Lead: environmental aspects (EHC 85, 1989)

INTERNATIONAL PROGRAMME ON CHEMICAL SAFETY

ENVIRONMENTAL HEALTH CRITERIA 85

LEAD - ENVIRONMENTAL ASPECTS

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experts and does not necessarily represent the decisions or the stated
policy of the United Nations Environment Programme, the International
Labour Organisation, or the World Health Organization.

Published under the joint sponsorship of
the United Nations Environment Programme,
the International Labour Organisation,
and the World Health Organization

World Health Orgnization
Geneva, 1989

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CONTENTS

ENVIRONMENTAL HEALTH CRITERIA FOR LEAD - ENVIRONMENTAL ASPECTS

1. SUMMARY AND CONCLUSIONS

1.1. Physical and chemical properties and sources of pollution
1.2. Uptake, loss, and accumulation in organisms
1.2.1. Model ecosystems
1.2.2. Uptake and accumulation by aquatic organisms
1.2.3. Uptake and accumulation by terrestrial organisms
1.2.4. Uptake of lead in the field
1.2.5. Uptake in the vicinity of highways and in urban areas
1.2.6. Uptake of lead from industrial sources
1.2.7. Intake of lead shot
1.3. Toxicity to microorganisms
1.4. Toxicity to aquatic organisms
1.5. Toxicity to terrestrial organisms
1.6. Toxic effects in the field

2. PHYSICAL AND CHEMICAL PROPERTIES

3. SOURCES OF LEAD IN THE ENVIRONMENT

4. UPTAKE, LOSS, AND ACCUMULATION IN ORGANISMS

4.1. Controlled experimental studies
4.1.1. Model ecosystems
4.1.2. Aquatic organisms
4.1.3. Terrestrial organisms
4.2. Accumulation in the field
4.2.1. General considerations
4.2.2. Highways and urban areas
4.2.3. Industrial sources
4.2.4. Lead shot

5. TOXICITY TO MICROORGANISMS

5.1. Toxicity of lead salts
5.2. Toxicity of organic lead

6. TOXICITY TO AQUATIC ORGANISMS

6.1. Toxicity to aquatic plants
6.2. Toxicity to aquatic invertebrates
6.2.1. Toxicity of lead salts
6.2.2. Toxicity of organic lead
6.3. Toxicity to fish
6.3.1. Toxicity of lead salts
6.3.2. Biochemical effects
6.3.3. Behavioural effects
6.4. Toxicity to amphibia

7. TOXICITY TO TERRESTRIAL ORGANISMS

7.1. Toxicity to plants

7.2. Toxicity to invertebrates
7.3. Toxicity to birds
7.3.1. Toxicity of lead salts
7.3.1.1 Toxicity to bird's eggs
7.3.1.2 Toxicity to adult and juvenile birds
7.3.1.3 Enzyme effects
7.3.1.4 Behavioural effects
7.3.2. Toxicity of metallic lead
7.3.2.1 Toxicity of powdered lead
7.3.2.2 Toxicity of lead shot
7.3.3. Toxicity of organolead compounds
7.4. Toxicity to non-laboratory mammals

8. EFFECTS OF LEAD IN THE FIELD

8.1. Tolerance of plants to lead
8.2. Highways and industrial sources of lead
8.3. Lead shot
8.4. Organic lead

9. EVALUATION

9.1. General considerations
9.2. The aquatic environment
9.3. The terrestrial environment

REFERENCES

WHO TASK GROUP ON ENVIRONMENTAL HEALTH CRITERIA FOR LEAD - ENVIRONMENTAL
ASPECTS

Members

Dr L.A. Albert, Environmental Pollution Programme, National Institute
for Research on Biotic Resources, Veracruz, Mexico
Dr R. Elias, Environmental Criteria and Assessment Office, US
Environmental Protection Agency, Research Triangle Park, North
Carolina, USA (Chairman)
Dr J.H.M. Temmink, Department of Toxicology, Agricultural University,
Biotechnion, Wageningen, Netherlands
Dr G. Roderer, Fraunhofer Institute for Environmental Chemistry and
Ecotoxicology, Schmallenberg-Grafschaft, Federal Republic of
Germany
Dr R. Koch, Division of Toxicology, Research Institute for Hygiene and
Microbiology, Bad Elster, German Democratic Republic
Dr Y. Kodama, Department of Environmental Health, University of
Occupational and Environmental Health, Kitakyushu, Japan
Professor P.N. Viswanathan, Ecotoxicology Section, Industrial Toxi-
cology Research Centre, Lucknow, India

Observers

Mr D.J.A. Davies, Department of the Environment, London, United
Kingdom
Dr I. Newton, Institute of Terrestrial Ecology, Monks Wood Experimen-
tal Station, Huntingdon, United Kingdom

Secretariat

Dr S. Dobson, Institute of Terrestrial Ecology, Monks Wood Experimen-
tal Station, Huntingdon, United Kingdom (Rapporteur)
Dr M. Gilbert, International Programme on Chemical Safety, World
Health Organization, Geneva, Switzerland (Secretary)
Mr P.D. Howe, Institute of Terrestrial Ecology, Monks Wood Experimen-
tal Station, Huntingdon, United Kingdom

NOTE TO READERS OF THE CRITERIA DOCUMENTS

Every effort has been made to present information in the criteria
documents as accurately as possible without unduly delaying their
publication. In the interest of all users of the environmental health
criteria documents, readers are kindly requested to communicate any
errors that may have occurred to the Manager of the International
Programme on Chemical Safety, World Health Organization, Geneva,
Switzerland, in order that they may be included in corrigenda, which
will appear in subsequent volumes.

* * *

A detailed data profile and a legal file can be obtained from the
International Register of Potentially Toxic Chemicals, Palais des
Nations, 1211 Geneva 10, Switzerland (Telephone No. 988400 - 985850).

ENVIRONMENTAL HEALTH CRITERIA FOR LEAD - ENVIRONMENTAL ASPECTS

A WHO Task Group on Environmental Health Criteria for Lead -
Environmental Aspects met at the Institute of Terrestrial Ecology,
Monks Wood, United Kingdom, from 7 to 11 December 1987. Dr B.N.K.
Davis welcomed the participants on behalf of the host institution, and
Dr M. Gilbert opened the meeting on behalf of the three co-sponsoring
organizations of the IPCS (ILO/UNEP/WHO). The Task Group reviewed and
revised the draft criteria document and made an evaluation of the risks
for the environment from exposure to lead.

The first draft of this document was prepared by Dr S. Dobson and
Mr P.D. Howe, Institute of Terrestrial Ecology. Dr M. Gilbert and
Dr P.G. Jenkins, both members of the IPCS Central Unit, were respon-
sible for the overall scientific content and editing, respectively.

* * *

Partial financial support for the publication of this criteria
document was kindly provided by the United States Department of Health
and Human Services, through a contract from the National Institute of
Environmental Health Sciences, Research Triangle Park, North Carolina,
USA - a WHO Collaborating Centre for Environmental Health Effects.

INTRODUCTION

There is a fundamental difference in approach between the
toxicologist and the ecotoxicologist concerning the appraisal of the
potential threat posed by chemicals. The toxicologist, because his
concern is with human health and welfare, is preoccupied with any
adverse effects on individuals, whether or not they have ultimate
effects on performance or survival. The ecotoxicologist, in contrast,
is concerned primarily with the maintenance of population levels of
organisms in the environment. In toxicity tests, he is interested in
effects on the performance of individuals - in their reproduction and
survival - only insofar as these might ultimately affect the population
size. To him, minor biochemical and physiological effects of toxicants
are irrelevant if they do not, in turn, affect reproduction, growth, or
survival.

It is the aim of this document to take the ecotoxicologist's point
of view and consider effects on populations of organisms in the
environment. No attempt has been made to link the conclusions reached
in this document with possible effects on human health, since a new
Environmental Health Criteria document examining the effects on human
health of lead compounds is in preparation. Due attention has been
given to persistence in the environment and bioaccumulation. These
will have implications for human consumption of the metal.

This document, although based on a thorough survey of the
literature, is not intended to be exhaustive in the material included.
In order to keep the document concise, only those data which were
considered to be essential in the evaluation of the risk posed by lead
to the environment have been included. Concentration figures for lead
in the environment, or in particular species of organism, have not been
included unless they illustrate specific toxicological points. "Snap
shot" concentration data, where a causal relationship between the
presence of the metal and an observed effect is not clearly
demonstrated, have been excluded.

The term bioaccumulation indicates that organisms take up chemicals
to a greater concentration than that found in their environment or
their food. "Bioconcentration factor" is a quantitative way of
expressing bioaccumulation: the ratio of the concentration of the
chemical in the organism to the concentration of the chemical in the
environment or food. Biomagnification refers, in this document, to the
progressive accumulation of chemicals along a food chain.

1. SUMMARY AND CONCLUSIONS

1.1. Physical and Chemical Properties and Sources of Pollution

Lead is a bluish or silvery-grey soft metal. With the exception of
the nitrate, the chlorate, and, to a much lesser degree, the chloride,
the salts of lead are poorly soluble in water. Lead also forms stable
organic compounds. Tetraethyllead and tetramethyllead are used exten-
sively as fuel additives. Both are volatile and poorly soluble in
water. Trialkyllead compounds are formed in the environment by the
breakdown of tetraalkylleads. These trialkyl compounds are less
volatile and more readily soluble in water. Lead is mined, most
usually as the sulfide, "galena". Pollution of the environment
occurs through the smelting and refining of lead, the burning of
petroleum fuels containing lead additives and, to a lesser extent, the
smelting of other metals and the burning of coal and oil. Metallic
lead deriving from shotgun cartridges or used as fishing weights is
lost in the environment and often remains available to organisms.

1.2. Uptake, Loss, and Accumulation in Organisms

Lead in the environment is strongly adsorbed onto sediment and soil
particles reducing its availability to organisms. Because of the low
solubility of most of its salts, lead tends to precipitate out of
complex solutions.

1.2.1. Model ecosystems

In aquatic and aquatic/terrestrial model ecosystems, uptake by
primary producers and consumers seems to be determined by the bio-
availability of the lead. Bioavailability is generally much lower
whenever organic material, sediment, or mineral particles (e.g., clay)
are present. In many organisms, it is unclear whether lead is
adsorbed onto the organism or actually taken up. Consumers take up
lead from their contaminated food, often to high concentrations, but
without biomagnification.

1.2.2. Uptake and accumulation by aquatic organisms

The uptake and accumulation of lead by aquatic organisms from water
and sediment are influenced by various environmental factors such as
temperature, salinity, and pH, as well as humic and alginic acid
content.

In contaminated aquatic systems, almost all of the lead is tightly
bound to sediment. Only a minor fraction is dissolved in the water,
even interstitial water between the sediment particles.

The lead uptake by fish reaches equilibrium only after a number of
weeks of exposure. Lead is accumulated mostly in gill, liver, kidney,
and bone.

Fish eggs show increasing lead levels with increased exposure con-
centration, and there are indications that lead is present on the egg
surface but not accumulated in the embryo.

In contrast to inorganic lead compounds, tetraalkyllead is rapidly
taken up by fish and rapidly eliminated after the end of the exposure.

1.2.3. Uptake and accumulation by terrestrial organisms

In bacteria, the majority of lead is associated with the cell
wall. A similar phenomenon is also noted in higher plants. Some lead
that passes into the plant root cell can be combined with new cell
wall material and subsequently removed from the cytoplasm to the cell
wall. Of the lead remaining in the root cell, there is evidence of
very little translocation to other parts of the plant because the con-
centration of lead in shoot and leaf tissue is usually much lower than
in root. Foliar uptake of lead occurs, but only to a very limited
extent.

In animals, there is a positive correlation between tissue and
dietary lead concentrations, although tissue concentrations are almost
always lower. The distribution of lead within animals is closely
associated with calcium metabolism.

Lead shot is typically trapped in the gizzard of birds where it is
slowly ground down resulting in the release of lead.

The tetravalent organic form of lead is generally more toxic than
the divalent, inorganic form, and its distribution in organisms may not
specifically follow calcium metabolism.

1.2.4. Uptake of lead in the field

Organisms have been found to incorporate lead from the environment,
generally in proportion to the degree of contamination. Lead depo-
sition in a region depends on the air concentrations of the metal,
which decrease with the distance from the source.

In shellfish, lead concentrations are higher in the calcium-rich
shell than in the soft tissue; they relate to the concentrations in
sediment.

Lead concentrations in some marine fish are higher in gills and
skin than in other tissues, but this may be largely due to adsorption.
Liver levels increase significantly with age.

In dolphins, lead is transferred from mothers to offspring during
fetal development and lactation. This might be related to the calcium
metabolism.

1.2.5. Uptake in the vicinity of highways and in urban areas

Lead concentrations are highest in soils and organisms close to
roads where traffic density is high. The lead measured is inorganic
and derives almost exclusively from alkyllead compounds added to
petrol.

The lead in the soil and in vegetation decreases exponentially with
the distance from the road. Lead is also found in the sediments of
streams in the vicinity of highways.

Lead contamination increases lead levels in plants and animals in
areas close to roads. These levels are positively correlated with
traffic volume and proximity of roads.

Most lead deposited is found within 500 m of the road and within
the upper few centimetres of soil. It can be assumed that lead levels
in soil and biota are not influenced by traffic at distances from roads
greater than this.

1.2.6. Uptake of lead from industrial sources

Terrestrial and aquatic plants accumulate lead in industrially
contaminated environments. In aquatic plant species, lead uptake can
occur from both water and sediment, although uptake from sediment
usually predominates. Lead levels decrease with distance from the
source and are lowest during the active growing season in terrestrial
plants. The role of foliar uptake is uncertain. Mosses accumulate
lead from the atmosphere and are often used as biological monitors of
airborne lead.

Elevated lead levels are also found in terrestrial invertebrates
and vertebrates from contaminated areas.

1.2.7. Intake of lead shot

Lead shot taken by birds into their gizzards is a source of severe
lead contamination. It results in high organ levels of lead in blood,
kidney, liver, and bone.

1.3. Toxicity to Microorganisms

In general, inorganic lead compounds are of lower toxicity to
microorganisms than are trialkyl- and tetraalkyllead compounds. Tetra-
alkyllead becomes toxic by decomposition into the ionic trialkyllead.

One of the most important factors which influence the aquatic
toxicity of lead is the free ionic concentration, which affects the
availability of lead to organisms. The toxicity of inorganic lead
salts is strongly dependent on environmental conditions such as water
hardness, pH, and salinity, a fact which has not been adequately
considered in most toxicity studies.

There is evidence that tolerant strains exist and that tolerance
may develop in others.

1.4. Toxicity to Aquatic Organisms

Lead is unlikely to affect aquatic plants at levels that might be
found in the general environment.

In the form of simple salts, lead is acutely toxic to aquatic
invertebrates at concentrations above 0.1 and >40 mg/litre for fresh-
water organisms and above 2.5 and >500 mg/litre for marine organisms.
For the same species, the 96-h LC₅₀s for fish vary between 1 and
27 mg/litre in soft water, and between 440 and 540 mg/litre in hard

water. The higher values for hard water represent nominal concen-
trations. Available lead measurements suggest that little of the total
lead is in solution in hard water. Lead salts are poorly soluble in
water, and the presence of other salts reduces the availability of lead
to organisms because of precipitation. Results of toxicity tests
should be treated with caution unless dissolved lead is measured.

In communities of aquatic invertebrates, some populations are more
sensitive than others and community structure may be adversely affected
by lead contamination. However, populations of invertebrates from
polluted areas can show more tolerance to lead than those from non-
polluted areas. In other aquatic invertebrates, adaptation to hypoxic
conditions can be hindered by high lead concentrations.

Young stages of fish are more susceptible to lead than adults or
eggs. Typical symptoms of lead toxicity include spinal deformity and
blackening of the caudal region. The maximum acceptable toxicant
limit (MATC) for inorganic lead has been determined for several species
under different conditions and results range from 0.04 mg/litre to
0.198 mg/litre. The acute toxicity of lead is highly dependent on the
presence of other ions in solution, and the measurement of dissolved
lead in toxicity tests is essential for a realistic result. Organic
compounds are more toxic to fish than inorganic lead salts.

There is evidence that frog and toad eggs are sensitive to nominal
lead concentrations of less than 1.0 mg/litre in standing water and
0.04 mg/litre in flow-through systems; arrested development and delayed
hatching have been observed. For adult frogs, there are no signifi-
cant effects below 5 mg/litre in aqueous solution, but lead in the diet
at 10 mg/kg food has some biochemical effects.

1.5. Toxicity to Terrestrial Organisms

The tendency of inorganic lead to form highly insoluble salts and
complexes with various anions, together with its tight binding to
soils, drastically reduces its availability to terrestrial plants via
the roots. Translocation of the ion in plants is limited and most
bound lead stays at root or leaf surfaces. As a result, in most
experimental studies on lead toxicity, high lead concentrations in the
range of 100 to 1000 mg/kg soil are needed to cause visible toxic
effects on photosynthesis, growth, or other parameters. Thus, lead is
only likely to affect plants at sites of very high environmental
concentrations.

Ingestion of lead-contaminated bacteria and fungi by nematodes
leads to impaired reproduction. Woodlice seem unusually tolerant to
lead, since prolonged exposure to soil or grass litter containing
externally added lead salts had no effect. Caterpillars maintained on
a diet containing lead salts show symptoms of toxicity leading to
impaired development and reproduction.

The information available is too meagre to quantify the risks to
invertebrates during the decomposition of lead-contaminated litter.

Lead salts are only toxic to birds at a high dietary dosage
(100 mg/kg or more). Almost all of the experimental work is on

chickens and other gallinaceous birds. Exposure of quail from hatching
and up to reproductive age resulted in effects on egg production at
dietary lead levels of 10 mg/kg. Although a variety of effects at high
dosage have been reported, most can be explained as a primary effect on
food consumption. Diarrhoea and lack of appetite, leading to anorexia
and weight loss, are the primary effects of lead salts. Since there is
no experimental evidence to assess effects on other bird species, it is
necessary to assume a comparable sensitivity. If this is so, then it
is highly improbable that environmental exposure would cause adverse
effects.

Metallic lead is not toxic to birds except at very high dosage when
administered in the form of powder. It is highly toxic to birds when
given as lead shot; ingestion of a single pellet of lead shot can be
fatal for some birds. The sensitivity varies between species and is
dependent on diet. Since birds have been found in the wild with large
numbers of lead shot in the gizzard (20 shot is not unusual), this
poses a major hazard to those species feeding on river margins and in
fields where many shot have accumulated.

There is little information on the effects of organolead compounds.
Trialkyllead compounds produced effects on starlings dosed at
0.2 mg/day; 2 mg/day was invariably fatal.

There are too few reports to draw conclusions about the effects of
lead on non-laboratory mammals. Wild rats showed similar effects to
their laboratory counterparts.

1.6. Toxic Effects in the Field

Most work on plant tolerance to lead has concentrated on plants
growing on mining wastes, naturally highly contaminated areas, and
roadside verges. Tolerance has only been found in populations of a few
plant species.

No effect on the reproduction of birds nesting near highways has
been observed. Toxic effects have been observed in pigeons in urban
areas, the kidneys being most frequently affected.

Lead poisoning, due to the ingestion of lead shot, is a cause of
death for large numbers of birds. In these cases, lead shot is found
in the gizzards, and lead levels are elevated in the liver, kidneys,
and bones.

A recurring incident of massive bird kills in estuaries near to
industrial plants manufacturing leaded "anti-knock" compounds has
been reported. The total lead content of the livers was sufficiently
high to cause mortalities: lead was mostly present in the alkyl form.

2. PHYSICAL AND CHEMICAL PROPERTIES

Details of the physical and chemical properties of lead are given
in Environmental Health Criteria 3: Lead (WHO, 1977).

Lead (atomic number, 82; atomic weight, 207.19; specific gravity,
11.34) is a bluish or silvery-grey soft metal. The melting point is
327.5 °C and the boiling point, at atmospheric pressure, is 1740 °C.
It has four naturally occurring isotopes: 208, 206, 207, and 204, in
order of abundance. The isotopic ratios for various mineral sources
are sometimes substantially different. This property has been used to
carry out non-radioactive tracer environmental and metabolic studies.

Although lead has four electrons in its valence shell, only two
ionize readily. The usual oxidation state of lead in inorganic
compounds is, therefore, +2 rather than +4. The inorganic compounds of
lead are generally poorly soluble, with the exception of the nitrate,
the chlorate, and, to a much lesser degree, the chloride. Some of the
salts formed with organic acids, e.g., lead oxalate, are also
insoluble.

Under appropriate conditions of synthesis, stable compounds are
formed in which lead is directly bound to a carbon atom. Tetraethyl-
lead and tetramethyllead are well-known organolead compounds. They are
of great importance owing to their extensive use as fuel additives.
Both are colourless liquids. Their volatility is lower than for most
fuel components. The boiling point of tetramethyllead is 110 °C and
that of tetraethyllead is 200 °C. By contrast, the boiling point range
for gasoline hydrocarbons is 20 to 200 °C. Evaporation of gasoline
tends to concentrate tetraethyllead and tetramethyllead in the liquid
residue.

Both tetramethyllead and tetraethyllead decompose at, or somewhat
below, the boiling point. Analysis of automobile exhaust gases shows
that the ratio of tetramethyllead to tetraethyllead increases as the
engine warms up, indicating that tetramethyllead is more thermostable
than tetraethyllead. These compounds are also decomposed by ultra-
violet light and trace chemicals in air such as halogens, acids, or
oxidizing agents.

3. SOURCES OF LEAD IN THE ENVIRONMENT

Details of the sources of lead are given in Environmental Health
Criteria 3: Lead (WHO, 1977). The relevant chapter is summarized
here.

The major sources of lead in the environment, of significance to
living organisms, arise from lead mining and the refining and smelting
of lead and other metals. The major dispersive, non-recoverable use of
lead is in the manufacture and application of alkyllead fuel addi-
tives.

From a mass balance point of view, the transport and distribution
of lead from stationary or mobile sources is mainly via air. Although
large amounts are probably also discharged into soil and water, lead
tends to localize near the points of such discharge. Lead that is
discharged into the air over areas of high traffic density falls out
mainly within the immediate metropolitan zone. The fraction that
remains airborne (about 20%, based on very limited data) is widely
dispersed. Residence time for these small particles is of the order of
days and is influenced by rainfall. In spite of widespread dispersion,
with consequent dilution, there is evidence of lead accumulation at
points extremely remote from human activity, for example in glacial
strata in Greenland. The concentration of lead in air varies from
2-4 µg/m³ in large cities with dense automobile traffic to less
than 0.2 µg/m³ in most suburban areas, and still less in rural
areas.

4. UPTAKE, LOSS, AND ACCUMULATION IN ORGANISMS

Lead is accumulated into many organisms, in many habitats. The follow-
ing is a selection rather than an exhaustive review. Examples of experiment-
ally determined bioaccumulation factors are given in Tables 1 and 2.

4.1. Controlled Experimental Studies

4.1.1. Model ecosystems

Appraisal

Vighi (1981) constructed a simple trophic chain model ecosystem
consisting of the alga Selenastrum capricornutum, the water flea
Daphnia magna, and the guppy Lebistes reticulatus, and introduced lead
as lead nitrate. Concentration factors for the various organisms are
given in Table 1. He calculated uptake rates and half-lives for loss
of lead. The time taken to reach half the equilibrium concentration
of lead in tissues of the organisms ("half-life of uptake") was
5.3 days for the alga, 7.7 days for the water flea, and 25.7 days for
total uptake into the fish. Uptake of lead into the guppy was split
into two components, that from water and that from food. Half-life of
uptake from water was 7.7 days whereas, from food, it was 33 days.
Half-lives for loss of lead were calculated as 9 days for fish which
had received their lead only from water, and as 40 days for fish which
had received lead from food.

Lu et al. (1975) established aquatic/terrestrial model ecosystems
based on three different soil types. At the beginning of the exper-
iment, lead chloride was incorporated into the soil. Sorghum seeds
were sown in the soil, and algae, daphnids, and pond snails were
introduced into the water. On day 7, salt marsh caterpillars were
introduced to feed on the sorghum, and, on day 27, mosquito larvae were
added to the water. On day 30, some mosquito larvae were removed for
analysis and mosquito fish were added to the water to eat the remaining
larvae. The experiment was terminated on day 33. Results differed
greatly according to the soil type. Using silica sand, with a natural
lead concentration of 0.122 mg/kg and 10 mg/kg lead chloride added,
the lead levels in organisms were higher than with other soils. With
this soil, lead levels were as follows: water 0.013, algae 275,
daphnids 187, snails 334, mosquito larvae 403, fish 13, sorghum leaves
497, and sorghum roots 695 mg/kg. Using silica sand with 10% of silty
clay loam (natural lead content 4.5 mg/kg) and 10 mg lead chloride/kg
added, uptake into all organisms was markedly less. Lead levels were
as follows: water 0.002, algae 114, daphnids 85, snails 56, mosquito
larvae 80, fish 1, sorghum leaves 1, and sorghum roots 5 mg/kg. Lead
appears to be very strongly bound to even small amounts of fine soil
material and, therefore, unavailable to organisms.

Table 1. Accumulation of lead into aquatic organisms
---------------------------------------------------------------------------------------------------------
Organism Life- Test/ Organ^b Tem- pH Compound Dura- Exposure Bioconcen- Refer-
stage/ type^a perature tion (µg/ tration ence
size ( °C) (days) litre) factor^c
---------------------------------------------------------------------------------------------------------
Green alga D WP 21.1-24.7 7.2-7.8 nitrate 7 4.5 70 000 Vighi
(Selenastrum D WP 21.1-24.7 7.2-7.8 nitrate 28 4.5 102 222 (1981)
capricornutum) D WP 21.1-24.7 7.2-7.8 nitrate 7 40.1 27 431
D WP 21.1-24.7 7.2-7.8 nitrate 28 40.1 32 419

Pondweed adult A WP 25 30 25 6200 Nakada
(Elodea et al.
nuttallii) (1979)

Water hyacinth adult A top 23-27 nitrate 16 1000 492 Muramoto
(Eichhornia adult A roots 23-27 nitrate 16 1000 6200 & Oki
crassipes) (1983)
adult A leaves 23-27 nitrate 28 1000 5.89 Kay &
Haller
(1986)

Oyster B WB chloride 21 100 13.4^d Watling
(Crassostrea (1983a)
gigas)

Oyster B WB chloride 21 100 17^d Watling
(Crassostrea (1983a)
margaritacea)

Marine mollusc B WB chloride 21 100 27.1^d Watling
(Perna perna) (1983a)

Mussel B WB chloride 21 100 31.7^d Watling
(Choromytilus (1983a)
meridionalis)

Mussel 6-7 cm A kidney 15 nitrate 13 100 3000 Coombs
(Mytilus 6-7 cm A kidney 15 citrate 13 100 10 000 (1977)
edulis)
---------------------------------------------------------------------------------------------------------

Table 1. (contd.)
---------------------------------------------------------------------------------------------------------
Organism Life- Test/ Organ^b Tem- pH Compound Dura- Exposure Bioconcen- Refer-
stage/ type^a perature tion (µg/ tration ence
size ( °C) (days) litre) factor^c
---------------------------------------------------------------------------------------------------------

Water flea D WB 21.1-24.7 7.2-7.8 nitrate 7 4.5 2905 Vighi
(Daphnia magna) D WB 21.1-24.7 7.2-7.8 nitrate 7 315 µg/g 0.04^e (1981)
D WB 21.1-24.7 7.2-7.8 nitrate 28 4.5 5140
D WB 21.1-24.7 7.2-7.8 nitrate 28 460 µg/g 0.05^e
D WB 21.1-24.7 7.2-7.8 nitrate 7 35.7 756
D WB 21.1-24.7 7.2-7.8 nitrate 7 1100 µg/g 0.025^e
D WB 21.1-24.7 7.2-7.8 nitrate 28 35.7 1903
D WB 21.1-24.7 7.2-7.8 nitrate 28 1300 µg/g 0.05^e

Snail 6-15 mm C WB 15 7.1-7.7 nitrate 28 32 3750 Spehar
(Physa et al.
integra) (1978)

Amphipod 5-7 mm C WB 15 7.1-7.7 nitrate 28 32 6250 Spehar
(Gammarus et al.
pseudolimnaeus) (1978)

Caddisfly naiad C WB 15 7.1-7.7 nitrate 28 32 8400 Spehar
(Brachycentrus 5-8 mm et al.
sp.) (1978)

Stonefly naiad C WB 15 7.1-7.7 nitrate 28 32 7800 Spehar
(Pteronarcys 20-40 mm et al.
dorsata) (1978)

Stonefly naiad D WB 3-9 7.0-7.2 nitrate 14 1080 656 Nehring
(Pteronarcys (1976)
californica)

Mayfly naiad D WB 3-9 7.0-7.2 nitrate 14 4900 14 913 Nehring
(Ephemerella (1976)
grandis)

Carp 10-14 g A viscera 14.5-16.5 6.9 nitrate 2 10 000 4200 Muramoto
(Cyprinus 10-14 g A gills 14.5-16.5 6.9 nitrate 2 10 000 304 (1980)
carpio)
---------------------------------------------------------------------------------------------------------

Pumpkinseed 10-20 g A WB 18-20 6.0 nitrate 8 40 4.88^f Merlini
sunfish & Pozzi
(Lepomis 12-21 g A WB 18-20 7.5 nitrate 8 40 1.86^f (1977)
gibbosus)

Goby 6-38 g A spleen 20-25 acetate 8 265 79.4 Somero
(Gillichthys 6-38 g A gills 20-25 acetate 8 265 78.5 et al.
mirabilis) 6-38 g A fins 20-25 acetate 8 265 78.5 (1977)

Guppy 150-200 mg D WB 21.1-24.7 7.2-7.8 nitrate 7 3.8 654 Vighi
(Lebistes 150-200 mg D WB 21.1-24.7 7.2-7.8 nitrate 7 4.6 1081^g (1981)
reticulatus) 150-200 mg D WB 21.1-24.7 7.2-7.8 nitrate 7 13 µg/g 0.38^e
150-200 mg D WB 21.1-24.7 7.2-7.8 nitrate 28 3.8 1072
150-200 mg D WB 21.1-24.7 7.2-7.8 nitrate 28 4.6 3459^g
150-200 mg D WB 21.1-24.7 7.2-7.8 nitrate 28 23 µg/g 0.7^e
150-200 mg D WB 21.1-24.7 7.2-7.8 nitrate 7 33.5 197
150-200 mg D WB 21.1-24.7 7.2-7.8 nitrate 7 35.5 367^g
150-200 mg D WB 21.1-24.7 7.2-7.8 nitrate 7 27 µg/g 0.48^e
150-200 mg D WB 21.1-24.7 7.2-7.8 nitrate 28 33.5 359
150-200 mg D WB 21.1-24.7 7.2-7.8 nitrate 28 25.5 1015^g
150-200 mg D WB 21.1-24.7 7.2-7.8 nitrate 28 68 µg/g 0.52^e

Rainbow trout 1.0 g D WB 14-15.8 7.7-8.1 tetra- 7 3.5 725.7^d Wong
(Salmo methyl et al.
gairdneri) (1981)
---------------------------------------------------------------------------------------------------------
^a A = static conditions (water changed for duration of study); B = water renewed daily;
C = flow-through conditions (lead concentration in water continuously maintained).
^b WB = whole body; WP = whole plant.
^c Bioconcentration factor = concentration in organism/concentration in medium (calculated on a dry
weight basis unless otherwise stated).
^d Wet weight.
^e Calculated on lead content of food source; alga for Daphnia and Daphnia for guppy.
^f Based on radioactive tracer.
^g Exposure period from sowing of seed to 30 days post-emergence.

Table 2. Accumulation of lead into terrestrial organisms
---------------------------------------------------------------------------------------------------------
Organism Age Route Organ^a Compound Dura- Exposure Bioconcen- Reference
tion (mg/kg) tration
(days) factor
---------------------------------------------------------------------------------------------------------
Corn soil shoots nitrate 30^b 4233 0.07 Zimdahl et al.
( Zea mays) soil roots nitrate 30^b 4233 0.33 (1978)
soil shoots sulfate 30^b 4564 0.05 Zimdahl et al.
soil roots sulfate 30^b 4564 0.08 (1978)

Sugarbeet soil shoots nitrate 30^b 4233 0.23 Zimdahl et al.
(Beta vulgaris) soil roots nitrate 30^b 4233 1.3 (1978)
soil shoots sulfate 30^b 4564 0.04 Zimdahl et al.
soil roots sulfate 30^b 4564 0.15 (1978)

Bean soil shoots nitrate 30^b 4233 0.08 Zimdahl et al.
(red kidney) soil roots nitrate 30^b 4233 1.0 (1978)
soil shoots sulfate 30^b 4564 0.01 Zimdahl et al.
soil roots sulfate 30^b 4564 0.07 (1978)

Wheat soil shoots nitrate 30^b 4233 0.02 Zimdahl et al.
(Tritium aestivum) soil roots nitrate 30^b 4233 0.2 (1978)
soil shoots sulfate 30^b 4564 0.009 Zimdahl et al.
soil roots sulfate 30^b 4564 0.07 (1978)

Earthworm sewage WB acetate 35 2500 0.07 Hartenstein
(Eisenia foetida) et al. (1980)

American kestrel nestling oral kidney metallic 10 25^d 0.084^c Hoffman
(Falco sparverius) nestling oral liver metallic 10 25^d 0.05^c et al. (1985a)

Starling adult oral kidney triethyl 11 2.85^d 0.65^c Osborn
(Sturnus vulgaris) adult oral kidney trimethyl 11 2.85^d 1.9^c et al. (1983)
---------------------------------------------------------------------------------------------------------
^a WB = whole body;
^b Exposure period from sowing of seed to 30 days post-emergence.
^c Wet weight;
^d mg/kg per day.
4.1.2. Aquatic organisms

Appraisal

In contaminated aquatic systems, almost all of the lead is tightly
bound to sediment. Only a minor fraction is dissolved in the water,
even in the interstitial water.

The lead uptake by fish reaches equilibrium only after a number of
weeks of exposure. Lead is accumulated mostly in gill, liver, kidney,
and bone.

Fish eggs show increasing lead levels with increased exposure
concentration, and there are indications that lead is present on the
egg surface but not accumulated in the embryo.

In contrast to inorganic lead compounds, tetraalkyllead is rapidly
taken up by fish and rapidly eliminated after the end of the exposure.

Aickin & Dean (1978) exposed 47 bacterial strains and 9 strains of
fungi to 300 mg lead/litre (as lead acetate), for 48 h and 7 days,
respectively, during the stationary phase of the growth cycle. The
uptake of lead was 0.1% to 36% of the dry weight in the bacterial
strains and 4% to 19% in the fungi, and was greater than the uptake of
copper or cadmium in comparable experiments. When the uptake in 10
bacterial strains was compared, using other, less soluble sources of
lead, a general reduction in the amount of lead accumulated was found.
Even with lead nitrate, which is soluble, there was reduced uptake in
seven out of ten strains. Very little lead was taken up when the metal
was added to the medium as lead tetraphenyl. Metallic lead was taken
up by many strains to a greater extent than either lead sulfide or the
lead oxides.

Four aquatic plant species were exposed by van der Werff & Pruyt
(1982) to lead nitrate at concentrations of 1 and 10 µmol/litre for
41 to 46 days and 70 to 73 days. They found that, at both harvest
times, the submerged Elodea nuttallii and partly-submerged Callitriche
platycarpa had a higher tissue lead content than the floating species
Spirodela polyrhiza and Lemna gibba. Lead was found in the shoots,
roots, and rosettes of Callitriche in descending order after 43 days
at both concentrations of lead. Roots contained 3 mg lead/kg
after exposure to 1 µmol/litre and 13.0 mg/kg after exposure to
10 µmol/litre. Shoots contained nearly 2.5 times more lead than
roots, and rosettes less than half as much. Nakada et al. (1979)
exposed the submerged plant Elodea nuttallii to lead concentrations of
0.025, 0.05, 0.1, and 0.5 mg/litre. After 30 days, the lead content of
the plants (with roots removed) was calculated on a dry weight basis.
Concentration factors were 6200, 4300, 2800, and 630, respectively.
When lead accumulation was monitored in a mixed solution of lead,
cadmium, copper, and zinc, the concentration factor was found to be
lower than when lead alone was given.

In studies by Kay et al. (1984) the water hyacinth (Eichhornia
crassipes) was exposed to solutions containing lead nitrate at 0 to
5 mg lead/litre for 6 weeks. The accumulation of lead was dose-related
and in the order of roots > stems > leaves. Lead concentrations at
similar levels of exposure were only slightly greater after 6 weeks
than after 3 weeks. The highest level (5467 mg/kg dry weight) was
observed in roots after 6 weeks, following exposure to 5 mg/litre. The
results were a compilation of two studies run in the spring and autumn
in Florida, USA, lead uptake being consistently higher in the autumn.
Kay & Haller (1986) found a concentration factor of 5.89 in water
hyacinth leaves at a water concentration of 1 mg/litre. The highest
concentration factor was observed at the lowest dose tested; this might
indicate that there is a limit on the maximum uptake of the metal by
this plant. In further studies, Kay & Haller (1986) exposed water
hyacinth to lead nitrate (0 to 5 mg lead/litre) for 4 weeks. Water
hyacinth weevils feeding on the leaves, which had been exposed to 5 mg
lead/litre, showed concentration factors of 8.89 and 4.5 over the water
and the leaves, respectively.

When Meyer et al. (1986) exposed dragonfly larvae to lead nitrate
at 20 µg lead/litre for 6 weeks at 15 °C, they found significant
accumulation of lead in the fat, midgut, and rectum (0.55, 0.38, and
0.42 mg/kg, respectively). No significant lead residues were found in
the brain. The highest levels (1 mg/kg wet weight) were found in the
integument. However, this result is not significantly different from
controls, which also showed high lead levels (0.8 mg/kg) in the
exoskeleton.

In studies by Pringle et al. (1968), mature eastern oysters were
exposed to lead in the water at 25, 50, 100, or 200 µg/litre for
49 days. The final concentrations of lead in soft tissues were 17,
35, 75, and 200 mg/kg, respectively. This represents a lead uptake of
0.35, 0.71, 1.50, and 4.00 mg/kg per day for the four exposure levels,
respectively.

Coombs (1977) exposed mussels ( Mytilus edulis ) to lead, either as
nitrate or complexed with citrate, humic and alginic acids, or pectin,
for 13 days at 0.1 mg lead/litre. All tissues showed increasing
absorption of the metal over time, but highest concentrations were
found in the kidney (Table 1). The uptake rate and total accumulation
of lead in all tissues were three to four times higher with lead
citrate than with nitrate. The other complexes were not so effective
in increasing lead uptake; at best they produced 1.5- to 2-fold
increases.

When Anderson (1978) exposed crayfish (Orconectes virilis) to lead
acetate at concentrations of 0, 0.5, 1, and 2 mg lead/litre, he found,
over the 40-day exposure period, a marked increase in the lead content
of both gills and exoskeleton as water concentration and exposure time
increased. There was also an increase in the lead concentration in
muscle and viscera, but this was not significantly affected by either
treatment concentration or length of exposure.

Ray et al. (1981) exposed three species of marine invertebrates,
Nereis virens, Crangon septemspinosa, and Macoma balthica, to two
sediments which contained different amounts of lead. The sediments had

no added lead but were collected from different areas; they also
contained different amounts of other metals (copper, zinc, and
cadmium). Sediment A (48% sand; lead at 96.2 mg/kg dry weight)
contained lower levels of all metals than did sediment B (33% sand;
lead at 243.9 mg/kg dry weight). Animals were exposed to the sediment
for 30 days. Although N. virens showed no increase in lead tissue
concentrations in sediment A, other species in sediment A, and all
species in sediment B, revealed tissue lead increases over time.
Concentration factors ranged from 0.01 to 0.06, higher tissues levels
being attained after exposure to sediment B.

In a similar study, Lewis & McIntosh (1986) exposed the freshwater
isopod Asellus communis to two contaminated sediments in water at three
different pH levels for 20 days. Sediment A, a clay loam, contained
higher metal levels (lead at 367 mg/kg dry weight) than sediment B, a
silt loam (lead at 266 mg/kg dry weight). Higher lead levels were
found in the corresponding interstitial water in sediment A (lead at
10.2 µg/litre and 5.1 µg/litre for A and B sediments, respect-
ively). Lead accumulation from sediment was significant in sediment A
at pH 4.5 and 5.5 but not at pH 7.5, and from water, only at pH 4.5.
After 20 days exposure to sediment A at pH 4.5, concentration factors
were 1.4, in terms of sediment, and 39 000, in terms of water,
corresponding to lead levels in Asellus of 510 mg/kg dry weight. There
was no significant accumulation from sediment B.

Maddock & Taylor (1980) investigated the uptake of organolead
compounds by shrimp, mussel, and dab (a flatfish) in short-term
experiments and in mussel and dab in long-term experiments. For the
short-term exposure, shrimps ( Crangon crangon ), mussels (Mytillus
e dulis), and dabs (Limanda limanda) were held in concentrations of
teramethyl-, tetraethyl-, trimethyl-, and triethyllead up to the level
of the 96-h LC₅₀. This experiment measured the lead content of
animals used in the tests to determine acute toxicity. Results should,
therefore, be treated with caution because of some mortality at the
higher end of the range. Bioconcentration factors were higher for
tetraalkyllead than for trialkyllead; they lay between 20 and 650 for
the two tetraalkyl compounds in the three species, and between 1 and 24
for the two trialkyl compounds in the same three species. Mussels
exposed to either 0.01, 0.05, or 0.10 mg trimethyllead chloride/litre
(96-h LC₅₀ = 0.5 mg/litre) showed maximum uptake of lead within
9 days, and further exposure over 35 days failed to cause any further
tissue accumulation of lead. Uptake was dose-related; the mean tissue
content after exposure at 0.10 mg/litre for 21 days was 68 mg/kg wet
weight, representing a bioconcentration factor of 90. The greatest
tissue concentration occurred in the gill with the digestive gland,
gonad, and foot containing progressively less lead. Loss of lead was
rapid when the animals were transferred to clean water, with a mean
half-time of 3 to 4 days. Results for triethyllead chloride uptake and
loss by mussels were very similar. The authors conducted a comparison
between uptake of organic and inorganic lead in mussels; it is clear
that inorganic lead is accumulated to a much greater extent. Dabs were
exposed to either 1.0 or 2.0 mg trimethyllead/litre (96-h LC₅₀ =
24.6 mg/litre), or to either 0.1 or 0.2 mg triethyllead/litre (96-h
LC₅₀ = 1.17 mg/litre) for 41 days. With the exception of liver
uptake of trimethyllead, where equilibrium was reached after 20-days,
uptake into liver and muscle was linear over this period. Uptake by

liver and muscle was similar, with average tissue lead levels at around
30 mg/kg wet weight; this represented bioconcentration factors of 2 for
trimethyl- and 12 for triethyllead. Loss was slow with half-times in
excess of 41 days where these could be determined.

In studies by Holcombe et al. (1976), brook trout (Salvelinus
fontinalis) were exposed to lead nitrate concentrations of 0.9 to
474 µg lead/litre for three generations over a period of 3 years.
Gill, liver, and kidney tissues of first and second generation trout
accumulated the greatest amount of lead. In the first generation
fish, these organs appear to reach equilibrium after 20 weeks
exposure to 235 and 474 µg lead/litre, but not at lower concen-
trations. A equilibrium of lead residues was reached in liver and
kidney tissue from second generation fish after 70 weeks of exposure to
119 µg lead/litre. Lead residues in gill tissue continued to
increase throughout the 100 weeks of the first and second generations.
In the third generation, samples of eggs at spawning, and alevins, 4
weeks after hatch, showed that lead residues increased with higher
exposure concentrations. Although eggs showed increasing lead levels
with increased exposure concentrations, newly hatched alevins had
negligible residues. This indicates that the lead was present in the
egg membrane but not accumulated by the embryo. Juvenile alevins
accumulated lead up to an age of 8 weeks and then showed a reduced
concentration of lead after 12 weeks. It is not clear from the results
whether this represents lead loss or simply a reduced rate of uptake in
the larger fish.

Merlini & Pozzi (1977) exposed the pumpkinseed sunfish to lead
nitrate (traced with ²⁰³Pb) at 40 µg lead/litre for up to 8 days at
pH 6.0 and 7.5. The fish accumulated nearly three times as much lead
from water at the lower pH (Table 1).

In studies by Hodson et al. (1978b), 4-month-old rainbow trout
were exposed to nominal concentrations of lead between 0 and
1000 µg/litre (at pHs of 6, 8, and 10 for 3 days, and 7, 8, and 9
for 2 days). It was found that blood lead levels increased as the pH
of the test water decreased from 10 to 6. The highest blood lead level
(approximately 10 000 µg/litre) was recorded after exposure at pH 6
and a water concentration of 180 µg lead/litre. This represents a
concentration factor of about 50 in blood over water. The authors
calculated that a decrease by a pH unit of 1.0, from any reference pH,
resulted in an increase of blood lead by a factor of 2.1. Blood lead
was found to be in equilibrium with lead in the water within 48 h of
exposure.

Somero et al. (1977) exposed the estuarine teleost Gallichthys
mirabilis to lead acetate concentrations of 2650 mg lead/litre for 36
days in 100% sea water (3.36% salinity) and in 75%, 50%, and 25% sea
water. The lead content of all tissues studied showed an increase with
decreasing salinity. Highest levels were in the spleen, ranging from a
concentration factor (on a dry weight basis) of 74.4, for 100% sea
water, to 137.7, for 25% sea water. The same authors also exposed the
fish to two different temperature regimes, 10 °C and 20-25 °C, for
42 days in normal sea water. They found that a higher temperature
resulted in a higher tissue lead content.

Muramoto (1980) held carp (Cyprinus carpio) for 48 h in lead
nitrate concentrations of between 0 and 20 mg lead/litre, with and
without one of the three complexans, EDTA, NTA, or DTPA. The accumu-
lation of lead in both viscera and gills was dose-related, with the
highest levels for viscera and gills being 86 000 and 4560 mg/kg dry
weight, respectively. The complexans reduced the uptake of lead at all
dose levels. Concentrations in viscera ranged from 399 to 690 mg/kg
for the three complexans, and 298 to 645 mg/kg in gills, after exposure
to 20 mg/litre lead (which had given the above levels without chelating
agents). It is not clear whether the levels of lead in the gills
represented uptake into the tissue or adsorption onto the exterior
surfaces.

Wong et al. (1981) exposed rainbow trout to tetramethyllead at
24 µg/litre for up to 10 days. Because of the high volatility and
low water solubility of the compound, the authors designed an apparatus
specifically for the test. The water was changed completely every 2 h
in a flow-through system to which the tetramethyllead was continuously
added. They found that most of the alkyllead was accumulated in the
intestinal lipid (concentrations of 63 to 140 mg/kg wet weight),
followed, in decreasing order, by gills, skin/head, and air bladder.
They also calculated uptake rates and depuration rates for tetra-
methyllead in rainbow trout. The uptake rate was greatest (1 µg/g of
fish/day) at the beginning of exposure, and reached equilibrium by
day 7. When exposure stopped and the fish were returned to clean
water, levels of tetramethyllead in the tissues decreased rapidly over
3 days and then declined more slowly. Concentrations of alkyllead in
tissues had returned to pre-exposure levels within 1 week. Rate
constants for loss were 0.58/day for intestinal lipid and 0.29/day for
skin and head.

In studies by Ireland (1977), toads (Xenopus laevis) were fed with
live earthworms containing 10, 308, or 816 mg lead/kg for 4 or 8 weeks.
Toads fed the diet containing 10 mg/kg for 8 weeks had significantly
less lead in kidney and liver than toads fed 308 mg/kg diet for 4 weeks
(or 308 mg/kg for 4 weeks, followed by 816 mg/kg for 4 weeks). Bone
and skin lead levels were significantly less after 4 weeks on 10 mg/kg
than after 4 weeks on 308 mg/kg diet. No other significant difference
was observed. Muscle lead levels did not vary significantly between
treatments. Individual organ analysis, within groups, showed high lead
levels in kidney, bone, and liver, but low values in skin and muscle.
The highest levels were found in kidney and were 19.1, 73.3, and
81.3 mg/kg dry weight at the three dose levels, respectively.

4.1.3 Terrestrial organisms

Appraisal

In bacteria, the majority of lead is associated with the cell wall.
A similar phenomenon is also noted in higher plants. Some lead that
passes into the plant root cell can be combined with new cell wall
material and subsequently removed from the cytoplasm to the cell wall.
Of the lead remaining in the root cell, there is evidence of very
little translocation to other parts of the plant because the concen-
tration of lead in shoot and leaf tissue is usually much lower than in
root. Foliar uptake of lead occurs, but only to a very limited
extent.

Lead shot is typically trapped in the gizzard of birds where it is
slowly ground down resulting in the release of lead.

The tetravalent organic form of lead is generally more toxic than
the divalent, inorganic form, and its distribution in organisms may not
aspecifically follow calcium metabolism.

When Tornabene & Edwards (1972) incubated two species of bacteria,
Micrococcus luteus and Azotobacter sp., in a medium with a suspended
dialysis bag containing lead bromide, the two species took up 490 and
310 mg lead/g whole cells (dry weight), respectively. The authors
analysed subcellular fractions of the bacteria and found 99.3% and
99.1%, for the two bacteria, respectively, in the cell wall plus
membrane fraction. The remainder of the lead was found in the
cytoplasm. The same authors, Tornabene & Edwards (1973), located
electron-dense inclusions in cell membranes of Micrococcus. Tornabene
& Peterson (1975) showed that the lead was not specifically bound to
lipid fractions in the cell membrane but that the membrane provided a
suitable substrate in which aggregations of lead could form.

Zimdahl et al. (1978) sowed maize (Zea mays), sugarbeet, bean, and
wheat in soil dosed with lead nitrate or sulfate (0 to 5000 mg
lead/kg). Lead uptake into shoots and roots (on a dry weight basis)
was measured 30 days after emergence. It was found that more lead was
taken up into the roots than into the shoots (Table 2). Although the
data are not conclusive, the authors suggest that less lead is taken up
when soil is treated with lead sulfate than with lead nitrate. In a
2-year study, Baumhardt & Welch (1972) grew Zea mays in the field where
lead acetate had been applied to the soil at rates of 0 to 3200 kg
lead/ha. The lead contents of the plants for the 0 and 3200 kg/ha
treatments were, respectively, 2.4 and 37.8 mg/kg for young whole
plants, 3.6 and 27.6 mg/kg for leaves at tasselling, and 4.2 and
20.4 mg/kg for whole plants at grain harvest. The lead content of
grain was unaffected by any of the applications.

Lane & Martin (1977) investigated the uptake of lead into the seed
and seedlings of the radish, the location of the lead being determined
histochemically. The intact testa prevented uptake of lead into the
embryo, but when the testa ruptured during germination, the radicle
took up lead readily, as did the rest of the tissues (endosperm). As
the seedling developed, lead was concentrated in the radicle and the
hypocotyl, with relatively little being transported to the shoot.

In studies by Malone et al. (1974), Zea mays was exposed to lead,
in either a hydroponic solution or in distilled water, in four
different forms: citrate, chloride, nitrate, or EDTA chelate. Lead
concentrations in the solutions ranged from 10 to 1000 mg/litre. The
uptake of lead was followed using phase-contrast light and electron
microscopy. Roots generally accumulated a surface precipitate of lead
salts as fairly large crystals. Lead was slowly absorbed into the
roots and appeared as much smaller crystals associated primarily with

the cell walls. The lead was taken up by dictyosome vesicles which
migrated towards the cell wall and ultimately formed extensions of the
cell wall. These vesicles fused together to encase the lead crystals
within the cell wall material. In some cases, these inclusions
projected into the cell cytoplasm. Similar deposits were found in
shoots and leaves as the lead was slowly transported throughout the
plant. Lead was never associated with the phloem or its companion
cells and never with the guard cells of the epidermal stomata. Thus
lead was excluded from the biochemically active plasmalemma.

Hemphill & Rule (1975) applied solutions of radioactively labelled
lead nitrate to the leaves of lettuce and radish for a period of 25
days and then grew the plants on for a further 25 days. The lead
content and distribution were assessed using scintillation counting and
autoradiography. There was some absorption of lead into the leaves and
some subsequent translocation, but this was very small. The percentage
translocation of applied lead (expressed in terms of the total lead
applied) was not more than 0.2%, and generally much less than this,
except where contamination with the applied material had possibly
occurred. It is not clear whether the total recovery of the labelled
lead was estimated.

Dollard (1986) conducted a similar experiment with glasshouse-grown
radish, carrot, and French bean plants. In radish, a small amount
(0.05% to 0.28%) of the applied lead was transported to the swollen
root. This movement occurred through intact or damaged cuticle, and
there was some indication that damage to the leaf surface enhanced lead
uptake. Carrot plants absorbed 0.43% of foliar-applied lead, but
transported it no further than the leaf petiole over the 8- to 12-week
period of the experiment. The transport of lead to the tap root was
<0.01% of that applied. For the French bean, no movement of lead into
pod or seed was detected. The author estimated that up to 35% of root
lead in radish could be accounted for by foliar absorption, whereas in
carrot this would be no more than 3% (based on lead deposition rates
from the atmosphere close to roads).

Beyer et al. (1982) monitored the uptake of metals into earthworms
from soil treated with sewage sludge. In all treatments, the concen-
tration of lead in the earthworms correlated with the concentration in
soil. There was, however, no bioconcentration of lead into worms,
concentration factors being consistently less than 1.0 for soil lead
levels ranging between 16 and 43 mg/kg.

In studies by Straalen & Meerendonk (1987), adult collembola
(Orchesella cincta), collected from an unpolluted pine forest and
cultured in the laboratory, were fed with green algae on paper disks.
Lead nitrate solution was added to the food suspension. The
concentration of lead in the food ranged from 1600 to 2200 mg/kg dry
weight. The study lasted for 8 weeks, contaminated food being fed for
the first 4 weeks and clean food for the second 4 weeks. Lead concen-
trations in the collembola fluctuated within wide limits during the
accumulation phase. An average steady state was achieved after
approximately 4 weeks, with lead concentrations of approximately
0.2 mg/kg dry weight. This value was obtained for worms with the gut
contents cleared. The authors identified three components to the body
lead content: gut contents, a "fast body burden", and a "slow body

burden". The fast component appeared to be lost during moults.
Calculated half-times for loss of lead from the three components were
as follows: 0.34 days for gut content, 7.37 days for "fast body
burden" and 21.66 days for "slow body burden".

Irwin & Karstad (1972) exposed adult mallard drakes to 17.8, 89, or
178 g of particulate lead per m² in a simulated marsh environment for
14 weeks. Lead shot (no. 5) were scattered over the penned area which
simulated a marsh area of puddled mud. The number of shot actually
ingested per bird is not clear. Lead levels in muscle, liver, and bone
increased with increasing exposure. Liver and bone contained higher
concentrations; after 14 days exposure to 178 g/m², lead concen-
trations in liver and bone were 28.4 mg/kg wet weight and 176 mg/kg dry
weight, respectively.

When Clemens et al. (1975) dosed adult mallard with five lead shot
(no. 6) and monitored tissue concentrations of lead over a period of
20 days, they found higher lead tissues levels in birds on a high-fibre
diet (12.5% fibre) than on a low-fibre (3%) diet. The highest levels
were found in the bone after 16 days (570 mg/kg dry weight) and in the
kidney after 12 days (225 mg/kg wet weight), both on the high-fibre
diet. In the birds on a low-fibre diet, lead levels peaked in all
tissues after 2 to 4 days and then declined. In birds on high-fibre
diets, the same was true only for blood. Lead levels did not peak
until 12 days in liver, kidney, leg muscle, and bone. In both groups,
the pectoral muscle, after an initial rise, showed fluctuating levels
with no consistent pattern.

Finley et al. (1976) dosed male and female mallard with either one
(no. 4) lead shot or one (no. 4) lead/iron combination shot (with 47%
lead). The birds were observed for 4 weeks. The lead levels in liver,
kidney, blood, and bone were twice as high in birds dosed with lead
alone, reflecting the relative amounts of the metal consumed. Females
had double the lead levels of males, except in bone, where the
difference was a factor of ten. The levels in females dosed with lead
shot were 1.15, 3.53, 0.71, and 112.27 mg/kg for liver, kidney, blood,
and bone, respectively. Similar trends were found in eggs laid during
the period, with the birds dosed with lead shot laying down more lead
in the eggs. The egg contents and shell contained 0.5 and 2.8 mg
lead/kg, respectively, after dosing with lead shot.

When mallard were dosed with one lead shot (no. 4), the pre-dosing
blood lead level was 83 µg/litre and rose to 317 µg/litre 1 month
after dosing. Four weeks after male and female mallard were dosed
similarly, lead accumulation was significantly greater in bones with a
high medullary content (femur and sternum) than in bones with a lower
content (ulna/radius and wing bones). Females always contained higher
bone residues than males. The femurs of laying females averaged
488.8 mg lead/kg dry weight compared with 113.6 mg/kg in non-laying
females and 9.4 mg/kg in males. When birds were dosed with a second
lead shot and analysed 4 weeks later, levels in laying females were
unchanged but levels in males had risen by a factor of three. The
authors suggested that a saturation level had been reached in the
females (Dieter & Finley, 1978; Finley & Dieter, 1978).

Buggiani & Rindi (1980) dosed adult domestic ducks with 24 lead
shot (no. 6) once a week for 5 weeks. A second group were dosed for
6 weeks with the same number of shot plus EDTA (1 mmol/kg body weight).
At the end of the experiment, lead concentrations were measured in the
blood and the nasal glands. Blood lead was three times higher than
control levels in both groups. The ratio of nasal gland lead to blood
lead was 1 for birds from both groups. Immediately after treatment
with lead shot, this ratio was 3 suggesting that the nasal gland is a
source of lead excretion in ducks.

In studies by Osborn et al. (1983), starlings (Sturnus vulgaris)
were orally dosed with solutions of triethyllead or trimethyllead
chlorides at concentrations of 0, 200, and 2000 µg/litre per day for
11 days, or until death. All the birds in the low-dose group survived
for the full 11 days; birds dosed with trimethyllead accumulated more
lead in the brain, kidney, and liver than did triethyllead-treated
birds. The highest lead levels were found in the kidney: triethyllead-
treated birds contained 1.85 mg/kg wet weight and trimethyllead-treated
birds contained 5.38 mg/kg wet weight in their kidneys. Birds given
the high dose all died within 6 days, and had higher lead levels in all
tissues than birds given the lower dose. In these dead birds, highest
lead concentrations were found in the liver of triethyllead-treated
birds (40.2 mg/kg wet weight) and the liver and kidney of trimethyl-
lead-treated birds, (32.4 and 30.2 mg/kg wet weight, respectively).
Osborn (1979) pointed out that metal levels in different tissues of
birds should be treated with caution since they depend on many
different factors. In particular, levels in dead or dying birds are
not comparable to those in healthy birds because of redistribution
prior to death. Also, it is not possible to compare exposure of birds
in the field with those in the laboratory simply by measuring tissues
levels.

4.2 Accumulation in the Field

4.2.1 General considerations

Appraisal

In shellfish, lead concentrations are higher in the calcium-rich
shell than in the soft tissue; they relate to the concentrations in
sediment.

Lead concentrations in some marine fish are higher in gills and
skin than in other tissues, but this may be largely due to adsorption.
Liver levels increase significantly with age.

In dolphins, lead is transferred from mothers to offspring during
fetal development and lactation. This might be related to the calcium
metabolism.

Ayling (1974) sampled the oyster Crassostrea gigas from the Tamar
River in Tasmania and found mean dry weight lead concentrations in
oysters and mud samples of 0 to 135 mg/kg and 4 to 1500 mg/kg,
respectively. The author stated that lead was not taken up through any
physiological demand, but was randomly incorporated at the sites
containing high concentrations in the mud. When analysing the bivalve
Elliptio complanata from the Great Lakes for lead levels, Dermott &
Lum (1986) found higher levels (10.2 to 25.2 mg/kg) in the shell than
in soft tissues (ND to 2.2 mg/kg). Lead was significantly higher in
the outer periostracum of the shell than in the inorganic prismatic
layer. In spite of high levels at one site contaminated by effluent,
lead was not deposited in the prismatic shell layer. Sediment levels
in the sampling areas ranged from 29 to 103.3 mg lead/kg. Pringle et
al. (1968) found low levels of lead (<0.2 mg/kg in soft tissues) in
estuarine molluscs. There was no seasonal variation in lead concen-
trations.

Enk & Mathis (1977) detected lead in all components of a stream
with no industrial contamination. The levels were as follows: water
(<0.5 mg/litre), fish (2.47 to 2.88 mg/kg), sediment (8.3 mg/kg),
aquatic insects (6.83 to 12.59 mg/kg), snails (13.64 mg/kg).

When Gilmartin & Revelante (1975) analysed anchovy and sardine from
the Adriatic Sea, the highest lead concentrations were found in the
gills (6.8 and 6.5 mg/kg wet weight, respectively) and skin (4.5 and
4.3 mg/kg wet weight, respectively). Higher liver lead concentrations
in anchovy occurred later in the year. For most of the period of
study, lead was not detectable in the muscle, digestive system, or
liver. Perttila et al. (1982) found that lead increased significantly
with age in the Baltic herring (Clupea harengus).

Van Hook (1974) calculated concentration factors for lead into
earthworms sampled from the field. Factors were below 1 (range 0.11 to
0.3) for soil lead levels ranging between 15 and 50 mg/kg dry weight.
Bagley & Locke (1967) analysed wild birds of many species, from the
eastern USA, for tissue lead levels. The majority of birds examined
were water-fowl. All the birds were healthy and contained no lead
shot. Mean liver residues of lead ranged from 0.5 to 3.7 mg/kg wet
weight and mean tibia residues from 2.0 to 13.0 mg/kg wet weight.
Martin (1972) and Martin & Nickerson (1973) analysed starlings in the
USA for lead and found residues ranging from 0.4 to 13.3 mg/kg in 1970
and 0.12 to 6.6 in 1971.

In studies on the common porpoise (Phocoena phocoena) from the east
coast of Scotland, Falconer et al. (1983) found that lead residues were
below detectable limits (0.5 mg/kg). The sampled animals had died
after becoming entangled in cod nets. The tissues analysed were the
brain, liver, kidney, heart, and spleen. Honda et al. (1986) sampled
striped dolphin (Stenella coeruleoalba) and found significant accumu-
lation of lead in the bone of offspring during the suckling period.
Significantly more lead was found in adult males than females. The
authors suggested that lead was removed from the mother via the milk
and as the result of parturition. Lead levels ranged between 0.09 and
0.74 mg/kg wet weight.

4.2.2 Highways and urban areas

Appraisal

The lead in the soil and in vegetation decreases exponentially with
the distance from the road. Lead is also found in the sediments of
streams in the vicinity of highways.

Lead contamination increases lead levels in plants and animals in
areas close to roads. These levels are positively correlated with
traffic volume and proximity of roads.

There is extensive documentation on the occurrence of lead in soil
and organisms close to roads.

Khalid et al. (1981) analysed soil samples from different areas of
Baghdad, and found that mean levels ranged from 36 mg/kg for an
industrial area to 308 mg/kg for north-east Baghdad. It was also found
that lead concentrations were highest in areas of high traffic volume
and the city centre had higher levels than other areas.

Chow (1970) established, by isotopic composition, that lead
detected in soil and dried grass derived exclusively from alkyllead
compounds added to petrol. Wheeler & Rolfe (1979) established a double
exponential relationship between lead levels in vegetation and distance
from the road. The two exponents were assumed to represent particles
of different size. Larger particles were deposited within about 5 m of
the edge of the road surface. Smaller particles settled more slowly
and were deposited within 100 m of the road, though beyond 50 m from
the road surface there was little more than a background level of lead.
Lead in the smaller particles was more soluble than in the larger.
Lead levels were very high close to the road. At a traffic density of
8100 vehicles/day, lead concentrations of 1225 mg/kg soil and 196 mg/kg
vegetation were found within 0.3 m of the road. This declined rapidly;
soil levels were 526 mg/kg at 1 m, 93 mg/kg at 5 m, and 55 mg/kg at
10 m from the road, with similar falls in vegetation levels. Although
the soil had a high capacity to adsorb lead, an estimated 72-76% of the
total lead deposited had been lost from the soil by leaching or run-
off.

In a similar study (Ward et al., 1975) of a road in New Zealand
with a traffic density of 1200 vehicles/day, a similar distribution of
lead was noted. All lead deposited could be found within 100 m of the
road and within the upper 5 cm of the soil. The authors calculated
that the total deposition of lead since the introduction of leaded

petrol was around 250 g/metre of road length. Of this, 140 g lead
could be accounted for within 250 m of the road side and in the upper
6 cm of soil. Cannon & Bowles (1962) reported that lead levels depend
on the traffic volume on the roads and rise to 3000 mg/kg in grass near
major road intersections. Lead is also found in streams close to major
roads.

Van Hassel et al. (1979) reported little or no difference between
the lead concentration in water of roadside streams and that of streams
away from highways. There was, however, a significant increase in the
lead content of the stream sediment, to which lead is readily adsorbed.
Similar results were found by Mudre & Ney (1986) who investigated the
lead content of the sediment in a series of tributary streams running
into the Chickahominy River in Virginia. The same highway crossed all
streams. Levels of lead close to the road were significantly higher
than in upstream samples from all streams. Samples taken some distance
downstream did not differ from upstream ones; lead contamination was
very localized. There were marked differences between streams due to
various factors including drain-off from vegetation into the stream,
weather, stream flow rates, and traffic density at different times of
year.

Ash & Lee (1980) monitored lead in earthworms (two species) from
sites close to roads and from low-traffic areas in the United Kingdom.
The earthworms were purged of gut contents before analysis, and all
results are expressed in terms of dry weight. The control site in
rural Scotland showed lead levels of 0.96 and 0.31 mg/kg dry weight in
the two species. Close to two major roads, levels were 130 and
341 mg/kg (for the A660 road) and 274 and 500 mg/kg (for the A1 road
with greater traffic density) for the two earthworm species, respect-
ively. A city recreational area gave levels of 32 and 76 mg/kg and a
site on farmland (300 m from the main A1 road) gave levels of 38 and
26 mg/kg for the two earthworm species, respectively. Goldsmith &
Scanlon (1977) measured lead concentrations (excluding gut contents) in
earthworms at 6, 12, and 18 m from two roads in Virginia, USA. The
roads had traffic volumes of 21 040 and 1085 vehicles/day, respect-
ively. Lead levels in earthworms were 51, 50, and 32 mg/kg dry weight
at 6, 12, and 18 m, respectively, from the busier road. At 12 and 18 m
from the less busy road, levels were 8.5 and 11.65 mg/kg, respect-
ively.

Price et al. (1974) found that sap-sucking, phytophagous, and
insectivorous insects contained, on average, 10.3, 15.5, and 25.0 mg
lead/kg, respectively, close to a road. In low-lead areas, the three
types of insect showed 4.7, 3.4, and 3.3 mg/kg, respectively. The
authors claim evidence for the concentration of lead through
food-chains. Giles et al. (1973) came to similar conclusions while
measuring lead in phytophagous and carnivorous insects. Beyer & Moore
(1980) reported that caterpillars feeding on black cherry leaves
contained 76% as much lead as did their food. More lead was found in
the insects close to the road than in those further away. Beyer (1986)
has questioned the bioconcentration of lead in road-side food-chains,
since no study has exhaustively monitored lead in prey and predators of
a recognized food-chain. Other explanations of the available data are
probable; different species of insects, both prey and predator, have
been shown to take up lead to very different degrees.

May & McKinney (1981) showed that lead concentrations in
Hawaiian fish, sampled from streams close to roads, ranged from 0.8 to
4.93 mg/kg wet weight in whole fish, with high levels corresponding to
high-traffic density. The species sampled included some bottom-
feeders, but were mainly fish of the open water. Ney & Van Hassel
(1983) measured the whole body lead content of six fish species sampled
from a stream flowing under a major highway. Fish were sampled close
to the bridge. The residues (means for species) ranged between 7.2 and
19.5 mg/kg dry weight, and species which live in the open water had
lower levels than bottom-feeding species. Levels of lead in sediments,
benthic invertebrates, and fish were higher at this site than upstream
or downstream of the road crossing, indicating localized binding of the
metal (Van Hassel et al., 1979, 1980).

Birdsall et al. (1986) measured lead concentrations in sediment and
in the tadpoles of bullfrogs (Rana catesbeiana) and green frogs (Rana
clamitans) taken from drains beside roads with different daily traffic
volumes and from ponds at least 0.4 km from the nearest road. Sediment
samples showed lead concentrations ranging from 7.8 to 40 mg/kg dry
weight for ponds and 18 to 940 mg/kg dry weight for highway drains.
These were usually 4 to 5 times greater than corresponding levels in
tadpoles. Levels in bullfrog tadpoles were 2.6 to 6.0 mg/kg and 0.7 to
270 mg/kg for the ponds and drains, respectively. Green frog tadpoles
contained 0.9 to 8.9 mg/kg in ponds and 4.8 to 240 mg/kg in drains.
There was a positive correlation between traffic volume and the lead
content of sediment and amphibians.

Ohi et al. (1974) determined lead levels in blood, femurs, and
kidneys of adult pigeons sampled from rural and urban sites in Japan.
Lead levels were highest in femurs, with means ranging from 16.5 to
31.6 mg/kg wet weight over three urban sites, while two rural sites
showed mean levels of 2.0 and 3.2 mg/kg. Blood levels showed a similar
trend; the urban sites gave 0.15, 0.33, and 0.33 mg/litre while the
rural sites showed 0.054 and 0.029 mg/litre. Kidney levels were lower,
and also showed a reduced lead level in rural areas. Hutton & Goodman
(1980) obtained similar results in pigeons in London, with differences
between central London, suburban London, and surrounding rural areas.
Getz et al. (1977) sampled four species of song birds from an urban
site and rural sites in Illinois, USA. The rural sites were chosen to
be at least 2 km from the nearest town and 50 m from any road. Highly
significant differences in lead content between urban and rural values
were found for all species and in all tissues (feathers, gut, liver,
kidney, and femur), except for lung and pectoral muscle, which showed
low lead content. Kidney levels in urban areas were 33.9, 98.5, 13.5,
and 25.0 mg/kg dry weight for house sparrow, starling, grackle, and
American robin, respectively, and in rural areas were 3.5, 3.6, 3.5,
and 7.3 mg/kg for the same species, respectively. Grue et al. (1986)
found lead levels 3 to 13 times higher in starlings (nestling and
adults) breeding near roads than in birds sampled from control sites.
There was a less pronounced, but still significant, difference between
similarly sited breeding colonies of swallows (Grue et al., 1984).

Jefferies & French (1972) measured the lead in the liver and whole
body of 101 small mammals of three species, Microtus agrestis,
Clethrionomys glareolus, and Apodemus sylvaticus, sampled from fields
or from roadside verges. The mean lead concentration of whole bodies

increased from 4.19 mg/kg dry weight for mammals trapped on woodland or
arable sites to 5.98 mg/kg on the verges of minor roads and 7.0 mg/kg
on the verges of a major road. Vegetation from the same sites averaged
33.4, 42.5, and 306.7 mg/kg dry weight, respectively. Goldsmith &
Scanlon (1977) trapped small mammals in three study areas of roadside
verges with different traffic densities. Significantly greater lead
levels were found in heavy traffic areas in individuals of three
species: Cryptotis parva, Microtus pennsylvanicus, and Peromyscus
leucopus. However, no significant difference between areas was found in
the shrew Blarina brevicauda. These species represent herbivores and
carnivores, with the shrews eating predominantly insect prey.
Carnivores had higher levels of lead than herbivores. Welch & Dick
(1975) found that lead levels in liver, kidney, and bone (but not those
in brain, lung, stomach, or muscle) of deer mice (Peromyscus
maniculatus) were related to the proximity to the road and to traffic
volume.

Quarles et al. (1974) found that the lead content of small mammals
increased with proximity to the road. In comparable areas, there was
22.7 mg/kg in the shrew Blarina, 16.3 mg/kg in the vole Microtus and
6.8 mg/kg in the mouse Peromyscus. The authors compared published
information on the food consumption, food choice, and habits of the
three species. The size of the home range was suggested as a
contributing factor to differing lead concentrations; the mouse has a
much more extensive range than the shrew or vole. Food type, with the
insectivorous vole taking in most lead, was also likely to be
important. Williamson & Evans (1972) analysed the lead content of a
wide variety of invertebrates from roadside verges and also of small
mammals which eat these invertebrates. They found no evidence to
indicate concentration of lead in food-chains. Although the
insectivorous shrews had higher lead levels than their herbivorous
neighbours, the shrews contained less lead per unit weight than did
their prey.

4.2.3 Industrial sources

Appraisal

Elevated lead levels are also found in terrestrial invertebrates
and vertebrates from contaminated areas.

Rains (1971) analysed the lead content of wild oats (Avena fatua)
growing in the vicinity of a smelter. The area had been subject for
more than 70 years to lead contamination from the smelter, which was
still in operation. The lead content of the plants increased
throughout the year, the lowest levels occurring during the active
growing season. Lead levels continued to rise after the ears were

fully formed and the upper portions of the plant were dry, and peaked
at 500 mg/kg dry weight. Some lead would be taken up from the soil,
but the predominant source of the lead would be atmospheric.

Mayes et al. (1977) measured lead uptake into a submerged aquatic
plant Elodea canadensis in two lakes. The control lake was far from
any industrial sources of metal, while the second lake received waste
water from an electroplating plant. Specimens of Elodea were anchored
in each lake into contaminated and non-contaminated sediments. Plants
grown in the same water, but in sediment from different sources, had
significantly greater lead content when grown on the contaminated
sediment. Similarly, Elodea accumulated more lead when grown in con-
taminated water, irrespective of the sediment. Thus, both water and
sediment are sources of lead for this plant. Samples grown in uncon-
taminated water and sediment accumulated lead concentrations of
5.2 mg/kg while those in contaminated water or sediment accumulated up
to 160.9 mg/kg.

Ruhling & Tyler (1970) analysed the lead content of mosses Hypnum
cupressiforme and Hylocomium splendens in different regions of
Scandinavia to monitor fall-out of industrial lead. They found
significantly higher levels of lead in H. splendens from southern
Sweden compared with northern Scandinavia (11 mg/kg) and also higher
levels in south-west (90 mg/kg) than in south-east Sweden (52 mg/kg).
The same pattern was found in lead levels of H. cupressiforme between
different areas of southern Sweden. The authors eliminated all other
possible sources of lead than anthropogenic ones.

Edelman et al. (1983) analysed earthworms (Lumbricus rubellus) and
soil samples, for lead content, near a zinc-smelting complex. Levels
in soil ranged from 14 to 430 mg/kg dry weight and in worms from 9 to
670 mg/kg dry weight. Although there was a significant correlation
between distance from smelter and levels of lead in soil and worms, it
was not as strong a relationship as for cadmium or zinc. Soil lead
content, soil pH, and soil organic matter together accounted for 70% of
variance in worm lead uptake. The authors found higher lead levels in
worms from soil of a lower pH and lower organic matter content. Lead
was estimated after clearing the worms of gut contents.

Bengtsson & Rundgren (1984) analysed ground-living invertebrates,
such as spiders, harvestmen, slugs, beetles, and ants, from metal-
polluted forest soils, at varying distances from a Swedish brass mill.
Mean lead levels were significantly higher in most of the species
within 650 m of the mill. Litter levels of lead of 600-1000 mg/kg were
found, dependent on the distance from the mill. Lead levels in litter
were 20-30 times less than zinc or copper levels.

Roberts et al. (1978) found significantly higher lead levels in
surface soil, vegetation, and invertebrates at two abandoned non-
ferrous mine spoil tips than in control areas. The two areas showed
lead at 8430 and 14 010 mg/kg dry weight soil, 120 and 249 mg/kg dry
weight vegetation, and 61.9 and 81.7 mg/kg dry weight invertebrates.
Four species of small mammals (Microtus agrestis, Apodemus sylvaticus,
Clethrionomys glareolus, and Sorex araneus) showed significantly
higher levels of lead when trapped in the contaminated areas. The
highest levels in M. agrestis were 45.3 and 42.8 mg/kg fresh weight,

for the two areas. When tissues of A. sylvaticus were analysed,
kidney, liver, bone, and brain contained significantly higher levels
than controls. Lead levels of 352 and 189 mg/kg dry weight for bone
compared with 11.5 and 21.1 mg/kg for the two control areas. Muscle
residues were not significantly different between areas. Similar
results were found in the tissues of A. sylvaticus living on smelter
waste (Johnson et al., 1978). Surface soil contained 4030 mg lead/kg
(control 76.1 mg/kg dry weight) and bone levels were 672 mg/kg dry
weight (control 34.2 mg/kg).

Cloutier et al. (1986) assayed the lead content of tissues from
meadow voles (Microtus pennsylvanicus), living on nickel or uranium
mine tailings. Soft tissue levels of lead were below detection limits
in most cases. Bone levels of lead were slightly elevated at the
uranium site, but not significantly. The highest levels were
21.9 mg/kg dry weight in sub-adults and 23 mg/kg in adults. No sex or
age differences were reported. There was a rise in bone lead levels
between winter and the following autumn in the lead-rich area (the
uranium site), but a fall over the same period at the nickel site and
control site.

Smith & Rongstad (1982) determined lead concentrations in the whole
body of Peromyscus maniculatus and M. pennsylvanicus from an active
zinc-copper mine and a proposed zinc-copper mine. P. maniculatus from
the proposed mining site showed lead concentrations of the same order
as in a non-mining control area. From the mining site, there were
consistently higher concentrations for both sexes and ages, juveniles
and adults. M. pennsylvanicus in the mining site showed no elevation in
lead content over controls.

4.2.4 Lead shot

Appraisal

Lead shot taken by birds into their gizzards is a source of severe
lead contamination. In organs, high levels of lead are found in blood,
kidney, liver, and bone.

Mudge (1983) analysed for lead 1620 livers and 1871 wing bones
from 23 species of British waterfowl (shot or found dead). The
highest levels of lead in the liver were found in birds with ingested
pellets in the gizzard. The species with the highest levels, excluding
those birds without ingested pellets in the gizzard, were gadwall
(11.3-22.0 mg/kg dry weight), mute swan (11.6-32.7 mg/kg), Bewick's
swan (73.0-109.9 mg/kg), and greylag goose (57.2-61.9 mg/kg). Of the
14 species of duck analysed for lead in the wing bone (and not
containing lead shot at the time of sampling), the highest levels were
in mallard (<5.0-472.9 mg/kg dry weight), teal (<5.0-298.8 mg/kg), and
wigeon (<5.0-175.9 mg/kg). The author also assayed 63 blood samples
from four species of waterfowl; the highest mean blood lead levels were
in whooper swan (4.6 mg/litre) and Bewick's swan (8.3 mg/litre).

Analysis for lead of mute swan blood samples, taken from swans
from contaminated and uncontaminated areas in the United Kingdom, has
revealed large differences between areas (NCC, 1981) (see also section
8.2). The highest level reported was from the River Trent, Nottingham

(3.75 mg/litre), and the lowest level was 0.08 mg/litre from the
Abbotsbury swannery, Dorset. Simpson et al. (1979) analysed various
organs of lead-poisoned mute swans found dead. The highest lead
levels were found in the kidney (350-6550 mg/kg dry weight), liver
(51-206 mg/kg), and bone (212-1255 mg/kg). These levels compared with
`healthy' control swan levels of 1-77 mg/kg kidney, 1-11 mg/kg liver,
and 21-41 mg/kg bone.

Anderson (1975) examined about 1500 waterfowl dying at Rice Lake,
Illinois, USA. When 96 lesser scaup, of which 75% had at least one
lead pellet in the gizzard, were analysed for lead, the mean levels
were 46 mg/kg, 66 mg/kg, and 40 mg/kg for liver, kidney, and wing bone,
respectively.

5. TOXICITY TO MICROORGANISMS

Appraisal

In general, inorganic lead compounds are of lower toxicity to
microorganisms than are trialkyl- and tetraalkyllead compounds.
Tetraalkyllead becomes toxic by decomposition into the ionic
trialkyllead.

One of the most important factors which influence the aquatic
toxicity of lead is the free ionic concentration, which affects the
availability of lead for organisms. The toxicity of inorganic lead
salts is strongly dependent on environmental conditions, such as water
hardness, pH, and salinity, a fact which has not been adequately
considered in most toxicity studies.

There is evidence that tolerant strains exist and that tolerance
may develop in others.

5.1 Toxicity of Lead Salts

Bringmann & Kuhn (1959a) reported a toxic threshold for lead, as
lead nitrate, of 1.3 mg/litre for the bacterium Escherichia coli,
related to cell numbers produced. Lead, as the nitrate or bromide,
had little effect on the growth of the human skin bacterium
Micrococcus luteus at a level of 600 µg/litre over 48 h (Tornabene &
Edwards, 1973). The latter authors recultured bacteria after the lead
treatment, with inocula transferred to fresh medium after 48 h of
growth. After 20 days of continuous growth, the cellular yield had
decreased to less than half that of a control culture. The pigmen-
tation of this characteristically yellow bacterium was reduced by this
time. Electron-microscopic examination of these cells indicated that
cytoplasmic material was leaking out. Lead is largely concentrated in
the cell membranes of bacteria, and could be seen as electron-dense
inclusions; membrane breakdown was usually seen in the area of lead
inclusions.

Gray & Ventilla (1971) cultured the ciliate Cristigera sp. on a
diet of bacteria (Pseudomonas sp.), both organisms having been isolated
from beach sand. Lead nitrate added to the cultures reduced the growth
rate, but did not inhibit growth at between 0.1 and 0.3 mg/litre. The
result was significant at the 5% level.

Monahan (1976) reported a 50% reduction in cell numbers of the
freshwater alga Selenastrum capricornutum after 7 days exposure to lead
in the culture medium at a concentration of 0.5 mg/litre medium.
Increasing the pH of the medium from acidic to alkaline levels reduced
the toxicity of lead to the alga. Christensen et al. (1979) used the
same freshwater alga and a second alga, Chlorella stigmatophora,
cultured in artificial sea water, in a study of the effects of
inorganic lead, alone and in combination with other metals.
Selenastrum was cultured in standard algal assay medium (SAAM). In an
initial range-finding test, Selenastrum and Chlorella were cultured
with lead, as lead acetate, in solution at concentrations of 0.01, 0.1,
1.0, 10.0, 100.0, and 1000.0 mg/litre. The effects were assessed in
terms of total cell volume, relative to controls, after 13 days for

Selenastrum and 24 days for Chlorella. Selenastrum was slightly
stimulated by lead acetate at 0.01 mg/litre, with a relative cell
volume of 1.05. At 0.1 mg lead/litre, Selenastrum showed a reduced
cell volume of 0.87 relative to controls, and, at 1.0 mg lead/litre, a
cell volume ratio of 0.12. At concentrations of 10.0 mg/litre or more,
lead killed the algal cells. Only at the highest exposure of
1000 mg/litre was there any visible precipitation of the lead acetate.
Chlorella was also killed by lead acetate at concentrations of
10.0 mg/litre or more in the artificial sea water. At 1.0 mg/litre,
lead reduced the cell volume of Chlorella to 0.71 relative to controls.
At lower lead concentrations, there was a stimulation of the alga with
cell volumes of 2.09 and 1.60 relative to controls for exposures to
0.01 and 0.1 mg lead/litre, respectively. In both Selenastrum and
Chlorella, lead increased the average cell volume of the algal cells
significantly at the same time as it reduced the growth rate. In these
experiments, Selenastrum was exposed to lead concentrations varying
between 0.09 and 1.44 mg/litre. Over this range, growth rate
decreased from 1.12 mm³/litre per day at 0.09 mg lead acetate/litre
to 0.5 mm³/litre per day at 1.44 mg lead acetate/litre. The
average volume of individual cells increased from 62 µm³ to
91 µm³ over the same dose range of lead in the culture medium.
When Chlorella was exposed to a range of lead concentrations in
artificial sea water between 0.36 and 5.76 mg/litre, the growth rate
declined from 0.68 mm³/litre per day to 0.4 mm³/litre per
day, and the average cell volume increased from 24 µm³ to
61 µm³. Values for EC50 for cell volume were calculated at
140 µg lead acetate/litre for Selenastrum and 700 µg/litre for
Chlorella. The authors suggest that the discrepancy between their own
result for Selenastrum and that reported by Monahan (1976) might be due
to the greater concentration (by about five times) of dissolved salts
in his culture medium.

Culturing the algae in a medium containing combinations of metals
showed that the presence of manganese or copper reduces the toxicity of
lead to these organisms (Christensen et al., 1979). Prasad & Prasad
(1982) exposed three freshwater green algae (Ankistrodesmus falcatus,
Scenedesmus obliquus, and Chlorococcum sp.) to lead chloride at concen-
trations of 0 to 10 mg lead/litre, and measured growth on the 10th day
after inoculation using an optical density method. There was no effect
on growth from 0.1 to 1.5 mg lead/litre, but at 2.0 mg/litre or more,
there was inhibition of growth in all three species. At 10 mg
lead/litre, A. falcatus was killed, and S. obliquus and Chlorococcum
sp. were reduced to 9% and 15%, respectively, of the mass of
controls.

Hongve et al. (1980) exposed a natural phytoplankton community
to concentrations of an unspecified inorganic lead salt ranging from
5 x 10^-7 to 5 x 10^-4 mol/litre. The community of organisms,
isolated from lake water, consisted mainly of diatoms: Tabellaria
flocculosa (53% by volume), Synedra sp. (13%), and Asterionella
formosa (7%). Other important constituent species were Cryptomonas
spp., Rhodomonas minuta variety lacustris, Dinobryon divergens, small
species of chryptomonades, Gymnodinium sp., and Mallomonas sp. The
authors monitored photosynthetic activity as uptake of ¹⁴C-labelled
hydrogen carbonate over a 20-h incubation. Photosynthetic carbon

fixation was reduced in a dose-dependent manner throughout the exposure
range of lead in the medium; the reduction was 90% relative to control
cultures at the highest exposure of 5 x 10^-4 mol lead/litre. The
addition of lake sediment to treated cultures reduced the toxic effect
of lead; addition to control cultures increased the photosynthetic
carbon uptake by 19%. A similar reduction in the toxic effects of lead
was seen on adding organic matter filtered out of the lake water and
after adding the chelating agent nitrilotriacetic acid (NTA) at non-
toxic levels. Neither of these two variables affected photosynthesis
in control cultures. The NTA had the greatest effect on lead toxicity,
virtually eliminating the effect of lead on photosynthesis.

Persoone & Uyttersprot (1975) examined the effects of lead chloride
on reproduction in the marine ciliate Euplotes vannus by estimating the
number of generations produced after culture for 48 h. The Euplotes
cultures were exposed to 0.001, 0.01, 0.1, 1, 10, or 100 mg lead
chloride/litre. Reproduction was unaffected by lead at concentrations
up to 0.1 mg/litre in the culture medium. At 1 mg/litre, the lead
caused approximately 15% inhibition of reproduction and, at
10 mg/litre, 30% inhibition. At 100 mg lead chloride/litre, all the
ciliates died.

Hessler (1974) exposed a marine unicellular green flagellate alga
(Platymonas subcordiformis) to lead chloride at concentrations of 100,
500, and 1000 mg lead/litre of sea-water medium. There was precipi-
tation of lead from solution and these doses gave corresponding values
for lead in solution of 2.5, 10, and 60 mg/litre, respectively. Log-
phase cells, growing exponentially, were more sensitive to lead than
stationary-phase cells. At 2.5 and 10 mg/litre, lead retarded popu-
lation growth by delaying cell division and daughter cell separation.
A concentration of 60 mg/litre caused complete inhibition of growth and
cell death. Normal wild-type cells were more sensitive to lead than
either cells sheared of their flagellae or cells of a mutant without
flagellae.

Hessler (1975) exposed Platymonas to the same range of lead
concentrations but in the presence of mutagenic agents (ultraviolet
irradiation or nitroguanidine). High levels of mutation were found but
were not increased in the presence of lead.

Malanchuk & Gruendling (1973) estimated EC₅₀s for reduction in
¹⁴CO₂-fixation in freshwater algae after exposure to lead nitrate.
Results were extrapolated from a graph of milligrams lead per litre
plotted against radioactivity per milligram dry cell weight. Results
varied with the size of the inoculum (cell density). For the
cyanophyte (blue-green alga) Anabaena sp. the EC₅₀ was 15 mg/litre
(1- and 2-ml samples) or 26 mg/litre (4-ml sample). For the chloro-
phytes Chlamydomonas reinhardti and Cosmarium botrytis, a desmid,
EC₅0s were 17 and 5 mg/litre, respectively. The chrysophyte
Navicula pelliculosa showed EC₅₀s of 17 mg/litre (1- and 2-ml
samples) and 28 mg/litre (4-ml sample). However, another chrysophyte,
Ochromonas malhamensis, was not inhibited by up to 30 mg lead/litre.

Whitton (1970) investigated the effects of lead chloride on a
variety of species of filamentous green alga isolated from flowing
streams in northern England; some were from metal-polluted streams and

some from unpolluted ones. Results were expressed semi-quantitatively
and a tolerance index was determined in terms of lead concentration.
This index is a geometric mean of codings indicating minimal and
maximal effects. These values varied between 3 and 60 mg lead/litre.
The most sensitive species, Cladophora, and the least sensitive
species, Microspora, were unusual; all others tested gave values
between 17 and 46 mg/litre.

Bringmann & Kuhn (1959b) reported a toxic threshold of 2.5 mg/litre
for the green alga Scenedesmus (related to cell division) and of
1.25 mg/litre for the protozoan Microregma (related to feeding).

Babich & Stotzky (1983) observed that hard water protected
Tetrahymena pyriformis from the effects of lead salts. Gray &
Ventilla (1973) exposed a sediment-living, bacterivorous ciliate
protozoan, Cristigera sp., to concentrations of 0.15 or 0.3 mg lead
nitrate/litre for 4 to 5 h. Lead reduced growth rates by 8.5% and
11.8% for the two doses, respectively. Analysis of variance showed the
effect to be significant at the 1% level. Apostol (1973) used another
ciliate protozoan, Paramecium caudatum, in acute and long-term tests to
examine the toxicity of lead acetate. In the 5-h acute test,
Paramecium showed a sharp threshold of toxic response at around
1000 mg lead acetate/litre, with survival times dropping steeply from
>300 min to 5 min or less. In chronic tests over 14 days, growth of
the population was delayed by lead acetate at 1, 10, and 100 mg/litre.
Peak population numbers were progressively reduced as lead concen-
trations increased. At the beginning of the test, the median survival
span at 1000 mg/litre was <5 min, whereas at the end of the test, the
survival span was >5 h at the same concentration. It is clear,
therefore, that there is considerable individual variation in the
population and scope for adaptation in the wild.

Ruthven & Cairns (1973) determined the minimal lethal concentration
and maximum tolerated concentration of lead, as lead nitrate, for six
different species of freshwater algae and protozoans. Two species,
Peranema and Euglena gracilis, tolerated 1000 mg lead/litre (nominal
concentration). Blepharisma tolerated 42 mg/litre; Tetrahymena and
Paramecium multimicronucleatum tolerated 24 mg/litre, while
Chilomonas tolerated only 5.6 mg/litre. Minimal lethal concentrations
for the four more sensitive species ranged from 56 mg/litre to
>100 mg/litre.

Rosenweig & Pramer (1980) examined the effects of lead nitrate on
seven species of nematode-trapping fungi from soil. Mycelial growth
was reduced in two species at lead concentrations of 100 mg/litre, and
in all but one species at 300 mg/litre. Reduced capacity to produce
traps (rings of mycelium which capture nematode worms) was correlated
with reduced growth, except in the case of one species where growth was
inhibited with no effect on trap production. Increasing pH reduced the
toxicity of lead to the fungi Aspergillus niger (Babich & Stotzky,
1979), Achyla sp. and Saprolegnia sp. (Babich & Stotzky, 1983).
Babich & Stotzky (1979, 1983) noted that the presence of carbonate or
phosphate ions reduced the toxic effect of lead on Aspergillus and
Fusarium growth, presumably by precipitating lead from the medium.

Crist et al. (1985) collected and dried green leaves from a variety
of tree species representative of central hardwood forests of the USA.
The leaf mixtures were treated with lead sulfate to give concentrations
of lead in the leaf litter ranging from 0 to 1000 mg/kg and incubated
in laboratory microcosms. Replicates were treated with different
amounts of sulfuric acid to give pH values in the incubates of between
3 and 5. Lead, at these concentrations, had no effect on leaf decompo-
sition at any of the pH values tested.

5.2 Toxicity of Organic Lead

Roderer (1980) investigated the effects of tetraethyllead on a
flagellated alga, Poterioochromonas malhamensis. After 3 days of
culture in darkness, there were no toxic effects of tetraethyllead even
at concentrations of 0.3 mmol/litre. At 0.25 mmol/litre in light, all
cells were killed by tetraethyllead. At concentrations below
0.25 mmol/litre, there was a dose-related effect on growth, mitosis,
and cytokinesis, resulting in the formation of giant polyploid cells.
Tetraethyllead was converted to a highly toxic derivative in light
with, or without, cells present. This toxic compound was produced in
toxic amounts within 3 to 6 h of illumination, but reached a maximum
concentration after 24 to 32 h. Free radicals, which are produced
during the photolysis of tetraethyllead, were shown not to be respon-
sible for the toxicity. Tetraethyllead is removed from water by
aeration because of its low water solubility and high volatility. In
the presence of light, this process is counteracted by the formation of
a stable, water-soluble material, which is toxic to algae. The authors
identified the toxic product as triethyllead.

Marchetti (1978) investigated the effects of tetraalkyllead in
natural sea water on mixed coastal marine bacteria using the biological
oxygen depletion method in a respirometer. Two commercial products
were used: tetramethyllead TML-CB and tetraethyllead TEL-CB. Tetra-
alkyllead solutions were prepared by adding 2 ml of each product to 1
litre of filtered natural sea water and slowly stirring magnetically
for 1 h. The upper quarter of the solution was used for preparing
experimental dilutions. The lag phase was related to the TML-CB lead
concentration up to lead concentrations in water of 3.2 mg/litre and to
the TEL-CB lead concentration up to 0.16 mg/litre. There was also a
relationship between lead concentration and respiration rate, starting
from 0.36 and 0.08 mg/litre, respectively, for TML-CB and TEL-CB.
Below these concentrations there was no significant effect on either
lag phase or oxygen consumption. The EC₀, EC₅₀, and EC₁₀₀ over
48 h were 0.9, 1.9, and 4.5 mg/litre, respectively, for TML-CB and
0.08, 0.2, and 2.0 mg/litre for TEL-CB. TML-CB is less toxic to
bacteria than TEL-CB, on the basis of total lead content, even if the
presence of some toluene in the TML formulation is taken into account.
The author speculates that the difference in toxicity depends on the
different speed of transformation from the tetraalkyl form, scarcely
soluble and toxic, to the more soluble and less toxic trialkyl form.
Tests with the same preparation on the photosynthesis of the alga
Dunialiella tertiolecta gave an EC₀, an EC₅₀, and an EC₁₀₀ of
0.45, 1.65, and 4.5 mg/litre, respectively, for TML-CB and 0.1, 0.15,
and 0.3 mg/litre for TEL-CB.

Silverberg et al. (1977) exposed the freshwater algae Scenedesmus
quadricaudata, Ankistrodesmus falcatus, and Chlorella pyrenoidosa to
tetramethyllead. Tetramethyllead is not soluble in water and is
volatile. The compound was biologically generated in a reaction
vessel using trimethyllead acetate and Aeromonas sp. or indigenous
microorganisms in Hamilton Harbour water and sediment. When
tetramethyllead was detected in air drawn off the reaction vessel, this
was bubbled through cultures of the algae. Exposure was, because of
the nature of the material, momentary because of conversion to
trimethyllead. The primary productivity of the cultures was estimated
using ¹⁴C-hydrogen carbonate uptake, and growth was estimated using
both cell dry weight and counts of cell numbers. Cells were also
harvested and fixed for electron-microscopic evaluation. Although the
exposure cannot be estimated exactly, the authors estimate that <0.5 mg
of tetramethyllead was passed through the cultures during the course of
the 7-day study. Chlorella was the most sensitive of the three
organisms showing a decrease of 74% in growth and 83% in photo-
synthesis. Scenedesmus showed a 32% decrease in growth and 85% decrease
in photosynthesis; the corresponding figures for Ankistrodesmus were
32% and 49%, respectively. The cultures showed loss of green
coloration, the green becoming semitransparent yellow with time. Cells
were enlarged and clumped into masses. After electron-microscopic
examination, it could be seen that the chloroplasts were most affected
by the tetramethyllead. Lead was detected inside the cells using
electron-microscopic analysis. The authors state that tetramethyllead
is twice as toxic as trimethyllead acetate and 20 times more toxic than
lead nitrate for the same organisms.

Roderer (1983) found that compounds used to alleviate lead
poisoning in man (Na₂EDTA, EDTA, DPA, DIZO, BAL) increased, rather
than decreased, the effects of inorganic and triethyllead on the
unicellular alga Poterioochromonas malhamensis. In a later, more
comprehensive investigation of factors affecting the toxicity of
triethyllead to Poterioochromonas (Roderer, 1986), the author studied
the protective action of thiol compounds, vitamins, trace elements, and
other agents. None of the tested thiol or disulfide compounds
protected the alga from triethyllead. Two vitamins, tocopheryl acetate
and ascorbic acid, one trace element, zinc, and ATP, cyclic AMP, and
concanavalin A, together with some combinations of agents, markedly
suppressed the growth-inhibiting effects of triethyllead. Zinc was the
most effective single agent, increasing growth of the algal cultures by
70 times in the presence of triethyllead at 10^-5 mol/litre. A
combination of 10 essential trace elements was even more effective and
almost totally eliminated the toxic effect of the lead compound.

6. TOXICITY TO AQUATIC ORGANISMS

Lead is unlikely to affect aquatic plants at levels that might be
found in the general environment.

In the form of simple salts, lead is acutely toxic to aquatic
invertebrates at concentrations between 0.1 and >40 mg/litre for
freshwater organisms and between 2.5 and >500 mg/litre for marine
organisms. The 96-h LC₅₀ for fish varies between 1 and 27 mg/litre,
in soft water, and between 440 and 540 mg/litre, in hard water, for the
same species. The higher values for hard water represent nominal
concentrations. Available lead measurements suggest that little of the
total lead is in solution in hard water. Lead salts are poorly soluble
in water, and the presence of other salts reduces the availability of
lead to organisms because of precipitation. Results of toxicity tests
should be treated with caution unless dissolved lead is measured.

There is little information on the effects of organic lead
complexes. Sublethal effects have been reported.

6.1. Toxicity to Aquatic Plants

Appraisal

There is little evidence for effects of lead on aquatic plants at
concentrations below 1 to 15 mg/litre. Many studies of aquatic plants
have been made in sediment-free systems. However, the addition of
uncontaminated sediment reduces the toxicity of lead to aquatic plants
by reducing its availability.

Van der Werff & Pruyt (1982) exposed four aquatic plants,
Elodea nuttallii, Callitriche platycarpa, Spirodela polyrhiza, and
Lemna gibba, to concentrations of lead nitrate of up to 10^-5 mol
lead/litre for 70 to 73 days. There was no observable toxicity, and
growth rates were unaffected. Brown & Rattigan (1979) exposed the
aquatic macrophyte Elodea canadensis (Canadian pond-weed) and the free-
floating duckweed Lemna minor to a range of lead acetate concentrations
for 28 days and 14 days, respectively. The authors assessed damage to
the plants visually on a coded scale from 0 (no damage) to 10
(complete plant kill). They reported that concentrations of 136 and
16.3 mg/litre produced 50% damage to the two plant species, respect-
ively. In a separate experiment, they exposed Elodea to lead for 24 h
in the dark, and then measured oxygen evolution in the light. Levels
of 47.6 and 99 mg lead/litre reduced photosynthetic oxygen evolution
by 50% and 90%, respectively. Kay et al. (1984) exposed the water
hyacinth Eichhornia crassipes to lead nitrate concentrations of 0.5 to
5.0 mg/litre for 6 weeks. There was no observed effect on root
development, leaf colour, development of new plantlets, flowering, or
total plant growth.

Stanley (1974) determined EC₅₀s for various growth parameters of
Eurasian watermilfoil (Myriophyllum spicatum) exposed to lead (salt
unspecified). Plants were grown in soil with water above. The
EC₅₀ for root weight was 363 mg/litre, for shoot weight was
808 mg/litre, for root length was 767 mg/litre, and for shoot length

was 725 mg/litre. The effects of adding the lead to the soil as
opposed to the water were investigated. There was less effect with
lead added to the soil because of adsorption to soil particles.
There was a ratio of 1.43 between root growth when lead was added to
soil over that when lead was added to water, following exposure to
20.7 mg lead/litre. For exposure to 207 mg lead/litre, the corre-
sponding ratio was 1.88.

6.2. Toxicity to Aquatic Invertebrates

Appraisal

The results of experiments on the toxicity of lead salts to aquatic
invertebrates are difficult to interpret due to the variations in
experimental conditions and the lack of a standardized method for
determining lead concentrations in water. In most studies, concen-
trations of lead in water are nominal; the contribution to toxicity of
factors, such as pH, water hardness, anions, and complexing agents
cannot be fully evaluated.

In communities, some populations of organisms are more sensitive
than others, and community structure may be adversely affected by lead
contamination. However, populations from polluted areas can show more
atolerance to lead than those from non-polluted areas. In other
organisms, adaptation to hypoxic conditions can be hindered by high
lead concentrations.

There is information on the toxicity of lead salts to aquatic
invertebrates, but little information on the effects of organic lead
compounds. The toxicity of lead to aquatic invertebrates is summarized
in Tables 3 and 4.

6.2.1. Toxicity of lead salts

Cleland (1953) found that lead nitrate at a concentration of
6 x 10^-4 mol lead/litre suppressed the development of a fertilization
membrane elevation in eggs of the sea urchin (Psammechinus miliaris).
Subsequent cleavage of the fertilized egg was generally normal.

Watling (1983b) reported that the larvae of the oyster Crassostrea
gigas grew less well, over a 14-day exposure period, with lead nitrate
in the water at 0.01 or 0.02 mg/litre. The exposed larvae showed a
mean length of 5.0 and 5.3 mm, respectively, in solutions of 0.01 and
0.02 mg lead/litre, after 14 days, compared to 6.3 mm for the controls.
After a further 14 days in clean water, most of the reduction in size
had been recovered. The treated larvae were 8.0 and 7.8 mm mean length
for the two dose levels, and the controls were 8.2 mm long. The author
also reported that lead, at both 0.01 and 0.02 mg/litre, reduced the
numbers of larvae settling and delayed the peak settlement time of the
population.

Table 3. Toxicity of lead salts to aquatic invertebrates
---------------------------------------------------------------------------------------------------------
Organism Life- Flow/ Temp. Alkali- Hard- pH Salt Parameter Water Reference
stage stat^a (°C) nity^c ness^c concent-
ration
(mg/litre)
---------------------------------------------------------------------------------------------------------
American oyster stat 25-27 25^d nitrate 48-h LC₅₀ 2.45
(Crassostrea (2.2-3.6) Calabrese
virginica) stat 25-27 25^d nitrate 48-h LC₀ 0.5 et al.
stat 25-27 25^d nitrate 48-h LC₁₀₀ > 6.0 (1973)

Hard clam stat 25-27 25^d nitrate 48-h LC₅₀ 0.78 Calabrese
(0.72-0.80) & Nelson
(Mercenaria stat 25-27 25^d nitrate 48-h LC₁₀₀ 1.20 (1974)
mercenaria)

Softshell clam stat 21.5- 29-31^d 7.8- nitrate 48-h LC₅₀ > 50 Eisler
(Mya arenaria) 22.5 8 (1977)
stat 21.5- 29-31^d 7.8- nitrate 96-h LC₅₀ 27
22.5 8
stat 21.5- 29-31^d 7.8- nitrate 168-h LC₅₀ 8.8
22.5 8

Cockle adult stat 15 nitrate 48-h LC₅₀ > 500 Portmann
(Cardium edule) & Wilson
(1971)
Pink shrimp adult stat 15 nitrate 48-h LC₅₀ 375 Portmann
(Pandalus & Wilson
montagui) (1971)

Neanthes juvenile stat 7.8 acetate 96-h LC₅₀ > 7.5^e Reish
arenacoe- adult stat 7.8 acetate 96-h LC₅₀ > 10^e et al.
dentata (1976)
juvenile stat 7.8 acetate 28-day LC₅₀ 2.5^e Reish
adult stat 7.8 acetate 28-day LC₅₀ 3.2^e et al.
(1976)
Capitella larva stat 7.8 acetate 96-h LC₅₀ 1.2^e Reish
capitella adult 7.8 acetate 96-h LC₅₀ 6.8^e et al.
(1976)
adult 7.8 acetate 28-day LC₅₀ 1.0^e Reish
et al.
---------------------------------------------------------------------------------------------------------

Table 3. (contd.)
---------------------------------------------------------------------------------------------------------
Organism Life- Flow/ Temp. Alkali- Hard- pH Salt Parameter Water Reference
stage stat^a (°C) nity^c ness^c concent-
ration
(mg/litre)
---------------------------------------------------------------------------------------------------------
(1976)
Crab stat 26.5- 7.0- nitrate 96-h LC₅₀ > 370 Krishnaja
(Scylla serrata) 29.5 7.2 et al.
(1987)
Mussel stat 28-32 7.7- 32-38 7.0- nitrate 48-h LC₅₀ > 40 Subbaiah
(Lamellidens 11.7 7.3 et al.
marginalis) (1983)

Freshwater crab stat 28-32 7.7- 32-38 7.0- nitrate 48-h LC₅₀ > 40 Subbaiah
(Oziotelphusa 11.7 7.3 et al.
senex senex) (1983)

Snail stat 28-32 7.7- 32-38 7.0- nitrate 48-h LC₅₀ > 40 Subbaiah
(Pila globosa) 11.7 7.3 et al.
(1983)
Copepod stat 9.5- 0.58 meq/ 7.2 acetate 48-h LC₅₀ 5.5 Baudouin
(Cyclops abyssorum) 10.5 litre (4.0-7.7) & Scoppa
(1974)
Copepod stat 9.5- 0.58 meq/ 7.2 acetate 48-h LC₅₀ 4.0 Baudouin
(Eudiaptomus 10.5 litre (2.5-6.4) & Scoppa
padanus) (1974)

Water flea stat 41-50 44-53 7.4- chloride 48-h LC₅₀ 0.45^f Biesinger &
(Daphnia magna) 8.2 Christensen
stat 41-50 44-53 7.4- chloride 21-day LC₅₀ 0.3 (1972)
8.2 (0.236-0.381)
stat 11.5- 390- 235- 7.4- acetate 24-h LC₅₀ 4.89 Khangarot
14.5 415 260 7.8 (4.19-5.89) & Ray
(1987)
stat 11.5- 390- 235- 7.4- acetate 48-h LC₅₀ 3.61 Khangarot
14.5 415 260 7.8 (2.83-4.4) & Ray
(1987)
Water flea stat 9.5- 0.58 meq/ 7.2 acetate 48-h LC₅₀ 0.60 Baudouin
(Daphnia hyalina) 10.5 litre (0.41-0.89) & Scoppa
(1974)
Table 3. (contd.)
---------------------------------------------------------------------------------------------------------
Organism Life- Flow/ Temp. Alkali- Hard- pH Salt Parameter Water Reference
stage stat^a (°C) nity^c ness^c concent-
ration
(mg/litre)
---------------------------------------------------------------------------------------------------------

Amphipod flow 15 40-44 44-48 7.1- nitrate 96-h LC₅₀ 0.124 Spehar
(Gammarus 7.7 et al.
pseudolimnaeus) (1978)

Crayfish flow^b 15-17 7.0 chloride 96-h LC₅₀ 2.6 Boutet &
(Austropotamobius flow^b 15-17 7.0 chloride 30-day LC₅₀ 1.5 Chaise-
pallipes pallipes) flow^b 15-17 7.0 chloride 30-day LC₅₀ 0.9^e martin
(1973)
Crayfish flow^b 15-17 7.0 chloride 96-h LC₅₀ 3.3 Boutet &
(Orconectes limosus) flow^b 15-17 7.0 chloride 30-day LC₅₀ 1.7 Chaise-
flow^b 15-17 7.0 chloride 30-day LC₅₀ 0.9^e martin
(1973)
Midge egg/ stat 21-23 43.9 46.8 7.5 nitrate 10-day LC₅₀ 0.258 Anderson
(Tanytarsus larva et al.
dissimilis) (1980)

Mayfly larva flow 7.0- nitrate 14-day LC₅₀ 3.5 Nehring
(Epherella 7.2 (1976)
grandis)

Stonefly larva flow 7.0- nitrate 14-day LC₅₀ > 19.2 Nehring
(Pteronarcys 7.2 (1976)
californica)
---------------------------------------------------------------------------------------------------------
^a Stat = static conditions (water unchanged for duration of test); flow = flow-through conditions
(lead concentration in water continuously maintained).
^b Intermittent flow-through conditions.
^c Alkalinity and hardness expressed as mg/litre CaCO₃.
^d These figures are values for salinity (expressed as ^o/_oo), not alkalinity.
^e With a food source.
^f Water fleas were fed during test.

Table 4. Toxicity of organolead to aquatic invertebrates
---------------------------------------------------------------------------------------------------------
Organism Mean Mean Flow/ Temp. Salinity Compound^b Parameter Water Reference
length weight stat^a (°C) (^o/_oo) concentration
(mm) (g) (mg/litre)
---------------------------------------------------------------------------------------------------------
Mussel 64 28.5 flow 15 34.9 TML 96-h LC₅₀ 0.27 Maddock
(Mytilus edulis) 64 28.5 flow 15 34.9 TEL 96-h LC₅₀ 0.1 & Taylor
64 28.5 stat 15 34.9 TriML 96-h LC₅₀ 0.5 (1980)
64 28.5 stat 15 34.9 TriEL 96-h LC₅₀ 1.1

Brown shrimp 48 1.1 flow 15 34.9 TML 96-h LC₅₀ 0.11 Maddock
(Crangon crangon) 48 1.1 flow 15 34.9 TEL 96-h LC₅₀ 0.02 & Taylor
48 1.1 stat 15 34.9 TriML 96-h LC₅₀ 8.8 (1980)
48 1.1 stat 15 34.9 TriEL 96-h LC₅₀ 5.8
---------------------------------------------------------------------------------------------------------
^a Stat = static conditions (water unchanged for duration of test); flow = flow-through conditions (lead
concentration in water continuously maintained).
^b TML = tetramethyl lead; TEL = tetraethyl lead; TriML = trimethyl lead chloride;
TriEL = triethyl lead chloride.
Calabrese et al. (1973) found that the EC₅₀ of lead chloride for
the development of larvae of the American oyster was 2.45 mg/litre; the
EC_o was 0.5 mg lead/litre. Lead nitrate is more toxic to the hard
clam (Mercenaria mercenaria) than to the American oyster
(Crassostrea virginica) (Calabrese & Nelson, 1974) (Table 3). Coombs
(1977) exposed batches of 20 mature mussels, Mytilus edulis, of
shell length 6-7 cm, to lead (added to the water as nitrate or
complexed with citrate, humic and alginic acids, or pectin). The
author showed that the uptake of lead was increased by complexation,
citrate being the most effective complexing agent for stimulating
absorption of the metal. Electron-microscopic examination of tissues
showed that the mussels were able to tolerate large amounts of lead in
their tissues, and to reduce its toxicity by enclosing the metal in
membrane-bound vesicles. Stromgren (1982) reported that lead citrate,
at water concentrations of up to 0.2 mg/litre, had no effect on the
growth rate of Mytilus edulis.

Lead nitrate in water, at concentrations of up to 0.565 mg/litre,
had no effect on the survival of the freshwater snail Physa integra
(Spehar et al., 1978). Borgmann et al. (1978) exposed the freshwater
snail Lymnaea palustris to various lead nitrate concentrations, in a
flow-through study, ranging from 3.8 to 54 µg/litre over 120 days.
There was no effect on survival at concentrations of 3.8 and
12 µg/litre, but mortality occurred at concentrations of
19 µg/litre or more. The growth rate of the survivors was not
affected at lead concentrations of 19 µg/litre. The authors observed
a 50% reduction in snail biomass production after exposure to lead at
36 µg/litre from hatching during the period of maximal growth.

Baudouin & Scoppa (1974) reported LC₅₀ results for two species of
freshwater copepod and for a water flea (Table 3). They failed to find
any indication of a lethal threshold for lead. Roberts & Maguire
(1976) added inorganic lead (salt unspecified) to sand, collected from
the surface and sub-surface below mid-tidal level, at concentrations of
0.001, 0.1, or 1 mg/litre in sea water. The populations of various
meiofauna were estimated with time, up to 410 h after adding the lead.
The most affected organisms in the surface sand were harpacticoid
copepods, whose numbers declined with time and increasing lead
concentrations. Measurement of lead in the interstitial water of the
test samples showed that much of the metal was strongly adsorbed to
sand particles very early in the experiment. Only for the first 10 h,
at the highest exposure, were significant amounts of lead detectable in
the water (ca. 0.1 mg/litre). Nematodes were the most sensitive organisms
in sub-surface sand.

Fraser et al. (1978) collected samples of the freshwater crustacean
Asellus aquaticus from various polluted and unpolluted sites in the
basin of the River Trent, United Kingdom. The different populations
were exposed for 24 h in the laboratory to lead nitrate solutions
at pH 4.5 and lead concentrations of 0, 100, 250, 500, 750, 1000, and
1500 mg/litre. The authors found a log-linear relationship between
lead concentration and survival of the Asellus. Animals less than
4 mm in length survived less well than larger animals at the higher
lead concentrations. Approximately 50% of both small and large
Asellus survived for 24 h after exposure to lead nitrate at
100 mg/litre. Those animals collected from an area with higher lead

levels were more tolerant to the metal in laboratory experiments,
suggesting some selection in the wild. Exposure in the wild, during
3 years of analysis, varied between 0 and 0.24 mg/litre in the high-
lead area and between 0 and 0.08 mg/litre in the low-lead area.

Spehar et al. (1978) exposed the freshwater amphipod Gammarus
pseudolimnaeus to lead nitrate solutions in lake water for 28 days.
The lead caused more than 50% mortality at water concentrations of
0.136 mg/litre or more, over 4 days. By the end of the study,
mortality was 60% at the lowest concentration of lead nitrate tested,
0.032 mg/litre. Survival curves showed a marked increase in slope
between test concentrations of 0.067 and 0.136 mg/litre. At higher
concentrations of lead nitrate, virtually all the final mortality
occurred within the first 7 days of exposure.

Freedman et al. (1980) investigated the effect of lead speciation
on the toxicity of the metal to the shrimp Hyallela azteca in arti-
ficial test media. Lead was added to the medium in association with
four different molarities of phosphate, 10^-3, 10^-4, 10^-5, and
10^-6 mol/litre, and at two different pH values, 6 and 8. Theor-
etical calculations were made of the concentration of free lead in the
solutions. At pH 6, very little free lead exists at high phosphate
concentrations, irrespective of the total lead concentration. Simi-
larly, little free lead is predicted at any phosphate concentration at
pH 8. Mortality figures related well to the predicted values for free
lead in the various solutions. At pH 6, a total lead concentration of
5 mg/litre, and a phosphate concentration of 10^-6 mol/litre, there
was 100% mortality after 48 h. For phosphate molarities of 10^-5 and
10^-4 mol/litre, toxicity was progressively reduced. At the highest
phosphate concentration, mortality only reached 25% after 120 h. Free
lead values predicted for the same three phosphate concentrations were
2.76, 2.24, and 0.11 mg/litre, respectively. Chinnayya (1971) found
that lead nitrate at 10^-3 mol/litre in fresh water reduced the oxygen
consumption of the shrimp Caridina rajadhari from a control level of
0.49 ml/h per g wet weight of shrimps to 0.38 ml/h per g. This
concentration of lead caused no mortality over 10 days. The lowest
concentration of lead nitrate causing mortality in this species was
5 x 10^-3 mol/litre.

Anderson (1978) maintained the crayfish Orconectes virilis in
natural river water, with lead acetate added to concentrations of 0.5,
1.0, or 2.0 mg lead/litre. The water was changed at 5-day intervals to
maintain the lead concentration, and at 10-day intervals, the oxygen
consumption of the crayfish was measured. There was a dose-related
reduction in oxygen consumption after 10 days of exposure to lead
acetate. After 20, 30, and 40 days of exposure, there was no
difference in oxygen consumption between control and treated crayfish;
the animals had acclimatized to the lead. The crayfish were found to
be compensating for the effect of the lead, which reduced the capacity
for oxygen uptake through the gills, by increasing the flow of water
over the gill surfaces. There was a dose-related relationship between
ventilation volume and lead concentration in the water over the range
0-2.0 mg/litre; the ventilation volume at 2.0 mg lead acetate/litre was
19 ml/min, compared with 12 ml/min for controls. Since the water in
the test tanks was kept saturated with oxygen, the crayfish were able
to restore fully their oxygen uptake.

Brown & Ahsanullah (1971) studied the effects of lead nitrate on
mortality and growth of the worm Ophryotrocha labronica and the brine
shrimp Artemia salina. When they were exposed to lead at 1 mg/litre,
the LT₅₀ was >600 h for Ophryotrocha and 576 h for Artemia. There
was no significant suppression of growth rate (measured as increase in
length) after exposure of the worm to 10 mg/litre for 8 days or
1 mg/litre for 10 days. However, a significant suppression of the
growth of 48-h brine shrimp larvae was reported after exposure to lead
nitrate at 5 and 10 mg/litre for 6 days.

Fischer et al. (1980) investigated the effects of lead chloride on
the tubifex worm Tubifex tubifex under aerobic and hypoxic conditions.
When worms were exposed for 6 days to lead chloride, at a concentration
of 10 mg/litre of tap water, there was no mortality. The authors
sectioned segments of the worms and measured the nuclear volume of the
chloragocytes. These cells are responsible for the synthesis of haemo-
globin and respond to hypoxic conditions by increasing their activity.
Nuclear volume correlates with available oxygen. In aerated water,
lead chloride caused an increase in nuclear volume of the chloragocytes
from 68.8 µm³, the control size, to 93.1 µm³. Under hypoxic
conditions, control nuclear volume increased to 137.9 µm³, but in
lead-treated animals increased only to 99.6 µm³. This physio-
logical response in compensating for hypoxia is essential to this
animal in its normal environment, where large changes in available
oxygen will be commonplace.

Biesinger & Christensen (1972) found that reproductive impair-
ment was a more sensitive measure of the toxicity of lead chloride
to water fleas (Daphnia magna) than survival. They determined an
EC₁₆ and EC₅₀ of 30 and 100 µg lead/litre, respectively, for a 3-
week exposure.

Warnick & Bell (1969) exposed nymphs of stonefly (Acroneuria
lycorias), mayfly (Ephemerella subvaria), and caddisfly
(Hydropsyche betteni) to lead sulfate in static bioassays. They
reported 50% survival times of >14 days at 64 mg/litre, 7 days at
16 mg/litre, and 7 days at 32 mg/litre. There was a considerable
decrease in the metal concentrations in solution over the 2-week
experimental period, and the authors considered that nominal concen-
trations were unreliable after 96 h. Spehar et al. (1978) found no
effect of their highest dose of 0.565 mg lead nitrate/litre on the
survival of nymphs of stoneflies or caddisflies (Pteronarcys dorsata,
Hydropsyche betteni, Brachycentrus sp., and Phemerella sp.).

Anderson et al. (1980) exposed the chironomid midge, Tanytarsus
dissimilis, to lead nitrate during two different stages of its life-
cycle. Exposure started with the eggs and continued for 10 days,
during which time the larvae had emerged. The average LC₅₀ from two
tests was 0.258 mg/litre. No significant effect on the growth of
surviving larvae was found until the LC₅₀ concentration was exceeded.
The authors emphasized that this species is particularly sensitive to
heavy metals. Chironomid midges are extremely plentiful in lakes and
streams, and represent a major food source for fish.

6.2.2. Toxicity of organic lead

Marchetti (1978) determined, in 48-h tests, no-observed-effect
levels (micrograms per litre) for tetraethyl- and tetramethyllead to
24-h nauplii of the brine shrimp Artemia salina, together with
LC₅₀ and LC₁₀₀.

--------------------------------------------------------------------------
Compound 0% 50% 100%
effect effect effect
--------------------------------------------------------------------------
Tetramethyllead 180 250 670

Tetraethyllead 25 85 260
--------------------------------------------------------------------------

6.3. Toxicity to Fish

Appraisal

The toxicity of lead-contaminated water to fish varies consider-
ably, depending on the availability and uptake of the lead ion.
Factors affecting this availability are water hardness (presence of
divalent anions), pH, salinity, and organic matter. Uptake is affected
by the presence of other cations and the oxygen content of the water.
Organic lead is taken up more readily than inorganic lead. The 96-h
LC₅₀ for inorganic lead in sensitive species can be as low as 1 mg
dissolved lead/litre; nominal concentrations being up to 100 times
higher in hard water. The few data available suggest that the toxicity
of organic lead may be 10 to 100 times higher than that of inorganic
lead. Long-term exposure of adult fish to inorganic lead induces
sublethal effects on morphology, amino levulinic acid dehydratase
(delta-ALAD) and other enzyme activities, and avoidance behaviour at
available lead concentrations of 10-100 mg/litre. Juvenile stages are
generally more sensitive than adults, but eggs are often less
sensitive because lead is adsorbed onto the egg surface and excluded
from the embryo.

The acute and subacute toxicity of lead to various species of fish
and various life stages is summarized in Tables 5 and 6.

6.3.1. Toxicity of lead salts

Jones (1938) exposed stickleback (Gasterosteus aculeatus) to
lead nitrate under static conditions, with the water replaced every
24 h, and observed the survival time over a range of doses. For adults
45-50 mm long, average survival times after exposure to 0.02, 0.5, and
20 mg lead nitrate/litre were 11 days, 81 h, and 6.5 h, respectively.
For smaller adults (18-20 mm long) average survival times were 14 days,
10 days, and 2 days after exposure to 0.1, 0.5, and 3.0 mg/litre,
respectively. The addition to the lead solutions (50 mg/litre) of
calcium chloride at 2 mg/litre considerably lengthened the survival
time; fish survived for more than 10 days, as long as the controls.

Davies et al. (1976) studied the acute toxicity of lead nitrate to
rainbow trout (Salmo gairdneri) in 96-h static tests in hard and soft
water. Lead salts tend to precipitate out in hard water and the
authors' results were given in terms of both dissolved and total lead.
For two bioassays in hard water, the 96-h LC₅₀ values obtained were
1.32 and 1.47 mg dissolved lead nitrate/litre. The corresponding total
lead values for the test water were 542 and 471 mg/litre, respectively.
In a flow-through test using soft water, the 96-h LC₅₀ was
1.17 mg/litre for both dissolved and total lead, since all of the salt
was in solution. High levels of dissolved divalent anions in hard
water, therefore, protect fish from lead by reducing its availability
to them. This is also reflected in the results of Pickering &
Henderson (1966), who conducted tests on a variety of fish species
using lead chloride and lead acetate. There is a clear difference in
their results between hard and soft water for the same species (Table
5). Results in this study are presented as total lead.

Lloyd (1961) pointed out that dissolved oxygen levels tend to be
low in polluted water, while toxicity tests are conducted in water
fully saturated with oxygen. He examined the effect of varying
dissolved oxygen at low levels of lead salts, in that range of concen-
trations important for determining safe levels in water. At 65%
oxygen, lead toxicity increased over that obtained using fully
saturated water by a factor of 1.2 (ratio of concentrations which were
equitoxic) and, at 40% saturation, by a factor of 1.45.

Davies et al. (1976) conducted long-term bioassays with rainbow
trout to establish a maximum acceptable toxicant limit (MATC) for
inorganic lead. The effects of lead nitrate on reproduction, egg
survival, hatching success, and growth of the hatched larvae were
assessed. In the first chronic test, fingerling trout were exposed to
nominal total lead concentrations of 0, 40, 120, 360, 1080, or
3240 mg/litre. Actual dissolved lead was measured and results were
expressed in terms of dissolved salt. A MATC of between 0.018 and
0.032 mg/litre was found in terms of the "black tail effect". A
similar bioassay with soft water suggested a MATC between
0.041 mg/litre, where no black tails occurred, and 0.076 mg/litre,
where 4.7% of fish showed the black tail effect. These fish had been
hatched from exposed eggs. When fingerlings from non-exposed eggs were
used, in a soft water bioassay, the MATC for the black tail effect was
between 0.072 mg/litre, when no black tails were seen, and
0.146 mg/litre, where 41.3% of fish had black tails. There were no
significant differences between the measured dissolved lead concen-
trations in the two tests, indicating that fish from exposed eggs and
sac fry were more sensitive to the effects of lead than those from non-
exposed eggs. A long-term bioassay on reproductive effects established
that reproductive females and eggs were relatively insensitive to lead.
Therefore, the MATC is more realistic if based on the effects of lead
on the sensitive fingerling stage. Brood fish in the reproductive test
were exposed to lead concentrations, measured in the water, of 0.0005,
0.060, 0.077, 0.104, 0.175, and 0.270 mg/litre. Eggs and fry of the
F₁ generation were exposed to measured lead at 0.0005, 0.060, 0.119,
0.238, 0.476, and 0.952 mg/litre. There was no mortality or effect on
egg hatchability. The "black tail effect" was noted as the first
stage of toxic symptoms to lead about 6 months after the hard-water

study began. The entire caudal region at, or posterior to, the first
caudal vertebra was blackened. Tail blackening of the tail was
followed by spinal curvature and eroded caudal fins. This effect was
noted in soft water tests at about 6 weeks. There was no effect in
these studies on the growth of young trout, except where spinal
curvature was so severe as to affect feeding.

Table 5. Toxicity of lead salts to fish
---------------------------------------------------------------------------------------------------------
Organisms Life- Flow/ Temp. Alkali- Hard- pH Salt Parameter Water Reference
stage/ stat^a (°C) nity^b ness^b concentration
size (mg/litre)
---------------------------------------------------------------------------------------------------------
Fathead minnow adult stat 25 18 20 7.5 chloride 24-h LC₅₀ 8.18
(Pimephales (6.72-10.5) Pickering
promelas) adult stat 25 300 360 8.2 chloride 24-h LC₅₀ 482 &
(426-562) Henderson
adult stat 25 18 20 7.5 chloride 48-h LC₅₀ 5.99
(4.31-8.69) (1966)
adult stat 25 300 360 8.2 chloride 48-h LC₅₀ 482
(426-562)
adult stat 25 18 20 7.5 chloride 96-h LC₅₀ 5.58
(3.94-7.89)
adult stat 25 300 360 8.2 chloride 96-h LC₅₀ 482
(426-562)
adult stat 25 18 20 7.5 acetate 24-h LC₅₀ 14.6
(10.7-63.9)
adult stat 25 18 20 7.5 acetate 48-h LC₅₀ 10.4
(7.21-16.7)
adult stat 25 18 20 7.5 acetate 96-h LC₅₀ 7.48
(4.86-11.8)

Bluegill sunfish adult stat 25 18 20 7.5 chloride 24-h LC₅₀ 25.9
(Lepomis (22.5-30.4) Pickering
macrochirus) adult stat 25 300 360 8.2 chloride 24-h LC₅₀ 482 &
(426-562) Henderson
adult stat 25 18 20 7.5 chloride 48-h LC₅₀ 24.5
(20.9-29.1) (1966)
adult stat 25 300 360 8.2 chloride 48-h LC₅₀ 468
(410-549)
adult stat 25 18 20 7.5 chloride 96-h LC₅₀ 23.8
(20.0-28.4)
adult stat 25 300 360 8.2 chloride 96-h LC₅₀ 442
(379-524)

Table 5. (contd.)
---------------------------------------------------------------------------------------------------------
Organisms Life- Flow/ Temp. Alkali- Hard- pH Salt Parameter Water Reference
stage/ stat^a (°C) nity^b ness^b concentration
size (mg/litre)
---------------------------------------------------------------------------------------------------------

Goldfish adult stat 25 18 20 7.5 chloride 24-h LC₅₀ 45.4
(Carassius (39.4-53.6) Pickering
auratus) adult stat 25 18 20 7.5 chloride 48-h LC₅₀ 31.5 &
(25.0-39.8) Henderson
adult stat 25 18 20 7.5 chloride 96-h LC₅₀ 31.5
(25.0-39.8) (1966)
40-80 stat 19-25 0 6.0- nitrate 48-h LC₅₀ 6.6 Weir &
mm 6.9 (4.7-9.2) Hine (1970)
40-80 stat 19-25 50 6.0- nitrate 48-h LC₅₀ 110 Weir &
mm 6.9 (100-121) Hine (1970)

Guppy adult stat 25 18 20 7.5 chloride 24-h LC₅₀ 24.5
(Lebistes (20.9-29.1) Pickering
reticulatus) adult stat 25 18 20 7.5 chloride 48-h LC₅₀ 24.5 &
(20.9-29.1) Henderson
adult stat 25 18 20 7.5 chloride 96-h LC₅₀ 20.6
(16.4-26.8) (1966)

Bluegill sunfish stat 20 nitrate 24-h LC₅₀ 6.3 Turnbull
(Lepomis stat 20 nitrate 48-h LC₅₀ 6.3 et al.
macrochirus) (1954)

Rainbow trout adult flow 10.3- 86-94 133- 7.7 nitrate 21-day 2.3 Hodson et
(Salmo 137 LC₅₀ (1.6-3.3) al. (1978a)
gairdneri) adult stat 14 267 385 8.15 nitrate 96-h LC₅₀ 1.32 Davies et
(measured; al. (1976)
= 542
total lead)
adult stat 10 30 32 6.85 nitrate 96-h LC₅₀ 1.17 Davies et
adult stat 7 29 30 6.85 nitrate 14-day 0.20 al. (1976)
LC₅₀
juve- flow 82-132 6.4- nitrate 96-h LC₅₀ 8.0 Hale
nile 8.3 (1977)

Brook trout adult flow 12 42.6 44.3 4.1 nitrate 96-h LC₅₀ 4.1 Holcombe et
(Salvelinus al. (1976)
fontinalis)
---------------------------------------------------------------------------------------------------------

Channel catfish 1.6 g stat 18 44 7.1 arsenate 24-h LC₅₀ > 100 Mayer &
(Ictalurus 1.6 g stat 18 44 7.1 arsenate 96-h LC₅₀ > 100 Ellersieck
punctatus) (1986)

Mosquito fish adult stat 22-24 < 100 7.7- nitrate 24-h LC₅₀ 240 Wallen et
(Gambusia 8.3 al. (1957)
affinis) adult stat 22-24 < 100 7.7- nitrate 48-h LC₅₀ 240
8.3
adult stat 22-24 < 100 7.7- nitrate 96-h LC₅₀ 240
8.3
adult stat 18-20 < 100 7.1- oxide 24-h LC₅₀ > 56 000
7.2
adult stat 18-20 < 100 7.1- oxide 48-h LC₅₀ > 56 000
7.2
adult stat 18-20 < 100 7.1- oxide 96-h LC₅₀ > 56 000
7.2

Grey mullet 0.3- flow 11-13 34.4-34.8^c 6.9- nitrate 96-h LC₅₀ > 4.5 Taylor et
(Chelon labrosus) 3.2 g 8.5 al. (1985)
---------------------------------------------------------------------------------------------------------
^a Stat = static conditions (water unchanged for duration of test); flow = flow-through conditions
(lead concentration in water continuously maintained).
^b Alkalinity and hardness expressed as mg/litre CaCO₃.
^c These figures are values for salinity (expressed in ^o/_oo), not alkalinity.

Table 6. Toxicity of organolead to fish
---------------------------------------------------------------------------------------------------------
Organisms Size Flow/ Temp. Alkali- Hard- pH Comp- Parameter Water Reference
(mm) stat^a (°C) nity^b ness^b ound^d concentration
(mg/litre)
---------------------------------------------------------------------------------------------------------
Bass (young) 6 stat 20 TML 48-h LC₅₀ 0.10 Marchetti
(Morone 6 stat 20 TEL 48-h LC₅₀ 0.065 (1978)
labrax)
Tidewater 40- stat 20 55 7.6- TML 96-h LC₅₀ 13.5 Dawson et
silverside 100 7.9 al. (1977)
(Menidia
beryllina)

Bluegill sunfish 33- stat 23 55 7.6- TML 96-h LC₅₀ 84 Dawson et
(Lepomis 75 7.9 al. (1977)
macrochirus) 50- stat 20 33-81 84-163 6.9- TEL 24-h LC₅₀ 2.0 Turnbull
110 7.5 et al.
50- stat 20 33-81 84-163 6.9- TEL 48-h LC₅₀ 1.4 (1954)
110 7.5
TEL 96-h LC₅₀ 0.02 Wilber
(1969)
Plaice 52 flow 15 34.9^c TML 96-h LC₅₀ 0.05 Maddock &
(Pleuronectes 52 flow 15 34.9^c TEL 96-h LC₅₀ 0.23 Taylor
platessa) 52 stat 15 34.9^c TriML 96-h LC₅₀ 24.6 (1980)
52 stat 15 34.9^c TriEL 96-h LC₅₀ 1.7 Maddock &
52 stat 15 34.9^c DML 96-h LC₅₀ 300 Taylor
52 stat 15 34.9^c DEL 96-h LC₅₀ 75 (1980)
---------------------------------------------------------------------------------------------------------
^a Stat = static conditions (water unchanged for duration of test); flow = flow-through conditions
(lead concentration in water continuously maintained).
^b Alkalinity and hardness expressed as mg/litre CaCO₃.
^c These figures are values for salinity (expressed in ^o/_oo), not alkalinity.
^d TML = tetramethyl lead; TEL = tetraethyl lead; TriML = trimethyl lead chloride; TriEL = triethyl
lead chloride; DML = dimethyl lead dichloride; DEL = diethyl lead dichloride.
Hodson et al. (1978a) similarly reported blackened tails in rainbow
trout exposed to lead in the water at 0.120 mg/litre. After 32 weeks,
30% of the fish that survived showed black tails. The authors also
reported that exposure to lead at concentrations as low as
0.013 mg/litre led to increases in red blood cell numbers, decreases in
red blood cell volume, decreases in blood cell iron content, and
decreases of red blood cell amino levulinic acid dehydratase activity
(delta-ALAD). No changes in haematocrit or whole blood iron content
were observed. The changes indicated increased production of red blood
cells to compensate for increased death of red cells and inhibition of
haemoglobin production. There was no significant uptake of lead from
dietary dosing, though dietary lead might decrease uptake of dietary
iron. All of the lead which causes toxic effects in fish is taken up
directly from the water via the gills. Johansson-Sjobeck & Larsson
(1979) also reported the depression of activity of delta-ALAD in
rainbow trout exposed for 30 days to lead nitrate solutions (0.010,
0.075, and 0.30 mg/litre). The enzyme was depressed in red blood
cells, spleen, and renal tissue. Fish exposed to the highest lead
concentration also showed anaemia and basophilic stippling of the
erythrocytes. White blood cells were not affected. Holcombe et al.
(1976) exposed three generations of brook trout (Salvelinus
fontinalis) to lead nitrate in the water. All second generation
trout exposed to 0.235 or 0.474 mg total lead/litre, and 34% of those
exposed to 0.119 mg/litre, developed spinal deformities. Scoliosis
developed in 21% of newly hatched third generation fish exposed to
0.119 mg lead nitrate/litre. The weights of these same third
generation fish were significantly reduced 12 weeks after hatching.
The authors calculated a MATC for brook trout, based on the scoliosis
effect, of between 0.058 and 0.119 mg total lead/litre (0.039 and
0.084 mg dissolved lead/litre) in soft water (hardness: 44 mg
CaCO₃/litre) at a pH of between 6.8 and 7.6.

Hodson et al. (1980) examined the possibility that the toxic effect
of lead on salmonids was due to ascorbic acid deficiency, since the
symptoms were similar. They found no interaction between lead and
ascorbic acid deficiency in their effects on the fish; thus, the
toxicity of lead is not connected with ascorbic acid metabolism.

Weis & Weis (1977) exposed killifish eggs to lead nitrate at 0.1,
1.0, or 10 mg/litre in water. The lead was added at the start of the
study and the solutions were not replaced. The authors stated that the
removal of lead from the solution would have probably amounted to 79%
over 96 h, the period of the test. The lead only slightly reduced axis
formation in the embryos at all dose levels. At hatching, the fish
were examined for malformations. Only 20% of the fry were normal when
the added lead concentration was 1 mg/litre, the remainder having
skeletal malformations. Some 40% of the fish could not uncurl from the
position they had in the chorion and remained inactive, lying on the
bottom of the dish. All fish exposed to 10 mg/litre lead were perma-
nently curled. The curled fish could respond to tactile stimulation
but returned to the curled position. Ozoh (1979) exposed eggs of the
zebrafish (Brachydario rerio) to lead nitrate at 0, 0.036, or
0.072 mg/litre (measured) and monitored hatching success and abnor-
malities in the embryos. Compared with a control hatch rate of 80%,

lead at 0.036 and 0.072 mg/litre gave rates of 19.8% and 27%, respect-
ively. The presence of lead also resulted in poor absorption of yolk,
erosion of the tail and fin, spinal curvature, and outgrowths from the
fry (which appeared to be epitheliomas).

6.3.2. Biochemical effects

Christensen (1975) examined a range of biochemical parameters in
brook trout embryos and alevins exposed to lead nitrate (0.057 mg/litre
to 0.53 mg/litre) at the egg stage for 16 to 17 days, and then for a
further 21 days as alevins. No effects on the eggs were seen. For
alevins there was a decrease in weight, an increase in alkaline
phosphatase activity, and an increase in acetylcholinesterase activity.
Christensen et al. (1977) exposed brook trout (Salvelinus fontinalis)
to lead nitrate at concentrations ranging from 0.009 mg/litre to
0.474 mg lead/litre for 2- or 8-week periods. They found no signifi-
cant effects on body weight, body length, or on blood plasma glucose or
lactic dehydrogenases. There were significant decreases in blood
haemoglobin levels after exposure to lead at 58 µg/litre or more
(after both 2 and 8 weeks). Plasma glutamic oxaloacetic transaminase
activity was decreased after exposure to 34 µg/litre for both 2 and
8 weeks. Plasma sodium was elevated after exposure to 0.474 mg/litre
for 8 weeks and chloride was elevated after exposure to 0.235 mg/litre
for 2 weeks.

Hodson (1976) found that lead, as lead nitrate, in flowing water at
concentrations as low as 13 µg/litre, caused a significant inhibition
in the activity of red blood cell delta-ALAD in rainbow trout after
4 weeks. In a later study (Hodson et al., 1977), significant effects
on delta-ALAD were observed within 2 weeks of exposure of rainbow
trout, brook trout, goldfish, and pumpkinseed sunfish to lead
concentrations of 10, 90, 470, and 90 µg/litre, respectively. The
goldfish were affected by disease during the course of the test (4%
mortality) and this result should be treated with caution. Jackim
(1973) exposed mummichog (Fundulus heteroclitus) and winter flounder
(Pseudopleuronectes americanus) to an initial concentration of 10 mg
lead (as lead nitrate)/litre under static conditions in sea water.
There were decreases of 22% and 18.5% in liver delta-ALAD activity in
mummichogs after 96 h and 2 weeks, respectively. Winter flounder
showed decreases of 66% and 58% in delta-ALAD activity in liver and
kidney, respectively, after 1 week. It should be noted that the
concentration of lead in solution at the end of the 2-week mummichog
study was 0.8 mg/litre, only 8% of the initial concentration.

Shaffi (1979) examined various biochemical parameters in nine
species of freshwater fish exposed to lead nitrate at nominal concen-
trations of 5, 10, 15, or 20 mg/litre. Lead caused glycogenolysis in
all fish studied. The effect was greatest on muscle levels of carbo-
hydrate, with lesser effects on liver, kidney, and brain. There was an
inverse relationship between muscle, liver, and brain glycogen levels
and the lead concentration in the water, and a direct relationship
between lead levels in water and blood levels of glucose and lactose.
Major carp were most affected, while various species of catfish and
murrel were less sensitive to lead.

Sastry & Gupta (1978a) exposed catfish (Channa punctata) to lead
nitrate at 3.8 mg/litre, previous tests having established that this
concentration was sublethal. The fish were exposed for either 15 or
30 days and then sacrificed. Preparations were made of the stomach,
intestine, pyloric caeca, and liver for the estimation of enzyme
activities. There was no change in alkaline phosphatase activity in
liver or stomach, but intestine and pyloric caeca enzyme activities
showed marked inhibition after 15 days exposure. After 30 days
exposure, alkaline phosphatase activity differed from the control level
only in the pyloric caeca; now there was a marked elevation of
activity. Alkaline phosphatase elevation is usually associated with
cellular damage. After both 15 and 30 days of exposure, there was an
elevation in acid phosphatase activity in all tissues. Three
carbohydrases examined all showed an initial increase in activity
followed by a marked decline. Proteases were elevated in activity
throughout the experiment. In a later study, the same authors (Sastry
& Gupta, 1978b) examined the effect of lead nitrate on digestive
enzymes in vitro. There was a dose-related effect of lead, over the
range 0.4, 0.8, and 1.6 µmol/litre, on the activities of alkaline
phosphatase, lipase, tripeptide aminopeptidase, and glycylglycine
dipeptidase. The inhibition caused by lead was reversed by the
addition of EDTA.

Varanasi et al. (1975) reported effects on the properties of the
epidermal mucus of rainbow trout exposed to lead chloride at concen-
trations between 0.1 and 1.0 mg/litre. Using electron spin resonance
(ESR), the mucus was found to be more fluid after exposure to lead, and
the effect persisted after removal of the lead from the water. The
mucous characteristics of the epidermis affect swimming efficiency.

6.3.3. Behavioural effects

Giattina & Garton (1983) conducted avoidance behaviour experiments
on rainbow trout using inorganic lead salts, and they concluded that
trout will avoid lead at approximately 0.026 mg/litre (water hardness:
26 to 31 mg/litre). The value obtained by Jones (1948) of 0.4 mg/litre
for the avoidance of lead in solution for the minnow Phoxinus
phoxinus and the three-spined stickleback Gasterosterus aculeatus
appears to be related to total lead rather than dissolved salt. No
avoidance of total lead at 10, 20, or 40 mg/litre was found for green
sunfish (Lepomis cyanellus) by Summerfelt & Lewis (1967). Weir &
Hine (1970) pre-trained goldfish (Carassius auratus) to avoid
electric shock with a light stimulus and then exposed them to
solutions of lead nitrate. The lowest concentration of dissolved lead
nitrate found to impair significantly the behavioural response was
0.07 mg/litre. The authors determined the lowest concentration of lead
causing mortality in the same conditions to be 1.5 mg/litre. Addition
of calcium carbonate to the test solution reduced the effect at higher
lead concentrations. After exposure to lead nitrate at 10 mg/litre,
the impairment of behavioural response was 70%. This was reduced to
25% by the addition of calcium carbonate at 50 mg/litre. The lead-
exposed groups were retained and kept in clean water after the tests,
and were re-tested for behavioural response at four weekly intervals.
The effect of lead was permanent.

Ellgaard & Rudner (1982) exposed bluegill sunfish (Lepomis
macrochirus) to concentrations of lead acetate ranging from 0.1 to
300 mg/litre. The LC₅₀ for this species was found to be 400 mg/litre.
Locomotor behaviour was monitored and no effects were noted. The
absence of such effects at sublethal concentrations of metals is
markedly unusual.

6.4. Toxicity to Amphibia

Appraisal

There is evidence that frog and toad eggs are sensitive to nominal
lead concentrations of less than 1.0 mg/litre in standing water and
0.04 mg/litre in flow-through systems; arrested development and delayed
hatching have been observed. For adult frogs, there are no significant
effects below 5 mg/litre in aqueous solution, but lead in the diet at
10 mg/kg food has some biochemical effects.

Kaplan et al. (1967) exposed tree frogs (Rana pipiens) for
30 days to solutions of lead nitrate at between 25 and 300 mg
lead/litre. They found sloughing of the integument, loss of postural
tone, and sluggishness at all concentrations tested. All symptoms
worsened with increasing lead concentration. Total red and white blood
cell counts decreased progressively with increasing lead concen-
trations. Neutrophils and monocytes decreased at lower lead
concentrations and all white cells at higher concentrations. The
estimated LC₅₀ was 105 mg/litre. Frogs exposed to lead nitrate at
500 mg/litre for 2 weeks, or 1000 mg/litre for 48 h, showed erosion of
the gastric mucosa.

Khangarot et al. (1985) reported LC₅₀ values for tadpoles of the
frog Rana hexadactyla of 100, 66.7, 41.3, and 33.3 mg/litre after 24,
48, 72, and 96 h, respectively, at a temperature of 13-16 °C and a pH
of 6.2-6.7.

Dilling & Healey (1926) exposed groups of one male and three female
common frogs (Rana temporaria) for 3 weeks to water containing lead
nitrate at concentrations between 16.5 and 3300 mg/litre. At the
beginning of the study, the females were in full reproductive condition
and gravid with eggs. The solutions were regularly changed, and pond
weed was present in the tanks. All adult frogs died when exposed to
concentrations of 330 mg/litre or more. At a lead nitrate concen-
tration of 165 mg/litre, two batches of spawn were laid but no develop-
ment of the embryos occurred. At 33 mg/litre, development commenced
but proceeded no further than the late gastrula stage (day 10 of normal
development). Only a few embryos developed when exposed to
16.5 mg/litre, and the tadpoles were 30% smaller than controls.
Control animals produced two batches of spawn and all eggs developed.
A further series of experiments, where only the spawn, and not the
adults, was exposed to lead nitrate solutions, showed that lead
affected development at concentrations much lower than those first
tried. At 0.7 mg/litre, the development of most eggs was arrested,
although those tadpoles which did develop were normal after a late
hatch.

Birge et al. (1979) exposed narrow-mouth toad (Gastrophryne
carolinensis) eggs to inorganic lead, in a continuous-flow bioassay,
from fertilization through to 4 days post hatch (7 days exposure).
They estimated an LC₅₀ value of 0.04 mg/litre. Toads were found to be
more sensitive to lead than goldfish or rainbow trout examined in
parallel assays.

Ireland (1977) fed lead-contaminated earthworms to the African
clawed toad (Xenopus laevis) for 8 weeks. The earthworm diet
contained 10, 308, or 816 mg lead/kg. No toads died as a result of
lead ingestion. There were no significant effects on growth rate,
haemoglobin, haematocrit, or reticulocyte values, but blood delta-ALAD
activity was significantly reduced.

7. TOXICITY TO TERRESTRIAL ORGANISMS

7.1. Toxicity to Plants

Appraisal

Translocation of the ion in plants is limited and most bound lead
stays at root or leaf surfaces. As a result, in most experimental
studies on lead toxicity, high lead concentrations in the range of 100
to 1000 mg/kg soil are needed to cause visible toxic effects on photo-
synthesis, growth, or other parameters. Thus, lead is only likely to
affect plants at sites of very high environmental concentrations.

Bazzaz et al. (1974a) grew sunflowers (Helianthus annuus) plants
in vermiculite in a controlled environment room. After 3 to 5 weeks,
when the plants were 45 to 60 cm tall, the top 15 cm of each plant was
excised and placed in a solution of lead salts (concentrations of 2,
20, 100, or 200 mg/litre) for 5 days. All doses caused a reduction in
net photosynthesis and respiration over the exposure period. A 50%
reduction in photosynthesis corresponded to a leaf tissue lead concen-
tration of 193 mg/kg. In a second study, leaf peels were exposed to
lead solutions ranging from 10 to 1000 µmol/litre, which caused
reductions in stomatal opening of between 31% and 64%. The authors
suggest that this effect accounts for the reduction of photosynthesis
in the whole plant. Bazzaz et al. (1974b) grew corn and soybean plants
in media. Nine days after germination, they were treated with lead
chloride at concentrations varying between 250 and 4000 mg lead/litre.
The photosynthetic rate, measured as carbon dioxide uptake, of leaves
from corn plants was reduced to approximately 80% of the control level
at concentrations of 500, 1000, or 2000 mg lead/litre, and was further
reduced to 48% of the control level at 4000 mg lead/litre. The
transpiration rate was reduced at all dose levels from approximately
67% of the control level at 250 mg lead/litre to an almost negligible
rate at 4000 mg lead/litre. In soybeans, photosynthetic and transpi-
ration rates were enhanced at 250 and 500 mg lead/litre. A reduction
in the photosynthetic rate was found only at 4000 mg/litre, while
transpiration was reduced at both 2000 and 4000 mg/litre.

Broyer et al. (1972) found no effect on the yield of commercial
beans, barley, or tomato plants exposed to lead nitrate via a hydro-
ponic culture solution at lead concentrations of up to 50 µg/litre.

Barker (1972) exposed explants of cauliflower inflorescence stem,
lettuce stem, carrot root, and potato tubers to lead acetate at
concentrations of between 0.005 and 50 mg/litre of medium over 20 days.
There was a significant reduction in mean fresh weight of lettuce and
carrot after exposure to lead concentrations of 0.005 mg/litre or more.
Cauliflower and potato, both slower growing, showed significant
reductions in yield only at 0.5 mg/litre or more.

Hooper (1937) studied the effect of lead sulfate in the hydroponic
medium on the growth of dwarf French beans at concentrations ranging
from 3 to 30 mg lead/litre. She adjusted the particular salts used in
the medium to avoid the problem of lead salt precipitation. There was
no effect on growth over a period of 1 month. Other species of plants
were sprayed with a lead sulfate solution of 5 mg lead/litre. There
was no effect on Ulex europeus or on Lupinus arboreus. Even
spraying with supersaturated solutions which left a white coating of
sulfate on the leaves had no appreciable effect.

Dilling (1926) exposed cress and mustard seeds to a solution of
lead acetate (ranging from 0.5 to 5 g/litre, in terms of lead ion) for
up to 25 days. A concentration of >0.5 g/litre delayed germination and
initial growth. The delay increased with increasing lead concentration
until, at 2.7 g/litre, only a few seeds germinated. At 5 g/litre, no
germination occurred. Similar results were found when the author used
lead nitrate solutions ranging from 0.01 to 10 g/litre, in terms of
lead ion. Delayed germination and initial growth occurred at
0.1 g/litre or more, with no germination at 10 g/litre. The transfer
of cress seeds to clean water after exposure to 0.7 or 1.5 g/litre for
18 days allowed germination and normal growth to take place.

Bell & Patterson (1926) started hyacinth bulbs over solutions of
lead acetate from 0.0001 to 10 g/litre and found a graded inhibition of
root growth over this concentration range. Bulbs developing in
solutions of 1 or 10 g/litre showed complete arrest of root development
and stunted flowers and leaves. The same bulbs showed stunting the
following year when regrown over tap water.

Davis & Barnes (1973) dosed growing seedlings of loblolly pine
(Pinus taeda) and red maple (Acer rubrum) with solutions of lead
chloride between 2 x 10^-4 and 5 x 10^-3 mol/litre twice weekly for
2.5 months. Following exposure to 10^-3 mol/litre or more, they
observed a significant reduction in height and root dry weight for
both species, and a reduction in stem dry weight for red maple.
There was a significant reduction in pine stem dry weight at
5 x 10^-3 mol/litre, and a significant increase in the maple leaf
anthocyanin content at 10^-3 mol/litre or more.

Keaton (1937) monitored the growth of pot-grown barley after the
addition of lead nitrate or carbonate to the soil. At concentrations
up to 3000 mg/kg soil, there were no deleterious effects on barley
growth, and at low lead application rates, there was a small
stimulation in barley growth (nitrate acts as a fertilizer at low
rates). This stimulation was most marked at lead concentrations of
between 0.1 and 0.4 mg/kg soil. Most of the lead was found to be
fixed to the soil particles. Soluble lead available to the plant
did increase with amount of salt added, but very little of the total
lead was soluble. Oberlander & Roth (1978) measured the uptake of
labelled potassium (⁴²K) into the roots and shoots of 7-day-old
barley plants from nutrient solutions containing lead. Uptake was
monitored over 5 h during exposure to lead at between 10-6 and
10^-4 mol/litre. Potassium uptake was reduced significantly to 48%
of the control level by a lead concentration of 10^-4 mol/litre.

Dijkshoorn et al. (1979) added lead acetate, to give concentrations
of between 11.4 and 1062 mg/kg, and fertilizer to sandy loam soil.
The soil was placed into pots and three successive crops of plants
were grown in the soil: plantain (Plantago lanceolata), clover
(Trifolium repens), and ryegrass (Lolium perenne). Lead had no
effect on plant yield even at the highest concentrations tested. The
"uptake" of lead into the plant was at a constant ratio of 0.1 to the
level in the soil. Lagerwerff et al. (1973) grew maize (Zea mays)
and alfalfa in a greenhouse in silt loam at two soil pH levels (5.2
and 7.2), with lead chloride added to 64, 113, and 212 mg/kg. Total
yield data (dry weight of plants) showed no effect of either lead or pH
in maize. For alfalfa, there was no effect of lead at pH 5.2, but at
pH 7.2 there was a significant increase in yield over controls with no
lead. Baumhardt & Welch (1972) found that emergence, plant height, and
grain yield of maize were not affected by a field application of lead
acetate at a rate of 50 to 3200 kg/ha. No effects were noted on
morphology, colour, maturity, or other growth parameters during the
2-year study. Carter & Wain (1964) investigated the use of lead
nitrate as a fungicide in broad bean plants. The salt was toxic to
fungi at sap concentrations of >0.1 mmol/litre, but was also toxic to
the plant.

7.2. Toxicity to Invertebrates

Appraisal

The information available is too meagre to quantify the risks to
invertebrates during the decomposition of lead-contaminated litter.

Doelman et al. (1984) incubated a mixed culture of bacteria in lead
nitrate solutions and grew the fungus Alternaria solani on malt agar
to which lead nitrate had been added. The cultures were used as food
for the nematodes Mesorhabditus monohystera and Aphelenchus avenae,
which were reared for up to 22 days on bacteria and fungus, respect-
ively. Lead was taken up by bacteria to give a range of doses to the
nematode of between 7.6 and 110 µg/g of food. All these exposures
had a significant inhibitory effect on the reproduction of
Mesorhabditus monohystera. A lead concentration of 2.47 µg/g in
fungus strongly inhibited the reproduction of Aphelenchus avenae but
variation was considerable; no statistics were presented for the fungal
study.

Beyer & Anderson (1985) exposed woodlice (Porcellio scaber) to
treated soil litter containing between 100 and 12 800 mg/kg dry weight
of lead, as lead oxide, over 64 weeks. No significant effect was found
on adult survival, number of young produced or on survival of young at
exposures up to 6400 mg lead/kg. There was a significant reduction in
all three parameters after exposure to 12 800 mg/kg. Beeby (1980) fed

woodlice (Porcellio scaber) during "gestation" on cocksfoot
grass (Dactylis glomerata) which had been dosed with lead (2911 or
16 483 mg/kg), as lead nitrate, and also on grass which had been
collected from roadside verges. The verge grass contained 110 or
407 mg/kg (having been collected from two different sites). There was
no deleterious effect at any of the exposure levels on the fertility of
the woodlice after oviposition had occurred. Lead levels in gravid
females correlated positively with body calcium levels and with the
number of days on the contaminated diet.

Weismann & Skrobak (1980) fed the caterpillar Scotia segetum on a
semisynthetic food to which lead had been added, and calculated
LT₅₀ values for lead chloride, at exposure levels of 250 and 500 mg/kg
diet, of 72.1 and 28.7 h, respectively. For lead acetate at levels of
250 and 500 mg/kg diet, the LT₅₀ values were 75.6 and 31.9 h, respect-
ively. An increased ascorbic acid content in the diet (1000 mg/kg)
reduced the lead toxicity by between 42% and 52%, but increased calcium
in the diet (1000 mg calcium carbonate/kg) had no effect on lead tox-
icity. Weismann & Svatarakova (1981) fed the same species of cater-
pillar on natural diets contaminated with lead at various doses (50,
100, 200, 400, or 800 mg/kg diet) throughout development. There were
reproductive effects at all doses, dependent on the instar of the
larvae at first exposure. Only 20% of third instar larvae exposed to
50 mg lead/kg developed to the adult stage. These adults were deformed
and the females failed to produce eggs. Only 40-73% of larvae fed on a
diet containing lead at 50 to 200 mg/kg, from the third instar,
produced pupae.

7.3. Toxicity to Birds

Appraisal

Lead salts are only toxic to birds at a high dietary dosage
(100 mg/kg or more). Almost all of the experimental work is on
chickens and other gallinaceous birds. Exposure of quail from hatching
and up to reproductive age resulted in effects on egg production at
dietary lead levels of 10 mg/kg. Although a variety of effects at high
dosage have been reported, most can be explained as a primary effect on
food consumption. Diarrhoea and lack of appetite, leading to anorexia
and weight loss, are the primary effects of lead salts. Since there is
no experimental evidence to assess effects on other bird species, it is
necessary to assume a comparable sensitivity. If this is so, then it
is highly improbable that environmental exposure would cause adverse
effects.

There is little information on the effects of organolead compounds.
Trialkyllead compounds produced effects on starlings dosed at
0.2 mg/day; 2 mg/day was invariably fatal.

The short-term and long-term dietary toxicity of lead salts and
organolead is summarized in Table 7.

7.3.1. Toxicity of lead salts

Lead salts have low to moderate acute and short-term toxicity to
birds. Lethal and severe sublethal effects have not been reported at
levels likely to be found in the wild. Some sublethal effects have
been noted after realistic exposure, but these are unlikely to affect
bird populations.

7.3.1.1 Toxicity to birds' eggs

Ridgway & Karnofsky (1952) injected lead nitrate solutions into
the yolk sac of chicken eggs, after 4 or 8 days of development, and
into the chorio-allantoic membrane after 8 days of development. The
LD₅₀ was 0.30 at 4 days and 4.50 at 8 days, expressed as molar
equivalents of lead, for the yolk sac route, and 3.00 molar equivalents
at 8 days for the chorio-allantoic route. The 4-day result is equiv-
alent to 0.10 mg lead nitrate/egg.

Haegele et al. (1974) dosed female mallards with 100 mg lead/kg
diet. This was added as a mixture of 43 mg/kg lead carbonate, 37 mg/kg
lead oxide, and 49 mg/kg lead sulfate, each salt contributing one-third
of the total lead. No significant effect on eggshell thickness was
found when it was measured on days 76 and 85 of treatment. When lead
was added to the diet along with DDE at 40 mg/kg, lead did not increase
the effect of the organochlorine on shell thickness.

7.3.1.2 Toxicity to adult and juvenile birds

Vengris & Mare (1974) exposed 6-week-old chickens to lead acetate
in drinking water for 35 days at doses ranging from 20 to 640 mg
lead/litre. The chickens were found to tolerate lead in the water at
concentrations up to 160 mg/litre without showing any clinical or
haematological signs, despite blood lead levels as high as
6.2 mg/litre. At a dose of 320 mg lead/litre, the chickens exhibited
early signs of lethargy and weakness, followed by anorexia, anaemia,
and loss of weight. Peripheral paralysis occurred prior to death. Six
out of twelve birds died within 30 days, and all surviving birds had
decreased haemoglobin levels at 30 days. At the highest dose of
640 mg/litre, similar clinical signs were observed but all birds died
within 34 days. Long-term exposure to lead at levels producing no
clinical symptoms had no effect on antibody production against
Newcastle disease virus.

Table 7. Acute and dietary toxicity of lead to birds
---------------------------------------------------------------------------------------------------------
Species Age Compound Parameter Concentration Reference
(mg/kg)
---------------------------------------------------------------------------------------------------------
Japanese quail 3-4 months tetraethyllead acute LD₅₀^a 24.6(14.7-41.3) Hudson et al. (1984)
(Coturnix coturnix 14 days powdered 5-day LC₅₀ > 5000 Hill & Camardese
japonica) metallic lead (1986)^b
14 days lead nitrate 5-day LC₅₀ > 5000 Hill & Camardese
14 days lead sub- 5-day LC₅₀ > 5000 (1986)^b
acetate
14 days lead arsenate 5-day LC₅₀ 2761(1622-4701) Hill & Camardese
(1986)^b

Mallard duck 3-4 months tetraethyllead acute LD₅₀^a 107(44.5-258) Hudson et al. (1984)
(Anas young lead nitrate < 100-day LC₅₀ > 500 DeWitt et al. (1963)
platyrhynchos) adult lead nitrate < 100-day LC₅₀ > 50 DeWitt at al. (1963)
---------------------------------------------------------------------------------------------------------
^a Single oral dose expressed as mg/kg body weight.
^b Hill & Camardese (1986) fed quail with a dosed diet for 5 days followed by a clean diet for 3 days.
In two separate studies, Damron et al. (1969) dosed 4-week-old
broiler chickens with dietary lead acetate at levels between 10 and
2000 mg lead/kg for 4 weeks. They report that, at dietary lead levels
of 100 mg/kg or less, there was no significant effect on body weight
gain or on food consumption. At dosing levels of 1000 and 2000 mg/kg,
there was a significant depression of body weight gain and food
consumption.

Morgan et al. (1975) dosed newly-hatched Japanese quail with lead
acetate in the diet at 10, 100, 500, and 1000 mg lead/kg for 5 weeks.
There was a significant effect on body weight after dosing with 500 and
1000 mg lead/kg diet (food consumption was not monitored), and blood
haemoglobin content was reduced in the same birds. A reduction in
haematocrit was found after dosing with 1000 mg lead/kg diet, but only
between weeks 4 and 5 of age. Relative weights of bursa, spleen,
liver, and heart were not affected. After 5 weeks of dosing, testis
size was reduced in birds fed 1000 mg lead/kg. All quail were able to
express a normal primary humoral immune response following antigenic
challenge with a saline suspension of sheep red blood cells, at 4 weeks
of age. Relative adrenal weights were significantly increased after
5 weeks on the diets containing 500 or 1000 mg lead/kg. A similar
experiment, but with the dosing beginning at 6 days of age, showed no
adrenal effect.

Edens et. al. (1976) investigated the effects of dietary lead
acetate on reproductive performance in Japanese quail (Coturnix
coturnix japonica). Chicks were reared from hatching on food to which
lead acetate had been added to give 0, 1, 10, 100, or 1000 mg lead/kg.
When chicks were 6 weeks old, they were transferred to a layer diet,
similarly dosed with lead, and housed in pairs. The lighting schedule
was continuous light for the first week, followed by 1 week on 10 h of
light. Thereafter, lighting was increased by 1 h per day each week
until the birds were receiving 14 h of light per day at 6 weeks of age.
At this point they were paired. The quail were killed at 12 weeks of
age. Records were kept of when females produced their first egg, rates
of egg production, and hatchability of artificially-incubated eggs,
together with body weights of adults. Only the highest dose rate
(1000 mg/kg diet) affected growth of the birds. For the first 6 weeks,
the body weight of treated birds, both males and females, was lower
than that of controls. By the age of 12 weeks, treated males had
caught up with controls but females were still significantly lighter.
Egg production by females was depressed even at the lowest dose, and
higher dose levels of lead acetate produced a greater effect. The
highest dose level almost completely suppressed egg production and the
few eggs produced at this dose level were soft-shelled or shell-less.
Maximum rate of egg production was reached at 8 weeks of age in both
control birds and females fed 1 mg/kg diet. This peak of egg
production was delayed until the birds were 12 weeks old in the groups
fed 10 or 100 mg/kg diet. There was also a delay in onset of egg
laying, relative to controls, in groups fed 10, 100, and 1000 mg/kg
diet. The highest dose also significantly delayed sexual maturity
relative to other dosed groups. The hatch rate of eggs laid by groups
fed 100 or 1000 mg/kg was significantly reduced.

Damron & Wilson (1975) conducted a series of studies to determine
the toxicity of lead, in various forms, to bobwhite quail (Colinus
virginianus). At dietary dose rates of lead acetate of up to
1500 mg/kg during 6 weeks, juvenile birds showed no effect on body
weight gain, food consumption, or mortality, and adult males showed no
effect on semen quality or organ weights. Feeding birds at a dietary
dose rate of 3000 mg/kg led to a significant depression in growth rate
and an increase in mortality. In a similar study using white Chinese
geese (Johnson & Damron, 1982), feeding lead acetate at dietary levels
up to 2000 mg/kg had no effect on body weight or food consumption. At
2000 mg/kg diet, there was a slight increase in the size of the liver
and some yellow discoloration.

Coburn et al. (1951) dosed adult mallard ducks (Anas
platyrhynchus) daily with aqueous solutions of lead nitrate, intro-
duced directly into the gizzard via a catheter. They found that a
daily dose of 6 mg/kg body weight had no effect on body weight, red
blood cell counts, or haemoglobin content over a period of 132 days,
but with daily doses of 8 or 12 mg/kg body weight, there was a decrease
in these parameters within 3 to 4 weeks. Kendall & Scanlon (1982)
dosed adult male ringed turtle doves (Streptopelia risoria) with lead
acetate, by intubation, at levels of 0, 25, 50, or 75 mg lead/kg body
weight, daily for 7 days. At the highest rate of dosing, the birds
lost 17% of their original body weight; weight loss was lower at the
other two dosing rates (5% and 8% for 25 and 50 mg/kg per day,
respectively). None of these weight changes was statistically signifi-
cant. Schafer et al. (1983) estimated an 18-h LD₅₀ for the red-winged
blackbird (Agelaius phoeniceus) of >111 mg lead toxicity/kg body
weight. This value was based on estimated intake from dosed food.

7.3.1.3 Enzyme effects

Dieter et al. (1976) established a correlation between the lead
levels in the blood of canvasback ducks and the activity of the enzyme
delta-ALAD. The ducks had taken up the lead from their natural
environment. A level of 0.20 mg lead/litre blood was associated with a
75% decrease in enzyme activity. Kendall & Scanlon (1982) reported a
similar correlation between lead residues in ring doves and delta-ALAD
activity.

7.3.1.4 Behavioural effects

Frederick (1976) fed mallard ducklings on a diet containing lead
nitrate (dissolved in propylene glycol) at 0, 5, 50, or 500 mg lead/kg
diet. There was no effect of any of the treatments on the general
activity of the ducklings after 3 and 8 days on these diets, but there
was a significant, dose-related effect on weight gain.

Barthalmus et al. (1977) dosed trained pigeons by gastric intu-
bation daily with 6.25, 12.5, or 25 mg lead acetate/kg body weight.
The pigeons had been trained to peck response keys for a food reward in
a complex system requiring multiple responses to obtain the reward.
The lowest dose produced no significant effect on behavioural perform-
ance. The highest dose led to mortality after 18-35 days, and there

were noticeable behavioural effects after 3-10 days. The middle dose
of 12.5 mg/kg produced no deaths, but did significantly alter
behavioural response after 30 days.

7.3.2. Toxicity of metallic lead

Lead shot taken into the gizzard of birds is highly toxic. Birds
are affected or killed by small numbers of shot. Powdered lead appears
to be less toxic, probably because it is not retained in the upper
gut.

7.3.2.1 Toxicity of powdered lead

Hill & Camardese (1986) dosed Japanese quail with powdered
metallic lead in the diet at doses ranging from 1000 to 5000 mg/kg
diet. There was no mortality after 5 days on the lead-containing diet,
or after a further 8 days of observation on a clean diet. At dose
levels of 1495 or 2236 mg/kg diet, food consumption was unaffected.

Pattee (1984) fed American kestrels (Falco sparverius) with
metallic lead in the diet at doses of 0, 10, or 50 mg/kg for 7 months.
Although lead levels were elevated in the bones and liver of birds on
treated diets, particularly at the highest dose level, no adverse
effects were found with respect to survival, egg laying, initiation of
incubation, fertility, or eggshell thickness. Hoffman et al. (1985a)
dosed 1-day-old nestling American kestrels for 10 days with powdered
metallic lead in corn oil daily (25, 125, or 625 mg/kg body weight per
day). The birds were fed on mice in the mornings prior to dosing by
intubation, and survivors were sacrificed on day 10. The only
mortality occurred at the highest dose rate; 4 out of 10 birds died
between days 6 and 8 of dosing. There was a significant effect on
weight gain, but only at the two highest doses. After 10 days of
dosing, birds given 625 mg/kg were 61% of control weight and birds
given 125 mg/kg were 84% of control weight. Birds dosed at 25 mg/kg
were 95% of control weight, not significantly different. In those
groups which were affected, weight was reduced after days 4 and 5 of
dosing. Mean brain weights of the groups given 625 and 125 mg/kg were
14% and 9%, respectively, lower than controls after 10 days. This
reflected a general lack of growth because brain weight to body weight
ratios were elevated relative to controls. There was also an effect
on the skeleton, in addition to effects on soft tissues. Growth in
both wing bones was reduced by 34-35% in the 625-mg/kg group and by
18-19% in the 125-mg/kg group. In a separate report (Hoffman et al.,
1985b), the effects on biochemical and haematological indicators were
given. Nestling American kestrels showed reduced haematocrit, haemo-
globin level, and plasma creatine phosphorylase activity after 10 days
of dosing with lead at 125 or 625 mg/kg body weight. Red blood cell
delta-ALAD activity was depressed by these dose levels and also at
25 mg/kg. Brain, liver, and kidney delta-ALAD activities were
inhibited by all lead treatments. Liver protein content and brain RNA
to protein ratio decreased after lead treatment, whereas liver DNA, DNA
to RNA ratio, and DNA to protein ratio increased. Brain monoamine
oxidase and ATPase activity was not significantly altered by lead at
these doses. The authors considered that these effects could explain,
in part, the delayed development of the nestlings.

7.3.2.2 Toxicity of lead shot

Clemens et al. (1975) dosed adult mallard with five no. 6 lead shot
and observed the birds over 20 days. The birds showed body weight loss
over this period, together with clinical signs including green
diarrhoea, anorexia, and weakness. High concentrations of lead in the
blood, kidney, liver, and bone were recorded but there were lower
concentrations in skeletal muscle. Birds on a high-fibre diet showed
more severe clinical signs and higher tissue lead concentrations than
birds on low-fibre diets. Mautino & Bell (1987) dosed mallard with two
no. 4 lead shot and observed signs of lead toxicosis within 24 h.
Varying degrees of paralysis, kinetic ataxia, or abnormal locomotor
function were shown by 14 out of 17 birds. These neurological signs
gradually disappeared and 8 days after dosing all birds appeared
normal. The blood lead level was highest after 1 week at 7.8 mg/litre
and remained significantly higher than the control value for a further
6 weeks. Blood samples were taken at weekly intervals. No lesions
were found in the birds after 7 weeks. The effect of lead on blood
delta-ALAD activity was maximal after 1 week, with 80% inhibition, and
gradually returned to normal over the 7-week study.

Irwin & Karstad (1972) exposed adult mallard drakes for 14 weeks
to concentrations of 17.8, 89, or 178 g of particulate lead/m² in a
simulated marsh area. The mortality was 17%, 57%, and 100% for the
three dose levels, respectively. All birds gave a positive fluorescent
erythrocyte test and showed chronic lead toxicosis. Birds exposed to
the highest concentration showed overt signs of lead poisoning and all
died within 23 days. Finley et al. (1976) dosed male and female
mallard with either one number 4 lead shot or one number 4 lead/iron
combination shot (with 47% lead), and observed the birds for 4 weeks.
No mortality was recorded and no tissue lesions were found. There was
a correlation between lead residues in the bone and the number of eggs
laid; the more eggs laid, the greater the residue of lead in the bone.
This presumably reflects the greater movement of calcium out of bone to
produce eggshells and its replacement from dietary calcium. After
Dieter & Finley (1978) dosed male and female mallard with a single
number 4 lead shot, two out of 60 birds died, showing signs typical of
lead poisoning at necropsy. One month after dosing, the blood lead
level was 0.317 mg/litre and the inhibition of erythrocyte delta-ALAD
activity was 53%. After 3 months, inhibition was 30% and after
4 months was 15%, due to removal of lead from the circulation.

Chasko et al. (1984) captured wild mallard (Anas platyrhynchos)
and black duck (Anas rubripes) and maintained them in captivity on a
"natural diet" consisting of millet and buckweed, available at all
times, together with duckweed, eelgrass, fish, sand shrimp, mussels,
crabs, and snails, available for some of the time. Groups of 10 ducks
(5 of each species) were dosed with 0, 2, or 5 lead shot or with 5 lead
shot given singly over a 2-week period. More lead was accumulated in
tissues from repeat dosing with single shot than with single dosing
with 5 shot. Mortality was similar for the two species, with the black
duck slightly more susceptible to lead. One out of 4 black ducks dosed
with 2 shot died; 2 out of 4 died after dosing with 5 shot. Weight
loss was also similar for both species. Birds with clear symptoms of
lead poisoning showed a weight loss of about 20%; ducks which died had

lost between 30% and 50% of body weight. Mortality generally increased
with dose rate of lead shot. Grandy et al. (1968) dosed 15 mallard
with 8 lead shot each and observed the effects over 30 days. Birds
were also dosed with shot containing less lead (an alloy of 40% lead
and 60% tin). Those mallard dosed with pure lead shot showed 100%
mortality; all died between 5 and 15 days after dosing. Birds fed the
alloy shot showed 27% mortality, with birds dying between 8 and 30 days
after dosing. Rozman et al. (1974) dosed adult female mallard ducks
with 8 lead shot, orally by gavage, and monitored serum enzyme
activities over the next 14 days. They reported significant increases
in the activity of serum glutamic pyruvic transaminase (SGPT) and
decreases in that of serum alkaline phosphatase (SAP) after lead treat-
ment. These enzyme changes were suspected to reflect tissue damage.

Chinese white geese dosed with a total of 200 lead shot over a
12-week period did not die (Johnson & Damron, 1982). This is in marked
contrast to studies in other species where only a few shot caused
death in a short time. Cook & Trainer (1966) exposed 10 Canada geese,
some adult males, some females, and some immatures, to lead pellets
(2-100 per bird) introduced directly into the oesophagus. The highest
recorded blood lead level was 16.8 mg/litre in an immature bird dosed
with 100 pellets. The lethal dose was found to be 4 to 5 pellets; the
two birds dosed with 5 pellets died within 39 and 72 days, respect-
ively. Regardless of the numbers of pellets introduced into the
gizzard, there was uniform erosion of lead from the pellets. The rate
of erosion of the pellets was initially very rapid, with a 65% to 70%
loss of lead within the first 5 days. The pellets had almost disap-
peared within 35 days. Gross signs of lead toxicity included weakness
and lethargy, anorexia, green diarrhoea, loss of weight, and oedematous
heads. The loss of weight was most noticeable in birds given lower
doses of lead, since those given high doses died while still retaining
good body condition. Necropsy findings included impaction of the
proventriculus, roughened and greenish staining of the gizzard lining,
severe enteritis, distended gall bladder, discoloured liver, and
flaccid heart. These pathological lesions were more noticeable in
birds which survived longer and, therefore, had experienced the effects
of the lead for longer periods.

Damron & Wilson (1975) found that dosing adult male bobwhite quail
with 10 or more lead shot per week for 4 weeks increased mortality.
More than 90% of males dosed with 30 lead shot per week died within
4 weeks.

Patee et al. (1981) dosed bald eagles with 10 lead shot each,
repeating the dose if the bird succeeded in regurgitating the shot.
Four out of five eagles died; the fifth was killed when it became blind
133 days after dosing. The time taken for the birds to die varied
between 10 and 125 days, though three birds died within 20 days. Body
weight loss varied between 16% and 23%, those birds which died quickly
losing less weight than those surviving longer. In a study lasting
60 days, Stendell (1980) fed American kestrels (Falco sparverius)
daily with either one number 9 shot (given in a dead mouse) or with
mallard which had died from lead poisoning and contained residues of 27
to 34 µg/kg body weight. No kestrels died or exhibited visible signs
of lead poisoning.

7.3.3. Toxicity of organolead compounds

Too few reports are available to demonstrate clearly the effects of
organolead compounds on birds. It is of moderate toxicity to birds
(Table 7). Tetraethyllead is readily converted to triethyllead in
water and in animals. Results suggest that trialkyllead compounds are
very toxic, but effects on only one species have been reported.

Haegele & Tucker (1974) showed that tetraethyllead had no effect on
eggshell thickness in mallard ducks or Japanese quail at a dose of
6.0 mg/kg body weight over 6 days. There was a transitory effect, but
normal thickness returned within the 6-day study.

Osborn et al. (1983) dosed starlings (Sturnus vulgaris) with
either trimethyl- or triethyllead in two separate experiments at doses
of 0, 0.2, or 2 mg/day for 11 days (approximately equivalent to
28 mg/kg body weight per day at the highest dose). All birds given the
highest dose of trimethyl- or triethyllead died within 6 days. Pre-
death symptoms were relatively mild in the case of triethyllead, con-
sisting of slightly slower respiratory rate and a tendency to squat,
rather than stand, with fluffed-out feathers as if cold. The effects
of trimethyllead were more dramatic. Within 24 h of the first dose,
one of the birds was so badly coordinated that it was unable to perch
or stand normally. Only a single bird, out of the group of six,
appeared normal at this stage. Within 6 h of the second dose, all
birds showed symptoms of lack of coordination. One was unable to
place accurately its bill in the feeder. There was considerable weight
loss. All birds died, or were killed for humanitarian reasons, within
the first 5 days. Birds on the highest dose of trimethyllead (but not
with triethyllead) had bright green watery droppings. Food consumption
was greatly reduced at the high-dose levels; this was not surprising
considering the lack of coordination. There was also an effect on the
feeding behaviour of birds receiving 0.2 mg/day. They ate approxi-
mately the same amount of food, on average, as the control birds, but
there was considerable variation from day to day in the amount eaten.
This was noticeable after very few doses, possibly occurring after a
single dose. Liver weights were significantly lower than those of
controls in the case of the high dose of triethyllead and both the high
and low doses of trimethyllead. Kidney weight was reduced only in
birds receiving the high dose of trimethyllead.

7.4. Toxicity to Non-Laboratory Mammals

There are many reports of lead levels in wild mammals but few
reports of toxic effects of the metal in the wild or in non-laboratory
species.

Kilham et al. (1962) captured wild rats from the area of a dump at
Hanover, New Hampshire, USA, which contained heavy metals. Nearly all
of the sampled animals showed intranuclear inclusion bodies in the
kidneys which were absent from populations of laboratory rats. These
inclusions were identical in staining and electron-microscopic charac-
teristics to similar bodies induced by lead in the laboratory. Renal
tumours were found in some of the rats associated with the inclusion
bodies. The livers of the trapped animals contained lead. Earlier

reports (Hindle & Stevenson, 1930; Hindle, 1932; Syveston & Larson,
1947) showed similar inclusion bodies in rats trapped in sewers in
London and New York. Zook et al. (1972) reported the killing of
34 simian primates and three fruit bats in Washington Zoo by lead in
paint on their cages, and reviewed other examples of zoo animals
poisoned by leaded paint.

8. EFFECTS OF LEAD IN THE FIELD

8.1. Tolerance of Plants to Lead

Plant tolerance to metals has been reviewed by Bradshaw et al.
(1965), Antonovics et al. (1971), and Wainwright & Woolhouse (1975).
Holl & Hampp (1975) and Peterson (1978) have reviewed the specific case
of tolerance to lead. Most work has concentrated on plants growing on
mining wastes rather than roadside verges.

The general conclusions are as follows. Metal tolerance is almost
always specific, i.e., tolerance to one metal does not confer tolerance
to others. There are degrees of tolerance, the metal content of par-
ticular soils correlating with the degree of tolerance of the local
plant population. Tolerance is inherited, i.e., tolerant parents
transmit tolerance to their offspring. Within a plant species, there
are tolerant and sensitive populations. Tolerance, therefore, develops
by selection, rather than by adaptation of individuals. Two possible
mechanisms for tolerance, to metals in general, have been identified;
an "external" mechanism prevents metal entering the plant, while an
"internal" mechanisms allows entry but prevents the metal from coming
into contact with sensitive processes within the organism.

Appraisal

Jowett (1958) studied lead tolerance in the grasses Agrostis
tenuis and A. stolonifera by measuring root growth in a culture sol-
ution containing lead nitrate at either 75 or 125 µmol/litre. Both
species, from either "control areas" or from areas rich in metals
other than lead, showed little tolerance. The growth, relative to
plants not exposed to lead, was <37% and <22%, respectively, for the
two lead levels. In A. tenuis from a mining area rich in zinc and
lead, the values were 80% and 62%, respectively. Bradshaw (1952) grew
A. tenuis taken from a disused lead mine (with 1% lead in the soil)
and from an uncontaminated site 100 metres away. In uncontaminated
soil, the plants from the mining area were smaller and grew more slowly
than the plants from the contaminated area. In soil from the mine,
plants collected from this site grew normally, whereas the others
showed no growth (50% of tillers were dead or dying within 3 months).

Briggs (1972) collected the liverwort Marchantia polymorpha from
city areas with soil lead concentrations of 252, 401, and 898 mg/kg dry
weight and from a control area (28 mg/kg dry weight). The plants were
exposed to lead nitrate in agar at a concentration of 400 mg lead/kg
for 7 days and increase in thallus length was monitored. There was no
effect on the plants from the city areas, but the control plant growth
was significantly reduced.

Malone et al. (1974) showed that lead was concentrated in the cell
walls of maize (Zea mays) and, therefore, excluded from interference
with biochemical processes. Lead also tended to be concentrated on the
surface of the roots of plants and excluded from the shoots.

8.2. Highways and Industrial Sources of Lead

Appraisal

No effect on the reproduction of birds nesting near highways has
been observed. Toxic effects have been observed in pigeons in urban
areas, the kidneys being most frequently affected.

In a report by Grue et al. (1984), swallows nesting near highways
accumulated significant amounts of lead, but there were no effects on
the number of eggs produced, number of nestlings, nestling body
weights, or body weights of adults. In a similar study (Grue et al.,
1986), starlings also accumulated lead but there were no effects on the
same reproductive parameters. In feral pigeons (Columba livia) in
London, Hutton (1980) detected effects including increased kidney
weight, presence of renal inclusion bodies, altered kidney mito-
chondrial structure, and function and depression of delta-ALAD activity
in blood, liver, and kidney. The effects were less than would have
been predicted from laboratory experiments. The author suggested that
factors, such as changes in the distribution of lead at the tissue and
organelle level, and the antagonistic action of zinc, might be respon-
sible.

Mierau & Favara (1975) measured lead in deer mouse populations
close to roads, and considered that the residues were 5 times too low
to cause any reproductive effects. Clark (1979) suggested that doses
of lead ingested by little brown bats, shrews, and voles from roadside
verges equalled or exceeded those which have caused mortality or
reproductive impairment in domestic mammals. Lead concentrations in
bats and shrews exceeded those concentrations found in mammals from
mining areas showing renal abnormalities.

8.3. Lead Shot

Appraisal

A report from the Nature Conservancy Council's Working Group in the
United Kingdom (NCC, 1981) discussed the problem of swan deaths
attributable to lead poisoning. Mute swans in the United Kingdom
showed 8% to 15% decreases in population numbers between 1955 and 1978.
During the period 1961-1978, there were large differences in swan
population changes in different parts of the country. Populations
increased in northern Scotland, north Wales, and parts of eastern and
southern England, whereas there were marked declines in central and
southern Scotland, North-West England, the Midlands, South Wales, and

the lower Thames Valley. Of the kills of swans reported between 1966
and 1978, 56% had no cause attributed, though some would have died from
natural causes. In the years 1980 and 1981, the Ministry of
Agriculture Fisheries and Food conducted postmortems on 288 mute swans.
They reported that 39.2% of the swans had died from lead poisoning, the
largest single cause of death found. Again there were regional
differences with 50% of English swans dying of lead poisoning, but none
of the Scottish swans. The source of the lead was either gun-shot or
anglers split shot. The two can be distinguished using antimony
content. Birds ingest particulate material, which may be contaminated
with lead shot, to grind food in the gizzard before digestion.
Postmortems on 299 mute swans carried out between 1973 and 1980
revealed gun-shot in only five birds. Other swan species are more
likely to contain gun-shot; two-thirds of the Whooper and Bewick swans
dying of lead poisoning on the Ouse Washes contained gun-shot. Lead
from petrol in pleasure boats has been discounted as a source of lead
in the birds. The report acknowledges the problem that lead use by
anglers has not changed appreciably in 150 years, yet the elevated swan
death rate is a recent phenomenon. The most likely explanation for
this is the distribution of aquatic plant life. In recent years,
marginal and submerged plants have been killed by pollution from boats
and, more significantly, by the use of herbicides to keep channels
clear. The lack of marginal plant life would make lead shot more
available to swans. Other species of waterfowl also contain lead shot
and are sometimes killed by it. These include greylag geese, mallard,
pochard, tufted duck, and goldeneye. The highest incidence of lead
shot contamination is in mallard in autumn at inland sites rather than
coastal ones.

Gun-shot is a more important source of lead in birds in North
America. Bagley et al. (1967) collected dead or dying Canada geese and
found that dying birds showed marked cephalic oedema, with subman-
dibular swellings, oedema of eyelids, and a profuse discharge from eyes
and nares. Shot was found in the gizzards of the geese and high lead
levels were recorded in liver, tibia, and kidney. Anderson (1975)
studied about 1500 waterfowl dying at Rice Lake, Illinois, USA, and
found that lesser scaup made up 75% of 394 birds collected dead or
dying. Of 96 scaup examined, 75% had lead, at least one pellet, in the
gizzard. Lead levels averaged 46 mg/kg in the liver and 66 and
40 mg/kg in kidney and wing bone, respectively. The incident occurred
following a period of drought which killed food plants. With a return
to normal water levels, plants began to grow again but lead pellets
were more readily available in the feeding sites.

Trainer & Hunt (1965) estimated that 1700 Canada geese succumbed to
lead poisoning in Wisconsin between 1940 and 1965. Other species were
also affected. Lewis & Ledger (1968) found that mourning doves taken
from a public field managed for shooting contained lead shot. Of 1949
gizzards examined, 1% contained between 1 and 24 shot. Examination of
the area revealed 10 890 pre-shooting and 43 560 post-shooting shot per
acre. Locke & Bagley (1967) found that gizzards from 4 out of 62 shot
birds contained lead and that lead levels in 43 livers ranged from 0.4
to 14 mg/kg.

8.4. Organic Lead

Appraisal

In the autumn of 1979, about 2400 birds were found dead or dying in
the Mersey estuary, United Kingdom, the majority being dunlin, a wader
(Bull et al., 1983). Smaller numbers were found in 1980 and 1981.
There is a plant manufacturing petrol additives in the vicinity.
Affected birds contained elevated lead levels, mostly as alkyllead.
The livers of dead birds from the incident contained an average of
11.14 mg total lead/kg wet weight, sick birds 8.85 mg/kg, apparently
healthy birds from the same area 4.5 mg/kg, and healthy birds from
another estuary 0.14 mg/kg. The authors note that Head et al. (1980)
found 1 mg lead/kg in Macoma balthica, a food source for the waders,
during the incident. Bull et al. (1983) concluded that total liver
lead was sufficiently high to result in death. It was mostly in the
form of alkyllead, which is at least as toxic as inorganic lead (Osborn
et al., 1983). Symptoms were similar to those of inorganic lead
poisoning and dissimilar to the effects of other pollutants present in
the area. The high liver concentration, compared to the kidney concen-
tration, was taken to indicate a recent acute exposure. There were no
other toxic chemicals in significant amounts in the area, and there was
no indication of disease. In the area discharging waste to the
estuary, there was an industrial source manufacturing anti-knock
compounds.

Gill et al. (1960) investigated effluent output from tetraethyllead
production plants to assess the likely environmental hazard of a new
plant. They measured 48-h LC₅₀s of the effluent, containing some
alkyllead, for three-spined stickleback and coho salmon at 14 g/litre,
and they concluded that the effluent would pose no hazard. No attempt
was made to assess indirect hazard caused to birds by food organisms
concentrating the lead.

In 1974, the 2000 ton cargo ship, "Cavtat", sank in a water depth
of 94 m, 5.6 km from the Adriatic coast of Italy. Its cargo consisted
of 325 tons of lead anti-knock compounds. At the time of recovery of
the vessel, a loss of 7% of this cargo was estimated. Tiravanti &
Boari (1979) concluded that the lead compounds were restricted to a
limited area around the wreck and, based on water concentrations of
<10 µg alkyllead/litre, had no significant environmental effect.

9. EVALUATION

9.1. General Considerations

In evaluating the environmental hazard of lead, it is necessary to
extrapolate from laboratory studies to ecosystems. This must be done
with extreme caution for the following reasons.

(a) The availability of lead to organisms in the environment is
limited by its strong adsorption to environmental components,
such as soil sediment, organic matter, and biota. It is
accepted that biomagnification of lead does not take place;
i.e., there is no increase in concentration of the metal in
food-chains. However, environmental contamination with lead
is widespread and organisms do accumulate high body burdens of
lead.

(b) Environmental variables such as temperature, pH and chemical
composition of water, soil type, and geology have been shown
in limited studies on a narrow range of species to affect both
the uptake and the effect of lead.

(c) Available, rather than nominal or total, lead is the determi-
nant parameter in assessing uptake by, and effects on,
organisms.

(d) There is limited data from controlled experimental studies on
the effects of mixtures of metals. Organisms in the environ-
ment are exposed to mixtures of pollutants. Acid deposition
can release various metals into the environment.

(e) Little experimental work has been carried out on species or
communities that are either representative or key components
of natural communities and ecosystems. Studies have not
considered all of the interactions between populations and all
of the environmental factors affecting these populations.

It is probable that subtle disturbances to the community would
occur at much lower concentrations than those suggested in laboratory
studies on acute effects. Much of the available information on lead
toxicity is based on experimental studies carried out at unrealisti-
cally high nominal concentrations and short-term exposure. This makes
it difficult to extrapolate to field conditions.

9.2. The Aquatic Environment

Lead enters the aquatic environment through surface runoff and
deposition of airborne lead. Adsorption to sediments occurs rapidly
and almost quantitatively.

The uptake and accumulation of lead by aquatic organisms from water
and sediments are influenced by various environmental factors. These
must be taken into consideration when evaluating the hazards of
environmental contamination by lead.

Lead uptake by aquatic organisms is slow and reaches equilibrium
only after prolonged exposure. Aquatic organisms at low trophic levels
show a much higher accumulation of lead than those at higher trophic
levels, reaching bioconcentration factors of up to 100 000. On the
other hand, biomagnification through food chains is very low, often
exhibiting values far below 1. However, this by no means indicates the
absence of hazard.

The toxicity of lead to aquatic organisms varies considerably
depending on availability, uptake, and species sensitivity; generally,
the earlier life stages are more vulnerable. Lead interferes with
biochemical, physiological, morphological, and behavioural parameters.

Organolead compounds are generally 10-100 times more toxic to
aquatic organisms than is inorganic lead. Tetraalkyllead becomes toxic
by conversion into trialkyllead.

9.3. The Terrestrial Environment

Lead is introduced to terrestrial communities by atmospheric
deposition on to exposed surfaces. There is insufficient evidence to
indicate a hazard to terrestrial organisms from airborne lead. Normal
concentrations of lead in soil range from 15 to 30 mg/kg; roadside
soils can reach 5000 mg/kg and soils from industrial sites may exceed
30 000 mg/kg. Although soil retards the movement of lead through
terrestrial communities, some lead may be leached from highly contami-
nated soils. Some soil lead is taken up by plants and passed to
animals, but a major fraction is accumulated at the surface of root
cells. Some of the factors that determine availability to plants are
pH, organic matter, and soil type. Generally, lead is not toxic to
plants at soil concentrations below 1000 mg/kg. Some plant populations
can tolerate higher concentrations, and some appear to develop a
genetic tolerance. Animals are exposed to lead through the ingestion
of water, food, soil, and dust. In all cases, the concentrations in
animals are related to environmental concentrations, and in most cases,
lead appears to accumulate preferentially in calcified tissues.
Certain bird populations are also exposed to lead shot.

It is improbable that environmental exposures cause acute adverse
effects in most terrestrial populations. However, lead shot is a major
hazard in certain bird populations that tend to ingest gravel into the
gizzard to grind food. Laboratory studies indicate that the expected
effects on animals would be changes in behaviour, disruption of
haematological metabolism, and inhibition of certain enzymes. There
may be a strong correlation with calcium metabolism.

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    See Also:
       Toxicological Abbreviations