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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

tion protective devices, plastics, and electronics equipment.

Although neither metallic lead nor the common lead minerals are classi

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

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

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

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

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

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

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

Interscience Publishers, New York.

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

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

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

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

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

Rahway, New Jersey.

A 2 Aquatic Life Toxicology*

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

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

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

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

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

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

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

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

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

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

Of the analytical measurements currently available, a water quality cri

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

water intercepts in Table 3.

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

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

No standard acute toxicity values for saltwater fish species are availa

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

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

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

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

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

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

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

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

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

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

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

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

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

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

tically significant differences could not be determined. After careful ex

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

tofish in turbid water.

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

respectively, at any time.

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

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

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

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

K 0 N A . ; ij t s t 2 L  8 I: I El I I a 4 a 3 I I 0 : a  I 2 I 2 t pl + + m m 8 m f s 4 f x s I d 3 f I I I I I I I I Q P f 0 LO " f a j t 2%  I I I I I I I I I REFERENCES Abou Donia, M. B. and 0 6. Menzel. 1967. Fish brain cholinesterase: Its in

hibition by carbamates and automatic assay. Comp. Biochem. Physiol. 21: 99.

Adams, E. S. 1975. Effect of lead and hydrocarbons from snowmobile exhaust on brook trout ( Salvelionus fontinalis). Trans. Am, Fish, Sot. 104: 363.

Anderson, R. L. et al. 1980. Survival and growth of Tanytarsus dissimilic ( cheronomidae) exposed to copper, cadmuim, zinc, and lead. Arch. Environ.

Contam. Toxicol. 9: 329.

Atchison, G. J., et al. 1977. Trace metal contamination of bluegill ( Lepomis macrochirus) from two Indiana lakes. Trans. Am. Fish. Sot.

106: 637.

Benijts Claus, C. and F. Benijts. 1975. In: J. H. Koeman and J. J. Strik teds.), Sublethal Effects of Toxic Chemicals on Aquatic Animals. Elsevier Pub. Co., Amsterdam. p. 43.

Biesinger, K. E. and G. M. Christensen. 1972. Effect of various metals on survival , growth, reproduction and metabolism of Daphnia magna. Jour. Fish,

Res. Board Can. 29: 1691.

Birge, W. J., et al. 1978. Embryo larva 1 Bioassays on Inorganic Coa 1 Ele

ments and In Situ Biomonitoring of Coal waste Effluents.  In: E. E. Samuel,

et a?. feds.), Surface Mining and Fish/ Wildlife Needs in the Eastern United States. U. S. Dept. Int., Fish Wildl. Serv., FWS/ OBS 78/ 81, Dec. 1978.

B 31 Borgman, II., et al. 1978. Rates of mortality, growth, and biomass produc

tion of Lymnaea palustris during chronic exposure to lead. Jour. Fish. Res.

Board Can. 35: 1109.

Brown, B. and M. Ahsanullah. 1971. Effect of heavy metals on mortality and growth. Mar. Pollut. Bull. 2: 182.

8ryan, G. W. 1976. Heavy metal contamination in the sea. In: R. Johnson

( ed.), Marine Pollution. Academic Press.

Buikema, A. L., et al. 1974. Evaluation of Philodina acuticornis ( Rotifera)

as a bioassay orpanism for heavy metals. Water Resour. Bull. 10: 648.

Cairns, J., et al. 1976. Invertebrate response to thermal shock following exposure to acutely sub lethal concentrations of chemicals. Arch. Hydro

biol. 77: 164.

Calabrese, A. and D. A. Nelson. 1974. Inhibition of embryonic development of the hard clam, Mercenaria mercenaria, by heavy metals. Bull. Environ.

Contam. Toxicol. 11: 92.

Calabrese, A., et al. 1973. The toxicity of heavy metals to embryos of the American oyster Crassostrea virginica. Mar. Biol. 18: 162.

Canterford, G. S., et al. 1978. Accumulation of heavy metals by the marine diatom Ditylum brightwellii ( West) C runow. Aust. J. Mar. Freshwater Res.

29: 613.

B 32 Chapman, G. A., et al. 1980. Effects of water hardness on the toxicity of metals to Daphnia magna. U. S. Environmental Protection Agency. Environ.

Res. Lab., Corvallis, Oregon. ( Manuscript)

Christensen, G. M. 1975. Biochemical effects of methylmercuric chloride,

cadmium chloride and lead nitrate on embryos and al. evins of the brook trout. Toxicol. Appl. Pharmacol. 32: 191.

Christensen, G., et al. 1977. The effect of methylmercuric chloride, cad

mium chloride, and lead nitrate on six biochemical factors of the brook trout ( Salvelinus fontinalis). Toxicol. Appl. Pharmacol. 42: 523.

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.

Eisler, R. 1977. Acute toxicities of selected heavy metals to the soft

shell clam, Mya arenaria. Bull. Environ. Contam. Toxicol. 17: 137.

Gray, J. S. and R. J. Ventilla. 1973. Growth rates of sediment living marine protozoan as a toxicity indicator for heavy metals. Ambio. 2: 118.

Hale, J. G. 1977. Toxicity of metal mining wastes. Bull. Environ. Contam.

Toxicol. 17: 66.

Hannan, P. J. and C. Patouillet. 1972. Effect of mercury on algal growth rates. Biotech. Bioen. 14: 93.

B 33 Hessler, A. 1974. Effects of lead on algae. I. Effects of Pb on viability and motility of Platymonas subcordiformis ( Chlorophyta: volvocales). Water Air Soil Pollut. 3: 371.

Hessler, A. 1975. Effects of lead on algae. Mutagenesis experiments on Platymonas subcordiformis ( Chlorophyta: valvocales). Mutat. Res. 31: 43.

Hodson, P. V. 1976. Amion levulinic acid dehydratase activity of fish blood as indictor of harmful exposure to lead. Jour. Fish. Res. Board Can.

33: 268.

Hodson, P. V., et al. 1977. Evaluation of erythrocyteamion levulinic acid dehydratase activity as a short term indicator in fish of a harmful exposure to lead. Jour. Fish. Res. Board Can. 34: 501.

Hodson, P. V., et al. 1978. Chronic toxicity of water borne and dietary lead to rainbow trout ( Salmo gairdneri) in Lake Ontario water. Water Res.

12: 869.

Hodson, P. V., et al. 1980. Effects of dietary ascorbic acid on chronic lead toxicity to young rainbow trout ( Salmo gairdneri). Can. Jour. Fish.

Aauat. Sci. 37: 170.

Holcombe, E. W., et al. 1976. Long term effects of lead exposure on three generations of brook trout ( Salvelinus fontinalis). Jour. Fish. Res. Board Can. 33: 1731.

B 34 Kaplan, H. M., et al. 1967. Toxicity of lead nitrate solutions for frogs ( Rana pipins). Lab. Anim. Care. 17: 240.

Kopfler, F. C. and J. Mayer. 1973. Concentration of five trace metals in the waters and oysters ( Crassostrea virginica) of Mobile Bay, Alabama.

Proc. Natl. Shellfish ASSOC. 63: 27.

Malanchuk, J. L. and G. K. Gruendling. 1973. Toxicity of lead nitrate to algae. Water Air and Soil Pollut. 2: 181.

Marchetti, R. 1978. Acute toxicity of alkyl lead to some marine organ

isms. Mar. Poll. Bull. 9: 206.

Monahan, T. J. 1976. Lead inhibition of chlorophycean microalgae. Jour.

Psycol. 12: 358.

Nehring, R. B. 1976. Aquatic insects as biological monitors of heavy metal pollution. Bull. Environ. Contam. Toxicol. 15: 147.

Pickering, Q. H. and C. Henderson. 1966. The acute toxicity of some heavy metals to different species of warmwater fishes. Air Water. Pollut. Int.

Jour. 10: 453.

Pringle, B. H., et al. 1968. Trace metal accumulation by estuarine mol

lusks. Proc. Am. Sot. Civil. Eng., Jour. San. Eng. Div. SA3, 455 475.

B 35 Reish, D. J. and R. S. Carr. 1978. The effect of heavy metals on the surviv

al, reproduction, development and life cycles for two species of polychaet

ous annelids. Mar. Pollut. Bull. 9: 24. ( Table 3 available from author.)

Reish, D. J., et al. 1976. The effect of heavy metals on laboratory popula

tions of two polychaetes with comparisons to the water quality conditions and standards in Southern California marine waters. Water Res. 10: 299.

Rivkin, R. B. 1979. Effects of lead on growth of the marine diatom Skele

tonema costatum. Mar. Biol. 239.

Sauter, S., et al. 1976. Effects of exposure to heavy metals on selected freshwater fish. Ecol. Res. Ser. EPA 600/ 3 76 105.

Schulz Baldes, M. 1972. Toxizitat und anreicherung von Blei bei der Mies

muschel Mytilis edulis im Laborexperiment. Mar, Biol. 16: 266.

Schulz Baldes, M. 1974. Lead uptake from seawater and food, and lead loss in the common mussel Mytilis edulis. Mar. Biol. 25: 177.

Schulz Baldes, M. 1976. Lead uptake in two marine phytoplankton organ

isms. Biol. Bull. 150: 118.

Shuster, C. N. and B. H. Pringle. 1969. Trace metal accumulation by the American Eastern oyster, Crassostrea virginica. Proc. Natl. Shellfish Assoc. 59: 91.

B 36 Spehar, R. L., et al. 1978. Toxicity and bioaccumulation of cadium and lead in aquatic invertebrates. Environ. Pollut. 15: 195.

Stewart, J. and M. Schulz Baldes. 1976. Long term lead accumulation in Abalone ( Haliotus sp.) fed on lead treated brown algae ( Egregia laevigata).

Mar. Biol. 36: 19.

Talbot, V., et al. 1976. Lead in Port Phillip Bay mussels. Mar. Pollut.

Bull. 7: 234.

Tarzwell, C. M.. and C. Henderson.

fishes. Ind. Wastes. 5: 12.

U. S. EPA. 1980. Unpublished Narragansett, Rhode Island. 1960. Tox icity of less common metals to laboratory data. Environ. Res. Lab.,

Wallen, I. E., et a?. 1957. Toxicity to Gambusia affinis of certain pure chemicals in turbid waters. Sewage Ind. Wastes. 29: 695.

Warnick, S. L and H. L. Bell. 1969. The acute toxicity of some heavy metals to different species of aquatic insects. Jour. Water Pollut. Control Fed.

41: 28C.

Waterman, A. J. 1937. Effects of salts of heavy metals on development of the sea urchin, Arbacia punctulata. Biol. Bull. 73: 401.

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

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

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

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

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

lutants. Hydrobiol. 32: 193.

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

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

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

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

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

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

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

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

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

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

EPA, 1977; Hammond, 1977].

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

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

land and over remote oceans.

Areas of abnormally high concentrations of lead occur in natu

ral ores, usually in conjunction with high concentrations of cad

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

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

Pigments and ceramics account for about 6 percent of annual pro

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

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

Children are known to ingest this type of deteriorating paint.

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

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

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

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

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

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

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

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

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

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

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

ly performed in unfiltered water and hence represent total lead.

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

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

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

cant source of lead in man's diet.

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

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

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

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

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

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

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

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

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

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

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

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

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

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

1972).

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

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

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

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

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

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

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

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

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

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

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

ated regarding the contribution of air to total daily lead absorp

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

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

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

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

nificant route of exposure.

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

eral sources of lead which may assume clinically significant pro

portions. Perhaps the most serious widespread problem is the con

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

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

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

Occupational exposure to lead may be quite excessive. Thus,

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

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

nificant.

The hazard of lead to children is of considerable concern.

The number of children excessively exposed to lead from miscella

neous sources is impressive. Thus, federally assisted lead screen

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

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

1977).

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

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

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

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

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

Soil along the shoulder of heavily traveled roadways also is heavi

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

pie, Laqerwerff and Specht, 1970).

The relative contribution of soil, automotive exhaust fallout,

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

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

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

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

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

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

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

Thus, the data are merely susqestive.

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

sure in eight children, hospitalized for excessive lead absorDtion,

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

vated fecal lead excretion as compared to nine control children,

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

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

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

20 O. ll( 2)

0.3( l)

0.3( l) 0.7( l)

0.1( l)

4.0( l)

1.9( l)

0.6( l) 0.12( 3)

0.06( 2)

0.07( l)

0.3( 2)

0.1( l)

0.2( l)

0.9( 2)

0.05( l)

0.1( 3)

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

CNumhers in parentheses indicate number of environmental samples.

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

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

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

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

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

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

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

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

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

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

particularly during pregnancy.

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

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

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

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

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

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

They provide a substantial amount of knowledoe concerning the sub

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

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

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

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

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

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

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

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

Recent studies using the nonradioactive tracer 204 Pb have con

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

ferred to the end of this section on lead metabolism.

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

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

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

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

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

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

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

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

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

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

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

sure conditions.

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

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

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

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

et al. 1967).

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

tery plant ( WHO, 1977).

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

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

So far as the soecific contribution of water is concerned,

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

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

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

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

Air Lead,

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

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

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

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

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

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

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

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

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

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

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

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

RI9rrrrion 01 blood lstd on mornin watmr loud in CmOrnmrlon~ <EMAIL_ADDRESS> FIGURE 1 Regression of Blood lead on Morning Water Lead in Caernarfonshire Source: Elwood, et al. 1976 C 28 BLOOD l po LEAD l/ l)

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

Interestingly, the relationship of early morning water Pb to run

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

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

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

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

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

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

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

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

It is known that calcium profoundly depresses lead absorDtion,

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

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

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

press lead absorption.

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

cinogenesis, mutagenesis, and teratoqenesis. Because of the Dauci

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

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

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

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

culating concentration of hemoglobin is not thoroughlv understood.

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

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

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

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

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

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

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

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

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

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

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

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

Succmyl  CoA 7 Heme + A&

Glyane Ferrocdelatose I I ALA Syntherase ( ALAS) Fe%

l< , +

8 Amir! olevulinic Protoporphyrin TX Acid ( ALA)

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

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

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

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

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

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

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

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

tion have not been studied extensively. However, substantial lead

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

field, et. al. 1972).

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

The minimal level of lead exposure resulting in lead encepha

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

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

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

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

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

The most common deficiencies encountered are overlap of lead expo

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

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

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

Two general approaches have been used in attackin the prob

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

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

normal children whose problems were clearly attributable to etiolo

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

( Bicknell, et al. 1968).

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

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

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

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

the experimental design was otherwise generally sound.

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

were found to be significantly higher than those of control sub

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

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

C 46 developed impaired visual discrimination learning behavior ( Car

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

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

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

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

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

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

behavioral effects of low level lead exposure in utero.

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

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

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

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

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

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

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

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

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

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

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

creased in the more recent years as working conditions have im

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

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

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

( orchardists and lesser exposed " intermediates" as separate cate

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

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

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

There was a definite gradation in blood and urine lead concentra

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

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

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

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

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

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

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

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

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

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

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

A good record of lead exposure history was considered important.

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

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

The only causes of death that showed a statistically signifi

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

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

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

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

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

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

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

tistically significant increase in deaths due to all malignant neo

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

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

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

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

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

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

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

genesis.

The first report of lead induced renal tumors ( Zollinqer,

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

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

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

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

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

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

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

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

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

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

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

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

bund animals were killed and examined morphologically. At equiva

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

Interstitial nephritis occurred in all grouns, including con

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

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

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

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

Neither of the two remaining reports concerning the carcino

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

C 58 TARLE IO SurllfnarY of Mortality Data Resulting from Lead Phsplratc Adlninistration to Ratsa Successive Time Intervals, Days  ___  _. __  .  __ o 100 101 200 201 300 301 400 401 500 50 l 600 601 700 __ _

Cb 2.6' 1.3' .17= c 2.6 I. 3 . I7 c 2.6 1.3 . I7 C 2.6 I. 3 . I7 C 2.6 1.3 . I7 C 2.6 1.3 . I7 C 2.6 1.3 . I7 ___ ~___~~~

n ill hc); inniiig 01 ihltet val 48 24 40 40 48 6 37 40 48 3 37 38 46 2 37 34 41 I 35 25 26  I4 18 II  6 _. . I_.~ _____ _______  ckud, IW rctul 0 I8 3 0 0 3 0 22 I 0 4 2 0 I 9 15 0 7 7 I5  5 I6 II 0 I lmlburs __._  _.__ _ .  . _  tlc.~ tl, rcful 0 0 0 0 0 0 0 00 00 0 0 I IO I I 14 0 I  3 0 0 5 turrwr5 __ _.._.__.  .   _______ .._____. .___.~ ..___ ill illlervitl dyurl; with 0 0 0 0 0 0 0 00 0 0 0 0 50 3 0 2 100 40 0 4  57 0 0  x3 turnul L  ______ ~ UIII1l` I1IVl! tno4 hlity, 0 82 IX 0 0 95 IX 5 4 100 I8 I5 9 100 24 37.5 41 100 65 55 74 IO0 94 ` 95 100 100 100 100 110 Itmw5 __  _. . . . .._ ___~ __ _____  __.. _.._.~~_

c:~ llll~ ll. rllvc iihorf. dily, 0 0 0 0 0 0 0 00 00 0 0 4 30 2 4 35 0 4  4 5 0 4  58 llllllW5 _ _ .__ _ __.__ I_~ _~. ~ ._.   . ._ _  _._  .

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

quired to produce tumors did not clearly result in increased mor

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