosage
schedule of lead is presented below, assuming an average
body weight of 400 g, and averaging the dose over the total
treatment period (Table 9).

Effect of Lead Exposure on the
Incidence of Renal Tumors in Ratsa
Successive Time Intervals, Days


0. lc
0 beginning of interval

dead, no renal tumors

dead, renal tumors
§ beginning of interval

dead, no renal tumors

dead, renal tumors

• 1
»H i

@ beginning of interval

dead, no renal tumors

dead, renal tumors

§ beginning of interval

dead, no renal tumors

dead, renal tumors

fvan Esch, et al. 1962.
C » Control
.0.1 = 0.1% basic lead acetate in diet
1.0 = 1% basic lead acetate in diet

Dosage Schedule used by Roe, et al. (1965) in their Study of the
Influence of Testosterone and Xanthopterin on the Induction of
Renal Neoplasms by Lead in Rats
Days on Pb
Pb alone
Pb alone
1. 25
Pb alone
Pb + testosterone
Pb + xanthopterin
Pb + testosterone
Pb + xanthopterin
No treatment
In analyzing the cancer data for these groups, it seems
reasonable to pool all the groups receiving the same dosage
of lead since neither testosterone nor xanthopterin influ-
enced the tumor incidence. However, xanthopterin alone
seemed to increase the mortality 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 interac-
tion 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 (Fig. 5). However, since
only one of these two remaining dosage levels was tumori-
genic, no dose-response relationship in regard to turaori-
genesis is calculable.

Summary of Mortality Data Resulting from Lead Phosphate Administration to Ratsa
Successive Time Intervals, Days
6 - 100	101 - ioo	ZOI - 360	361 - 466 ~ 101-506	561-466	601 - 700"

c 2.6
n at beginning
of interval

dead, no renal
11 0

dead, renal

in interval
dying with

no tumors
100 100

*Roe, el al. 1965
= controls
""average dose of lead phosphate, mg/kg/day

Figure 5. Cumulative mortality among rats not having renal
tumors (Roe, et al. 1965).

Interstitial nephritis occurred in all groups, including
controls. Unfortunately, other manifestations of toxicity,
e.g., anemia, reduced body weight gains, and food consump-
tion were not reported. In keeping with the observations
of Van Esch, et al.; Boyland, et al.; Mao and Molnar; and
Zollinger, very few of the affected animals exhibited metas-
tasis and no elevated incidence of other types of tumors
was noted.
Neither of the two remaining reports concerning the
carcinogenic effects of lead in rats (Mao and Molnar, 1967;
Oyasu, et al. 1970) involved more than one level of lead
administration. The results obtained by Mao and Molnar
serve to confirm the results of Van Esch, et al. in that
both groups used the same regimen of lead in the diet (1
percent lead acetate} and got similar incidences of renal
tumors (50 percent by Van Esch vs. 77.5 percent by Mao and
Molnar). Both also noted that the first appearance of tumors
was at about 300 days following initiation of lead administra-
tion. Mao and Molnar are the only authors who conducted
any lead analyses. They reported 19.3 to 54.2 jugPb/g kidney
cortex as compared to 3.1 /igPb/g in a single normal speci-
men. By way of comparison to man, Barry (1975) reported
a mean of 0.66 pg/g in kidney cortex of 10 occupationally-
exposed adult males, with a standard deviation of + 0.56

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 acetylaimofluorene (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-acetate in the
diet eventually developed renal tumors. This report, inci-
dentally, 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
Van Esch and Kroes (1969) have reported that basic
lead acetate causes renal tumors in mice, but not in ham-
sters. These were lifetime studies with lead being incorpor-
ated into the diet beginning at 5 weeks of age for the mice
and 3 to 4 weeks of age for the hamsters. Two levels of
lead were used, 0.1 percent and 1 percent, cut back to 0.5
percent early in the study owing to toxicity. Only one
renal tumor was found at the high 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 developed renal tumors. There were no renal
tumors in hamsters at either dosage level of lead. Mortal-
ity was somewhat increased at both levels of lead administration.

Another report of experimental carcinogenesis 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 predominant-
ly 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 speculate may be a preneoplastic change. BP
alone caused a very low incidence of alveolar metaplasia.
All treatments, including the methylcellulose injection
vehicle alone caused some deaths. The amount of lead per
dose was 1 mg as the oxide, or 0.92 mg as Pb. Assuming
a body weight of 100 g, this represents 92 mg/kg over 70
days or 1.3 mg/kg/day. By way of perspective, an adult
human breathing lead constantly at 6 >jg/m , which is a typi-
cal freeway air concentration would receive by tracheo-bron-
3	3
chial and alveolar deposition 6 /ig/ra x 20 m x 0.3 or 36
jjg/d or 36/70 or 0.5 jag/kg body weight per day. The ham-
sters were receiving 2600 times this amount in their lungs
on an equivalent body weight basis.
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 and Figure 6).

L.., f.-;.
r	• .1	\-
' f 1 ' ! 1 j i 'T7[ = ¦! I ¦1 [ : [ li; T ?.].
LJ-i-1—44-' ;1 f; : 'I —ij-/
5 99
0.5	1
ppm Dietary Pb x 10
Figure 6. Probit Plot of Incidence of Renal Tumors in Male
Rats (Azar, et al. 1973).

Mortality and Kidney Tumors in Rats Fed
Lead Acetate for Two Years
Dietary Pba No. of rats	% Mortality'3	% Kidney tumors
(ppm)	of each sex Male	Female	Male	Female
5	100	37	34	0	0
18	50	36	30	0	0
62	50	36	28	0	0
141	50	36	28	0	0
548	50	52	36	10	0
3	20	50	35	0	0
1130	20	50	50	50	0
2102	20	80	35	80	35
^Measured concentration of Pb in diet.
Includes rats that died or were sacrificed in extremis.
The dose of lead required to produce tumors did not clearly
result in increased mortality among the animals; however/
at dietary lead intake above 1000 ppm, weight gains were
reduced. Since a clear dose-response relationship was evi-
dent , it was possible to use the data to calculate a risk
assessment for cancer in man, utilizing the procedure sug-
gested by the EPA Cancer Assessment Group (see Methodology
Document). The safe dose for man was estimated to be 29
ug Pb/1 of water.

Summary and Evaluation of Animal Data
There is little doubt that lead is a carcinogen or
at least a co-carcinogen in some species of experimental
animals. There is some suggestion that bronchial adenomas
can be induced in combination with benzo-a-pyrene and that
oral administration of lead alone can induce gliomas. These
observations need confirmation. Furthermore, in both cases
there is no way of constructing a dose-response curve for
the effect since only one dosage regimen of lead was used.
Even in the case of the numerous studies that have demon-
strated the occurrence of renal tumors with lead administra-
tion, in only three cases has more than one dosage level
of lead been used. Among these three studies, two involved
rats and one involved mice. The study by Roe, et. al. 1965,
is not amenable to dose-response analysis. Three dosage
levels of lead were utilized, initially with a substantial
number of animals in each group (Table 9). Unfortunately,
75 percent of the animals at the high dose died of lead
toxicity early in the study, before any tumors were formed.
This reduced the population to six animals at 100 days and
to only two animals at the time the first tumor was observed.
Of the two remaining dosage levels, one was a no effect
level (no tumors), thereby leaving only one dosage level
at which some tumor incidence could be meaningfully assessed.
No slope could be estimated.
The second of the three potentially useful studies
for dose-response estimates was the one reported by Van
Esch, et al. 1962. These data are summarized in Table 8.
The data for the two lead-treated groups (0.1 percent and
1.0 percent) are comparable only up to day 730. Beyond

that point the groups were handled differently. The 0.1
percent lead group was carried on to day 985 before the
survivors were killed, whereas in the case of the 1.0 per-
cent lead group, the survivors were killed at about day
730. If one calculates mortality data up to the time at
which the 1.0 percent lead group was killed, a set of cumu-
lative mortality curves can be depicted (Pig. 4). Although
a comparison of mortality and tumor incidence is possible
at the two doses, the significance is uncertain because
of the small numbers of animals involved. The mouse data
of Van Esch's group (Van Esch and Kroes, 1969) are not use-
ful for constructing dose-response curves because only one
dose level of lead caused tumors. In summary, only the
study by Van Esch, et al. (1962) provides even the barest
hint of a dose-response relationship with reference to carcin
Teratogenici ty
The delivery into the world of a physically or mentally
abnormal child is as great a tragedy as cancer. There is
little information in the literature to suggest that lead
has a teratogenic effect in man. Although there were numer-
ous reports of a high incidence of stillbirths and miscar-
riages among women working in the lead trades, fetal anoma-
lies were not described. It must be pointed out too that
these women were probably exposed to much higher concentra-
tions of lead than for even occupationally exposed men today.
The more recent literature also is devoid of any references
to teratogenic effects in man.
In experimental animals, on the other hand, lead has
been shown repeatedly to have teratogenic effects. Early

studies demonstrated this in chick embryos by injection
of lead into the yolk sac (Catzione and Gray, 1941; Karn-
ofsky and Ridgway, 1952). Teratogenesis has also been observed
in rodents. These studies were done using high doses of
lead given intravenously or intraperitoneally. For example,
McClain and Becker (1975) used single interaperitoneal doses
of 25 to 70 mg/kg in rats. They found that teratologic
effects occurred when administration was on day 9. Admini-
stration later in pregnancy resulted in embryotoxicity (fetal
resorption) but not in teratogenic effects. Carpenter and
Ferm (1977) observed teratologic effects in hamsters fol-
lowing the administration of 50 mg/kg PbtNO^^ intravenously
on day 8. Chronic administration of lead in the drinking
water of pregnant rats at very high concentrations (up to
250 mg/1 resulted in delayed fetal development and fetal
resorption without teratologic effects (Kimmel, et al. 1976).
In summary, it seems that, in man, embryotoxicity pre-
cedes teratogenicity in the lead sensitivity scale. This
is supported by historical experience in occupationally
exposed women and by animal studies.
No information is available concerning mutagenicity
of lead.
Reproductive Effects
As was indicated in the previous section, lead has
been known to cause miscarriage and stillbirths in women
working in the lead trades during the latter half of the
19th century and probably on into the early part of the
20th century. It is very difficult to estimate minimally
toxic exposure for stillbirth and miscarriages because expo-

sure data, e.g., PbB are lacking for women who experienced
this problem. The minimally toxic level of exposure may
actually be quite low. Lane (1949) reported on the outcome
of 15 pregnancies incurred among 150 women working in an
unspecified lead 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 /jg/m^ and urinary lead excretion in men working
with these women being 75 to 125 }ig/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 in a lead mining area of Missouri (17 per-
cent) than in a Missouri urban area remote from lead mining
activities (0.41 percent) (Fahim, et al. 1976). Maternal
and fetal PbB's at birth also differed significantly for
normal births vs. births with premature membrane rupture.
Maternal and fetal PbB's for the normal deliveries were
about 14 and 4 jjg/dl respectively whereas they were about
26 and 13 respectively for mothers and infants with membrane
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 concept. Lancranjan, et al. (1975) reported
that the significant levels of teratospermia occurred among

men working in a lead storage battery factory. Their PbB's
were 30 to 80 jig/dl. Although many studies have attempted
to correlate semen quality with fertility, the extent to
which abnormally-shaped sperms participate in fertilization
is unclear. Experimental animal studies have shown reduced
fertility of both maternal and paternal origin (Stowe and
Goyer, 1971; Jacquet, et al. 1975).
There have been numerous conflicting reports concerning
the occurrence of chromosomal aberrations in lymphocytes
of lead-exposed workers. See, for example, O'Riordan and
Evans (1974) and 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 separated lead-using plants, one being a secondary
lead smelter and the other being a plant manufacturing "tin"
dishes. Lead exposures were roughly comparable PbB's were
of the order of 45 - 100 pq/61. Severe aberrations were
found in one plant whereas no such effects were seen in
the other. They further point out that no severe aberra-
tions have been seen in at least some animal studies in
which lead exposure was heavy and nutrition apparently ade-
quatre (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 reproductive casualty.

Renal Effects
There is considerable information in man concerning
the renal effects of lead. Two distinctive effects occur,
in both adults and children. One is reversible proximal
tubular damage, which is seen mainly with short-term expo-
sure. The other effect is reduced glomerular function which
has generally been considered to be of a slow, progressive
Tubular damage is manifested as the Panconi triad of
glycosuria, hypophosphatemia with phosphaturia, and gener-
alized aminoaciduria. The last-named manifestation appears
to occur more consistently 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 PbB's below
70 ^ig/dl. Thus, in a recent series of seven workers, all
of whom had PbB's 70 jjg/dl, with a range of 71-109, none
had aminoaciduria or glycosuria. Significantly, five had
hemoglobins below 12 g/dl (Cramer, et al. 1974). Similarly,
in a series of 15 infants hospitalized for lead poisoning,
and all having PbB's J 100 jjg/dl at entry only three had
aminoaciduria, with PbB's of 246, 299 and 798 /jg/dl (Chisolm, 1968).
Reduced glomerular filtration with attendant rise in
serus urea concentration is generally considered to be a
progressive disease, 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 ( 90 ml/min/1.73m ). In another study of eight men
with occupational lead exposure (PbB's = 29-98), four had
reduced glomerular filtration rates (Wedeen, et al. 1975).
Of these four cases, one had a PbB of 48 jag/dl at entry.
TmpAH was also reduced, indicating coexistent tubular damage.
Among the other three cases, 2 had only a marginal depression of Tin
From these and other studies, it appears that the kid-
ney is sensitive to glomerular-vascular damage, with an
imprecisely known threshold for effect which may be below
PbB = 50 pg/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 quar-
ter of the 20th century when lead exposure was considerably
higher than it has been more recently. There was no similar
elevated mortality among men employed 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 vasculopathy
(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 (Kosminder and
Pentelenz, 1962) attributable to lead exposure. However,
these cases have always involved clinical lead intoxication.
It does not seem likely, therefore, that the heart is a
critical target for lead effects.

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 reported
only at high exposure levels and only by one or two investi-
gators. For example, Dodic, et al. (1971) reported signs
of impaired liver function in 11 of 91 patients hospitalized
for lead poisoning. No information was provided as to indices
of lead exposure. Impairment of thyroid function 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 must have
been high, judging from the fact that the men were hospital-
ized. 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 PbB's as low as about 40 jjg/dl.
This seems unlikely since the cases he reported also were
anemic, a condition associated with the considerably higher
PbB's. A number of studies have suggested that a relation-
ship exists between lead exposure and amyotrophic lateral
sclerosis (ALS). The most recent report on this examined
plasma lead levels in 16 cases of ALS and in 18 controls
and found significant differences at the 0.05 level (Conraid,
et al. 1978).
Finally, animal studies indicate that relatively high
levels of lead exposure interfere with resistance to infec-
tious disease (Hemphill, et al. 1971; Gainer, 1974). There
are no reports of an abnormal infectious disease incidence
among people with high lead exposure, however.

Existing Guidelines and Standards
Since lead is ubiquitous in the environment, several
government agencies have become involved in regulating its
use. The most recent action was taken by the Consumer Pro-
duct Safety Commission (CPSC). In 1977 the CPSC lowered
the maximum allowable concentration of lead in housepaint
to 0.06 percent. At present the Occupational Saftey and
Health Administration (OSHA) is preparing a set of regula-
tions regarding occupational lead exposure. Similarly,
the U.S. EPA has set an ambient air lead standard. The
U.S. Food and Drug Administration also is due to hand down
in September, 1978 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 tapwater samples have been
found to exceed the current standard of 50 >jg/l. This is
generally a problem in softwater areas, particularly where
lead pipes convey the water supply to the tap from the sur-
face connection. The contribution of the diet is approxi-
mately 200 jug/day for adults. For children (ages 3 months
to 9 years) the diet contributes 40 to 200 ^ig 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
ambient air ranges from approximately 0.1 jjg/m in rural
areas to as much as 10 jug/m in areas of heavy automotive
Special Groups at Risk
In addition to these usual levels of exposure from
environmental media, there exist miscellaneous sources which
are hazardous. The level of exposure resulting from contact
is highly variable. Children with pica for paint chips
or for soil may experience elevation in blood lead ranging
from marginal to sufficiently great to cause clinical illness.
Certain adults may also be exposed 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 hematopoietic system is more
lead-sensitive than in men.
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 internal dose of lead that
can be tolerated without injury to the target 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 concentration 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 measure-
ment is necessary because of the multi-media character of
lead intake. It is virtually impossible to measure total
lead input in people in any meaningful way. To do so would
require long-term balance studies because past experience
has shown that intake and output fluctuate greatly from
day to day. Furthermore, it would be necessary to conduct
such studies on large numbers of free-living subjects, given
the influences of chemical and physical variables in the
numerous environmental forms of lead. 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
biological response to lead. As such, it is subject to
influences other than lead, notably iron deficiency. Tooth
lead is a potentially useful index of lead exposure, but
with the present state of art being what it is, tooth lead
is difficult to interpret. 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 spontane-
ously only in childhood. Beyond all that is the fact that
we have only a very small data base for dose-effect and
dose-response using any measure of dose other than PbB.
The use of PbB as a measure of internal dose is widely ac-
cepted, simply because nothing better is available.

Having specified that PbB is the best measure of internal
dose currently available, the next question concerns the
least PbB at which adverse health effects occur. Two recent
documents (U.S. EPA, 1977a; WHO, 1977a) have been published
in which judgments were rendered in this regard (Table 12).
It will be noted that the estimates are strikingly similar.
The estimated no-effects levels are based on limited popula-
tions and probably are lower to some undefinable degree
in the total population at risk. The expert panels that
made these estimates were largely composed of different
individuals, although there was some overlap.
Slightly more information was available to the U.S. EPA
panel than to the World Health Organization panel since
it reviewed literature only through mid-1977 whereas the
World Health Organization expert groups reviewed literature
through much of 1976. In addition, 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 hemo-
globin 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 PR 630979). It focused on the lead effects
in children since they are more sensitive than adults.

Summary of Lowest PbB's Associated with Observed
Biological Effects in Various Population Groups
Lowest observed
effect level	Effect	Population group
(pq Pb/100 ml blood)

ALAD inhibition
Children and adults

15 - 20
Erythrocyte protoporphyrin
Women and children

25 - 30
Erythrocyte protoporphyrin
Adult males

Increased urinary ALA excretion
Children and adults


Coproporphyrin elevation
Adults and children

- 60
Cognitive (CNS) deficits
- 60
Peripheral neuropathies
Adults and children
- 100
Encephalopathy symptoms
- 120
Encephalopathic symptoms
No-detected effect levels in terms of PbB
No detected
effect level

(jjgPb/100 ml blood)

Erythrocyte ALAD inhibition
adults, children
adult, female
adult, male
Erythrocyte ATPase inhibition
ALA excretion in urine
adults, children
CP excretion in urine
Peripheral neuropathy
Minimal brain dysfunction
Minimal brain dysfunction
aU.S. EPA, 1977
bWorld Health Organ., 1977

Quite properly, 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 jig Pb/dl).
In considering how much lower this limit should be, the
U.S. EPA cited the opinion of the Center for Disease Con-
trol, as endorsed by the American Academy of Pediatrics,
that the maximum safe blood lead level for any given child
should be 30 jjg/dl. Based upon epidemiological and statis-
tical considerations, the U.S. EPA estimated that if the
geometric mean PbB were kept at 15 jjg/dl, 99.5 percent of
children would have PbB's < 30 pg/dl. This position seems
prudent and reasonable. It provides a substantial margin
of safety which accommodates minor excursions in lead expo-
sure due to adventitious sources. Controls on lead in obli-
gatory media (e.g., air and water) do not of course 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. These,
however, are separate problems which must be dealt with
appropriately by responsible agencies.
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 10 to 12 jag/dl. This is
presumably composed overwhelmingly of dietary sources which,
in turn, is composed of both food and water.
The next question concerns the contribution of water
to lead exposure. It is unfortunate that only three useful
studies of the interrelationship between PbB and lead in
drinking water are available. There is an obvious need
for more such work. Overall, the Moore, et al. (1977a)

study, the one by Hubermont, et al. (1978), and the calcula-
tions 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 PbB and air Pb. The implication of
the equation 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. PbB and dietary lead vs. PbB (see section on "Contributions
to Lead from Diet Versus Air to PbB"). 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 independent of the contribution of
water lead to PbB. Thus, regardless of whether one starts
with a baseline PbB of 11 jug/dl, as was indicated in the
Moore, et al. study or whether one starts at some other
PbB level, e.g., 20 jig/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 Section on "Contribu-
tions of Lead from Diet vs. Air to PbB"). Here, the higher
the baseline, the less the contribution of any specific
air Pb. This is because log PbB (not PbB) is proportional
to baseline PbB + log air concentration. Future research
may provide better insight into whether this discrepancy
is real and, if so, why. The question is of some practical
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

PbB of 10 jjg/dl. An Azar-type model would suggest a lesser
impact starting from the higher baseline PbB.
So far as a specific recommendation regarding a revised
water standard for lead is concerned, one is tempted to
duck the whole issue by simply recommending more research.
However, that might defer the recommendation indefinitely.
A position must be taken using 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 /ig/1, approximately 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 ^g/1 instead of at 10 ;jg/l,
an additional contribution of approximately 3.4 ^jg/dl to
PbB would result (8.57 - 5.13 from Table 6). This would
yield a total PbB of 12 + 3.5 or 15.4 ^ug/dl, the approximate
maximum geometric mean PbB compatible with keeping 99.5
percent of the population under PbB » 30 ^ig/dl. Thus, based
on most recent data, the present water standard of 50 pg
Pb/1 may be viewed as representing the upper limit of accept-
ability. This criteria is based on empirical observation
of blood lead in human population groups consuming their
normal amount of water and food daily, specific amounts
of foods or dinking water consumed were not quantified,
but it can be assumed that they reflect an average consumption
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 Commission of the European Communities suggest
that the impact of lead in water on PbB is appreciably less
than has been estimated from published data used in this
document (personal communication from Alexander Berlin,
Commission of the European Communities, Luxembourg)'". Fur-
thermore, data (Table 13) 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 indicate that among children whose water supply contained
50 to 180 pq Pb/1, PbB's averaged 17.2 jig/dl (P.J. Landrigan,
Center for Disease Control, Atlanta, Georgia, personal commu-
nication ).
Relation of PbB to Lead in water
Among Children in Bennington, Vermont
Concentration of Lead	Number of	Mean Blood
in Water (pg/1)	Children	Lead Level (juq/dl)
50-59	14	18.9
60-69	8	16.9
70-79	10	15.0
80-89	4	18.5
100-109	3	16.0
110-119	2	21.5
130-139	2	16.0
170-179	_3	15.0
Total	46	17.2
aPhilip J. Landrigan, Center for Disease Control,
Atlanta, personal communication.
^Subsequent to the writing of this report, these data were
submitted to the EPA by Dr. Berlin. They were studied and
judged not to alter the conclusions arrived at in this docu-
ment concerning PbB vs. lead in water (see Appendix I).

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 level of lead
intake which would predict an incidence of cancer in 1:100,000
people. This calculated level of lead intake, 29 /ig/kg
of diet, poses some problems which must be confronted by
the EPA Carcinogen Assessment Group. Since this estimate
includes lead from all sources, its implications are beyond
the scope of this document. It should be noted, however,
that the International Agency for Research on Cancer, (IARC),
Lyon, France considers the experimental animal evidence
to be of dubious significance with regard to man (IARC,
1972). The IARC summary statement, quoted in part earlier
in this document, is as follows:
There is no evidence to suggest that
exposure to lead salts causes cancer
of any site in man. However, only one
epidemiological study of the relation-
ships between exposure to lead and the
occurrence of cancer has been reported.
It must be noted that the level of human
exposure equivalent to the levels of
lead acetate producing renal tumors
in rats is 810 mg per day (550 mg Pb).
This level appears to exceed by far
the maximum tolerated dose for man.

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Summary of "Research of PbB vs. Lead in Drinking Water in
Europe" as Presented by A. Berlin, et al. Commission of
the European Communities.
Results of the research examined in the Commission
of the European Communities (CEC) paper are summarized in
Tables 1-1 and 1-2. The data presented in Table 1-1 are
equations developed by the authors concerning the relation-
ship of blood lead (PbB) to water lead (PbW). Table 1-2
consists of calculations of the contribution of 100 ug Pb/1
of water to PbB (as ug/dl). some of these calculations
were made by Berlin, et al. 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 » a + 3 NpbW
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.

Relationships between PbW and PbBa


0.018 PbW + 22.9

r = 0.417
Pbw = jig/1
PbB = jig/100 ml
First morning flush
Addis and
Moore (1974)
0.76 + 0.15 PbW

Pbw and PbB in jimol/1
r = 0.58, first morning flush
Moore (1977a)
0.80 + 0.20 PbW

r = 0.52, running sample

0.533 + 0.675 3
Pbw and PbB in pmol/1
first morning flush
Moore, et al.
0.304 + 1.036 3
running sample

9.62 + 1.74 log
Pbw in jig/1 PbB in pg/lOOml
first morning flush
Lauwerys, et
al. (1977)
0.8 + 0.19 PbW

PbW and PbB in pmol/1
first morning flush
Moore (1977b)
0.8 + 0.53 PbW

full flush (paired samples)

19.6 + 7.2 PbW

PbW in ppm, PbB in /ig/100ml
first morning flush
Elwood, et al.
20.7 + 12.6 PbW

As above.
Re-evaluated data
Beattie, et
al. (1976)
aBerlin, et al. (1978).

Increment in PbB for an Increase of 100 pg/1
(for Concentrations around 100 pg/1)
in PbW
Increment in PbB Remarks
1.3	jig/lOOml
1.2	^g/100ml
3.4	jig/lOOml
3.3	/ig/100ml
1.8	jag/lOOml
2.0 jug/lOOml
6.0 jig/lOOml
3.9	/ig/100ml
0.83 /jg/100ml
1.9 ^ig/100ml
5.3 ^ig/100ml
0.72 /ig/lOOml
1.3 /ig/100ml
For running sample (linear
interpolation) 20-1040 jjg/l PbW
First flush (linear inter-
polation) 10-250 jug/1 PbW
For running sample (linear
interpolation) 10-250 pg/l PbW
For first flush (linear inter-
polation) 35-350 /jg/1 PbW
Using the linear equation derived
by the authors
Using the linear equation derived
by the author for running water
Using the non-linear equation
derived by the authors for
running water samples.
Using the non-linear equation de-
rived by the authors for first
morning flush.
Using the log equation derived
by the authors
In view of the low PbW value,
the extrapolation is uncertain.
Using the linear equation derived
by the authors for morning flush
Using the linear equation derived
by the authors for full flush
Using the linear equation derived
by the authors for morning flush.
Using the re-evaluated linear
equation derived by the authors
for morning flush.
De Graeve, et al.
Beattie, et al.
Cove11 (1975)
Addis, et al.
Addis, et al.
Moore (1977a)
Moore, et al.
Moore, et al.
Lauwerys, et al.
Vos, et al. (1977)
Moore, et al.
Moore, et al.
Elwood, et al.
Beattie, et al.
^Berlin, et al. (1978).

PbW using the EPA data were for females aged 20 to 50, a
sub-population which probably gets a larger proportion of
its water from 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.
In summary, of the studies cited in the CEC document,
most weight should probably be given to the Moore, et al.
(1977)	citation, 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
So far as the actual calculations in Table 1-2 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 jag/1
would result in a PbB contribution of 0.83 jig/dl. The error
is obvious. In the equation, the PbB contribution 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 1-2, only
5 could be verified. These were Addis, et al. (1974) (inter-
polation), Addis, et al. (1974) using authors' equation,
Moore (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 paper was unavailable (Covell, 1975;
Elwood, 1976), or because the necessary data were not in
the paper (De Graeve, et al. 1975; Moore, 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
actually 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 document and rejected
on the basis of the seemingly inappropriate use of a linear
regression model (see section on Contributions of Lead
from Diet vs. Air to PbB). It is therefore concluded that
information provided by CEC does not alter the evaluations
made in the criterion document.

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