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    INTERNATIONAL PROGRAMME ON CHEMICAL SAFETY


    ENVIRONMENTAL HEALTH CRITERIA 85





    LEAD - ENVIRONMENTAL ASPECTS














    This report contains the collective views of an international group of
    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 g/m3      in large cities  with dense automobile  traffic to  less
than  0.2 g/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

     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.

    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/ Organb  Tem-       pH       Compound  Dura-  Exposure  Bioconcen- Refer-    
                 stage/    typea         perature                      tion   (g/      tration    ence
                 size                    ( C)                         (days) litre)    factorc
---------------------------------------------------------------------------------------------------------
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       13.4d      Watling
 (Crassostrea                                                                                       (1983a)
  gigas)

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

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

Mussel                       B   WB                          chloride  21     100       31.7d      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/ Organb  Tem-       pH       Compound  Dura-  Exposure  Bioconcen- Refer-
                 stage/    typea         perature                      tion   (g/      tration    ence
                 size                    ( C)                         (days) litre)    factorc
---------------------------------------------------------------------------------------------------------

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 g/g  0.04e      (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 g/g  0.05e
                             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 g/g 0.025e
                             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 g/g 0.05e

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) 
---------------------------------------------------------------------------------------------------------

Table 1. (contd.)
---------------------------------------------------------------------------------------------------------
Organism         Life-     Test/ Organb  Tem-       pH       Compound  Dura-  Exposure  Bioconcen- Refer-    
                 stage/    typea         perature                      tion   (g/      tration    ence
                 size                    ( C)                         (days) litre)    factorc
---------------------------------------------------------------------------------------------------------

Pumpkinseed      10-20 g     A   WB      18-20      6.0      nitrate   8      40        4.88f      Merlini 
sunfish                                                                                            & Pozzi
 (Lepomis         12-21 g     A   WB      18-20      7.5      nitrate   8      40        1.86f      (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       1081g      (1981)
  reticulatus)    150-200 mg  D   WB      21.1-24.7  7.2-7.8  nitrate   7      13 g/g   0.38e
                 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       3459g
                 150-200 mg  D   WB      21.1-24.7  7.2-7.8  nitrate   28     23 g/g   0.7e
                 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      367g
                 150-200 mg  D   WB      21.1-24.7  7.2-7.8  nitrate   7      27 g/g   0.48e
                 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      1015g
                 150-200 mg  D   WB      21.1-24.7  7.2-7.8  nitrate   28     68 g/g   0.52e

Rainbow trout    1.0 g       D   WB      14-15.8    7.7-8.1  tetra-    7      3.5       725.7d     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   Organa   Compound   Dura-   Exposure  Bioconcen-   Reference
                                                          tion    (mg/kg)   tration
                                                          (days)            factor
---------------------------------------------------------------------------------------------------------
Corn                          soil    shoots   nitrate    30b     4233      0.07         Zimdahl et al.
( Zea mays)                    soil    roots    nitrate    30b     4233      0.33         (1978)
                              soil    shoots   sulfate    30b     4564      0.05         Zimdahl et al.
                              soil    roots    sulfate    30b     4564      0.08         (1978)

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

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

Wheat                         soil    shoots   nitrate    30b     4233      0.02         Zimdahl et al.
 (Tritium aestivum)            soil    roots    nitrate    30b     4233      0.2          (1978)
                              soil    shoots   sulfate    30b     4564      0.009        Zimdahl et al.
                              soil    roots    sulfate    30b     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      25d       0.084c       Hoffman
 (Falco sparverius)  nestling  oral    liver    metallic   10      25d       0.05c        et al. (1985a)

Starling            adult     oral    kidney   triethyl   11      2.85d     0.65c        Osborn
 (Sturnus vulgaris)  adult     oral    kidney   trimethyl  11      2.85d     1.9c         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

     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 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 mol/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 mol/litre    and  13.0 mg/kg after  exposure  to
10 mol/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 g    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 g/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 g/litre    and  5.1 g/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 g    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 g 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 g    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 g   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 g/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 g/litre)   was recorded after exposure at  pH  6
and  a water concentration of  180 g   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 g/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 g/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.

     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
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 g/litre   and rose to 317 g/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 g/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

     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.

    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

     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.

    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

     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.

    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 g/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 m3      to
91 m3      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 m3       to
61 m3.       Values  for  EC50 for   cell  volume  were  calculated at
140 g    lead  acetate/litre  for  Selenastrum and  700 g/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    stata  (C)   nityc    nessc                              concent-     
                                                                                    ration    
                                                                                    (mg/litre)
---------------------------------------------------------------------------------------------------------
American oyster           stat   25-27  25d                   nitrate   48-h LC50   2.45 
 (Crassostrea                                                                        (2.2-3.6)   Calabrese
  virginica)               stat   25-27  25d                   nitrate   48-h LC0    0.5         et al.
                          stat   25-27  25d                   nitrate   48-h LC100  > 6.0      (1973)

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

Softshell clam            stat   21.5-  29-31d           7.8- nitrate   48-h LC50   > 50       Eisler
 (Mya arenaria)                   22.5                    8                                      (1977)
                          stat   21.5-  29-31d           7.8- nitrate   96-h LC50   27              
                                 22.5                    8
                          stat   21.5-  29-31d           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   > 7.5e     Reish 
 arenacoe-        adult    stat                           7.8  acetate   96-h LC50   > 10e      et al. 
 dentata                                                                                         (1976)
                 juvenile stat                           7.8  acetate   28-day LC50 2.5e        Reish 
                 adult    stat                           7.8  acetate   28-day LC50 3.2e        et al. 
                                                                                                (1976)
 Capitella        larva    stat                           7.8  acetate   96-h LC50   1.2e        Reish 
 capitella        adult                                   7.8  acetate   96-h LC50   6.8e        et al. 
                                                                                                (1976)
                 adult                                   7.8  acetate   28-day LC50 1.0e        Reish 
                                                                                                et al. 
---------------------------------------------------------------------------------------------------------

Table 3.  (contd.)
---------------------------------------------------------------------------------------------------------
Organism         Life-    Flow/  Temp.  Alkali-  Hard-   pH   Salt      Parameter   Water       Reference
                 stage    stata  (C)   nityc    nessc                              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   0.45f       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    stata  (C)   nityc    nessc                              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                  flowb  15-17                   7.0  chloride  96-h LC50   2.6         Boutet &
 (Austropotamobius         flowb  15-17                   7.0  chloride  30-day LC50 1.5         Chaise-
  pallipes pallipes)       flowb  15-17                   7.0  chloride  30-day LC50 0.9e        martin
                                                                                                (1973)
Crayfish                  flowb  15-17                   7.0  chloride  96-h LC50   3.3         Boutet &
 (Orconectes limosus)      flowb  15-17                   7.0  chloride  30-day LC50 1.7         Chaise-
                          flowb  15-17                   7.0  chloride  30-day LC50 0.9e        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  Compoundb  Parameter   Water          Reference
                  length weight  stata  (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
ECo 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 g/litre    over 120 days.
There   was  no  effect  on  survival  at  concentrations  of  3.8  and
12 g/litre,      but   mortality   occurred   at   concentrations   of
19 g/litre    or  more.   The growth  rate  of  the survivors  was not
affected at lead concentrations of 19 g/litre.    The authors observed
a 50% reduction in snail biomass production after exposure to  lead  at
36 g/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 m3,     the control  size, to 93.1 m3.       Under hypoxic
conditions, control nuclear volume increased to 137.9 m3,      but  in
lead-treated  animals  increased  only to  99.6 m3.       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 g   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/ stata (C)   nityb    nessb                             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/ stata (C)   nityb    nessb                             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)
---------------------------------------------------------------------------------------------------------

Table 5. (contd.)
---------------------------------------------------------------------------------------------------------
Organisms         Life-  Flow/ Temp.  Alkali-  Hard-  pH    Salt      Parameter  Water         Reference
                  stage/ stata (C)   nityb    nessb                             concentration
                  size                                                           (mg/litre)
---------------------------------------------------------------------------------------------------------
 Sarotherodon             stat  28-32  7.7-11.7 32-           nitrate  24-h LC50  > 40         Subbaiah et 
  mossambicus                                   38                                              al. (1983)

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  34.4-34.8c      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)  stata  (C)  nityb    nessb           oundd              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    34.9c                    TML     96-h LC50  0.05        Maddock &
 (Pleuronectes      52    flow   15    34.9c                    TEL     96-h LC50  0.23        Taylor
 platessa)          52    stat   15    34.9c                    TriML   96-h LC50  24.6        (1980)
                   52    stat   15    34.9c                    TriEL   96-h LC50  1.7         Maddock &
                   52    stat   15    34.9c                    DML     96-h LC50  300         Taylor
                   52    stat   15    34.9c                    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 g/litre   or  more
(after both 2 and 8 weeks).  Plasma glutamic  oxaloacetic  transaminase
activity was decreased after exposure to 34 g/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 g/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 g/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 mol/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

     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 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 mol/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 g/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

     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.

    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 g/g   of food.  All  these exposures
had   a   significant  inhibitory   effect   on  the   reproduction  of
 Mesorhabditus   monohystera.   A lead  concentration of 2.47 g/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.

     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.

    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 LD50a       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 LD50a       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 g/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

     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.

    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 mol/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

     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 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

     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.

    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 g 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