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Chapter 7 Contents

Chapter 7. Ecological Effects of Lead ___________________________________________________________________ 7-1

7.1. Introduction to Ecological Concepts ____________________________________________________________ 7-1

7.1.1. Ecosystem Scale, Function, and Structure ________________________________________________ 7-2

7.1.2. Ecosystem Services _________________________________________________________________ 7-3

7.1.3. Critical Loads as an Organizing Principle for Ecological Effects of Atmospheric Deposition ________ 7-4

7.1.4. Ecosystem Exposure, Lag Time and Re-entrainment of Historically Deposited Lead _______________ 7-6

Table 7-1. Comparison among several metals: Time to achieve 95% of steady state metal 

concentration in soil; example in a temperate system _______________________7-7

7.2. Terrestrial Ecosystem Effects__________________________________________________________________ 7-8

7.2.1. Introduction to Terrestrial Ecosystem Effects _____________________________________________ 7-8

7.2.2. Soil Biogeochemistry and Chemical Effects ______________________________________________ 7-9

7.2.2.1. pH and CEC _______________________________________________________________ 7-9

7.2.2.2. Organic Matter ____________________________________________________________ 7-10

7.2.2.3. Aging 7-11

7.2.3. Bioavailability in Terrestrial Systems __________________________________________________ 7-12

7.2.3.1. Plants 7-13

7.2.3.2. Invertebrates ______________________________________________________________ 7-19

7.2.3.3. Terrestrial Vertebrates ______________________________________________________ 7-22

7.2.3.4. Food Web ________________________________________________________________ 7-23

Table 7-2. Soil-to-tissue bioaccumulation factors for various terrestrial plant, 

invertebrate, and vertebrate species ___________________________________7-25

7.2.4. Biological Effects__________________________________________________________________ 7-26

7.2.4.1. Plants and Lichen __________________________________________________________ 7-26

7.2.4.2. Invertebrates ______________________________________________________________ 7-30

7.2.4.3. Terrestrial Vertebrates ______________________________________________________ 7-33

7.2.5. Exposure and Response of Terrestrial Species____________________________________________ 7-35

7.2.6. Community and Ecosystem Effects ____________________________________________________ 7-37

7.2.7. Critical Loads in Terrestrial Systems ___________________________________________________ 7-41

7.2.8. Soil Screening Levels_______________________________________________________________ 7-42

7.2.9. Characterization of Sensitivity and Vulnerability _________________________________________ 7-43

7.2.9.1. Species Sensitivity _________________________________________________________ 7-43

7.2.9.2. Nutritional Factors _________________________________________________________ 7-43

7.2.9.3. Soil Aging and Site-Specific Bioavailability _____________________________________ 7-44

7.2.9.4. Ecosystem Vulnerability ____________________________________________________ 7-44

7.2.10. Ecosystem Services ________________________________________________________________ 7-45

7.2.11. Summary of Effects in Terrestrial Systems ______________________________________________ 7-46

7.2.11.1. Biogeochemistry and Chemical Effects _________________________________________ 7-46

7.2.11.2. Bioavailability and Uptake ___________________________________________________ 7-47

7.2.11.3. Biological Effects __________________________________________________________ 7-48

7.2.11.4. Exposure Response_________________________________________________________ 7-49

7.2.11.5. Community and Ecosystem Effects ____________________________________________ 7-49

7.2.11.6. Critical Loads, Sensitivity and Vulnerability _____________________________________ 7-50

7.3. Aquatic Ecosystem Effects __________________________________________________________________ 7-50

7.3.1. Introduction to Aquatic Ecosystem Effects ______________________________________________ 7-50

7.3.2. Biogeochemistry and Chemical Effects _________________________________________________ 7-52

7.3.2.1. Other Metals______________________________________________________________ 7-54

7.3.2.2. Biofilm __________________________________________________________________ 7-55

7.3.2.3. Carbonate ________________________________________________________________ 7-55
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7.3.2.4. Dissolved Organic Matter (DOM) _____________________________________________ 7-56

7.3.3. Bioavailability in Aquatic Systems ____________________________________________________ 7-57

7.3.3.1. Plants and Algae___________________________________________________________ 7-61

Table 7-3. Bioconcentration factors for Pb in aquatic plants ___________________________7-65

7.3.3.2. Invertebrates ______________________________________________________________ 7-65

Table 7-4. Bioaccumulation factors for Pb in aquatic invertebrates______________________7-69

7.3.3.3. Vertebrates _______________________________________________________________ 7-69

Table 7-5. Bioaccumulation factors for Pb in fish____________________________________7-75

7.3.3.4. Food Web ________________________________________________________________ 7-76

7.3.4. Biological Effects__________________________________________________________________ 7-78

7.3.4.1. Plants and Algae___________________________________________________________ 7-78

7.3.4.2. Invertebrates ______________________________________________________________ 7-81

7.3.4.3. Vertebrates _______________________________________________________________ 7-85

7.3.5. Exposure and Response of Aquatic Species______________________________________________ 7-94

7.3.6. Community and Ecosystem Effects ____________________________________________________ 7-98

7.3.7. Critical Loads in Aquatic Systems ____________________________________________________ 7-101

7.3.8. Characterization of Sensitivity and Vulnerability ________________________________________ 7-102

7.3.8.1. Seasonally-Affected Physiological Changes ____________________________________ 7-102

7.3.8.2. Increased Nutrient Uptake __________________________________________________ 7-103

7.3.8.3. Temperature and pH_______________________________________________________ 7-103

7.3.8.4. Life Stage _______________________________________________________________ 7-104

7.3.8.5. Species Sensitivity ________________________________________________________ 7-104

7.3.8.6. Ecosystem Vulnerability ___________________________________________________ 7-106

7.3.9. Ecosystem Services _______________________________________________________________ 7-106

7.3.10. Summary of Aquatic Effects ________________________________________________________ 7-108

7.3.10.1. Biogeochemistry and Chemical Effects ________________________________________ 7-108

7.3.10.2. Bioavailability ___________________________________________________________ 7-108

7.3.10.3. Biological Effects _________________________________________________________ 7-110

7.3.10.4. Exposure and Response ____________________________________________________ 7-111

7.3.10.5. Community and Ecosystem Effects ___________________________________________ 7-111

7.3.10.6. Critical Loads, Sensitivity and Vulnerability ____________________________________ 7-112

7.4. Causality Determinations for Lead in Terrestrial and Aquatic Systems________________________________ 7-113

Table 7-6. Summary of Pb causal determinations for plants, invertebrates and vertebrates___7-113

7.4.1. Bioaccumulation of Lead in Terrestrial and Aquatic Biota as it Affects Ecosystem Services _______ 7-113

7.4.2. Mortality________________________________________________________________________ 7-115

7.4.3. Growth Effects ___________________________________________________________________ 7-116

7.4.4. Physiological Stress _______________________________________________________________ 7-117

7.4.5. Hematological Effects _____________________________________________________________ 7-118

7.4.6. Developmental and Reproductive Effects ______________________________________________ 7-118

7.4.7. Neurobehavioral Effects____________________________________________________________ 7-120

7.4.8. Other Physiological Effects _________________________________________________________ 7-121

7.4.9. Community and Ecosystem Level Effects ______________________________________________ 7-121

Chapter 7. References ___________________________________________________________________________ 7-124
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Chapter 7. Ecological Effects of Lead

1 This chapter synthesizes and evaluates the most policy-relevant science to help form the foundation 

2 for the review of the secondary (welfare-based) NAAQS for Pb. The Clean Air Act definition of welfare 

3 effects includes, but is not limited to, effects on soils, water, wildlife, vegetation, visibility, weather, and 

4 climate, as well as effects on materials, economic values, and personal comfort and well-being. This 

5 chapter discusses the effects of Pb on ecosystem components and processes and is organized into four 

6 sections. The introduction (Section 7.1) presents the organizing principles of this chapter and several 

7 basic concepts of metal ecotoxicology and ecosystem services. Section 7.2 reviews the effects of Pb on 

8 terrestrial ecosystems; how soil biogeochemistry affects Pb bioavailability, biological effects of Pb 

9 exposure and subsequent vulnerability of particular ecosystems, and critical loads for soils. A similar 

10 discussion of the effects of Pb on aquatic ecosystems is presented in Section 7.3, including water-only 

11 exposures and sediment related effects. Both the terrestrial and aquatic system sections conclude with a 

12 discussion of alterations in ecosystem service functions as a consequence of Pb deposition. Finally, an 

13 integrative synthesis of effects of Pb across biota and causal determinations for Pb in both terrestrial and 

14 aquatic systems are presented in Section 7.4. Areas not addressed here include literature related to Pb shot 

15 or pellets and studies that examine human health-related endpoints which are described in other chapters

16 of this document.

7.1. Introduction to Ecological Concepts

17 Metals, including Pb, occur naturally in the environment at measurable concentrations in soils, 

18 sediments, and water. Organisms have developed adaptive mechanisms for living with metals, some of 

19 which are required micronutrients (but not Pb). However, anthropogenic enrichment can result in 

20 concentrations that exceed the capacity of organisms to regulate internal concentrations, causing a toxic 

21 response and potentially death. Differences in environmental chemistry may enhance or inhibit uptake of 

22 metal from the environment, thus creating a spatial patchwork of environments that are at greater risk 

23 than other environments. Similarly, organisms vary in their degree of adaptation to, or tolerance of, the 

24 presence of metals. These fundamental principles of how metals interact with organisms and ecosystems 

25 are described in detail in EPA’s Framework for Metals Risk Assessment (Fairbrother et al., 2007). This 

26 section introduces critical concepts for understanding how Pb from atmospheric deposition may affect 

 

Note: Hyperlinks to the reference citations throughout this document will take you to the NCEA HERO database (Health and Environmental 

Research Online) at http://epa.gov/hero. HERO is a database of scientific literature used by U.S. EPA in the process of developing science 

assessments such as the Integrated Science Assessments (ISA) and the Integrated Risk Information System (IRIS).
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1 organisms, communities, and ecosystems. The sections that follow provide more detail for how aquatic 

2 and terrestrial ecosystems respond to Pb and how environmental chemistry interacts with organisms to 

3 affect exposure and uptake. 

7.1.1. Ecosystem Scale, Function, and Structure

4 An ecosystem is defined as the interactive system formed from all living organisms (biota) and 

5 their abiotic (chemical and physical) environment within a given area. Ecosystems cover a hierarchy of 

6 spatial scales and can comprise the entire globe, biomes at the continental scale, or small, well7 circumscribed systems such as a small pond (U.S. EPA, 2008). A pond may be a small but complex 

8 system with multiple trophic levels ranging from phytoplankton to several feeding guilds of fish plus fish9 eating birds or mammals. A large lake, on the other hand, may be a very simple ecosystem, such as the 

10 Great Salt Lake in Utah that covers approximately 1,700 square miles but contains only bacteria, algae, 

11 diatoms, and two invertebrate species. All ecosystems, regardless of size or complexity, share the 

12 commonality of multiple interactions between biota and abiotic factors, and a reduction in entropy 

13 through energy flow from photosynthetic organisms to top predators. This includes both structural (e.g., 

14 soil type and food web trophic levels) and functional (e.g., energy flow, decomposition, nitrification) 

15 attributes. 

16 Ecosystems are most often defined by their structure, and are based on the number and type of 

17 species present. Individual organisms of the same species are similar in appearance and genetics, and can 

18 interbreed and produce fertile offspring. Interbreeding groups of individual organisms within the same 

19 species form populations, and populations of different species form communities. The community 

20 composition may also define an ecosystem type, such as a pine forest or a tall grass prairie. Pollutants can 

21 affect the ecosystem structure at any of these levels of biological organization (Suter et al., 2005). 

22 Individual plants or animals may exhibit changes in metabolism, enzyme activities, hormone function, or 

23 overall growth rates or may suffer gross lesions, tumors, deformities, or other pathologies. Effects on the 

24 nervous system of animals may cause behavioral changes that alter breeding behaviors or predator 

25 avoidance. However, effects on individuals must result in changes to their survival or reproductive output 

26 to have any effect on the population. Population level effects of pollutants include changes over time in 

27 abundance or density (number of individuals in a defined area), age or sex structure, and production or 

28 sustainable rates of harvest (Barnthouse, 2007). Community level attributes affected by pollutants include 

29 species richness and abundance (also known as biodiversity), dominance of one species over another, or 

30 size (area) of the community. Pollutants may affect communities in ways that are not observable in 

31 organisms or populations (Bartell, 2007), including: (1) effects resulting from interactions between 

32 species, such as altering predation rates or competitive advantage; (2) indirect effects, such as reducing or 
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1 removing one species from the assemblage and allowing another to emerge (Petraitis & Latham, 1999); 

2 and (3) alterations in trophic structure.

3 Alternatively, ecosystems may be defined on a functional basis, such as rates of photosynthesis, 

4 decomposition, nitrification, or carbon cycling. Pollutants may affect abiotic conditions (e.g., soil 

5 chemistry), which indirectly influences biotic structure and function (Bartell, 2007). Feedback loops or

6 networks influence the stability of the system, and can be mathematically described through simplistic or 

7 complex process, or energy flow, models (Bartell, 2007). For example, the Comprehensive Aquatic 

8 Systems Model (CASM) is a bioenergetics-based multi compartment model that describes the daily 

9 production of biomass (carbon) by populations of aquatic plants and animals over an annual cycle 

10 (DeAngelis et al., 1989). CASM, originally designed to examine theoretical relationships between food 

11 web structure, nutrient cycling, and ecosystem stability, has since been adapted for risk assessments and 

12 has been applied to numerous lakes with a variety of pollutants (Bartell, 2007). Likewise, other theoretical 

13 ecosystem models are being modified for use in assessing ecological risks from pollutant exposures 

14 (Bartell, 2007). 

15 Some ecosystems, and some aspects of particular ecosystems, are less vulnerable to long-term 

16 consequences of pollutant exposure. Other ecosystems may be profoundly altered if a single attribute is

17 affected. Thus, spatial and temporal definitions of ecosystem structure and function become an essential 

18 factor in defining impacted ecosystem services and critical loads of particular pollutants, either as single 

19 pollutants or in combination with other stressors. Both ecosystem services (Section 7.1.2) and critical 

20 loads (Section 7.1.3) serve as benchmarks or measures of the impacts of pollutants on ecosystems. 

7.1.2. Ecosystem Services

21 Ecosystem structure and function may be translated into ecosystem services (Daily, 1997). 

22 Ecosystem services are the benefits people obtain from ecosystems (Millennium Ecosystem Assessment, 

23 2003). Ecosystem services are defined as the varied and numerous ways that ecosystems are important to 

24 human welfare and how they provide many goods and services that are of vital importance for the 

25 functioning of the biosphere. This concept has gained recent interest and support because it recognizes 

26 that ecosystems are valuable to humans, and are important in ways that are not generally appreciated 

27 (Daily, 1997). Ecosystem services also provide a context for assessing the collective effects of human 

28 actions on a broad range of the goods and services upon which humans rely. 

29 In general, both ecosystem structure and function play essential roles in providing goods and 

30 services. Ecosystem processes provide diverse benefits including absorption and breakdown of pollutants, 

31 cycling of nutrients, binding of soil, degradation of organic waste, maintenance of a balance of gases in 

32 the air, regulation of radiation balance and climate, and fixation of solar energy (Daily, 1997; Westman, 

33 1977; WRI, 2000). These ecological benefits, in turn, provide economic benefits and values to society 
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1 (Costanza et al., 1997; Pimentel et al., 1997). Goods such as food crops, timber, livestock, fish and clean 

2 drinking water have market value. The values of ecosystem services such as flood control, wildlife 

3 habitat, cycling of nutrients and removal of air pollutants are more difficult to measure (Goulder & 

4 Kennedy, 1997). 

5 Particular concern has developed within the past decade regarding the consequences of decreasing 

6 biological diversity (Ayensu et al., 1999; Chapin et al., 1998; Hooper & Vitousek, 1997; Tilman, 2000; 

7 Wall, 1999). Human activities that decrease biodiversity also alter the complexity and stability of 

8 ecosystems and change ecological processes. In response, ecosystem structure, composition and function 

9 can be affected (Chapin et al., 1998; Daily & Ehrlich, 1999; Levlin, 1998; Peterson et al., 1998; Pimm, 

10 1984; Tilman, 1996; Tilman & Downing, 1994; Wall, 1999). Biodiversity is an important consideration at 

11 all levels of biological organization, including species, individuals, populations, and ecosystems. Human12 induced changes in biotic diversity and alterations in the structure and functioning of ecosystems are two 

13 of the most dramatic ecological trends of the past century (U.S. EPA, 2004; Vitousek et al., 1997). 

14 Hassan (2005) identified four broad categories of ecosystem services: 

15  Supporting services are necessary for the production of all other ecosystem services. Some 

16 examples include biomass production, production of atmospheric O2, soil formation and 

17 retention, nutrient cycling, water cycling and provisioning of habitat. Biodiversity is a 

18 supporting service in that it is increasingly recognized to sustain many of the goods and 

19 services that humans enjoy from ecosystems. These supporting services provide a basis for an 

20 additional three higher-level categories of services. 

21  Provisioning services such as products (Gitay et al., 2001) i.e., food (including game meat, 

22 roots, seeds, nuts, and other fruit, spices, fodder), fiber (including wood, textiles) and 

23 medicinal and cosmetic products.

24  Regulating services that are of paramount importance for human society such as 

25 (1) carbon sequestration, (2) climate and water regulation, (3) protection from natural hazards 

26 such as floods, avalanches, or rock-fall (4) water and air purification, and (5) disease and pest 

27 regulation.

28  Cultural services that satisfy human spiritual and aesthetic appreciation of ecosystems and 

29 their components. 

7.1.3. Critical Loads as an Organizing Principle for 

Ecological Effects of Atmospheric Deposition

30 Critical loads were first defined for regulating emissions of sulfur and nitrogen oxides, but have 

31 since been applied to exposure of other pollutants, including metals (Adams & Chapman, 2007). A critical 
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1 load is defined as, “a quantitative estimate of an exposure to one or more pollutants below which 

2 significant harmful effects on specified sensitive elements of the environment do not occur according to 

3 present knowledge” (Nilsson & Grennfelt, 1988). Because critical loads for Pb differ by ecosystem 

4 (aquatic-water; aquatic-sediment; terrestrial), differ by environmental chemistry properties that impact 

5 bioavailability, differ by species, and differ by endpoint of concern, they can be used as an organizing 

6 principle for linking atmospheric deposition with ecological impairment at multiple spatial scales. There 

7 are two aspects to consider: (1) the critical load at a steady state in the environment (i.e., how much input 

8 is required to balance the rate of output), and (2) the time required to reach the critical load (i.e., the lag 

9 time between onset of exposure and induction of measurable effects). This is particularly true for 

10 terrestrial ecosystems where changes in soil geochemistry, as a result of either changing land use or 

11 ecological succession, may significantly alter the amount of sequestration of Pb, thus changing its 

12 bioavailability and critical load if based on total metal. Ideally, therefore, critical loads for metals should 

13 be defined on the basis of bioavailable metal rather than total metal. This approach is being used in 

14 aquatic systems through the application of the biotic ligand model (BLM) (Di Toro et al., 2001; Di Toro et 

15 al., 2005), but is proving to be more difficult for modeling terrestrial systems. 

16 For aquatic systems, a dynamic equilibrium exists between the surface water, the water column,

17 and the sediment compartments (which must be defined when determining the critical load for each 

18 compartment). Although the sediment generally acts as a sink for pollutants in the water column, 

19 especially metals in particle form or as insoluble metal complexes, there may be some re-entrainment 

20 from the sediments into the water column. This sediment-water interface may change the solubility or 

21 bioavailability of the metal, thereby altering the critical load for the water column, particularly if 

22 expressed in the form of total metal. Note, however, that while sedimentation processes may change the 

23 time to steady state, they will not affect the ultimate critical load once steady state is achieved.

24 The following pieces of information are required to calculate a critical load, each of which is 

25 discussed in more detail in the subsequent sections of this chapter:

26  Ecosystem at risk;

27  Receptors of concern (plants, animals, etc.);

28  Endpoints of concern (organism, population or community responses, changes in ecosystem 

29 services or functions); 

30  Dose (concentration) - response relationships and threshold levels of effects;

31  Bioavailability and bioaccumulation rates; 

32  Naturally occurring (background) Pb (or other metal) concentrations; and

33  Biogeochemical modifiers of exposure.
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1 As stated in the 2008 ISA for Oxides of Nitrogen and Sulfur-Ecological Criteria there is no single 

2 “definitive” critical load for a pollutant, partly because critical load estimates reflect the current state-of3 knowledge and policy priorities, and also because of local or regional differences among ecosystems 

4 (U.S. EPA, 2008). Changes in scientific understanding may include, for example, expanded information 

5 about dose-response relationships, better understanding of bioavailability factors, and improved 

6 quantitative models for effects predictions. Changes in policy may include new mandates for resource 

7 protection, inclusion of perceived new threats that may exacerbate the effects of the pollutant of concern 

8 (e.g., climate change), and a better understanding of the value of ecosystem services.

9 In the short term, metal emissions generally have greater adverse effects on biota in aquatic 

10 systems than in terrestrial systems because metals are more readily immobilized in soils than in sediment. 

11 However, over the longer term, terrestrial systems may be more affected particularly by those metals with 

12 a long soil residence time, such as Pb. Thus, for a particular locale, either the terrestrial or the aquatic 

13 ecosystem at that site may have the lower critical load. Given the heterogeneity of ecosystems affected by 

14 Pb, and the differences in expectations for ecosystem services attached to different land uses, it is 

15 expected that there will be a range of critical load values for Pb for soils and waters within the U.S. 

7.1.4. Ecosystem Exposure, Lag Time and Re-entrainment 

of Historically Deposited Lead

16 Ecosystem exposure from atmospheric emissions of Pb depends upon the amount of Pb deposited 

17 per unit time. Ecosystem response will also depend upon the form in which the Pb is deposited, the areal 

18 extent of such deposition, and the modifying factors listed in the previous section. However, there is 

19 frequently a lag time between when metals are emitted and when an effect is seen, particularly in 

20 terrestrial ecosystems and, to a lesser extent, in aquatic sediments; water exposures result in more 

21 immediate system responses. This is because the buffering capacity of soils and sediments permits Pb to 

22 become sequestered into organic matter, making it less available for uptake by organisms. The lag time 

23 from start of emissions to achieving a critical load can be calculated as the time to reach steady state from

24 the time when the Pb was initially added to the system. Excluding erosion processes, the time required to 

achieve 95% of steady state is about 4 half-lives (t1/2)

1 25 (Smolders et al., 2007). Conversely, once 

26 emissions cease, the same amount of time is required to reduce metal concentrations to background 

27 levels.

28 Time to steady state for metals in soils is dependent upon rates of erosion, uptake by plants, and 

29 leaching or drainage from soils. Ignoring erosion, half-life of metals can be predicted (Smolders et al., 

30 2007) for a soil as:

 

1 Time required to reduce the initial concentration by 50% if metal input is zero.
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Kd

R

y TF

d

t

ρ

× +

× × = 0.69 10000

1/ 2

Equation 7-1

1 where:

2 d is the soil depth in meters (m)

3 y is the annual crop yield (tons/ha·yr)

4 TF is the ratio of the metal concentration in plant to that in soil

R is the net drainage loss out of the soil depth of concern (m3 5 /ha·yr)

ρ is the bulk density of soil (kg(dry weight) 6 /L)

7 Kd is the ratio of the metal concentration in soil to that in soil pore solution (L/kg)

8 Metals removed by crops (or plants in general) comprise a very small fraction of the total soil metal 

9 and can be ignored for the purpose of estimating time to steady state. Thus, equation 7-1 is 

10 simplified to:

Kd

R

d

t

ρ

0.69 10000

1/ 2

× × =

Equation 7-2

11 and becomes a function of soil depth, the amount of rainfall, soil density, and soil properties that 

12 affect Kd. Pb has a relatively long time to steady state compared to other metals, as shown in 

13 Table 7-1.

Table 7-1. Comparison among several metals: Time to achieve 95% of steady state metal 

concentration in soil; example in a temperate system

Metal Loading rate (g/Ha/Yr) Kd (L/Kg) Time (yr)

Se 100 0.3 1.3

Cu 100 480a 1,860a

Cd 100 690a 2,670a

Pb 100 19,000a 73,300a

Cr 100 16,700a 64,400a

a

Mean Kd (ratio of total metal concentrations in soils to that in soil pore water); and Time to achieve 95% of steady-state concentration in soil. (49 Dutch soils) (de Groot et al., 

1998). 

Note: Based on a soil depth of 23 cm, a rain infiltration rate of 3,000 m3

/ha·yr, and the assumption that background was zero at the start of loading.

Source: Smolders, Fairborother et al. (2007)

14 In aquatic systems, t1/2 for Pb in the water column depends on the ratio of the magnitudes of the 

15 fluxes coming from and going into the sediment, the ratio of the depths of the water column and sediment, 
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1 and the sediment half-life. Sediment t1/2 is dependent upon the particulate and dissolved fractions and is 

2 calculated as for soils (Equation 7-2). 

3 Re-entrainment of Pb particles via windblown dust from surface soils or dry sediments may occur. 

4 Amount and distance of re-entrained particles and deposition rates are dependent upon wind velocity and 

5 frequency; size, density, shape, and roughness of the particle; soil or sediment moisture; and terrain 

6 features including openness (including amount of vegetation), aspect relative to wind direction, and 

surface roughness. Resuspension is defined in terms of a resuspension factor, K, with units of m-1 7 , or a 

resuspension rate (Λ), with units of sec-1 8 (Equation 7-3). The resuspension rate, Λ, is the fraction of a 

9 surface contaminant that is released per time and is defined by:

C

R Λ =

Equation 7-3

10 where:

R is the upward resuspension flux (μg/m2 11 /sec)

C is the soil (or dry sediment) Pb concentration (μg/m2 12 )

13 Such emissions may have local impacts, but are not likely to have long-range effects, as particles 

14 generally remain low to the ground and are not lifted into the upper atmosphere. Although re15 entrainment may alter the particle size distribution in a local area, it generally does not alter the 

16 bioavailable fraction, and deposited particles will be subject to the same biogeochemical forces 

17 affecting bioavailability. Therefore, exposure via re-entrainment should be considered additive to 

18 exposure from atmospheric particulate deposition in terrestrial and aquatic ecosystems. 

7.2. Terrestrial Ecosystem Effects

7.2.1. Introduction to Terrestrial Ecosystem Effects

19 Numerous studies of the effects of Pb on components of terrestrial systems were reviewed in the 

20 2006 Pb AQCD. The literature on terrestrial ecosystem effects of Pb, published since the 2006 Pb AQCD,

21 is considered with brief summaries from the AQCD where relevant. Section 7.2 is organized to consider 

22 uptake of Pb and effects at the species level, followed by community and ecosystem level effects. Soil 

23 biogeochemistry of Pb in terrestrial systems is reviewed in Section 7.2.2. Section 7.2.3. considers the 

24 bioavailability and uptake of Pb by plants, invertebrates, and wildlife in terrestrial systems. Biological 

25 effects of Pb on terrestrial ecosystem components including plants and lichen, invertebrates, and 

26 vertebrates (Section 7.2.4) are followed by data on exposure and response of terrestrial species (Section 

27 7.2.5). Effects of Pb at the ecosystem scale are discussed in Section 7.2.6. Section 7.2 concludes with a 
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1 discussion of critical loads in terrestrial systems (Section 7.2.7), characterization of sensitivity and 

2 vulnerability of ecosystem components (Section 7.2.8), and effects on ecosystem services (Section 7.2.9). 

7.2.2. Soil Biogeochemistry and Chemical Effects

3 According to data presented in the 2006 Pb AQCD, the fraction of soil metal that is directly 

4 available to plants is the fraction found in soil pore water, even though the concentration of metals in pore 

5 water is small relative to bulk soil concentration. The amount of Pb dissolved in soil solution is controlled 

6 by at least six variables: (1) solubility equilibria; (2) adsorption-desorption relationship of total Pb with 

7 inorganic compounds (e.g., oxides of Al, Fe, Si, Mn; clay minerals); (3) adsorption-desorption reactions 

8 of dissolved Pb phases on soil organic matter; (4) pH; (5) cation exchange capacity (CEC); and (6) aging. 

9 Adsorption-desorption of Pb to soil solid phases is largely controlled by total metal loading. Therefore, 

10 areas with high Pb deposition will exhibit a lower fraction of total Pb partitioned to inorganic and organic 

11 matter. Decreasing soil pH, CEC, and organic matter have been strongly correlated to increases in the 

12 concentration of dissolved Pb species. Aging of metals in soils results in decreased amounts of labile 

13 metal as the Pb becomes incorporated into the soil solid phase (McLaughlin et al., 2010). Data from 

14 recent studies have further defined the impact of pH, CEC, organic matter (OM), and aging on Pb 

15 mobilization and subsequent bioavailability in soils.

7.2.2.1. pH and CEC 

16 Models of metal bioavailability calibrated from 500+ soil toxicity tests on plants, invertebrates, and 

17 microbial communities indicated that soil pH and CEC are the most important factors governing metal 

18 solubility and toxicity (Smolders et al., 2009). The variability of derived EC50 values was most closely 

19 associated with CEC. Smolders et al. (2007) determined that 12 to 18 months of artificial aging of soils 

20 amended with metal decreased the soluble metal fraction by about one order of magnitude. 

21 Miretzky et al. (2007) also showed that the concentration of mobile Pb was increased in acidic 

22 soils, and discovered that Pb adsorption to sandy loam clay was a function of both (1) Fe and Mn oxide 

23 interactions; and (2) the formation of weak electrostatic bonds with charged soil surfaces. Similarly, the 

24 mobility of smelter-produced metals in forest soils was found to be greater than in adjacent agricultural 

25 lands (Douay et al., 2009). The higher solubility was caused by the decreased soil pH of the forest 

26 environments. Further, decreasing the soil pH via simulated acid rain events increased naturally occurring 

27 Pb bioavailability in field tests (X. Hu et al., 2009).

28 A sequential extraction procedure was employed by Ettler et al. (2005) to determine the relative 

29 bioavailability of different Pb fractions present in soils collected from a mining and smelting area in the 

30 Czech Republic. Five Pb fraction categories were identified: (Fraction A) exchangeable; (Fraction B) acid 

31 extractable (bound to carbonates); (Fraction C) reducible (bound to Fe and Mn oxides); (Fraction D) 
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1 oxidizable (complexed with organic carbon); and (Fraction E) residual (silicates). Tilled agricultural soils 

2 were found to have decreased Pb, likely as a result of repeated cultivation, with the majority of Pb 

3 represented as the reducible Fraction C. Pb concentration in undisturbed forest soils, however, was largely 

4 present as the exchangeable fraction (A), weakly bound to soil OM. 

7.2.2.2. Organic Matter 

5 Organic matter (OM) decreases bioavailability of Pb, but as it is turned over and broken down, 

6 pedogenic minerals become more important in Pb sequestration (Schroth et al., 2008). Shaheen and 

7 Tsadilas (2009) noted that soils with higher clay content, OM, total calcium carbonate equivalent, and 

8 total free sesquioxides also exhibited higher total Pb concentration, indicating that less Pb had been 

9 removed by resident plant species. Huang et al. (2008) examined the re-mobilization potential of Pb in 

10 forest soils, and determined that mobilization of total Pb was strongly associated with dissolved organic 

11 matter (DOM). Groenenberg et al. (2010) used a non-ideal competitive adsorption Donnan model to 

12 explain the variability of OM binding affinity and uncertainties associated with metal speciation. They 

13 found that natural variations in fulvic acid binding properties were the most important variable in 

14 predicting Pb speciation. Guo et al. (2006) determined that the -COOH and -OH groups associated with 

15 soil OM were important factors in Pb sequestration in soil, and Pb sorption was increased as pH was 

16 raised from 2 to 8. Because organic content increased the Pb sequestration efficiency of soils, OM content 

17 had an inhibitory effect on Pb uptake by woodlouse species Oniscus asellus and Porcellio scaber (Gál et 

18 al., 2008). Vermeulen et al. (2009) demonstrated that invertebrate bioaccumulation of Pb from 

19 contaminated soils was dependent on pH and OM, but that other unidentified habitat-specific differences 

20 also contributed. The relationship of bioaccumulation and soil concentration was modified by pH and

21 OM, and also by habitat type. Kobler et al. (2010) showed that the migration of atmosphere-deposited Pb 

22 in soil matrices was strongly influenced by soil type, indicating that certain soil types may retain Pb for 

23 longer periods of time. In humic forest soils, the highest Pb concentrations were measured in the humified 

24 bottom layer, whereas in soils characterized by well-drained substrate and limestone bedrock, Pb 

25 concentration decreased over time, likely as a result of water drainage and percolation. The authors 

26 theorized that the most significant Pb migration route was transportation of particulate-bound Pb along 

27 with precipitation-related flow through large soil pores.

28 A number of recent laboratory studies have further defined the relationship of soil biogeochemical 

29 characteristics and Pb uptake by plants. Dayton et al. (2006) established significant negative correlations 

30 between log-transformed Pb content of lettuce plants (Lactuca sativa), soil organic content, and CEC, and 

31 similar negative relationships were also confirmed for soil pH and amorphous Fe and Al oxide content. As 

32 part of a metal partitioning study, (Kalis et al., 2007) determined that not only did metal concentration in 

33 the soil solution decrease as pH increased, but pH-mediated metal adsorption at the root surface of Lolium 
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1 perenne determined root Pb concentration, with concentration in the shoot correlated with root 

2 concentration. Interestingly, Kalis et al.(2007) and Lock et al. (2006) also observed that the influx of Pb in 

3 the water-soluble fraction had an impact on soil pH. In addition, 1 μM humic acid decreased root Pb 

4 concentration in L. perenne plants grown in 0.1 and 1 μM Pb solution, likely as a result of Pb 

5 complexation and sequestration with the added OM (Kalis et al., 2006). 

7.2.2.3. Aging

6 Smolders et al. (2007) reviewed the effects of aging of Pb in soils on the toxicity of Pb to plants 

7 and soil invertebrates, with aging defined primarily as leaching following initial influx, but also as 

8 binding and complexation. In nearly half of the Pb soil studies reviewed, observed dose-response curves 

9 could not be established following soil leaching, indicating that aged soils likely contain less bioavailable 

10 Pb. The authors concluded that competitive binding between soil ligands and biotic ligands on soil roots 

11 or invertebrate guts can be used to model the relationship of observed availability and toxicity of metals 

12 in soils. Because this concept is the basis of the Biotic Ligand Model (BLM) (Section 7.3.3), the authors 

13 proposed a terrestrial BLM approach to estimate the risk of metals to terrestrial organisms. However, 

14 Antunes et al. (2006) noted that there were several key challenges involved in development of a terrestrial 

15 BLM applicable to plants, particularly the reliable measurement of free ion activities and ligand 

16 concentration in the rhizosphere, the identification of the organisms’ ligands associated with toxicity, and 

17 the possible need to incorporate kinetic dissolution of metal-ligand complexes as sources of free ion. 

18 Further, Pb in aged field soils has been observed to be less available for uptake into terrestrial organisms, 

19 likely as a result of increased sequestration within the soil particles (Antunes et al., 2006). Magrisso et al. 

20 (2009) used a bioluminescent strain of the bacterium Cupriavidus metallidurans to detect and quantify Pb 

21 bioavailability in soils collected adjacent to industrial and highway areas in Jerusalem, Israel, and in 

22 individual simulated soil components freshly spiked with Pb. The bacterium was genetically engineered 

23 to give off the bioluminescent reaction as a dose-dependent response, and was inoculated in soil slurries 

24 for three hours prior to response evaluation. Spiked soil components induced the bioluminescent 

25 response, and field-collected components did not. However, the comparability of the simulated soils and 

26 their Pb concentration with the field-collected samples was not entirely clear. Lock et al. (2006) compared 

27 the Pb toxicity to springtails (Folsomia candida) from both laboratory-spiked soils and field-collected Pb28 contaminated soils of similar Pb concentrations. Total Pb concentrations of 3,877 mg Pb/kg dry weight 

29 and higher always caused significant effects on F. candida reproduction in the spiked soils. In field soils, 

30 only the soil with the highest Pb concentration of 14,436 mg Pb/kg dry weight significantly affected 

31 reproduction. When expressed as soil pore-water concentrations, reproduction was never significantly 

32 affected at Pb concentrations of 0.5 mg/L, whereas reproduction was always significantly affected at Pb 

33 concentrations of 0.7 mg/L and higher, independent of the soil treatment. Leaching soils prior to use in 
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1 bioassays had only a slight effect on Pb toxicity to resident springtails, suggesting that among the 

2 processes that constitute aging of Pb in field soils, leaching is not particularly important with respect to 

3 bioavailability. 

4 Red-backed salamanders (Plethodon cinereus) exposed to Pb-amended soils (553, 1,700, 4,700, 

5 and 9,167 mg Pb/kg) exhibited lowered appetite and decreased white blood cell counts at the two highest 

6 concentrations, as compared to controls {B, 2010, 379076}. However, salamanders tolerated field7 collected, aged soils containing Pb concentration of up to 16,967 mg Pb/kg with no significant deleterious 

8 effects.

9 In summary, studies published during the past 5 years continue to substantiate the important role 

10 that soil geochemistry plays in sequestration or release of Pb. Soil pH and CEC have long been known to 

11 be the primary controlling factors for amount of bioavailable Pb in soils, and a recent review of more than 

12 500 studies corroborates these findings (Smolders et al., 2009). Fe and Mn oxides are now known to also 

13 play an important role in Pb sequestration in soils. Pb binds to OM, although relatively weakly, and as the 

14 OM is broken down the Pb may be released into soil solution. Leaching of metal through soil pores may 

15 be the primary route for loss of bioavailable soil Pb; OM may reduce leaching and thus appear to be 

16 associated with Pb sequestration. Aging of Pb in soils through incorporation of the metal into the 

17 particulate solid phase of the soil results in long term binding of the metal and reduced bioavailability of 

18 Pb to plants and soil organisms. 

7.2.3. Bioavailability in Terrestrial Systems

19 Bioavailability was defined in the 2006 Pb AQCD as “the proportion of a toxin that passes a 

20 physiological membrane (the plasma membrane in plants or the gut wall in animals) and reaches a target 

21 receptor (cytosol or blood).” In 2007, EPA took cases of bioactive adsorption into consideration and 

22 revised the definition of bioavailability as “the extent to which bioaccessible metals absorb onto, or into, 

23 and across biological membranes of organisms, expressed as a fraction of the total amount of metal the 

24 organism is proximately exposed to (at the sorption surface) during a given time and under defined 

25 conditions” (Fairbrother et al., 2007). Characteristics of the toxicant that affect bioavailability are: (1) 

26 chemical form or species; (2) particle size; (3) lability; and (4) source. New information on sources of Pb 

27 in terrestrial ecosystems, and their influence on subsequent bioavailability, was reviewed in Chapter 3, 

28 while new information on the influence of soil biogeochemistry on speciation and chemical lability was 

29 presented in Section 7.2.2. This section summarizes new literature on uptake and subsequent presence of 

30 Pb in tissues. Bioaccumulation factors (BAF’s) (i.e., the ratio of Pb concentrations in tissues of terrestrial 

31 biota to concentrations in soil or in food items) reported in this section are summarized in Table 7-2. The 

32 2006 Pb AQCD extensively reviewed the methods available for quantitative determination of the 
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1 mobility, distribution, uptake, and fluxes of atmospherically delivered Pb in ecosystems, and they are not 

2 reviewed in this section.

7.2.3.1. Plants

3 The 2006 Pb AQCD noted that terrestrial plants accumulate atmospheric Pb primarily via two 

4 routes: direct stomatal uptake into foliage, and incorporation of atmospherically deposited Pb from soil 

5 into root tissue, followed by variable translocation to other tissues. Foliar Pb may include both 

6 incorporated Pb (i.e., from atmospheric gases or particles) and surficial particulate Pb deposition. 

7 Although the plant may eventually absorb the surficial component, its main importance is its likely 

8 contribution to the exposure of plant consumers. This section will first review recent studies on uptake of 

9 Pb by plants through foliar and soil routes, and their relative contribution, followed by the consideration 

10 of translocation of Pb from roots to shoots, including a discussion of variability in translocation among 

11 species.

Leaf and Root Uptake

12 Field studies carried out in the vicinity of Pb smelters have determined the relative importance of 

13 direct foliar uptake and root uptake of atmospheric Pb deposited in soils. Hu and Ding (2009) analyzed 

14 ratios of Pb isotopes in the shoots of commonly grown vegetables and in soil at three distances from a 

15 point source (0.1, 0.2, 5.0 km). Pb isotope ratios in plants and soil were different at two of those locations, 

16 leading the authors to the conclusion that airborne Pb was being assimilated via direct leaf uptake. Soil Pb 

17 concentration in the rhizosphere at the three sites ranged between 287 and 379 mg Pb/kg (Site I), 155 and 

18 159 mg Pb/kg (Site II), and 58 and 79 mg Pb/kg (Site III, selected as the control site). The median shoot 

19 and root Pb concentrations at each site were 36 and 47 mg Pb/kg, 176 and 97 mg Pb/kg, and 1.3 and 7 mg

20 Pb/kg, respectively, resulting in shoot:root Pb ratios exceeding 1.0 in Site I (for Malabar spinach [Basella 

21 alba], ratio = 1.6, and amaranth [Amaranthus spinosus], ratio = 1.1), and in Site II (for the weeds 

22 Taraxacum mongolicum, ratio = 1.9, and Rostellaria procumbens, ratio = 1.7). However, the two species 

23 studied at Site II were not studied at Site I or Site III. In the control site (Site III), no plant was found with 

24 a Pb shoot:root ratio greater than 1.0. Hu and Ding (2009) concluded that metal accumulation was greater 

25 in shoot than in root tissue, which suggested both high atmospheric Pb concentration and direct stomatal 

26 uptake into the shoot tissue. 

27 Cui et al. (2007) studied seven weed species growing in the vicinity of an old smelter (average soil 

28 Pb concentration of 4,020 mg Pb/kg) in Liaoning, China, to measure Pb accumulation rates in roots and 

29 shoots. Cutleaf groundcherry (Physalis angulata) accumulated the most Pb, with root and shoot 

30 concentration of 527 and 331 mg Pb/kg, respectively, and velvetleaf (Abutilon theophrasti) was the 

31 poorest absorber of Pb (root and shoot concentration of 39 and 61 mg Pb/kg, respectively). In all cases, 
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1 weed species near the smelter accumulated more Pb than plants from non-polluted environments (5 mg

2 Pb/kg), indicating that aerially deposited Pb produced by smelting is bioavailable to plants. However, the 

3 ratio of root:shoot Pb concentration varied by species, and the authors presented no data to differentiate 

4 Pb taken up from soil from Pb incorporated via foliar uptake. 

5 Chrastny et al. (2010) characterized the Pb contamination of an agricultural soil in the vicinity of a 

6 shooting range. Pb was predominantly in the form of PbO and PbCO3, and Pb was taken up by plants 

7 through both atmospheric deposition onto the plant and by root uptake.

8 Because of their long life spans, trees can provide essential information regarding the sources of 

9 bioavailable Pb. A Scots pine forest in northern Sweden was found to incorporate atmospherically derived 

10 Pb pollution directly from ambient air, accumulating this Pb in bark, needles, and shoots (Klaminder et 

11 al., 2005). Nearly 50% of total tree uptake was determined to be from direct adsorption from the 

12 atmosphere. Further, Aznar et al. (2009) found that the Pb content of black spruce (Picea mariana) 

13 needles collected along a metal contamination gradient emanating from a Canadian smelter in 

14 Murdochville, Quebec, showed a significant decrease in Pb concentration with increasing distance from 

15 the smelter. Interestingly, older needles were determined to accumulate larger quantities of Pb than 

16 younger ones, which the authors attributed to “incidental processing” of atmospheric Pb. Foliar damage 

17 and growth reduction were also observed in the trees (Aznar, Richer-Lafleche, et al., 2009). They were 

18 significantly correlated with Pb concentration in the litter layer, where Pb comes from atmospheric 

19 deposition and closely reflects it. In addition, there was no correlation between diminished tree growth 

20 and Pb concentration in the deeper mineral soil layers, strongly suggesting that only current atmospheric 

21 Pb was adversely affecting trees (Aznar, Richer-Laflèchea, et al., 2009). Similarly, Kuang et al. (2007)

22 noted that the Pb concentration in the inner bark of Pinus massoniana trees growing adjacent to a Pb-Zn 

23 smelter in the Guangdong province of China was much higher (1.87 mg Pb/kg dry weight) than in 

24 reference-area trees. Because concentration in the inner bark was strongly correlated with concentration in 

25 the outer bark, they concluded that the origin of the Pb was atmospheric.

26 Dendrochronology (tree ring analysis) has become an increasingly important tool for measuring the 

27 response of trees to Pb exposure (Watmough, 1999). The advent of laser ablation inductively coupled 

28 plasma mass spectrometry has made measurement of Pb concentration in individual tree rings possible 

29 (Watmough, 1999; Witte et al., 2004). This allows for close analysis of the timing of Pb uptake relative to 

30 smelter activity and/or changes in soil chemistry. For example, Aznar et al. (2008) measured Pb 

31 concentration in black spruce tree rings to determine the extent and timing of atmospheric deposition near 

32 the Murdochville smelter. Variability in tree-ring Pb content seemed to indicate that trees accumulated 

33 and sequestered atmospheric Pb in close correlation with the rates of smelter emission, but that 

34 sequestration lagged about 15 years behind exposure. However, the ability to determine time of uptake 

35 from the location in growth rings is weakened in species that transfer Pb readily from outer bark to inner 

36 bark. Cutter and Guyette (1993) identified species with minimal radial translocation from among a large 
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1 number of tree species, and recommended the following temperate zone North American species as 

2 suitable for metal dendrochronology studies: white oak (Quercus alba), post oak (Q. stellata), eastern red 

3 cedar (Juniperus virginiana), old-growth Douglas fir (Pseudotsuga menziesii), and big sagebrush 

4 (Artemisia tridentata). In addition, species such as bristlecone pine (Pinus aristata), old-growth 

5 redwood (Sequoia sempervirens), and giant sequoia (S. gigantea) were deemed suitable for local 

6 purposes. Patrick and Farmer (2006) determined that European sycamore (Acer pseudoplatanus) are not 

7 suitable for this type of dendrochronological analysis because of the formation of multiple annual rings. 

8 Pb in sapwood and heartwood is more likely a result of soil uptake than of direct atmospheric 

9 exposure (Guyette et al., 1991). Differentiation of geogenic soil Pb in tree tissue from Pb that originated 

10 in the atmosphere requires measurement of stable Pb isotope ratios (L. Patrick, 2006). Tree bark samples 

11 collected from several areas of the Czech Republic were subjected to stable Pb isotope analysis to 

12 determine the source and uptake of atmospheric Pb (Conkova & Kubiznakova, 2008). Results indicated 

13 that beech bark is a more efficient accumulator of atmospheric Pb than spruce bark. A decrease in the 

206Pb/207 14 Pb ratio was measured in bark and attributed to increased usage of leaded gasoline between 1955 

and 1990; an increased 206Pb/207 15 Pb ratio was ascribed to coal combustion (Conkova & Kubiznakova, 

2008). Similarly, Savard et al. (2006) compared isotope ratios of 206Pb/207Pb and 208Pb/206 16 Pb in tree rings 

17 from spruce trees sampled at a control site near Hudson Bay, with those sampled near the Horne smelter 

18 active since 1928, in Rouyn-Noranda, Canada. The concentration of total Pb showed a major increase in 

1944 and a corresponding decrease of the 206Pb/207 19 Pb ratios, suggesting that the smelter was responsible 

20 for the increased Pb uptake (Savard et al., 2006). The authors suggested that the apparent delay of 14 

21 years may have been attributable to the residence time of metals in airborne particles the buffering effect 

22 of the soils and, to a lesser extent, mobility of heavy metals in tree stems. Furthermore, through the use of 

23 the two different isotope ratios, Savard et al. (2006) were able to differentiate three types of Pb in tree 

24 rings: natural (derived from the mineral soil horizons), industrial (from coal burning urban pollution), and 

25 mining (typical of the volcanogenic massive sulfide ore deposits treated at the Horne smelter). 

26 Devall et al. (2006) measured Pb uptake by bald-cypress trees (Taxodium distichum) growing in a 

27 swamp near a petroleum refinery and along a bank containing Pb-contaminated dredge spoils. They 

28 measured Pb in tree cores and showed greater uptake of Pb by trees in the swamp than by trees growing 

29 on the dredge spoil bank, attributing the difference to exposure source (refinery versus dredge spoils) and 

30 differences in soil chemistry between the swamp and the dredge spoil bank (Devall et al., 2006). 

31 Similarly, Gebologlu et al. (2005) found no correlation between proximity to roadway and accumulated 

32 Pb in tomato and bean plants at sites adjacent to two state roads in Turkey (average Pb concentration 5.4 

33 and 6.0 mg Pb/kg), indicating that uptake may be influenced by multiple factors, including wind 

34 direction, geography, and soil chemistry. Average Pb levels in leaves were 0.6 and 0.5 mg Pb/kg for 

35 tomato and bean plants, respectively, while fruit concentration averaged 0.4 mg Pb/kg for both species. 

36 Conversely, if foliar contamination is due primarily to dust deposition, distance from a source such as a 
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1 road may be easily correlated with Pb concentration on the plants. For example, Ai-Khlaifat and Al2 Khashman (2007) collected unwashed date palm (Phoenix dactylifera) leaves at 3-m trunk height from 

3 trees in Jordan to assess the extent of Pb contamination from the city of Aqaba. Whereas relatively low 

4 levels of Pb were detected in leaves collected at background sites (41 mg Pb/kg), leaves collected 

5 adjacent to highway sites exhibited the highest levels of Pb (177 mg Pb/kg). The authors determined that 

6 Pb levels in date palm leaves correlated with industrial and human activities (e.g., traffic density) (Ai7 Khlaifat & Al-Khashman, 2007). However, decreases in tissue Pb concentration with increasing distance 

8 from point sources can also follow from decreasing Pb in soil. Bindler et al. (2008) used Pb isotopes to 

9 assess the relative importance of pollutant Pb versus natural Pb for plant uptake and cycling in Swedish 

10 forested soils. The Pb isotopic composition of needles/leaves and stemwood of different tree species and 

11 ground-cover plants indicated that the majority of Pb present in these plant components was derived from 

12 the atmosphere, either through aerial interception or actual uptake through the roots. For the ground-cover 

plants and the needles/leaves, the 206Pb/207 13 Pb isotopic ratios (1.12 to 1.20) showed that the majority of Pb 

was of anthropogenic origin. Stemwood and roots have higher 206Pb/207 14 Pb ratio values (1.12 to 1.30) 

15 which showed the incorporation of some natural Pb as well as anthropogenic Pb. For pine trees, the 

16 isotopic ratio decreased between the roots and the apical stemwood suggesting that much of the uptake of 

17 Pb by trees is via aerial exposure. Overall, it was estimated that 60-80% of the Pb in boreal forest 

18 vegetation originated from pollution; the Pb concentrations were, however, quite low – not higher than 1 

19 mg Pb/kg plant material, and usually in the range of 0.01-0.1 mg Pb/kg plant material (while soils had a 

20 range of 5 to 10 mg Pb/kg in the mineral horizons and 50 to 150 mg Pb/kg in the O horizons). Overall, the 

21 forest vegetation recycles very little of the Pb present in soils (and thus does not play a direct role in the 

22 Pb biogeochemical cycle in boreal forest soils).

Translocation and Sequestration of Lead in Plants

23 Although Pb is not an essential metal, it is taken up from soils through the symplastic route, the 

24 same active ion transport mechanism used by plants to take up water and nutrients and move them across 

25 root cell membranes (U.S. EPA, 2006). As with all nutrients, only the proportion of a metal present in soil 

26 pore water is directly available for uptake by plants. In addition, soil-to-plant transfer factors in soils 

27 enriched with Pb have been found to better correlate with bioavailable Pb soil concentration, defined as 

28 DTPA-extractable Pb, than with total Pb concentration (U.S. EPA, 2006).

29 The 2006 Pb AQCD stated that most of the Pb absorbed from soil remains bound in plant root 

30 tissues either because (1) Pb may be deposited within root cell wall material, or (2) Pb may be 

31 sequestered within root cell organelles. Sequestration of Pb may be a protective mechanism for the plant. 

32 Recent findings have been consistent with this hypothesis: Han et al. (2008) observed Pb deposits in the 

33 cell walls and cytoplasm of malformed cells of Iris lactea exposed to 0 to 10 mM Pb for 28 days. They 
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1 hypothesized that preferential sequestration of Pb in a few cells, which results in damage to those cells, 

2 helps in maintaining normal overall plant activities through the sacrifice of a small number of active cells. 

3 Similarly, macroscopic analysis of the roots of broad bean (Vicia faba) cultivated in mine tailings 

4 (average Pb concentration of 7,772 mg/kg) by Probst et al. (2009) revealed dark ultrastructural 

5 abnormalities that were demonstrated to be metal-rich particles located in or on root cell walls. It is 

6 unclear whether the presence of these structures had any effect on overall plant health.

7 Clark et al. (2006) investigated Pb bioavailability in garden soils in Roxbury and Dorchester, MA. 

8 The sources of Pb were considered to be Pb from paints and from leaded gasoline additives, with 40 to

9 80% coming from paint. The average Pb concentration in foliar tissue of bean plants was 14 ± 5 mg Pb/kg 

10 while the concentration in the bean pod was only 20.6 mg Pb/kg. For mustard plants, there was a linear 

relationship (R2 11 =0.85) between Pb concentration in plant tissues and Pb concentration in the soil (both for 

12 plants grown in situ and those grown under greenhouse conditions).

13 Murray et al. (2009) investigated the uptake and accumulation of Pb in several vegetable species 

14 (carrot [Daucus carota], radish [Raphanus sativus], lettuce [Lactuca sativa], soybean [Glycine max], and 

15 wheat [Triticum aestivum]) from metal-contaminated soils, containing 10 to 40 mg Pb/kg and 

16 demonstrated that most Pb remained in the roots. No Pb was measured in the above-ground edible 

17 soybean and wheat tissues, while carrots, the most efficient accumulator of Pb, contained a maximum Pb 

18 tissue concentration of 12 mg/kg dry mass. Similarly, (Cho et al., 2009) showed that green onion (Allium 

19 fistulosum) plants also take up little Pb when planted in soil spiked with Pb nitrate. No plant tissues 

20 contained a Pb concentration greater than 24 mg Pb/kg when grown for 14 weeks in soils of up to 3,560 

21 mg Pb/kg, and the majority of bioavailable Pb was determined to be contained within the roots. Chinese 

22 spinach (Amaranthus dubius) also translocates very little Pb to stem and leaf tissue, and uptake from Pb23 containing soils (28 to 52 mg Pb/kg) is minimal (Mellem et al., 2009). Sonmez et al. (2008) reported that 

24 Pb accumulated by three weed species (Avena sterilis, Isatis tinctoria, Xanthium strumarium) grown in 

25 Pb-spiked soils was largely concentrated in the root tissues, and little was translocated to the shoots 

26 (Sonmez et al., 2008). 

27 Recent research has shown that Pb translocation to stem and leaf tissues does occur at significant 

28 rates in some species, including the legume Sesbania drummondii (Peralta-Videa et al., 2009) and

29 buckwheat (Fagopyrum esculentum) (Tamura et al., 2005). Wang et al. (2006) noted that Pb soil-to-plant 

30 transfer factors were higher for leafy vegetables (Chinese cabbage, pak-choi, and water spinach) than for 

31 the non-leafy vegetables tested (towel gourd, eggplant, and cowpea). Tamura et al. (2005) demonstrated 

32 that buckwheat is an efficient translocator of Pb. Buckwheat grown in Pb-containing soils collected from 

33 a shooting range site (average 1M HCl extractable Pb= 6,643 mg Pb/kg) preferentially accumulated Pb in 

34 leaves (8,000 mg Pb/kg) and shoots (4,200 mg Pb/kg), over root tissues (3,300 mg Pb/kg). Although plant 

35 growth was unaffected, this level of leaf and shoot accumulation is likely to have significant implications 

36 for exposure of herbivores. Similarly, Shaheen and Tsadilas (2009) reported that vegetables (pepper, okra, 
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1 and eggplant) grown in soils containing 24 to 30 mg Pb/kg total Pb were more likely to accumulate Pb in 

2 leaves (range: undetected to 25 mg Pb/kg) rather than in fruits (range: undetected to 19 mg Pb/kg); 

3 however, no significant correlation between soil Pb concentration and plant tissue Pb concentration could 

4 be established (Shaheen & Tsadilas, 2009).

5 There is broad variability in uptake and translocation among plant species, and interspecies 

6 variability has been shown to interact with other factors such as soil type. By studying multiple species in 

7 four Pb-Zn mining sites in Yunnan, China, Li et al. (2009) demonstrated not only significant differences 

8 in uptake and translocation among the species studied, but also modification of the effect on species by 

9 type of soil. Plants sampled represented nine species from four families—Caryophyllaceae, Compositae, 

10 Cruciferae, and Pteridaceae. Overall, soil Pb concentration averaged 3,772 mg Pb/kg dry weight, with the 

11 highest site average measured at the Minbingying site (5,330 mg Pb/kg), followed by Paomaping (2,409 

12 mg Pb/kg), Jinding (1,786 mg Pb/kg), and Qilinkeng (978 mg Pb/kg). The highest average shoot Pb 

13 concentration (3,142 mg Pb/kg) was detected in Stellaria vestita (Caryophyllaceae) collected at 

14 Paomaping, while shoot concentration of Sinopteris grevilloides (Pteridaceae) collected from 

15 Minbingying exhibited the lowest shoot Pb concentration (69 mg Pb/kg). A similar trend was detected in 

16 root tissues. S. vestita root collected from the Paomaping area contained the maximum Pb concentration 

17 measured (7,457 mg Pb/kg), while the minimum root Pb levels were measured in Picris hieraciodides

18 (Pteridaceae) tissues collected from Jinping. These results indicate significant interspecies differences in 

19 Pb uptake, as well as potential soil-specific differences in Pb bioavailability. S. vestita, in particular, was 

20 determined to be an efficient accumulator of Pb, with a maximum enrichment coefficient of 1.3. 

21 Significant correlations between soil Pb concentration and average shoot and root Pb levels were also 

22 established (Y. Li et al., 2009). Within plant species, the variability in uptake and translocation of Pb may 

23 extend to the varietal level. Antonious and Kochhar (2009) determined uptake of soil-associated Pb for 23 

24 unique genotypes from four species of pepper plants (Capsicum chinense, C. frutescens, C. baccatum, and 

25 C. annum). Soil Pb concentration averaged approximately 0.6 mg Pb/kg dry soil. No Pb was detected in 

26 the fruits of any of the 23 genotypes, except two out of seven genotypes of C. baccatum, which had 0.9 

27 and 0.8 mg Pb/kg dry weight Pb in fruit. 

28 Fungal species, as represented by mushrooms, accumulate Pb from soils to varying degrees. Based 

on the uptake of naturally occurring 210 29 Pb, Guillen et al. (2009) established that soil-associated Pb was 

bioavailable for uptake by mushrooms, and that the highest 210 30 Pb accumulation was observed in Fomes 

31 fomentarius mushrooms, followed by Lycoperdon perlatum, Boletus aereus, and Macrolepiota procera, 

32 indicating some species differences. Benbrahim et al. (2006) also showed species differences in uptake of 

33 Pb by wild edible mushrooms, although they found no significant correlations between Pb content of 

34 mushrooms and soil Pb concentration. Pb concentrations in mushroom carpophores ranged from 0.4 to 

35 2.7 mg Pb/kg from sites with soil concentrations ranging from 3.6 and 7.6 mg Pb/kg dry soil. 
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1 Recent studies substantiated findings from the 2006 Pb AQCD that plants store a large portion of 

2 Pb in root tissue. Pb soil-to-plant transfer factors are higher for leafy vegetables than for the non-leafy 

3 vegetables (G. Wang et al., 2006) and buckwheat has recently been shown to be an efficient translocator 

4 of Pb from soil to above-ground shoots (Tamura et al., 2005). However, there is broad variability in Pb 

5 uptake and translocation rates among plant species, and interspecies variability has been shown to interact 

6 with other factors such as soil type. Field studies carried out in the vicinity of Pb smelters (X. Hu et al., 

7 2009) show that Pb may accumulate in shoot tissue through direct stomatal uptake rather than by soil8 root-shoot translocation. Dendrochronology has become more advanced in recent years and is a useful 

9 tool for monitoring historical uptake of Pb into trees exposed to atmospheric or soil Pb. Trees accumulate 

10 and sequester atmospheric Pb in close correlation with the rate of smelter emissions, although one study 

11 indicated that sequestration can lag behind exposure from emissions by 15 years. Pb in the outer woody 

12 portion of the tree is more likely the result of direct atmospheric exposure, while Pb in sapwood is more 

13 likely a result of soil uptake. This difference provides an important tool for analyzing source 

14 apportionment of Pb accumulation in plants (Guyette et al., 1991).

7.2.3.2. Invertebrates

15 At the time of publication of the 2006 Pb AQCD, little information was available regarding the 

16 uptake of atmospheric Pb pollution (direct or deposited) by terrestrial invertebrate species. Consequently, 

17 few conclusions could be drawn concerning the Pb uptake rate of particular species although there was 

18 some evidence that dietary or habitat preferences may influence exposure and uptake. Recent literature 

19 indicates that invertebrates can accumulate Pb from consuming a Pb-contaminated diet and from exposure 

20 via soil, and that uptake and bioaccumulation of Pb by invertebrates is lower than that observed for other 

21 metals. 

Snails

22 Cantareus asperses snails exposed to dietary Pb at 3.3, 86, and 154 mg/kg of diet (spiked with Pb

23 sulfate) for up to 64 days were found to assimilate a significant proportion of Pb, and feeding rates were 

24 unaffected by the presence of the metal (Beeby & Richmond, 2010). While bioconcentration factors 

25 (BCF’s) for Cd were observed to increase over the 64-day study period, the rate of Pb assimilation 

26 remained consistent over time and no evidence for a regulatory mechanism for Pb was observed. The 

27 authors observed that, for additional Pb to be retained, snails would have to grow additional soft tissue. 

28 Helix aspersa snails rapidly accumulated Pb from contaminated soil (1,212 mg Pb/kg) and from eating 

29 contaminated lettuce (approximately 90 mg Pb/kg after 16 weeks’ growth on Pb-contaminated soil) 

30 during the first 2 weeks of exposure, at which point snail body burdens reached a plateau (Scheifler, De 

31 Vaufleury, et al., 2006). There were no observed effects of Pb exposure or accumulation on survival or 
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1 growth in C. asperses or H. aspersa. In another study (Ebenso & Ologhobo, 2009b), juvenile Achatina 

2 achatina snails confined in cages on former Pb-battery waste dump sites were found to accumulate Pb 

3 from both plant and soil sources. Soil Pb concentration averaged 20, 200, and 1,200 mg Pb/kg at the three 

4 main waste sites, while leaf tissues of radish (Raphanus sativus) grown at these sites averaged 7, 30, and 

5 68 mg Pb/kg dry weight, respectively. Although plant concentrations were low, they were correlated with 

6 elevated snail Pb tissue concentration. Pb concentration in snail tissues averaged 12, 91, and 468

7 mg Pb/kg, respectively, at the three sites, which the authors stipulated were above the maximum 

8 permissible concentration of Pb for human consumption of mollusks, mussels, and clams (1.5 μg Pb/g 

9 tissue). Pb concentration in snail tissues generally is much lower than that of the soil substrates upon 

10 which they were reared, but higher than in other soil-dwelling organisms. De Vaufleury et al. (2006)

11 exposed Helix aspera snails to standardized (1999 European International Organization for 

12 Standardization methodology [ISO 11267:1999]) artificial-substrate soils containing 13, 26, 39, or 52 mg

13 Pb/kg for 28 days without supplemental food. After the exposure period, snail foot tissue contained 

14 increased levels of Pb—1.9, 1.7, and 1.5 μg Pb/g dry weight versus concentration averaging 0.4 mg/kg in 

15 control organisms. Viscera also exhibited increased Pb levels at the two highest exposures, with measured 

16 tissue concentration of 1.2 and 1.1 mg Pb/kg, respectively, as compared with control tissue Pb levels of 

17 0.4 mg Pb/kg. However, there was no significant increase in snail-tissue Pb concentration when natural 

18 soil was used in place of ISO medium, and there was no relationship between soil Pb concentration and 

19 snail tissue concentration, strongly suggesting the presence of soil variables that modify bioavailability. 

20 Notten et al. (2008) investigated the origin of Pb pollution in soil, plants, and snails by means of Pb 

21 isotope ratios. They found that a substantial proportion of Pb in both plants and snails was from current 

22 atmospheric exposure.

Earthworms

23 Soil characteristics that interact with bioavailability of Pb may include biogeochemistry associated 

24 with different soil horizons, source of Pb, and proportion of soil:leaf litter. These have been studied 

25 principally with various species of earthworms. Bradham et al. (2006) examined the effect of soil 

26 chemical and physical properties on Pb bioavailability. Eisenia andrei earthworms were exposed to 21 

27 soils with varying physical properties that were freshly spiked with Pb to give a standard concentration of 

28 2,000 mg/kg dry weight. Both internal earthworm Pb concentration and mortality rates increased with 

29 decreasing pH and CEC although the apparent role of CEC may only have been due to its correlation with 

30 other soil characteristics. These data corroborate that Pb bioavailability and toxicity are increased in 

31 acidic soils and in soils with a low CEC (Section 7.2.2). This finding was confirmed by Gandois et al 

32 (2010), who determined that the free-metal-ion fraction of total Pb concentration in field-collected soils 

33 was largely predicted by pH and soil iron content.
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1 The role of soil profile and preferred depth was studied using eight species of earthworms from 27 

2 locations in Switzerland, representing three ecophysiological groups (Ernst et al., 2008): epigeic (surface3 dwelling worms), endogeic (laterally burrowing worms that inhabit the upper soil layers), and anecic 

4 (vertically burrowing worms that reach depths of 6 inches). For epigeic and anecic earthworms, the total 

5 concentration of Pb in leaf litter and in soil, respectively, were the most important drivers of Pb body 

6 burdens. By contrast, the level of Pb in endogeic earthworms was largely determined by soil pH and CEC. 

7 As a result of these differences, the authors suggested that atmosphere-sourced Pb may be more 

8 bioavailable to epigeic than endogeic species, because it is less dependent on modifying factors. Suthar et 

9 al. (2008), on the other hand, found higher Pb bioaccumulation in the endogeic earthworm Metaphire 

10 posthuma than in the anecic earthworm species Lampito mauritii, and speculated that differences in Pb 

11 tissue level arose from differing life-history strategies, such as feeding behaviors, niche preferences, and 

12 burrowing patterns, all of which exposed the endogeic species to greater Pb concentration. Accumulation 

13 studies conducted with Eisenia fetida earthworms documented the difficulty of extrapolating 

14 accumulation kinetic constants from one soil type to another, and showed that many soil physiochemical 

15 properties, including pH, OM, and CEC, among others, work in conjunction to affect metal bioavailability 

16 (Nahmani et al., 2009). However, once taken up from the environment, more than half of the 

17 bioaccumulated Pb appears to be contained within earthworm tissue and cell membranes (Li et al., 2008).

18 Despite significant Pb uptake by earthworms, Pb in earthworm tissue may not be bioavailable to 

19 predators. Pb in the earthworm Aporrectodea caliginosa was determined to be contained largely in the 

20 granular fraction (approximately 60% of total Pb), while the remaining Pb body burden was in the tissue, 

21 cell membrane, and intact cell fractions (Vijver et al., 2006). From this, the authors concluded that only a 

22 minority of earthworm-absorbed Pb would be toxicologically available to cause adverse effects in the 

23 earthworms or in their predators. Earthworm activity can alter Pb bioavailability and subsequent uptake 

24 by earthworms themselves and other organisms. The presence of earthworms may increase soil pH though 

25 the secretion of cutaneous mucus, and worm activity is generally associated with increased bioavailability 

26 of Pb. Sizmur and Hodson (2009) speculated that earthworms affect Pb mobility by modifying the 

27 availability of cations or anions. However, Coeurdassier et al. (2007) found that snails did not have a 

28 higher Pb content when earthworms were present, and that unexpectedly, Pb was higher in earthworm 

29 tissue when snails were present.

Arthropods

30 Cicadas pupating in historically Pb-arsenate-treated soils accumulated Pb at concentrations similar 

31 to those reported previously for earthworms (Robinson et al., 2007). Likewise, tissue Pb levels measured 

32 in Coleoptera specimens collected from areas containing average soil concentration of 45 and 71 mg

33 Pb/kg exhibited a positive relationship with soil Pb content, although abundance was unaffected (Schipper 
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1 et al., 2008). By contrast, the Pb sequestration rates that were observed in two woodlouse species, O. 

2 asellus and P. scaber, were species-dependent (Gál et al., 2008). Both species were field collected at Pb3 contaminated sites (average concentration, 245 mg Pb/kg dry weight; range, 21-638 mg Pb/kg dry 

4 weight), with O. asellus Pb levels averaging 43 mg Pb/kg over all sites, while P. scaber contained no 

5 detectable Pb residues. Pb concentration measured in granivorous rough harvester ants (Pogonomyrmex 

6 rugosus), in the seeds of some plant species they consume, and in surface soil, were all shown to decline 

7 with increasing distance from a former Pb smelter near El Paso, Texas, where soil leachable Pb at the 

8 three sites of ant collection ranged from 0.003 to 0.117 mg Pb/kg (Del Toro et al., 2010). Ants 

9 accumulated approximately twice as much Pb as was measured in seeds, but the study did not separate the 

10 effects of dietary exposure from those of direct contact with soil or respiratory intake.

7.2.3.3. Terrestrial Vertebrates

11 Tissue Pb residues in birds and mammals associated with adverse toxicological effects were 

12 presented in the 2006 Pb AQCD. In general, avian blood, liver, and kidney Pb concentrations of 0.2-3 μg

13 Pb/dL, 2-6 mg Pb/kg wet weight, and 2-20 mg Pb/kg wet weight, respectively, were linked to adverse 

14 effects. A few additional studies of Pb uptake and tissue residues in birds and mammals conducted since 

15 2006 are reviewed here. 

16 In a study of blood Pb levels in wild Steller’s eiders (Polysticta stelleri) and black scoters 

17 (Melanitta nigra) in Alaska, the authors compiled avian blood Pb data from available literature to develop 

18 reference values for sea ducks (Brown et al., 2006). The background exposure reference value of blood 

19 Pb was <20 μg Pb/dl, with levels between 20 and 59 μg Pb/dl as indicative of Pb exposure. Clinical 

20 toxicity was in the range of 60-99 μg Pb/dl in birds while >100 μg Pb/dl results in acute, severe toxicity. 

21 In measurement of blood Pb with a portable blood Pb analyzer, only 3% of birds had values indicating 

22 exposure and none of the birds had higher blood Pb levels or clinical signs of toxicity. Tissue distribution 

23 of Pb in liver, kidney, ovary and testes of rain quail (Coturnix coramandelicus) following oral dosing of 

24 0.5 mg/kg, 1.25 mg/kg or 2.5 mg/kg Pb acetate for 21 days indicated that Pb uptake was highest in liver 

25 and kidney and low in ovary and testes (Mehrotra et al., 2008). Resident feral pigeons (Columba livia) 

26 captured in the urban and industrial areas of Korea exhibited increased lung Pb concentration, ranging 

27 from 1.6 to 1.9 mg Pb/kg wet weight (Nam & Lee, 2006). However, tissue concentration did not correlate 

28 with atmospheric Pb concentration, so the authors concluded that ingestion of particulate Pb (paint chips, 

29 cement, etc.) in the urban and industrial areas was responsible for the pigeons’ body burden. Similarly, 

30 70% of American woodcock (Scolopax minor) chicks and 43% of American woodcock young-of-year 

31 collected in Wisconsin, U.S., exhibited high bone Pb levels of 9.6-93 mg Pb/kg dry weight and 1.5-220 

32 mg Pb/kg, respectively, even though radiographs of birds’ gastrointestinal tracts revealed no evidence of 
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1 shot ingestion (Strom et al., 2005). Authors hypothesized that unidentified anthropogenic sources may 

2 have caused the observed elevated Pb levels. 

3 In addition to birds, soil-dwelling mammals can also bioaccumulate atmospherically-sourced Pb: 

4 Northern pocket gophers (Thomomys talpoides) trapped within the Anaconda Smelter Superfund Site 

5 were shown to accumulate atmospherically deposited Pb. Gopher liver and carcass Pb concentration 

6 averaged 0.3 and 0.4 mg Pb/kg wet weight on low Pb soils (47 mg Pb/kg), 0.4 and 0.9 mg Pb/kg wet 

7 weight in medium Pb soils (95 mg Pb/kg) and 1.6 and 3.8 mg Pb/kg wet weight in high Pb soils (776.5 

8 mg Pb/kg) (Reynolds et al., 2006). 

9 Casteel et al. (2006) found that bioavailability of Pb from environmental soil samples in swine (Sus 

10 domestica) depended on Pb form or type, with high absorption of cerussite and manganese-Pb oxides and 

11 poor absorption of galena and anglesite. Juvenile swine (approximately 5-6 weeks old and weighing 8-11 

12 kg) were fed Pb-contaminated soils collected from multiple sources for 15 days (concentration range of 

13 1,270 to 14,200 mg Pb/kg) to determine the relative bioavailability. While Pb concentrations were 

14 roughly equivalent in blood, liver, kidney, and bone tissues, individual swine exhibited different uptake 

15 abilities (Casteel et al., 2006).

16 Interestingly, dietary Ca deficiency (0.45 mg Ca daily versus 4 mg under normal conditions) was 

17 linked to increased accumulation of Pb in zebra finches (Taeniopygia guttata) that were provided with 

18 drinking water containing 20 mg/L Pb (Dauwe et al., 2006). Liver and bone Pb concentration were 

19 increased by an approximate factor of three, while Pb concentration in kidney, muscle, and brain tissues 

20 were roughly doubled by a Ca-deficient diet. However, it is not known whether this level of dietary Ca 

21 deficiency is common in wild populations of birds.

7.2.3.4. Food Web

22 In addition to the individual factors reviewed above, understanding the bioavailability of Pb along a 

23 simple food chain is essential for determining risk to terrestrial animals. While the bioavailability of 

24 ingested soil or particulates is relatively simple to measure and model, the bioavailability to secondary 

25 consumers of Pb ingested and sequestered by primary producers and primary consumers is more complex. 

26 Kaufman et al. (2007) caution that the use of total Pb concentration in risk assessments can result in 

27 overestimation of risk to ecological receptors, and they suggest that the bioaccessible fraction may 

28 provide a more realistic approximation of receptor exposure and effects. This section reviews recent 

29 literature that estimates the bioaccessible fraction of Pb in dietary items of higher order consumers, and 

30 various studies suggesting that Pb may be transferred through the food chain but that trophic transfer of 

31 Pb results in gradual attenuation, i.e., lower concentration at each successive trophic level.

32 Earthworm and plant vegetative tissue collected from a rifle and pistol range that contained average 

33 soil Pb concentration of 5,044 mg Pb/kg were analyzed for Pb content and used to model secondary 
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1 bioavailability to mammals (Kaufman et al., 2007). Earthworms were determined to contain an average of 

2 727 mg Pb/kg, and the Pb content of unwashed leaf tissues averaged 2,945 mg Pb/kg. Canonical 

3 correspondence analysis detected no relationship between earthworm and soil Pb concentration, but did 

4 show correlation between unwashed vegetation and soil concentration. The authors noted that the 

5 relatively high Pb concentration of unwashed as opposed to washed vegetation indicated the potential 

6 importance of aerial deposition (or dust resuspension) in determining total vegetative Pb concentration. 

7 Based on the mammalian gastric model, they noted that 50% of vegetation tissue Pb and 77% of 

8 earthworm tissue Pb was expected to be bioavailable to consumers. The avian gizzard model indicated 

9 that 53% of soil Pb and 73% of earthworm Pb was bioaccessible to birds, and, for both mammals and 

10 birds, the bioaccessible fraction of Pb was a function of total Pb concentration. 

11 The transfer of Pb from soils contaminated by a Pb-zinc mine was limited along a soil-plant-insect12 chicken food chain (Zhuang et al., 2009). In soils averaging 991 mg Pb/kg, Rumex K-1 plants sequestered 

13 an average of 1.6 mg/kg wet weight Pb in the shoot tissue, while larvae of the leafworm Spodoptera litura

14 accumulated an average Pb concentration of 3.3 mg Pb/kg wet weight S. litura-fed chickens (Gallus 

15 gallus domesticus) accumulated 0.58 mg Pb/kg and 3.6 mg Pb/kg in muscle and liver tissue, respectively, 

16 but only liver Pb burden was increased significantly relative to controls. A large proportion of ingested Pb 

17 was excreted with the feces. Likewise, an insectivorous bird species, the black-tailed godwit (Limosa 

18 limosa) was shown to accumulate Pb from earthworms residing in Pb-contaminated soils (Roodbergen et 

19 al., 2008). Pb concentration in eggs and feathers was increased in areas with high soil and earthworm Pb 

20 concentration (336 and 34 mg Pb/kg, respectively): egg Pb concentration averaged 0.17 mg Pb/kg and 

21 feather concentration averaged 2.8 mg Pb/kg. This suggests that despite a residence breeding time of only 

22 a few months, this bird species could accumulate Pb when breeding areas are contaminated.

23 Rogival et al. (2007) showed significant positive correlations between soil Pb concentration along a 

24 gradient (approximately 50 to 275 mg Pb/kg) at a metallurgical plant, and Pb concentration in both acorns 

25 (from Quercus robur) and earthworms (primarily Dendrodrilus rubidus and Lumbricus rubellus) collected 

26 on site. Acorn and earthworm Pb contents were, in turn, positively correlated with the Pb concentration in 

27 the liver, kidney, and bone tissues of locally trapped wood mice (Apodemus sylvaticus). 

28 The uptake and transfer of Pb from soil to native plants and to red deer (Cervus elaphus) was 

29 investigated in mining areas of the Sierra Madrona Mountains in Spain (Reglero et al., 2008). The authors 

30 reported a clear pattern between plant Pb concentration and the Pb content of red deer tissues with 

31 attenuation (i.e., decreasing concentration) of Pb up the food chain. Interestingly, soil geochemistry likely 

32 was affected by mining activity as holm oak (Quercus ilex), gum rockrose (Cistus ladanifer), elmleaf 

33 blackberry (Rubus ulmifolius), and grass (Graminae) tissues collected from mining areas exhibited 

34 increased Pb levels (up to 98 mg Pb/kg in grasses and 21 mg Pb/kg in oak) despite the fact that total soil 

35 Pb concentration were not significantly greater than those of the non-mining areas. 
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1 Positive relationships were observed between Cepaea nemoralis snail tissue Pb levels and Pb 

2 concentration measured in Urtica dioica leaves in field-collected samples from areas characterized by 

3 metal soil contamination (approximately 200 to 400 mg Pb/kg) (Notten et al., 2005). Inouye et al (2007)

4 found that several invertebrate prey of fence lizards, including Acheta domestica crickets, Tenebrio 

5 molitor beetles, and P. scaber isopods, accumulate Pb from dietary exposures (10, 50, 100, 250, 500, 750, 

6 and 1,000 mg Pb/kg) lasting between 44 and 72 days. By day 44, Pb body burdens of crickets were 31, 50 

7 and 68 mg/kg (wet weight) at the three highest dietary exposures, respectively. Isopods and beetle larvae 

8 accumulated significantly less Pb, with average body burdens of 10, 15, and 14 mg Pb/kg following 56 

9 days of exposure, and 12, 14, and 31 mg Pb/kg following 77 days of exposure, respectively. For all 

10 invertebrates tested, Pb was sequestered partly in the exoskeleton, and partly in granules. Exoskeleton Pb 

11 may be available to predators, but periodically returns to background level with each shedding. Granular 

12 Pb is unavailable. Trophic attenuation is thus likely.

13 Overall, studies of Pb transfer in food webs have established the existence of pervasive trophic 

14 transfer of the metal, but no consistent evidence of trophic magnification. It appears that on the contrary, 

15 attenuation is common as Pb is transferred to higher trophic levels. However, many individual transfer 

16 steps, as from particular plants to particular invertebrates, result in concentration, which may then be 

17 undone when stepping to the next trophic level. It is possible that whether trophic transfer is magnifying 

18 or attenuating depends on Pb concentration itself. Kaufman et al. (2007) determined that, at low 

19 concentrations of soil Pb, risk to secondary consumers (birds and mammals) was driven by the 

20 bioavailability of Pb in worm tissues, while at high soil concentrations, bioavailability of soil-associated 

21 Pb was more critical. The authors concluded that incorporation of bioavailability/bioaccessibility 

22 measurements in terrestrial risk assessments could lead to more accurate estimates of critical Pb levels in 

23 soil and biota. Finally, while trophic magnification does greatly increase exposure of organisms at the 

24 higher levels of the food web, these studies establish that atmospherically deposited Pb reaches species 

25 that have little direct exposure to it. For those species, detrimental effects are not a function of whether 

26 they accumulate more Pb than the species they consume, but of the absolute amount they are exposed to, 

27 and their sensitivity to it. 

Table 7-2. Soil-to-tissue bioaccumulation factors for various terrestrial plant, 

invertebrate, and vertebrate species

Species Soil Concentration Tissue 

Type

Tissue 

Concentration

Bioaccumulation 

Factor Reference

Vegetable 379 mg/kg Root 47.3 mg/kg 0.12 X. Hu & Ding (2009)

Vegetable 379 mg/kg Shoot 36.2 mg/kg 0.1 X. Hu & Ding (2009)

P. angulata 4020 mg/kg Root 527.1 mg/kg 0.13 Cui et al. (2007)

P. angulata 4020 mg/kg Shoot 331.1 mg/kg 0.08 Cui et al. (2007)

A. theophrasti 4020 mg/kg Root 38.7 mg/kg 0.01 Cui et al. (2007)

A. theophrasti 4020 mg/kg Shoot 61.4 mg/kg 0.02 Cui et al. (2007)
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Species Soil Concentration Tissue 

Type

Tissue 

Concentration

Bioaccumulation 

Factor Reference

S. vestita 2409 mg/kg Shoot 3142 mg/kg 1.3 Y. Li et al. (2009)

S. vestita 2409 mg/kg Root 7457 mg/kg 3.1 Y. Li et al. (2009)

S. grevilloides 5330 mg/kg Shoot 69 mg/kg 0.01 Y. Li et al. (2009)

D. carota 40.1 mg/kg Root 12.2 mg/kg 0.3 Murray et al. (2009)

A. fistulosum 3560 mg/kg Whole plant 23.7 mg/kg 0.007 Cho et al. (2009)

C. baccatum 0.6 mg/kg Fruit 0.86 mg/kg 1.4 Antonious & Kochhar (2009)

A. achatina 1200 mg/kg Whole body 468.5 mg/kg 0.39 Ebenson & Ologhobo (2009b)

A. achatina 200 mg/kg Whole body 91.4 mg/kg 0.46 Ebenson & Ologhobo (2009b)

A. achatina 20 mg/kg Whole body 12.2 mg/kg 0.61 Ebenson & Ologhobo (2009b)

H. aspersa 39 mg/kg Viscera 1.1 mg/kg 0.03 De Vaufleury et al. (2006)

H. aspersa 52 mg/kg Viscera 1.2 mg/kg 0.02 De Vaufleury et al. (2006)

L. terrestris 710 mg/kg Whole body 20.6 mg/kg 0.03 Darling & Thomas (2005)

O. asellus 245 mg/kg Whole body 43 mg/kg 0.18 Gál et al. (2008)

T. talpoides 777 mg/kg Liver 1.63 mg/kg 0.002 Reynolds et al. (2006)

T. talpoides 777 mg/kg Whole body 3.85 mg/kg 0.005 Reynolds et al. (2006)

L. limosa 336 mg/kg Egg 0.17 mg/kg 0.0005 Roodbergen et al. (2008)

L. limosa 336 mg/kg Feather 2.79 mg/kg 0.008 Roodbergen et al. (2008)

7.2.4. Biological Effects

1 Various adverse effects can be observed in exposed terrestrial species following uptake and 

2 accumulation of Pb. While many of the responses are specific to organism type, induction of antioxidant 

3 activities in response to Pb exposure has been reported in plants, invertebrates, and vertebrates. In this 

4 section, the observed biological effects caused by exposure to atmosphere-derived Pb will be discussed, 

5 while the results of dose-response experimentation will be addressed in Section 7.2.5. Because 

6 environmental releases of Pb often include simultaneous release of other metals, it can be difficult to 

7 identify Pb-specific effects in field studies, with the exception of effects from leaded gasoline and some 

8 Pb smelter deposition. Many laboratory studies that expose organisms to natural soils (or to biosolids9 amended soils) also include exposure to multiple metals. There is some information about mechanisms of 

10 metal interactions, such as through competition for binding locations on specific enzymes or on cellular 

11 receptors, but generally such interactions (particularly of multiple metals) are not well understood 

12 (ATSDR, 2004). Despite a few well-known examples of metal antagonism (e.g., Cu and Mo or Cd and 

13 Zn), it is common practice to assume additivity of effects (Fairbrother et al., 2007). Because this review is 

14 focused on effects of Pb, studies reviewed for this section and the following include only those for which 

15 Pb was the only, or primary, metal to which the organism was exposed. 

7.2.4.1. Plants and Lichen

16 Pb exposure has been linked to decreased photosynthesis in affected plants, significant induction of 

17 antioxidant activities, genetic abnormalities, and decreased growth. Induction of antioxidant responses in 

18 plants has been shown to increase tolerance to metal soil contamination, but at sufficiently high levels, 
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1 antioxidant capacity is exceeded, and metal exposure causes peroxidation of lipids and DNA damage, 

2 eventually leading to accelerated senescence and potentially death (Stobrawa & Lorenc-Plucinska, 2008). 

Effects on Photosystem and Chlorophyll

3 The effect of Pb exposure on the structure and function of plant photosystem II was studied in giant 

4 duckweed, Spirodela polyrrhiza (Ling & Hong, 2009). Although this is an aquatic plant, photosystem II is 

5 present in all plants, and effects on photosystem II observed in any plant species are likely to occur in all 

6 of them. The Pb concentration of extracted photosystem II particles was found to increase with increasing 

7 environmental Pb concentration, and increased Pb concentration was shown to decrease emission peak 

8 intensity at 340 nm, amino acid excitation peaks at 230 nm, tyrosine residues, and absorption intensities. 

9 This results in decreased efficiency of visible light absorption by affected plants. The authors theorized 

that Pb2+ may replace either Mg2+ or Ca2+ 10 in chlorophyll or the oxygen-evolving center, inhibiting 

11 photosystem II function through an alteration of chlorophyll structure. Consistently with these results, Wu

12 et al. (2008) demonstrated that Pb exposure interfered with and decreased light absorption by spinach 

13 (Spinacia oleracea) plants. Spinach seeds were soaked in 5, 12, or 25 mM PbCl2 for 48 hours prior to 

14 germination, and following 42 days of growth, plants were sprayed with PbCl2 solutions. Chloroplast 

15 absorption peak intensity, fluorescence quantum yield at 680 nm, and whole-chain electron transport rate 

16 all decreased with Pb exposure, as did photosystem II photoreduction and oxygen evolution. Liu et al. 

17 (2010) observed that chlorophyll a and b content in wheat grown in soils spiked with Pb nitrate rose with 

18 length of exposure until 14 days, at which point chlorophyll decreased. At exposures of 0.1 and 0.5 mM 

19 Pb in hydroponic solution for 50 days, concentration of chlorophyll a and b was decreased in radish (R. 

20 sativus) (Kumar & Tripathi, 2008). Changes in chlorophyll content in response to Pb were also observed 

21 in lichen and moss species following exposures intended to simulate atmospheric deposition (Carreras & 

22 Pignata, 2007). Usnae amblyoclada lichen was exposed to aqueous Pb solutions of 0.5, 1, 5, and 10 mM 

23 Pb nitrate; chlorophyll a concentration was shown to decrease with increasing Pb exposure. However, the 

24 ratio of lichen dry weight to fresh weight increased following Pb exposures. As compared to other metals, 

25 Pb caused less physiological damage, which the authors attributed to the metal’s high affinity for binding 

26 to and sequestration within cell walls.

27 The effect of Pb exposure on chlorophyll content of the moss and liverwort species Thuidium 

28 delicatulum, T. sparsifolium, and Ptychanthus striatus was investigated following simulated atmospheric 

exposures of 10-10 to 10-2 29 M Pb (Shakya et al., 2008). Both chlorophyll a and total chlorophyll content of 

30 the mosses (T. delicatulum and T. sparsifolium) decreased with increasing Pb exposure, but the effect was 

31 not statistically significant. For the liverwort, Pb exposure resulted in significant decreases in content of 

32 chlorophyll a, chlorophyll b, and total chlorophyll. Further, the total chlorophyll content of Hypnum 
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1 plumaeforme mosses was decreased by 5.8% following exposure to 10 mM Pb, while lower exposures 

2 slightly elevated chlorophyll content. 

Response of Antioxidants

3 Increased antioxidant activity is a common response to Pb exposure, although this endpoint may 

4 not necessarily be an indication of deleterious effects on plant vitality. Increases in reactive oxygen 

5 species with increasing exposure to Pb have been demonstrated in broad bean (Vicia faba) (C.-R. Wang et 

6 al., 2010; C.-R. Wang, Wang, Tian, Yu, et al., 2008; C. Wang, Y. Tian, et al., 2010) and tomato 

7 (Lycopersicon esculentum) (C.-R. Wang, Wang, Tian, Xue, et al., 2008), where they were accompanied by 

8 proportional increases in superoxide dismutase (SOD), glutathione, guaiacol peroxidase, as well as lipid 

9 peroxidation, and decreases in catalase. Reddy et al. (2005) found that horsegram (Macrotyloma 

10 uniflorum) and bengalgram (Cicer arietinum) plants exposed to Pb solutions of 200, 500, and 800 mg

11 Pb/kg exhibited increased antioxidant activity: at exposures of 800 mg Pb/kg, root and shoot SOD activity 

12 increased to 2–3 times that of controls, and induction was slightly higher in M. uniflorum. Similarly, 

13 catalase, peroxidase, and glutathione-S-transferase activities were elevated in Pb-stressed plants, but were 

14 again higher for M. uniflorum. Antioxidant activities were also markedly greater in the root tissues than 

15 the shoot tissues of the two plants, and were very likely related to the higher Pb accumulation rate of the 

16 roots. The effectiveness of the up-regulation of antioxidant systems in preventing damage from Pb uptake 

17 was evidenced by lower malondialdehyde (MDA) (a chemical marker of lipid peroxidation) concentration 

18 in M. uniflorum versus C. arietinum, indicating a lower rate of lipid peroxidation in response to M. 

19 uniflorum’s higher antioxidant activity.

20 Gupta et al. (2010) contrasted responses of two ecotypes of Sedum alfredii (an Asian perennial 

21 herb), one an accumulator of Pb and the other not. Glutathione level was increased in both, and root and 

22 shoot lengths were decreased following long-term exposures to Pb up to 200 μM. However, the 

23 accumulator plants exhibited greater SOD and ascorbate peroxidase activity, likely as a result of greater 

24 Pb uptake and a concurrent increased detoxification capacity. Similar results were reported by Islam et al. 

25 (2008): following Pb exposures of 200 μM, catalase, ascorbic acid, and glutathione levels of another 

26 Chinese herb, Elsholtzia argyi, were increased, while SOD and guaiacol peroxidase activities decreased. 

27 Microscopic analysis also showed that affected plants exhibited abnormal chloroplast structures. The 

28 response of glutathione was further confirmed in wheat (Liu et al., 2010) grown in soils spiked with Pb 

29 nitrate. Evidence of increasing lipid peroxidation (MDA accumulation) with increasing Pb exposure was 

30 also found in mosses (Sun et al., 2009) and lichens. Lichens field-collected from the trunks of poplar 

31 (Populus tremula) trees in eastern Slovakia were chemically analyzed for metal concentration arising 

32 from exposure to smelter pollution (Dzubaj et al., 2008). These concentrations (ranging from 13 to 1,523 

33 mg Pb/kg dry weight) were assessed in relation to physiological variables, including chlorophyll a and b, 
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1 carotenoids, photosystem II activity, CO2 gas exchange (respiration), and MDA content. Lichen Pb levels 

2 were significantly correlated only with MDA content.

Growth

3 There is evidence of effects of Pb on higher growth processes as well. Both growth and carotenoid 

4 and chlorophyll content of Brassica juncea (mustard) plants were negatively affected by Pb exposure 

5 (John et al., 2009). Pb treatments of 1,500 μM (as Pb acetate solution) decreased root lengths and stem 

6 heights by 50% after 60 days. Exposure to 600 μM Pb and greater decreased carotenoid content, while 

7 chlorophyll a was decreased at Pb exposures of 450 μM and higher. However, no effects were seen in 

8 growth or chlorophyll production of maize (Zea mays) following growth in smelter ash–spiked soils 

9 containing 1,466 mg Pb/kg (and 18.6 mg Cd/kg) (Komarek et al., 2009). Pb concentrations of 7,331 

10 mg/kg (98.0 mg/kg Cd) were required to elicit chlorosis and the expected decreased in growth.

11 Chinese cabbage (Brassica pekinensis) exposed to Pb-containing soils (0, 4, and 8 mM/kg dry 

12 weight) exhibited depressed nitrogen assimilation as measured by shoot nitrite content, nitrate reductase 

13 activity, and free amino acid concentration (Xiong et al., 2006). The authors planted germinated cabbage 

14 seeds in soils spiked with Pb acetate to give final soil concentration of 0.2, 0.4, and 0.8 mg/kg dry weight 

15 total Pb and collected leaf samples for 11 days. At exposures of 0.4 and 0.8 mg/kg, leaf nitrite content was 

16 decreased by 29% and 20%, while nitrate content was affected only at the highest Pb exposure (70% of 

17 control levels). Free amino acid content in exposed plants was 81% and 82% of control levels, 

18 respectively. B. pekinensis shoot biomass was observed to decrease with increasing Pb exposures, with 

19 biomass at the two highest Pb exposures representing 91% and 84% of control growth, respectively.

Genetic and Reproductive Effects

20 Exposure to Pb also resulted in genetic abnormalities, including bridges, condensed bivalents, and 

21 laggards, in the meiotic cells of pea plants (Lathyrus sativus) (Kumar & Tripathi, 2008). Seeds were 

22 germinated in soils amended with Pb nitrate at concentrations of 25, 50, 100, 200, and 300 mg Pb/kg, and 

23 concentrations of 100 mg Pb/kg and greater were found to be genotoxic or detrimental to pea viability. 

24 Cenkci et al. (2010) exposed fodder turnip (B. rapa) to 0.5 to 5 nM of Pb nitrate for 6 days and showed 

25 decreased genetic template stability (as quantified by random amplified polymorphic DNA profiles) and 

26 decreased photosynthetic pigments. 

27 Pb exposure also decreased germination rate and growth, and increased pollen sterility in radish 

28 grown for 50 days in hydroponic solutions containing 0.5 mM Pb (Kumar & Tripathi, 2008) Plants 

29 decreased growth, curling and chlorosis of young leaves, and decreased root growth. In addition, Gopal 

30 and Rizvi (2008) showed that Pb exposure increased uptake of phosphorus and iron and decreased sulfur 

31 concentration in radish tops.
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1 Interestingly, as in zebra finch (Section 7.2.3.3), Ca was found to counteract the toxic effects of Pb 

2 in both monocotyledon and dicotyledon plant seedlings, with tomato (Lycopersicon esculentum), rye 

3 (Lolium sp.), mustard, and maize plants exhibiting increased tolerance to Pb exposures of 5, 10, and 20 

4 mg/L in the presence of Ca concentration of 1.2 mM and higher (Antosiewicz, 2005).

7.2.4.2. Invertebrates

5 Exposure to Pb also causes antioxidant effects, reductions in survival and growth, as well as 

6 decreased fecundity in terrestrial invertebrates as summarized in the 2006 Pb AQCD. In addition to these 

7 endpoints, recent literature also indicates that Pb exposure can cause significant neurobehavioral 

8 aberrations, and in some cases, endocrine-level impacts. Second-generation effects have been observed in 

9 some invertebrate species. 

10 The morphology of γ-aminobutyric acid (GABA) motor neurons in Caenorhabditis elegans

11 nematodes was affected following exposure to Pb nitrate for 24 hours (Du & Wang, 2009). The authors 

12 determined that exposures as low as 2.5 μM Pb nitrate could cause moderate axonal discontinuities, and 

13 observed a significant increase in the number of formed gaps and ventral cord gaps at Pb nitrate 

14 exposures of 75 and 200 μM. Younger C. elegans larvae were more likely to exhibit neurobehavioral 

15 toxicity symptoms in response to Pb exposure (2.5 μM) (Xing, Guo, et al., 2009). Neural degeneration, as 

16 demonstrated by dorsal and ventral cord gaps and neuronal loss was also more pronounced in young 

17 larval C. elegans than in older larvae and adults (Xing, Rui, et al., 2009). C. elegans nematodes exposed 

18 to Pb concentration as low as 2.5 μM for 24 hours also exhibited significantly altered behavior 

19 characterized by decreased head thrashes and body bends. Exposures of 50 μM Pb and greater decreased 

20 the number of nematode forward turns (D. Y. Wang & Xing, 2008). Chemotaxis towards NaCl, cAMP, 

21 and biotin was also decreased in C. elegans nematodes exposed to Pb concentration greater than 2.5 μM 

22 (Xing, Du, et al., 2009). This evidence suggests that Pb may exert neurotoxic action in invertebrates as it 

23 does in vertebrates. However, it is unclear how these behavioral aberrations would affect fitness or 

24 survival (D. Y. Wang & Xing, 2008). 

25 Younger individuals also appear to be more sensitive to the reproductive effects of Pb exposure. 

26 Guo et al. (2009) showed that concentrations of 2.5, 50, and 100 μM Pb had greater significant adverse 

27 effects on reproductive output when early-stage larval C. elegans were exposed. Adult C. elegans

28 exhibited decreased brood size only when exposed to the highest Pb concentration.

29 The progeny of C. elegans nematodes exposed to 2.5, 75, and 200 μM Pb nitrate exhibited 

30 significant indications of multi-generational toxicity (D. Y. Wang & Peng, 2007). Life spans of offspring 

31 were decreased by increasing parental Pb exposure, and were comparable to the reductions in parental 

32 life-spans. Similarly, diminished fecundity was observed in the progeny of C. elegans exposed to Pb (9%, 

33 19%, and 31% reductions of control fecundity, respectively), although the decrease was smaller than in 
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1 the exposed parental generation (reductions of 52%, 58%, and 65%, respectively). Significant behavioral 

2 defects affecting locomotion were also observed in the offspring, but these impacts were not determined 

3 to be concentration-dependent.

4 E. andrei earthworms exposed to 21 different soils, each containing 2,000 mg/kg freshly added Pb, 

5 for 28 days exhibited highly variable mortality, ranging from 0% to 100%, (Bradham et al., 2006). Pb 

6 body burden of exposed worms ranged from 29 to 782 mg Pb/kg. Internal Pb concentration was also 

7 negatively correlated to reproductive output. CEC and pH were found to be the principal soil 

8 characteristics determining the differences in those effects, although the apparent role of CEC may only 

9 have been due to its correlation with other soil characteristics. Low soil Pb concentration (5 mg/kg) also 

10 decreased the protein content of E. fetida earthworms during a 7-day exposure (M. Li et al., 2009). Higher 

11 Pb concentration had no effect on protein production. However, cellulase activity was increased by the 7-

12 day exposures to Pb at all exposure concentrations (31%, 13%, and 23% of control activity at exposures 

13 of 5, 50, and 500 mg Pb/kg, respectively), indicating that Pb exerted a detrimental effect on worm 

14 metabolism. By contrast, Svendsen et al. (2007) found that L. rubellus earthworms exposed for 42 days to 

15 field-collected smelter-polluted soils containing average Pb concentration of 106, 309, and 514 mg Pb/kg 

16 dry weight exhibited normal survival and cocoon production rates, even though they accumulated more 

17 Pb with increased environmental concentration. The much smaller effect may be explained by the 

18 increased aging time undergone by field soil, causing a larger fraction of the total Pb to be complexed and 

19 sequestered by organic and inorganic compounds.

20 As in plants, induction of antioxidant activity is affected by exposure to Pb in invertebrates. The 

21 induction of antioxidant activity was correlated to standard toxicity measurements in Theba pisana snails 

22 (Radwan et al., 2010). Topical application of Pb solutions (500 to 2,000 μg Pb per animal) to snails 

23 resulted in decreased survival, increased catalase and glutathione peroxidase activities, and decreased 

24 glutathione concentration. The 48-hour LD50 concentration was determined to be 653 μg per snail. Snail 

25 glutathione content was decreased at exposures of 72.2% of the 48-hour LD50 value, while Pb exposure at 

26 40% of the 48-hour LD50 value induced catalase and glutathione peroxidase activities. Further, decreased 

27 food consumption, growth, and shell thickness were observed in juvenile A. achatina snails exposed to 

28 Pb-contaminated (concentration greater than 134 mg/kg) diet for 12 weeks (Ebenso & Ologhobo, 2009a). 

29 A similar depression of growth was observed in sentinel juvenile A. achatina snails deployed at Pb30 polluted sites in the Niger Delta region of Nigeria. Although snail mortality was not increased 

31 significantly by exposure to soil Pb up to 1,200 mg/kg, a concentration-dependent relationship was 

32 established for growth, with significant reduction observed at 12-week exposures to 20 mg Pb/kg (Ebenso 

33 & Ologhobo, 2009b). However, consumption of field-collected Pb-polluted U. dioica leaves containing 3 

34 mg/kg stopped all reproductive output in C. nemoralis. Snails also exhibited diminished food 

35 consumption rates when offered leaves with both low (1.5 mg Pb/kg) and high Pb content, but the 

36 mechanism of the dietary aversion was not defined (Notten et al., 2006). 
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1 Chronic dietary exposure to Pb was also examined in post-embryonic oribatid mites (Archegozetes 

2 longisetosus) (Kohler et al., 2005). Both algae and bark samples were soaked in 100 mg/L Pb as Pb nitrate 

3 and provided as diet and substrate, respectively, to larval mites. In addition to elevated heat shock proteins 

4 (hsp70), 90.8% of the protonymphs exhibited significant leg deformities, including abnormal claws, 

5 shortened and thickened legs, and translocated setae. Although not specifically discussed, it is very likely 

6 that these deformities would decrease mite mobility, prey capture, and reproductive viability.

7 Lock et al. (2006) compared the toxicity of both laboratory-spiked soils and field-collected Pb8 contaminated soils to springtails (F. candida). The 28-day EC50 values derived for F. candida ranged from 

9 2,060 to 3,210 mg Pb/kg in leached and unleached Pb-spiked soils, respectively, whereas field-collected 

10 soils had no significant effect on springtail reproduction up to (but not including) 14,436 mg Pb/kg (Lock 

11 et al., 2006). Consequently, leaching soils prior to use in bioassays had only a slight effect on Pb toxicity 

12 to resident springtails, and did not provide an appropriate model for field-weathered, Pb-contaminated 

13 soils. This indicates that physiochemical factors other than leaching may be more important determinants 

14 of Pb bioavailability. A 4-week exposure to Pb-amended soils containing up to 3,200 mg Pb/kg had no 

15 significant adverse effect on Sinella curviseta springtail survival or reproduction (Xu et al., 2009).

16 Carabid beetles (Pterostichus oblongopunctatus) inhabiting soils contaminated by pollution from a 

17 Pb-Zn smelter (containing 136 to 2,635 mg Pb/kg) were field-collected and then laboratory-reared for two 

18 generations (Lagisz & Laskowski, 2008). While fecundity was positively correlated to soil metal 

19 concentration (e.g., more eggs were produced by females collected from contaminated areas), the 

20 hatching rate of eggs diminished with increasing soil metal contamination. For the F1 generation, females 

21 produced by parents inhabiting highly polluted areas exhibited decreased body mass. The authors stated 

22 that these results indicate that invertebrates inhabiting metal- (or Pb-) contaminated soils could face 

23 “significantly altered life-history parameters.”

24 Several studies suggest that Pb may disrupt hormonal homeostasis in invertebrates. Shu et al. 

25 (2009) reported that vitellogenin production in both male and female S. litura moths was disrupted 

26 following chronic dietary exposure to Pb. Adult females reared on diets containing 25, 50, 100, or 200 

27 mg Pb/kg exhibited decreased vitellogenin mRNA induction, and vitellogenin levels were demonstrated 

28 to decrease with increasing Pb exposure. Conversely, in a study by Zheng and Li (2009), vitellogenin 

29 mRNA was detected at higher levels in males exposed to 12 and 25 mg Pb/kg, although vitellogenin 

30 levels were not affected. Similarly, the sperm morphology of the Asian earthworm (Pheretima guillelmi) 

31 was found to be altered significantly following 2-week exposure to soils containing 1,000, 1,400, 1,800, 

32 and 2,500 mg Pb/kg (Zheng & Li, 2009). Common deformities were swollen head and head helices, while 

33 head bending was also recorded in some cases. These deformities were observed following exposures to 

34 concentration below the 14-day LC50 (3,207 mg Pb/kg) and below the concentration at which weight was 

35 diminished (2,800 mg Pb/kg). Experimentation with the model organism Drosophila indicates that Pb 
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1 exposure may increase time to pupation and decrease pre-adult development, both of which are 

2 endocrine-regulated (Hirsch et al., 2010).

7.2.4.3. Terrestrial Vertebrates

3 According to the 2006 Pb AQCD, commonly observed effects of Pb on avian and mammalian 

4 wildlife include decreased survival, reproduction, and growth, as well as effects on development and 

5 behavior. More recent experimental data presented here expand and support these conclusions, and also 

6 indicate that Pb can exert other effects on exposed terrestrial vertebrates, including alteration of hormones 

7 and other biochemical variables.

8 Red-backed salamanders (Plethodon cinereus) exposed to Pb-amended soils (553, 1,700, 4,700, 

9 and 9,167 mg Pb/kg) by Bazar et al. (2010) exhibited lowered appetite and decreased white blood cell 

10 counts at the two highest concentrations, but tolerated field-collected, aged soils containing Pb 

11 concentrations of up to 16,967 mg Pb/kg with no significant deleterious effects. The white blood cell 

12 count of adult South American toads, Bufo arenarum was also decreased by weekly sublethal 

13 intraperitoneal injections of Pb acetate at 50 mg Pb/kg body weight (Chiesa et al., 2006). The toads also 

14 showed altered serum profiles and increased number of circulating blast cells. Final toad blood Pb levels 

15 were determined to be 8.6 mg Pb/dL, although it is unclear whether this is representative of Pb 

16 concentration observed in field B. arenarum populations exposed to Pb. The authors suggested that, based 

17 on these findings, long-term environmental exposure to Pb could affect toad immune response. In western 

18 fence lizards (S. occidentalis), sub-chronic (60-day) dietary exposure to 10 to 20 mg Pb/kg per day 

19 resulted in significant sublethal effects, including decreased cricket consumption, decreased testis weight, 

20 decreased body fat, and abnormal posturing and coloration (Salice et al., 2009). Long-term dietary Pb 

21 exposures are thus likely to decrease lizard fitness.

22 Even in cases of high environmental Pb exposures, however, linking Pb body burdens to adverse 

23 biological effects can be difficult. Pb tissue concentration in field-collected urban blackbirds (Turdus 

24 merula) were determined to be 3.2 mg Pb/kg, 4.9 mg Pb/kg, and 0.2 mg Pb/kg wet mass in breast 

25 feathers, washed tail feathers, and blood, respectively (Scheifler, Coeurdassier, et al., 2006). Although 

26 these levels were significantly higher than those measured in rural blackbirds, elevated Pb tissue 

27 concentration was not significantly correlated to any index of body condition. On the other hand, 

28 Hargreaves et al. (2010) showed that Pb tissue concentration of female arctic shorebirds was negatively 

29 correlated with reproductive success. Maternal blood Pb levels were negatively associated with hatching 

30 success in black bellied plovers (Pluvialis squatarola) and ruddy turnstones (Arenaria interpres), and 

31 with nest duration in all species tested. There was no significant correlation between adult whole-blood or 

32 feather Pb concentration and Pb levels in produced eggs.
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1 The long-term effect of atmospheric Pb deposition on pied flycatcher (Ficedula hypoleuca) 

2 nestlings was determined in native communities residing in the Laisvall mining region of Sweden 

3 (Berglund et al., 2010). Moss samples indicated that Pb deposition in study areas ranged between 100 and 

4 2,000 mg Pb/kg dry weight during operations and 200 and 750 mg Pb/kg when operations ceased. A 

5 simultaneous slight reduction was observed in pied flycatcher blood Pb levels, from 0.4 to 0.3 mg Pb/kg). 

6 However, clutch size was decreased in pied flycatchers inhabiting the mining area both during and after 

7 mining operations, and mean nestling mortality was 2.5 and 1.7 higher after mining operations in the 

8 mining region than in reference areas, respectively. The authors noted that Pb deposition in the mining 

9 region remained elevated even after mining operations ceased, and that stable Pb isotope analysis 

10 suggested that smelter Pb remained available to pied flycatcher through the transfer of historically 

11 deposited Pb in soil to prey items.

12 The level of corticosteroid hormones in field populations of white stork nestlings (Ciconia ciconia) 

13 in a mining area affected by Pb and other metals was positively correlated with blood Pb levels (Baos et 

14 al., 2006). The effect was more pronounced for single nestlings than for multiple-chick broods. 

15 Surprisingly, average blood Pb levels in chicks inhabiting reference areas was 91 μg/L ( ± 51), which was 

16 higher than blood Pb levels from the mining area (44 ± 34 μg/L). However, the correlation between blood 

17 Pb levels and the corticosteroid stress response in white stork nestlings was observed in both groups of 

18 birds. Burger and Gochfeld (2005) exposed herring gull (Larus argentatus) chicks to Pb acetate via an 

19 intraperitoneal injection of 100 mg Pb/kg body weight, to produce feather Pb concentration approximately 

20 equivalent to those observed in wild gulls. Pb-exposed gulls exhibited abnormal behaviors, including 

21 decreased walking and food begging, erratic behavioral thermoregulation, and diminished recognition of 

22 caretakers.

23 Again, dietary or other health deficiencies unrelated to Pb exposure are likely to exacerbate the 

24 effects of Pb. Ca-deficient female zebra finches (T. guttata) had a suppressed secondary humoral immune 

25 response following 28-day exposures to 20 mg/L Pb in drinking water (Snoeijs et al., 2005). This 

26 response, however, was not observed in birds fed sufficient Ca. Although a significant finding, these data 

27 are difficult to interpret under field conditions where the overall health of avian wildlife may not be easily 

28 determined. 

29 Chronic Pb exposures were also demonstrated to adversely affect several mammalian species. 

30 Young adult rats reared on a diet containing 1,500 mg/kg Pb acetate for 50 days demonstrated less 

31 plasticity in learning than non-exposed rats (McGlothan et al., 2008), indicating that Pb exposure caused 

32 significantly altered neurological function. Yu et al. (2005) showed that dietary Pb exposure affected both 

33 the growth and endocrine function of gilts (S. domestica). Consumption of 10 mg Pb/kg diet resulted in 

34 lower body weight and food intake after 120 days of dietary exposure; Pb exposure decreased final weight 

35 by 8.2%, and average daily food intake of Pb-exposed pigs was decreased by 6.8% compared to control 

36 intake. Additionally, concentration of estradiol, lutenizing hormone, and pituitary growth hormone were 
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1 decreased (by 12%, 14%, and 27% versus controls, respectively), while blood Pb level was increased by 

2 44% to an average 2.1 μg/dL. In cattle grazing near Pb-zinc smelters in India, blood Pb levels were 

3 positively correlated with plasma levels of the thyroid hormones thyroxine (T4) and tri-iodothyronine 

4 (T3) and the hepatic biomarkers alanine transaminase and aspartate transaminase (Swarup et al., 2007). 

5 Total lipids, total protein and albumin levels were decreased in the same animals. 

6 Pb-treated oocytes of buffalo (Bubalus bubalis) assessed in vitro at concentrations ranging from 0.5 to 

7 1,000 μg/dL in one-day cultures indicated a significant decline in viability of oocytes at 100 μg/dL (Nandi 

8 et al., 2010). Dose-dependent effects on oocyte viability, morphological abnormalities, cleavage, 

9 blastocyst yield and blastocyst hatching were observed in Pb-treated oocytes with maturation significantly 

10 reduced at 250 μg/dL and 100% oocyte death at 3,200 μg/dL. Similarly, the reproductive viability of red 

11 deer (C. elaphus) inhabiting a Pb-contaminated mining area of Spain was shown to be altered, with 11% 

12 and 15% reductions in spermatozoa and acrosome integrity observed in male deer from the mining area 

13 compared with those residing in reference areas (Reglero et al., 2009).

7.2.5. Exposure and Response of Terrestrial Species

14 Given that exposure to Pb may adversely affect organisms at the individual, population, or 

15 community level, determining the rate and concentration at which these effects occur is essential in 

16 predicting the overall risk to terrestrial organisms. However, data from controlled studies using a single 

17 compound are scarce relative to field studies, which in turn often investigate effects of multiple metal 

18 contaminants and afford too little control on interacting variables to be of use in establishing a general 

19 dose-response function. This section updates available information derived since the 2006 Pb AQCD, 

20 summarizing several dose-response studies with soil invertebrates. No new exposure-response 

21 information was available for plants, birds, or mammals.

22 Dose-dependent responses in antioxidant enzymes were observed in adult L. mauritii earthworms 

23 exposed to soil-associated Pb contamination (75, 150, 300 mg Pb/kg) (Maity et al., 2008). By day seven 

24 of exposure, glutathione-S-transferase activity and glutathione disulfide concentration were positively 

25 correlated with increasing Pb exposures, while glutathione concentration exhibited a negative dose26 response relationship with soil Pb concentration. However, these trends had become insignificant by the 

27 end of the total exposure period (28 days), as a result of normalization of antioxidant systems following 

28 chronic exposure. This strongly suggests that changes to earthworm antioxidant activity are an adaptive 

29 response to Pb exposures. 

30 Both survival and reproductive success of E. fetida earthworms showed concentration-dependent 

31 relationships with soil Pb concentration during the course of standard 14- and 56-day toxicity tests (Jones 

32 et al., 2009). Five levels of Pb soil concentration were prepared for the acute 14-day study via spiking 

33 with Pb nitrate—0, 300, 711, 1,687, and 2,249 mg Pb/kg, while soil concentration of 0, 355, 593, 989, and 
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1 1,650 mg Pb/kg were used in chronic (56-day) earthworm bioassays. A 14-day acute LC50 of 2,490 mg

2 Pb/kg was determined from the dose-response relationship, while the approximate 56-day NOEC (no 

3 observed effect concentration) and EC50 values were about 400 mg/kg and 1,000 mg/kg Pb, respectively. 

4 Currie et al. (2005) observed mortality of E. fetida after 7 and 14 days in spiked field soil at seven levels 

5 of Pb (0 to 10,000 mg Pb/kg). They reported LC50 values of 2,662 mg Pb/kg at 7 days and 2,589 

6 mg Pb/kg at 14 days or 2,827 mg Pb/kg at both 7 and 14 days, depending on the number of worms in the 

7 experimental enclosure.

8 Other studies have shown no correlation between Pb concentration in either earthworm tissue or 

9 soil, and earthworm survival rate. Although the Pb content of E. fetida held in metal-contaminated soils 

10 containing between 9.7 and 8,600 mg Pb/kg was positively correlated with Pb concentration of soil, there 

11 was no statistical relationship with earthworm survival during a 42-day exposure period (Nahmani et al., 

12 2007). However, Pb concentration in soil leachate solution was significantly correlated with decreased 

earthworm survival and growth (linear regression: R2 13 = 0.64, p< 0.0001). The 42-day Pb EC50 for E. fetida

14 growth was 6,670 mg Pb/kg. 

15 Langdon et al. (2005) exposed three earthworm species (E. andrei, L. rubellus, and A. caliginosa) 

16 to Pb nitrate-amended soils at concentrations of 1,000 to 10,000 mg Pb/kg to determine species variability 

17 in uptake and sensitivity. Twenty-eight-day LC50 values for the three species were 5,824 mg Pb/kg, 2,867 

18 mg Pb/kg, and 2,747 mg Pb/kg, respectively, indicating that L. rubellus and A. caliginosa are significantly 

19 more vulnerable to Pb contamination than E. andrei, a common laboratory species. This is comparable to 

20 previous findings by Spurgeon et al. (1994) who reported 14-day LC50 of 4,480 mg Pb/kg and 50-day 

21 LC50 of 3,760 mg Pb/kg for E. fetida, another standard laboratory test species. In the more recent study of 

22 E. fetida sensitivity summarized above, Jones (2009) reported LC50 values for E. fetida that are similar to 

23 those for L. rubellus and A. caliginosa. It is likely that these apparent species differences are a result of 

24 differential bioavailability of the Pb in test soils. However, the Pb body burden of all three species in the 

25 study by Langdon et al. (2005) increased with increasing environmental concentration, and there were no 

26 species differences in Pb tissue content. When given a choice between treated and untreated soils, all 

27 worm species exhibited significant avoidance of Pb-contaminated soils, and altering pH (and, 

28 consequently, Pb bioavailability) had no impact on avoidance (Langdon et al., 2005). Field earthworms 

29 may thus be able to reduce their exposure to Pb through behavior.

30 The individual and population-level responses of the springtail Paronychiurus kimi to Pb were 

31 determined by Son et al. (2007) using artificial soils, following the 1999 European ISO methodology. The 

32 7-day Pb LC50 was determined to be 1,322 mg Pb/kg dry weight, while the 28-day reproduction EC50 was 

33 established as 428 mg Pb/kg. The intrinsic rate of population increase was lower at a Pb soil concentration 

34 of 1,312 mg/kg, and the authors estimated that, at this level, P. kimi populations would be extirpated. The 

35 authors noted that, in this case, the reproductive endpoint overestimated the population-level risk for P. 

36 kimi springtails exposed to Pb, and proposed that more specific measures of population-level endpoints 
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1 (such as the reduction in intrinsic rate of increase) be used to determine risk to populations. Menta et al. 

2 (2006) showed that a nominal soil concentration of 1,000 mg Pb/kg decreased the reproductive output of 

3 two collembolans, Sinella coeca and F. candida. Pb concentration of 50, 100, and 500 mg Pb/kg slightly 

4 but significantly depressed S. coeca adult survival, while F. candida survival was statistically unaffected 

5 by Pb exposure.

6 In addition to species variability, physical and chemical factors affecting Pb bioavailability were 

7 also demonstrated to significantly influence the toxicity of Pb to terrestrial species. As noted previously in 

8 Section 7.2.2, laboratory-amended artificial soils provide a poor model for predicting the toxicity of Pb9 contaminated field soils, because aging and leaching processes, along with variations in physiochemical 

10 properties (pH, CEC, OM), influence metal bioavailability. Consequently, toxicity values derived from 

11 exposure-response experimentation with laboratory-spiked soils probably overestimate true 

12 environmental risk, with the possible exception of highly acidic sandy soils. Because toxicity is 

13 influenced by bioavailability of soil biogeological and chemical characteristics, extrapolation of toxic 

14 concentrations between different field-collected soils will be difficult. Models that account for those 

15 modifiers of bioavailability, such as the terrestrial biotic ligand model proposed by Smolders et al. (2009), 

16 have proven difficult to develop due to active physiological properties of soil organisms affecting either 

17 uptake (such as root phytochelatins) or sequestration of Pb (such as granule formation in root tissues and 

18 earthworms, or substitution of Pb for calcium in bones).

7.2.6. Community and Ecosystem Effects

19 According to the 2006 Pb AQCD, natural terrestrial ecosystems near significant Pb point sources 

20 (such as smelters and mines) exhibited a number of ecosystem-level effects, including decreased species 

21 diversity, changes in floral and faunal community composition, and decreasing vigor of terrestrial 

22 vegetation. These findings are summarized in Table AX7-2.5.2 of the Annex to the 2006 Pb AQCD (U.S. 

23 EPA, 2006). More recent literature explored the interconnected effects of Pb contamination on soil 

24 bacterial and fungal community structure, earthworms, and plant growth, in addition to impacts on soil 

25 microbial community function. 

26 Inoculation of maize plants with Glomus intraradices arbuscular mycorrhizal fungi isolates 

27 decreased Pb uptake from soil, resulting in lower shoot Pb concentration and increased plant growth and 

28 biomass (Sudova & Vosatka, 2007). Similarly, Wong et al. (2007) showed that the presence of arbuscular 

29 mycorrhizal fungi improved vetiver grass (Vetiveria zizanioides) growth, and while Pb uptake was 

30 stimulated at low soil concentration (10 mg Pb/kg), it was depressed at higher concentration (100 and 

31 1,000 mg Pb/kg). Bojarczuk and Kieliszewska-Rokicka (2010) found that the abundance of 

32 ectomycorrhizal fungi was negatively correlated with the concentration of metals, including Pb, in the 

33 leaves of silver birch seedlings. Arbuscular mycorrhizal fungi may thus protect plants growing in Pb-
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1 contaminated soils. Microbes too may dampen Pb uptake and ameliorate its deleterious effects: biomass 

2 of plants grown in metal-contaminated soils (average Pb concentration 24,175 mg Pb/kg dry weight) 

3 increased with increasing soil microbial biomass and enzymatic activity (Epelde et al., 2010). However, 

4 above certain Pb concentration, toxic effects on both plants and microbial communities may prevent these 

5 ameliorating effects. R.Y. Yang et al. (2008) found that both the mycorrhizal colonization and the growth 

6 of Solidago canadensis were negatively affected by soil Pb contamination. They suggested that, more 

7 generally, Pb-mediated alterations in plant-fungal dynamics may be the cause of ecological instability in 

8 terrestrial vegetative communities exposed to metals. 

9 The presence of both earthworms and arbuscular mycorrhizal fungi decreased the mobility of Pb in 

10 mining soils undergoing phytoremediation (Ma et al., 2006). Inoculation with both earthworms and fungi 

11 increased plant growth at sites contaminated with mine tailings compared to that observed at sites with 

12 75% less Pb contamination. Most likely, this was a result of the decrease in bioavailable (DTPA13 extractable and ammonium acetate-extractable) Pb to 17% to 25% of levels in areas without the 

14 earthworm and arbuscular mycorrhizal fungi amendments. The presence of earthworms in metal15 contaminated soils decreased the amount of water-soluble Pb (Sizmur & Hodson, 2008), but despite this 

16 decrease, ryegrass accumulated more Pb from earthworm-worked soils than soils without worms present. 

17 Sizmur and Hodson speculated that increased root dry biomass may explain the increased uptake of Pb in 

18 the presence of earthworms. By contrast, Coeurdassier et al. (2007) found that Pb was higher in 

19 earthworm tissue when snails were present, but that snails did not have a higher Pb content when 

20 earthworms were present.

21 Microbial communities of industrial soils containing Pb concentrations of 61, 456, 849, 1,086, and 

22 1,267 mg Pb/kg dry weight were also improved via revegetation with native plants, as indicated by 

23 increased abundances of fungi, actinomycetes, gram-negative bacteria, and protozoa, as well as by 

24 enhanced fatty acid concentration (C. B. Zhang et al., 2006). Increased plant diversity ameliorated the 

25 effects of soil Pb contamination (300 and 600 mg Pb/kg) on the soil microbial community (R. Y. Yang et 

26 al., 2007).

27 The effect of Pb on microbial community function has been quantified previously using functional 

28 endpoints such as respiration rates, fatty acid production, and soil acid phosphatase and urease activities, 

29 which may provide a better estimate of ecological impacts than microbial diversity or abundance 

30 measurements. When the microbial properties of metal-contaminated urban soils were compared to those 

31 of rural soils, significant differences were detected in basal community respiration rates and microbial 

32 abundance (Y. Yang et al., 2006). The urban soils studied contained multiple metal contaminants, but 

33 microbial biomass was the only measured endpoint to be significantly and negatively correlated to Pb 

34 concentration. This suggests that anthropogenic Pb contamination may adversely affect soil microbial 

35 communities, and alter their ecological function. Most studies of metal-induced changes in microbial 

36 communities have been conducted using mixtures of metals. Akerblom et al. (2007) tested the effects of 
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1 six metals (Cr, Zn, Mo, Ni, Cd, and Pb) individually. All tested metals had a similar effect on the species 

2 composition of the microbial community. Exposure to a high Pb concentration (52 mg Pb/kg) negatively 

3 affected respiration rates. Total phospholipid fatty acid content was determined to negatively correlate 

4 with increasing Pb exposure, indicating alteration of the microbial community. (S. Khan et al., 2010), 

5 found that following a 2-week exposure to three levels of Pb (150, 300, and 500 mg Pb/kg), both acid 

6 phosphatase and urease levels (measures of soil microbial activity) decreased significantly, although they 

7 had recovered by the ninth week. In addition, the number of culturable bacteria was also decreased, but 

8 only at the highest exposure concentration tested. 

9 Soil bacteria community structure and basal respiration rates were examined in natural soils with 

10 pH values ranging from 3.7 to 6.8 (Lazzaro et al., 2006). Six soil types of differing pH were treated with 

11 Pb nitrate concentrations of 0.5, 2, 8, and 32 mM. Basal respiration was adversely affected in two soil 

12 types tested at the highest Pb treatment (32 mM), and in a third at the two highest Pb treatments (8 and 32 

13 mM). Terminal Restriction Fragment Length Polymorphism analysis indicated that bacterial community 

14 structure was only slightly altered by Pb treatments. While pH was correlated with the amount of water15 soluble Pb, these increases were apparently not significant enough to affect bacterial communities, 

16 because there were no consistent relationships between soil pH and respiration rate or microbial 

17 community structure at equivalent soil Pb concentration.

18 Pb exposure negatively affected the prey capture ability of certain fungal species. The densities of 

19 traps constructed by nematophagous fungi decreased in soils treated with 0.15 mM Pb chloride (Mo et al., 

20 2008). Nematophagous fungi are important predators of soil-dwelling nematodes, collecting their prey 

21 with sticky nets, branches, and rings. This suppression caused a subsequent reduction in fungal nematode 

22 capturing capacity, and could result in increased nematode abundance. 

23 High concentration of soil metals were linked to a significant reduction in soil microorganism 

24 abundance and diversity. Soil columns spiked with Cu, Zn, and Pb acetate (total Pb concentration of 278 

25 to 838 mg Pb/kg, depending on depth) exhibited a 10- to 100-fold decrease in microbial abundance, with 

26 specific microbe classes (e.g., actinomycetes) seemingly more affected than others (Vaisvalavicius et al., 

27 2006). Concurrently, decreases in soil enzymatic activity were also observed, with saccharase activity 

28 decreased by 57–77%, dehydrogenase activity by 95–98%, and urease activity 65–97%. Although this 

29 suggests that Pb contamination may alter the nutrient cycling capacity of affected soil communities, it is 

30 difficult to separate the impact of Pb from the contributions of copper and zinc that were added with the 

31 Pb.

32 The microbial communities of soils collected from a Pb-Zn mine and a Pb-Zn smelter were 

33 significantly affected by Pb and other metals (e.g., Cd) (Q. Hu et al., 2007). At a mine site, Pb 

34 concentration of 57 to 204 mg Pb/kg and Cd concentration of 2.4 to 227 mg Cd/kg decreased the number 

35 of bacteria-forming colonies extracted from soils. Principal component analysis of microbial community 

36 structure demonstrated that different communities were associated with different metal soil concentration. 
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1 Similarly, soil microbial communities exposed to metal contamination from a smelter site (soil Pb 

2 concentration ranging from 30 to 25,583 mg Pb/kg dry weight) showed decreased bacterial functional 

3 diversity (although fungal functional diversity increased) and no effects on soil respiration rates were 

4 observed (Stefanowicz et al., 2008). This led the authors to conclude that bacterial diversity is a more 

5 sensitive endpoint and a better indicator of metal exposure than fungal diversity or microorganism 

6 activity. In a similar study, Kools et al. (2009) showed that soil ecosystem variables measured after a 6-

7 month exposure to metal-contaminated soil indicated that Pb concentration (536 or 745 mg Pb/kg) was an 

8 important driver of soil microbial species biomass and diversity.

9 Pb-resistant bacterial and fungal communities were extracted regularly from soil samples at a 

10 shooting range site in southern Finland (Hui et al., 2009). While bioavailable Pb concentration averaged 

11 100 to 200 mg Pb/kg as determined by water extraction, the total Pb concentrations measured on site were 

12 30,000 to 40,000 mg Pb/kg. To determine Pb tolerance, bacterial colonies extracted and cultured from 

13 shooting range and control soils were grown on media containing either 0.4 or 1.8 mM Pb. While bacteria 

14 isolated from control soil did not proliferate on high-Pb media, shooting-range soil microbe isolates grew 

15 on high-Pb media and were deemed Pb tolerant. The authors noted that bacterial species common in 

16 control samples were not detected among the Pb-tolerant species isolated from shooting-range soils. Thus, 

17 it was concluded that even long-term exposure to minimally bioavailable Pb can alter the structure of soil 

18 decomposer communities, which could in turn decrease decomposition rates.

19 Microbial communities associated with habitats other than soils are also affected by exposure to 

20 atmospherically deposited Pb. Alder (Alnus nepalensis) leaf microorganism populations were greater in 

21 number at non-affected sites than at sites adjacent to a major Indian highway with increased Pb pollution 

22 (Joshi, 2008). The density, species richness, and biomass of testate amoebae communities grown on 

23 Sphagnum fallax mosses were significantly decreased following moss incubation in Pb solutions of either 

24 625 or 2,500 μg Pb/L (Nguyen-Viet et al., 2008). More importantly, species richness and density were 

25 negatively correlated with Pb concentration accumulated within the moss tissue. The structure of 

26 microbial communities associated with lichen surfaces was affected by lichen trace-element 

27 accumulation, including Pb content. Lichens collected from industrial areas had elevated Pb concentration 

28 (10 to 20 mg Pb/kg versus 5 to 7 mg Pb/kg in urban and rural areas, respectively) and housed bacterial 

29 communities characterized by increased cyanobacteria biomass (Meyer et al., 2010). 

30 Following a 28-day exposure to field-collected soils contaminated with metals (including Pb at 426 

31 mg Pb/kg), both population growth and individual growth of the earthworm L. rubellus were diminished 

32 (Klok et al., 2006). The authors proposed that, although these reductions were unlikely to result in 

33 extirpation, avian predators such as the godwit (Limosa limosa) that feed heavily on earthworms may be 

34 affected adversely by a reduction of available earthworm biomass. 

35 During the past 5 years, there has been increasing interest in the effects of Pb and other metals on 

36 the functional aspects of soil microbial communities. Most studies show that Pb decreases diversity and 
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1 function of soil microorganisms. However, in an example of ecological mutualism, plant-associated 

2 arbuscular mycorrhizal fungi protect the host plant from Pb uptake and fungal viability seems to be 

3 protected by the plants. Similarly, soil microbial communities (bacterial species as well as fungi) in Pb4 contaminated soils are improved by revegetation. A few studies have reported on effects of Pb to 

5 populations of soil invertebrates. They demonstrated that Pb can decrease earthworm population density, 

6 although not to levels that would result in local extinction. There have been no recently reported studies 

7 on the potential effects of Pb on terrestrial vertebrate populations or communities, or possible indirect 

8 effects through reduction of prey items such as earthworms. However, it is well known from historical 

9 data that Pb can have a widespread and dramatic effect on populations of waterfowl exposed to spent shot 

10 (Beyer et al., 1996) and may be negatively affecting loons in the northeastern U. S. through ingestion of 

11 Pb fishing sinkers (Scheuhammer & Norris, 1996). Studies at shooting ranges and downwind of smelters 

12 previously reported in the 2006 Pb AQCD demonstrate effects of dispersed Pb on terrestrial soil and plant 

13 communities with resulting decreases in secondary consumer vertebrate species. 

7.2.7. Critical Loads in Terrestrial Systems

14 The critical load is the environmental concentration predicted or estimated from available literature 

15 data, above which adverse effects to organisms are likely to occur. It is based on the relative toxicity of 

16 the compound to species or ecosystem processes of concern, and an estimate of residence time in the soil 

17 environment. The concept of critical load is discussed in more detail in Section 7.1.3 of this chapter and 

18 in Section 7.3 of the 2006 Pb AQCD (U.S. EPA, 2006).

19 Hall et al. (2006) used the critical load approach to conduct a national risk assessment of 

20 atmospheric Pb deposition for the U. K. While specific regions were determined to have low critical load 

21 values for Pb (central England, the Pennines, and southern Wales), the authors noted that this approach 

22 can be significantly biased, as available ecotoxicological data used in the modeling were from studies that 

23 were not conducted in soils representative of all U.K. soils. De Vries et al. (2009) similarly observed that 

24 the uncertainty inherent in a critical load approach to Pb risk assessment is influenced by the critical 

25 concentration of dissolved metal and the absorption coefficients of exposed soils. However, this approach 

26 did indicate that for forest soils in the Netherlands, 29% of the areas would be expected to exceed the 

27 critical load, based on currently available toxicity data and Pb pollution data (de Vries & Groenenberg, 

28 2009). Similarly, although Pb soil concentration in the Bologna Province of Italy were far below 

29 concentrations harmful to soil organisms, current atmospheric Pb deposition rates suggest that critical 

30 load exceedances are likely in the future, unless annual Pb emissions are decreased (Morselli et al., 2006).

31 In some cases, risk assessment for Pb predicts risks to terrestrial animals at environmental 

32 concentrations that fall below natural background levels, and uncertainties associated with standard 

33 toxicity testing can become magnified through the risk assessment process, degrading the reliability of 
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1 estimates of hazardous environmental concentrations. Smolders et al. (2009) concluded from their 

2 comparison of field soils, spiked soils, and artificially aged spiked soils that corrections for aging and 

3 interacting soil properties in spiked soils will make predicted-no-effect concentrations rise above 

4 background. Buekers et al. (2009) proposed the use of a Tissue Residue Approach as a risk estimation 

5 method for terrestrial vertebrates that does not predict risks at background levels, and has smaller 

6 uncertainty. This approach eliminates the need for quantitative estimation of food intake or Pb species 

7 bioavailability. Blood Pb no-adverse-effect concentration (NAEC) and lowest adverse effects 

8 concentration (LAEC) derived from 25 literature studies examining the effects of Pb exposure on growth, 

9 reproduction, and hematological endpoints were used to construct a series of species sensitivity 

10 distributions for mammals and birds. For mammals, the hazardous concentration for 5% of species (HC5) 

11 values obtained from the species sensitivity distributions ranged from 11 to 18 μg Pb/dL blood; HC5 

12 values for birds ranged from 65 to 71 μg/dL. The authors proposed the use of 18 and 71 μg/dL as critical 

13 threshold values for mammals and birds, respectively, which are below the lowest NAEC for both data 

14 sets used, and are above typical background Pb values. 

15 Short of conducting expensive in vivo toxicity studies, it is difficult to determine environmental Pb 

16 toxicity given the variation of physiochemical and soil properties that alter bioavailability and toxicity. 

17 Furman et al. (2006) proposed the use of a physiologically based extraction test to predict risks posed to 

18 waterfowl from environmental Pb contamination. The extraction process was modeled after gastric and 

19 intestinal conditions of waterfowl, and was used to gauge the bioavailability of Pb from freshly amended 

20 and aged contaminated soils. The concentration of Pb extracted through the use of the physiologically 

21 based extraction test was demonstrated to be significantly correlated to Pb tissue concentration in 

22 waterfowl exposed via in vivo studies of the same soils. 

7.2.8. Soil Screening Levels

23 Developed by EPA, ecological soil screening levels (Eco-SSLs) are maximum contaminant 

24 concentrations in soils that are predicted to result in little or no quantifiable effect on terrestrial receptors. 

25 These conservative values were developed so that contaminants that could potentially present an 

26 unacceptable hazard to terrestrial ecological receptors are reviewed during the risk evaluation process 

27 while removing from consideration those that are highly unlikely to cause significant effects. The studies 

28 considered for the Eco-SSLs for Pb and detailed consideration of the criteria for developing the Eco-SSLs 

29 are provided in the 2006 Pb AQCD (U.S. EPA, 2006). Preference is given to studies using the most 

30 bioavailable form of Pb, to derive conservative values. Soil concentration protective of avian and 

31 mammalian diets are calculated by first converting dietary concentration to dose (mg/kg body weight per 

32 day) for the critical study, then using food (and soil) ingestion rates and conservatively derived uptake 

33 factors to calculate soil concentration that would result in unacceptable dietary doses. This frequently 
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1 results in Eco-SSL values below the average background soil concentration, as is the case with Pb for 

2 birds. The Pb Eco-SSL was completed in March 2005 and has not been updated since. Values for 

3 terrestrial birds, mammals, plants, and soil invertebrates are 11, 56, 120, and 1,700 mg Pb/kg soil (dry 

4 weight), respectively.

7.2.9. Characterization of Sensitivity and Vulnerability

5 Research has long demonstrated that Pb affects survival, reproduction, growth, metabolism, and 

6 development in a wide range of species. The varying severity of these effects depends in part upon 

7 species differences in metabolism, sequestration, and elimination rates. Dietary factors also influence 

8 species sensitivity to Pb. Because of effects of soil aging and other bioavailability factors discussed above 

9 (Section 7.2.2), in combination with differing species assemblages and biological accessibility within 

10 prey items, ecosystems may also differ in their sensitivity and vulnerability to Pb. The 2006 Pb AQCD 

11 reviewed many of these factors which are updated herein by reference to recent literature.

7.2.9.1. Species Sensitivity 

12 There is wide variation in sensitivity of terrestrial species to Pb exposure, even among closely 

13 related organisms. Langdon et al. (2005) showed a two-fold difference in LC50 values among three 

14 common earthworm species, with the standard laboratory species, E. andrei, being the least sensitive. 

15 Similarly, 28-day EC50 values derived for F. candida collembola (springtails) were between 2,060 and 

16 3,210 mg/kg in Pb-spiked soils (Lock et al., 2006), while the springtail species S. curviseta exhibited no 

17 response to a 28-day exposure to 3,200 mg/kg Pb-spiked soil (Xu et al., 2009). Mammalian NOEC values 

18 expressed as blood Pb levels were shown to vary by a factor of 8, while avian blood NOECs varied by a 

19 factor of 50 (Buekers et al., 2009). Age at exposure, in particular, may affect sensitivity to Pb. For 

20 instance, earlier instar C. elegans were more likely than older individuals to exhibit neurobehavioral 

21 toxicity following Pb exposure (Xing, Guo, et al., 2009), and also demonstrated more pronounced neural 

22 degeneration than older larvae and adults (Xing, Rui, et al., 2009).

7.2.9.2. Nutritional Factors 

23 Dietary factors can exert significant influence on the uptake and toxicity of Pb in many species of 

24 birds and mammals. The 2006 Pb AQCD describes how Ca, Zn, Fe, vitamin E, Cu, thiamin, P, Mg, fat, 

25 protein, minerals, and ascorbic acid dietary deficiencies increase Pb absorption and its toxicity. For 

26 example, vitamin E content was demonstrated to protect against Pb-induced lipid peroxidation in mallard 

27 ducks. Generally, Pb exposure is more likely to produce adverse behavioral effects in conjunction with a 

28 nutrient-deficient diet. As previously reported in the 2006 Pb AQCD, Ca deficiencies may increase the 

29 susceptibility of different terrestrial species to Pb, including plant (Antosiewicz, 2005), avian (Dauwe et 
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1 al., 2006; Snoeijs et al., 2005) and invertebrate species. Antosiewicz determined that, for plants, Ca 

2 deficiency decreased the sequestration capacity of several species (tomato, mustard, rye, and maize), and 

3 that this likely resulted in an increased proportion of Pb at sites of toxic action. Because Pb ions can 

4 interact with plant Ca channel pores, in the presence of low Ca and high Pb concentration, a higher 

5 proportion of Pb can interact with these channels and be taken up by plants. A similar phenomenon has 

6 been observed in invertebrates, where the metabolic pathway of metals mimics the metabolic pathway of 

7 Ca (Simkiss et al. (1982), as cited in Jordaens et al. (2006)). Hence, in environments with 

8 disproportionately high Pb versus Ca concentration, accumulation of Pb may be accelerated, as in plants. 

9 Ca deficiency in birds was demonstrated to stimulate the production of Ca-binding proteins in the 

10 intestinal tract, which extract more Ca from available diet; however, this response also enhances the 

11 uptake and accumulation of Pb from diet and drinking water (Fullmer et al. (1997), as cited in Dauwe et 

12 al. (2006)). 

7.2.9.3. Soil Aging and Site-Specific Bioavailability 

13 Total soil Pb concentration is a poor predictor of hazards to avian or mammalian wildlife, because 

14 site-specific biogeochemical and physical properties (e.g., pH, OM, metal oxide concentration) can affect 

15 the sequestration capacity of soils. Additionally, soil aging processes have been demonstrated to decrease 

16 the bioavailable Pb fraction; as such, laboratory toxicity data derived from spiked soils often overestimate 

17 the environmental risk of Pb. Smolders et al. (2009) compared the toxicity of freshly Pb-spiked soils to 

18 experimentally aged spiked soils and field-collected Pb-contaminated soils. Experimental leaching and 

19 aging was demonstrated to increase invertebrate Pb EC50 values by factors of 0.4 to greater than 8; in 

20 approximately half the cases, the proportionality of toxicity to Pb content disappeared following 

21 experimental aging of freshly spiked soils through leaching. The leaching-aging factor for Pb was 

22 determined to be 4.2, and represented the ratio of ED10 values derived in aged soils to freshly spiked soils 

23 (factors greater than one indicate decreased toxicity in aged field soils relative to laboratory spiked soils). 

24 Consequently, the sensitivity of terrestrial vertebrates to environmental Pb exposures will be heavily 

25 dependent on the relative rate of aging and site-specific bioavailability. 

7.2.9.4. Ecosystem Vulnerability

26 Relative vulnerability of different terrestrial ecosystems to effects of Pb can be inferred from the 

27 information discussed above on species sensitivity and how soil geochemistry influences the 

28 bioavailability and toxicity of Pb. Soil ecosystems with low pH, particularly those with sandy soils, are 

29 likely to be the most sensitive to the effects of Pb. Examples of such systems are forest soils, including 

30 oak, beach, and conifer forests. The Pine Barrens in southern New Jersey (also known as the Pinelands) is 

31 an example of a highly vulnerable ecosystem: it is a dense coniferous (pine) forest with acidic, sandy, 
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1 nutrient poor soil. As agricultural areas are taken out of production and revert to old fields and eventually 

2 forests, their vulnerability to Pb is likely to increase as a result of decreasing OM and acidification of soils 

3 (from discontinuation of fertilizing and liming). On the other hand, increasing density of native or 

4 invasive plants with associated arbuscular mycorrhizal fungi will likely act to ameliorate some of the 

5 effects of Pb (see previous discussion of studies by Sudova and Vostka (2007) and Wong et al. (2007). It 

6 is, however, difficult to categorically state that certain plant or soil invertebrate communities are more 

7 vulnerable to Pb than others, as the available toxicity data have not yet been standardized for differences 

8 in bioavailability (because of use of different Pb salts, different soil properties, and different lengths of 

9 aging of soil prior to testing), nutritional state, or organism age, or other interacting factors. Data from 

10 field studies are complicated by the co-occurrence of other metals and alterations of pH, such as 

11 acidification from SO2 in smelter emissions, that are almost universal at sites of high Pb exposure, 

12 especially at mine or smelter sites. However, because plants primarily sequester Pb in the roots, uptake by 

13 soil invertebrates is the most likely pathway for Pb exposure of higher trophic level organisms. 

14 Invertivores are likely at higher risk than herbivores. In fact, estimations of Pb risk at a former Pb smelter 

15 in northern France indicated that area Pb concentration presented the greatest threat to insectivorous bird 

16 and mammal species, but only minimal risk to soil invertebrate and herbivorous mammals (Fritsch et al., 

17 2010). By extension, birds and mammals in ecosystems with a richer biodiversity of soil invertebrates 

18 may be more vulnerable to Pb than those in ecosystems with fewer invertebrates (e.g., arid locations). 

19 Regardless, the primary determinant of terrestrial ecosystem vulnerability is soil geochemistry, notably

20 pH, CEC, and amount of OM.

7.2.10. Ecosystem Services

21 There are no publications at this time that specifically focus on the ecosystem services affected by 

22 Pb in terrestrial systems. The evidence reviewed in this ISA illustrates that Pb can cause ecological effects 

23 in each of the four main categories of ecosystem services (Section 7.1.2) as defined by Hassan et al. 

24 (2005). These effects are sorted into ecosystem services categories and summarized here: 

25  Supporting: altered nutrient cycling, decreased biodiversity, decline of productivity, food 

26 production for higher trophic levels

27  Provisioning: plant yields

28  Regulating: decline in soil quality, detritus production

29  Cultural: ecotourism and cultural heritage values related to ecosystem integrity and 

30 biodiversity, impacts to terrestrial vertebrates.
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1 A few studies since the 2006 Pb AQCD consider the impact of metals in general on ecosystem 

2 services. In a review of the effects of metals on insect behavior, ecosystem services provided by insects 

3 such as detritus reduction and food production for higher trophic levels were evaluated by considering 

4 changes in ingestion behavior and taxis (Mogren & Trumble, 2010). Pb was shown in a limited number of 

5 studies to affect ingestion by insects. Crickets (Chorthippus spp) in heavily contaminated sites reduced 

6 their consumption of leaves in the presence of increasing cadmium and Pb concentrations (Migula & 

7 Binkowska, 1993). Decreased feeding activity in larval and adult Colorado potato beetle (Leptinotarsa 

8 decemlineata) were observed as a result of dietary exposures of Pb and copper (Kwartirnikov et al., 

9 1999), while no effects were found in ingestion studies of Pb with willow leaf beetle, Lochmaea caprae

10 (Rokytova et al., 2004) mottled water hyacinth weevil, Neochetina eichhorniae (Kay & Haller, 1986) and 

11 hairy springtail, Orchesella cincta (Van Capelleveen et al., 1986). 

12 Soil health for agricultural production and other soil-associated ecosystem services is dependent on 

13 the maintenance of four major functions: carbon transformations, nutrient cycles, soil structure 

14 maintenance, and the regulation of diseases and pests and these parameters may be altered by metal 

15 deposition (Kibblewhite et al., 2008). Pb impacts to terrestrial systems reviewed in the previous sections 

16 provide evidence for impacts to supporting, provisioning, and regulating ecosystem services provided by 

17 soils. For example, earthworms were shown to impact soil metal mobility and availability, which in turn 

18 resulted in changes to microbial populations (biodiversity), pH, dissolved organic carbon, and metal 

19 speciation (Sizmur & Hodson, 2009), all of which may directly affect soil fertility. 

20 There is limited evidence of Pb impacts to plant productivity. Productivity of gray birch (Betula 

21 populifolia) was impaired in soils with elevated As, Cr, Pb, Zn and V (Gallagher et al., 2008). Tree growth 

22 measured in both individuals and at the assemblage level using satellite imagery and field spectrometry 

23 was significantly decreased with increasing metal load in soil.

7.2.11. Summary of Effects in Terrestrial Systems

24 This summary of the effects of Pb on terrestrial ecosystems covers information from the 

25 publication of the 2006 Pb AQCD to present. Refer to Section 7.4: Causality determinations for Pb in 

26 Terrestrial and Aquatic Systems for a synthesis of all evidence dating back to the 1977 AQCD considered 

27 to determine causality. 

7.2.11.1. Biogeochemistry and Chemical Effects

28 The amount of Pb dissolved in pore water determines the impact of soil Pb on terrestrial 

29 ecosystems to a much greater extent than the total amount present. It has long been established that the 

30 amount of Pb dissolved in soil solution is controlled by at least six variables: (1) solubility equilibria; (2) 

31 adsorption-desorption relationship of total Pb with inorganic compounds; (3) adsorption-desorption 
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1 reactions of dissolved Pb phases on soil OM; (4) pH; (5) CEC; and (6) aging. Since 2006, further details 

2 have been contributed to the understanding of the role of pH, CEC, OM, and aging. Smolders et al. (2009)

3 demonstrated that the two most important determinants of both solubility and toxicity in soils are pH and 

4 CEC. However, they had previously shown that aging, primarily in the form of initial leaching following 

5 deposition, decreases soluble metal fraction by approximately one order of magnitude (Smolders et al., 

6 2007). Since 2006, OM has been confirmed as an important influence on Pb sequestration, leading to 

7 longer-term retention in soils with higher OM content, and also creating the potential for later release of 

8 deposited Pb. Aging, both under natural conditions and simulated through leaching , was shown to 

9 substantially decrease bioavailability to plants, microbes, and vertebrates.

7.2.11.2. Bioavailability and Uptake 

Plants

10 Studies with herbaceous species growing at various distances from smelters added to the existing 

11 strong evidence that atmospherically transported Pb is taken up by plants. These studies did not establish 

12 the relative proportion that originated from atmospheric Pb deposited in the soil, as opposed to that taken 

13 up directly from the atmosphere through the leaves. Multiple new studies showed that in trees, this 

14 proportion is likely to be very substantial. One study attempted to quantify it, and suggested that 50% of 

15 the Pb contained in Scots Pine in Sweden is taken up directly from the atmosphere. Studies with 

16 herbaceous plants found that in most species tested, soil Pb taken up by the roots is not translocated into 

17 the stem and leaves. Studies with trees found that soil Pb is generally translocated from the roots. 

Invertebrates

18 Since the 2006 Pb AQCD, various species of terrestrial snails have been found to accumulate Pb 

19 from both diet and soil. New studies with earthworms have found that both internal concentration of Pb 

20 and mortality increase with decreasing soil pH and CEC. In addition, tissue concentration differences 

21 have been found in species of earthworms that burrow in different soil layers. The rate of accumulation in 

22 each of these species may result from layer differences in interacting factors such as pH and CEC. 

23 Because earthworms often sequester Pb in granules, some authors have suggested that earthworm Pb is 

24 not bioavailable to their predators. There is some evidence that earthworm activity increases Pb 

25 availability in soil, but it is inconsistent. In various arthropods collected at contaminated sites, recent 

26 studies found gradients in accumulated Pb that corresponded to gradients in soil with increasing distance 

27 from point sources. 
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Vertebrates 

1 There were few new studies of Pb bioavailability and uptake in birds since the 2006 Pb AQCD. A 

2 study of two species of sea ducks in Alaska found that 3% of the birds had tissue levels of Pb that 

3 indicated exposure above background. Urban pigeons in Korea were found to accumulate 1.6 to 1.9 

4 mg/kg wet weight Pb in the lungs, while in Wisconsin 70% of American woodcock chicks and 43 % of 

5 young-of-year had elevated bone Pb (9.6 to 93 mg Pb/kg dry weight in chicks, 1.5 to 220 mg Pb/kg dry 

6 weight in young-of-year). None of the locations for these studies was in proximity to point sources, and 

7 none was able to identify the origin of the Pb. A study at the Anaconda Smelter Superfund site found 

8 increasing Pb accumulation in gophers with increasing soil Pb around the location of capture. A study of 

9 swine fed various Pb-contaminated soils showed that the form of Pb determined accumulation.

Food web

10 New studies were able to measure Pb in the components of various food chains that included soil, 

11 plants, invertebrates, arthropods and vertebrates. They confirmed that trophic transfer of Pb is pervasive, 

12 but no consistent evidence of trophic magnification was found.

7.2.11.3. Biological Effects

Plants 

13 Experimental studies have added to the existing evidence of photosynthesis impairment in plants 

14 exposed to Pb, and have found damage to photosystem II due to alteration of chlorophyll structure, as 

15 well as decreases in chlorophyll content in diverse taxa, including lichens and mosses. A substantial 

16 amount of evidence of oxidative stress in response to Pb exposure has also been produced. Reactive 

17 oxygen species were found to increase in broad bean and tomato plants exposed to increasing 

18 concentrations of soil Pb, and a concomitant increase in superoxide dismutase, glutathione, peroxidases, 

19 and lipid peroxidation, as well as decreases in catalase were observed in the same plants. Monocot, dicot, 

20 and bryophytic taxa grown in Pb-contaminated soil or in experimentally spiked soil all responded to 

21 increasing exposure with increased antioxidant activity. In addition, reduced growth was observed in 

22 some experiments, as well as genotoxicity, decreased germination, and pollen sterility.

Invertebrates 

23 Recently published studies have shown neuronal damage in nematodes exposed to low 

24 concentrations of Pb (2.5 μM), accompanied by behavioral abnormalities. Reproductive adverse effects 

25 were found at lower exposure in younger nematodes, and effects on longevity and fecundity were shown 
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1 to persist for several generations. Increased mortality was found in earthworms, but was strongly 

2 dependent on soil characteristics including pH, CEC, and aging. Snails exposed to Pb through either 

3 topical application or through consumption of Pb-exposed plants had increased antioxidant activity, 

4 decreased food consumption, growth, and shell thickness. Effects on arthropods exposed through soil or 

5 diet varied with species and exposure conditions, and included diminished growth and fecundity, 

6 endocrine and reproductive anomalies, and body deformities. Increasing concentration of Pb in the 

7 exposure medium generally resulted in increased effects within each study, but the relationship between 

8 concentration and effects varied between studies, even when the same medium, e.g., soil, was used. 

9 Evidence suggested that aging and pH are important modifiers.

Vertebrates

10 Effects on amphibians and reptiles included decreased white blood cell counts, decreased testis 

11 weight, and behavioral anomalies. However large differences in effects were observed at the same 

12 concentration of Pb in soil, depending on whether the soil was freshly amended, or field-collected from 

13 contaminated areas. As in most studies where the comparison was made, effects were smaller when field14 collected soils were used. In some birds, maternal elevated blood Pb level was associated in recent studies 

15 with decreased hatching success, smaller clutch size, high corticosteroid level, and abnormal behavior. 

16 Some species show little or no effect of elevated blood Pb level. Effects of dietary exposure were studied 

17 in several mammalian species, and cognitive, endocrine, immunological, and growth effects were 

18 observed.

7.2.11.4. Exposure Response

19 Evidence reviewed in previous sections demonstrates clearly that increased exposure to Pb is 

20 generally associated with increases in observed effects in terrestrial ecosystems. It also demonstrates that 

21 many factors, including species and various soil physiochemical properties, interact strongly with 

22 concentration to modify those effects. In these ecosystems, where soil is generally the main component of 

23 the exposure route, Pb aging is a particularly important factor, and one that may be difficult to reproduce 

24 experimentally. Without quantitative characterization of those interactions, characterizations of exposure25 response relationships would likely not be transferable outside of experimental settings. Since the 2006 

26 Pb AQCD, a few studies of exposure-response have been conducted with earthworms, and results have 

27 been inconsistent.

7.2.11.5. Community and Ecosystem Effects

28 New evidence of effects of Pb at the community and ecosystem scale include several studies of the 

29 ameliorative effects of mycorrhizal fungi on plant growth, attributed to decreased uptake of Pb by plants, 
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1 although both mycorrhizal fungus and plant were negatively affected. The presence of both earthworms 

2 and mycorrhizal fungi decreased solubility and mobility of Pb in soil in one study, but the presence of 

3 earthworms was associated with higher uptake of Pb by plants in another. The presence of snails 

4 increased uptake of Pb by earthworms, but not vice-versa. Most recently published research on 

5 community and ecosystem scale effects of Pb has focused on soil microbial communities, which have 

6 been shown to be impacted in both composition and activity. Many recent studies have been conducted 

7 using mixtures of metals, but have tried to separate the effects of individual metals when possible. One 

8 study compared the effects of 6 metals individually (Akerblom et al., 2007), and found that their effects 

9 on community composition were similar. In studies that included only Pb, or where effects of Pb could be 

10 separated, soil microbial activity was generally diminished, but in some cases recovered over time. 

11 Species and genotype composition were consistently altered, and those changes were long-lasting or

12 permanent.

7.2.11.6. Critical Loads, Sensitivity and Vulnerability

13 Exploratory studies of critical load approaches for risk assessment for Pb have been recently 

14 conducted in the U. K., the Netherlands, and Italy. Their authors suggested that the main limitations of

15 critical loads approaches in those countries were gaps and uncertainty in both ecotoxicological and Pb 

16 deposition data. The most visible indication of the need for improvement was that critical load values 

17 were often below background values. Smolders (2009) suggested that correcting for aging and other 

18 interacting factors would likely raise predicted-no-effect concentrations, and others proposed basing risk 

19 management on tissue residue in organisms, or creating extraction methods that more closely mimic 

20 uptake and accumulation.

21 Recent studies have addressed differences in sensitivity explicitly, and clearly demonstrated high 

22 variability between related species, as well as within larger taxonomic groupings. Mammalian NOEC 

23 values expressed as blood Pb levels were shown to vary by a factor of 8, while avian blood NOECs varied 

24 by a factor of 50 (Buekers et al., 2009). Protective effects of dietary Ca have been found in plants, birds, 

25 and invertebrates.

7.3. Aquatic Ecosystem Effects

7.3.1. Introduction to Aquatic Ecosystem Effects

26 This section of the Pb ISA reviews the new literature since the 2006 Pb AQCD (U.S. EPA, 2006)

27 on the effects of Pb on aquatic systems. The focus is on the effects of Pb, with particular focus on ambient 

28 level, to aquatic organisms including algae, aquatic plants, invertebrates, vertebrates, and other biota with 
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1 an aquatic life stage (e.g., amphibians). The current mean and range of Pb concentrations in surface 

2 waters (mean 0.66 μg Pb/L, range 0.04 to 30 μg Pb/L), sediments (mean 120 μg Pb/g dry weight, range 

3 0.5 to 12,000 μg Pb/g dry weight) and fish tissues (mean 1.03 μg Pb/g dry weight, range 0.08 to 23 μg

4 Pb/g dry weight [whole body]) in the U.S. based on a synthesis of National Water Quality Assessment 

5 (NAWQA) data was reported in the previous 2006 Pb AQCD (U.S. EPA, 2006). The 2006 Pb AQCD 

6 provided an overview of regulatory considerations for water and sediments in addition to consideration of 

7 biological effects and major environmental factors that modify the response of aquatic organisms to Pb 

8 exposure. Regulatory guidelines for Pb in water and sediments have not changed since the 2006 Pb 

9 AQCD and are summarized below with consideration of limited new information on these criteria since 

10 the last review. This section is followed by new information on biogeochemistry, bioavailability and 

11 biological effects of Pb since the 2006 Pb AQCD. 

12 The most recent ambient water quality criteria (AWQC) for Pb were released in 1985 (U.S. EPA, 

13 1985) by the EPA Office of Water which employed empirical regressions between observed toxicity and 

14 water hardness to develop hardness-dependent equations for acute and chronic criterion. These criteria are 

15 published pursuant to Section 304(a) of the Clean Water Act and provide guidance to states and tribes to 

16 use in adopting water quality standards for the protection of aquatic life and human health in surface 

17 water. The ambient water quality criteria for Pb are currently expressed as a criteria maximum 

18 concentration (CMC) for acute toxicity and criterion continuous concentration (CCC) for chronic toxicity

19 (U.S. EPA, 2010). In freshwater, the CMC is 65 μg Pb/L and the CCC is 2.5 μg Pb/L at a hardness of 100 

20 mg/L. In saltwater, these values are 210 μg Pb/L CMC and 8.1 μg Pb/L CCC, respectively. 

21 A draft document intended to update the AWQC for Pb (Great Lakes Environmental Center, 2008)

22 was recently prepared for the EPA. This draft included significantly revised values for both acute and 

23 chronic endpoints that were also based on a hardness equation; the revised chronic AWQC in particular is 

24 much higher than in previous AWQC due to a substantial reduction in the acute to chronic ratio (ACR). 

25 Recent studies have suggested that the ACRs used in the existing criteria documents were too high, 

26 possibly because of the age and size of fish used in those studies (Mebane et al., 2008).

27 The 2006 Pb AQCD summarized two approaches for establishing sediment criteria for Pb based on 

28 either bulk sediment or equilibrium partitioning (Section 7.2.1, Table 7-2 and Section AX7.2.1.4). The 

29 first approach is based on empirical correlations between metal concentrations in bulk sediment and 

30 associated biological effects to derive threshold effect concentrations (TEC) and probable effects 

31 concentrations (PEC) (MacDonald et al., 2000). The TEC/PEC approach derives numeric guidelines to 

32 compare against bulk sediment concentrations of Pb. The other approach in the 2006 Pb AQCD was the 

33 equilibrium partitioning procedure published by the EPA for developing sediment criteria for metals (U.S. 

34 EPA, 2005). The equilibrium partitioning approach considers bioavailability by relating sediment toxicity 

35 to pore water concentration of metals. The amount of simultaneously extracted metal (SEM) is compared 
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1 with the metals extracted via acid-volatile sulfides (AVS) since metals that bind to AVS (such as Pb) 

2 should not be toxic in sediments where AVS occurs in greater qualities than SEM. 

3 Since the publication of the 2006 Pb AQCD both of these methods, for estimating sediment criteria 

4 for metals, have continued to be used and refined. The SEM approach was further refined in the 

5 development of the sediment BLM (Di Toro et al., 2005). The BLM is discussed further in Section 7.3.3. 

6 Comparison of empirical approaches with AVS-SEM in metal contaminated field sediments shows that 

7 samples where either method predicted there should be no toxicity due to metals, no toxicity was 

8 observed in chronic amphipod exposures (J. A. Besser et al., 2009; MacDonald et al., 2009). However,

9 when the relationship between invertebrate habitat (epibenthic and benthic) and environmental Pb 

10 bioaccumulation was investigated, De Jonge et al. (2010) determined that different environmental 

11 fractions of Pb were responsible for invertebrate uptake and exposure. Pb uptake by benthic invertebrate 

12 taxa was not significantly correlated to AVS Pb levels, but rather to total sediment concentrations 

13 (De Jonge et al., 2009). Conversely, epibenthic invertebrate Pb body burdens were better correlated to 

14 AVS concentrations, rather than total Pb sediment concentrations (De Jonge et al., 2010).

15 In the following sections, new information since the 2006 Pb AQCD on Pb in aquatic ecosystems 

16 will be presented. Throughout the sections, brief summaries of conclusions from the 2006 Pb AQCD are 

17 included where appropriate. Section 7.3 is organized to consider uptake of Pb and effects at the species 

18 level, followed by community and ecosystem level effects. Section 7.3.2 considers the biogeochemistry 

19 and chemical effects of Pb in aquatic systems. New research on the bioavailability and uptake of Pb into 

20 aquatic organisms including plants, invertebrates and vertebrates is presented next (Section 7.3.3). Effects 

21 of Pb on the physiology of aquatic fauna and biota (Section 7.3.4) are followed with data on exposure and 

22 response of aquatic organisms (Section 7.3.5). Ecosystem-scale responses are reviewed in Section 7.3.6 

23 followed by a brief consideration of critical loads in aquatic systems (Section 7.3.7), characterization of 

24 sensitivity and vulnerability of ecosystem components (Section 7.3.8) and the effects of Pb on ecosystem 

25 services associated with aquatic environments (Section 7.3.9).

7.3.2. Biogeochemistry and Chemical Effects

26 Quantifying Pb speciation in aquatic environments is critical for determining the toxicity of the 

27 metal to aquatic organisms. As reviewed in the 2006 Pb AQCD and discussed in detail in Section 3.3. of 

28 this assessment (Fate and Transport), the speciation process is controlled by many environmental factors. 

29 Although aerially deposited Pb largely consists of the labile Pb fraction, once the atmospherically-derived 

30 Pb enters surface waters its fate and bioavailability are influenced by Ca concentration, pH, alkalinity, 

31 total suspended solids, and dissolved organic carbon (DOC), including humic acids. In sediments, Pb is 

32 further influenced by the presence of sulfides and iron and manganese oxides. For instance, in neutral to 

acidic aquatic environments, Pb is typically present as PbSO4, PbCl4, Pb2+ 33 , cationic forms of Pb 
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1 hydroxide, and ordinary hydroxide [Pb(OH)2], while in alkaline waters, common forms of Pb include Pb 

2 carbonates [Pb(CO3)] and hydroxides [Pb(OH)2]. In freshwater systems, Pb complexes with inorganic 

OH- and CO3

2− and forms weak complexes with Cl− 3 ; conversely, Pb speciation in seawater is a function of 

chloride concentration and the primary species are PbCl3, PbCO3, PbCl2, and PbCl+ 4 . In many, but not all 

5 aquatic organisms, Pb dissolved in water can be the primary exposure route to gills or other biotic ligands. 

6 The toxicity associated with Pb in the water column or sediment pore waters is directly affected by the 

7 competitive binding of Pb to the anions listed above. 

8 Currently, national and state ambient water quality criteria for Pb attempt to adjust measured 

9 concentrations to better represent the bioavailable free ions, and express the criteria value as a function of 

10 the hardness (i.e., amount of calcium and magnesium ions) of the water in a specific aquatic system. 

11 Models such as the BLM (Paquin et al., 2002) include an aquatic speciation model (WHAM V; see 

12 below) combined with a model of competitive binding to gill surfaces, and provides a more

13 comprehensive method for expressing Pb concentrations at specific locations in terms of the bioavailable 

14 metal While the BLM is not yet used in setting regulatory criteria for Pb in the U.S., its application in risk 

15 analysis has become widely recognized (Fairbrother et al., 2007). Sediment quality criteria have yet to be 

16 adopted by EPA, although an equilibrium partitioning procedure is now available to predict which 

17 sediments have metal concentrations that are not toxic to aquatic organisms (U.S. EPA, 2005). The 

18 approach is based on the ratio of the sum of simultaneously extracted metals and amount of AVS, adjusted 

19 for the fraction of organic carbon present in the sediments, and is reviewed in detail in the 2006 Pb AQCD 

20 (2006). It is important to note that this method cannot accurately predict which sediments are toxic or 

21 which metal is the primary risk driver. 

22 A more detailed understanding of the biogeochemistry of Pb in aquatic systems (both the water 

23 column and sediments) is critical to accurately predicting toxic effects of Pb to aquatic organisms. It 

24 should be recognized, however, that in addition to exposure via sediment and water, chronic exposures to 

25 Pb also include dietary uptake, even though the toxicokinetics of this exposure pathway are not yet well 

26 understood in aquatic organisms and the influence of the bioavailability factors described above is 

27 unknown. Furthermore, changes in environmental factors that reduce the bioaccessible Pb fraction can 

28 result in either sequestration in sediments or subsequent release as mobile, bioaccessible forms. This 

29 section provides updated information about the influence of chemical parameters that affect Pb 

30 bioaccessibility in the aquatic environment (in sediments and the water column).

31 Several models are available for estimating the speciation of dissolved Pb. These models were 

32 tested by Balistrieri and Blank (2008) by comparing the speciation of dissolved Pb in aquatic systems 

33 affected by historical mining activities with that predicted by several models, including Windermere 

34 humic aqueous model (WHAM VI), non-ideal competitive absorption Donnan-type model (NICA35 Donnan), and Stockholm humic model (SHM). Accurate prediction of labile Pb concentrations was 

36 achieved only with SHM, although other metal concentrations were better described by the WHAM 
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1 model. Whereas both WHAM VI and NICA-Donnan predicted that the bulk of Pb contamination would 

2 be complexed with iron, SHM predicted Pb speciation predominantly characterized by both iron and 

3 inorganic Pb complexes. Predicted dynamic Pb concentrations developed with the WHAM VI and NICA4 Donnan methods overestimated Pb concentrations measured using diffusive gradients in thin-films in 

5 Lake Greifen (Switzerland), but underestimated concentrations in Furbach stream (both in the Coeur 

6 D’Alene River Basin in Idaho), indicating that such models may not be able to accurately describe metal 

7 speciation under all environmental conditions (Balistrieri & Blank, 2008). 

8 Quantification of different sediment metal-binding phases, including sulfide, organic C, Fe, and Mn 

9 phases, is important to fully understand the bioaccessible fraction of Pb and the toxicity to benthic 

10 organisms (Simpson & Batley, 2007). However, physical disturbance, pH change, and even the biota 

11 themselves also alter sediment binding or release of Pb. Atkinson et al. (2007) studied the effects of pH on 

12 sequestration or release of Pb from sediments. Although high and circumneutral water pH (8.1 and 7.2) 

13 did not affect the release of sequestered Pb from sediments, lowering the pH to 6 increased the 

14 concentration of Pb in overlying waters from less than 100 μg Pb/L to 200-300 μg Pb/L. Physical 

15 sediment disturbance also increased the amount of sediment-bound Pb released into the aqueous phase. 

16 When Pb-contaminated sediment was physically disturbed, the dissolved oxygen content of the overlying 

17 water was observed to significantly impact Pb mobilization, with greater Pb mobilization at lower 

18 dissolved oxygen levels (3 to 9 mg/L O2) (Atkinson et al., 2007). In addition, although Pb concentrations 

19 in the sediments of a mine-impacted wetland in Hezhang, China, were determined to be strongly 

20 associated with organic/sulfide and residual fractions (e.g., 34 to 82% of total Pb), the presence of aquatic 

21 macrophytes altered the Pb speciation, increasing the fraction of Pb bound to Fe-Mn oxides (42% to 47% 

22 of total Pb) (Bi et al., 2007). This phenomenon was investigated in greater depth by Sundby et al. (2005), 

23 who determined that release of oxygen from macrophyte roots resulted in the oxidation of sediment24 bound Pb, leading to the release of bioaccessible Pb fractions (Sundby et al., 2005).

7.3.2.1. Other Metals

Komjarova and Blust (2008) looked at the effect of the presence of Cd2+ 25 on the uptake of Pb by the 

26 freshwater cladoceran Daphnia magna. While Pb uptake rates were not affected by Cu, Ni or Zn, 

27 enhanced Pb accumulation was observed in the presence of 0.2 μM Cd. The highest Pb concentrations 

28 (0.25 μM) in turn facilitated Cu uptake. Area-specific and whole organism Pb transport rates were 

29 greatest in the mid-intestine. It was concluded that Pb-induced disruptions of ion homeostasis and metal 

30 absorption processes might be a possible explanation of stimulated Pb uptake in the presence of Cd, as 

31 well as the increase in Cu uptake rates provoked by presence of Pb at its highest studied concentration. 

32 Komjarova and Blust (2009b) then considered the effect of Na, Ca and pH on simultaneous uptake of Cd, 

33 Cu, Ni, Pb and Zn. Cd and Pb showed increased uptake rates at high Na concentration. It was thought that 
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1 increased Na uptake rates promoted Pb entrance to the cell. With respect to the effect of pH, reduced 

2 proton competition begins to influence Pb uptake in waters with high pH. A clear suppression of Cd, Ni, 

3 Pb and Zn uptake was observed in the presence of Ca (2.5 mM). Ca has been reported to have a protective 

4 effect in other studies (involving other organisms). The presence of other metals may also affect the 

5 uptake of Pb by fish. At low concentrations, Cd in a Pb-Cd mixture out-competed Pb at gill tissue binding 

6 sites in rainbow trout (Oncorhynchus mykiss), resulting in a less-than additive toxicity when fish were 

7 exposed to both metals in tandem (Birceanu et al., 2008).

7.3.2.2. Biofilm

8 Farag et al. (2007) measured Pb concentrations in various media (water, colloids, sediment, 

9 biofilm) as well as invertebrates and fish collected within the Boulder River watershed. They concluded 

10 that the fraction of Pb associated with Fe-oxides was most frequently transferred to biofilms and the other 

11 biological components of the sampled systems (Farag et al., 2007). Consequently, an increase in the Pb 

12 Fe-oxide fraction could signify a potential increase in the bioaccessible pool of Pb. The authors also noted 

13 that this fraction may promote downstream transport of Pb contamination. Ancion et al. (2010)

14 investigated whether urban runoff metal contaminants could modify biofilm bacterial community 

15 structure and diversity and therefore potentially alter the function of biofilms in stream ecosystems. They 

16 found that accumulation rates for metals in biofilm were maximal during the first day of exposure and 

17 then decreased with time. Equilibrium between metal concentrations in the water and in the biofilm was 

18 reached for all metals after 7-14 days of exposure. The affinity of the biofilm for Pb was, however, much 

19 greater than for Cu and Zn. With respect to recovery, the release of metals was slow and after 14 days in 

20 clean water 35% of Pb remained in the biofilm. By retaining and releasing such metal pollutants, biofilms 

21 may play a key role in determining both the concentration of the dissolved metals in the water column and 

22 the transfer of the metals to invertebrates and fish grazing on them. An enrichment factor of 6,000:1 for 

23 Pb between the biofilm and the water was measured after 21 days exposure to synthetic urban runoff. The 

24 relatively slow release of such metal may greatly influence the transfer of Pb to organisms feeding on the 

25 biofilms. This may be of particular importance during storm events when large amounts of Pb are present 

26 in the urban runoff. It was suggested that biofilms constitute an integrative indicator of metal exposure 

27 over a period of days to weeks.

7.3.2.3. Carbonate

28 An investigation of heavy metal concentrations in an industrially impacted French canal (Deule 

29 canal) indicated that total extractable Pb in sediments ranged from 27 to 10,079 mg Pb/kg, with 52.3% 

30 present in Fe-Mn oxide fractions, 26.9% as organic sulfide fraction, 10.7% in carbonates, and 10.1% in 

31 the residual fraction (Boughriet et al., 2007). The relatively high fraction of Pb associated with carbonates 
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1 was not observed at other sites, as sediments in these areas contained low proportions of carbonates. 

2 Hence, addition of carbonates (either from anthropogenic or natural sources) can significantly impact Pb 

3 speciation in sediments, and potential bioavailability to resident organisms. In addition, increased surface 

4 water carbonate concentrations also reduced the bioaccessible Pb fraction as measured by chronic Pb 

5 accumulation in the fathead minnow, (Pimephales promelas) (Mager et al., 2010), and by Pb toxicity to 

6 fathead minnow and the cladoceran Ceriodaphnia dubia (Mager et al., 2011).

7.3.2.4. Dissolved Organic Matter (DOM)

7 Uptake of Pb by water-column organisms is affected by the concentration of DOM (Mager et al., 

8 2010). The specific composition of DOM has been shown to affect the bioaccessibility of environmental 

9 Pb. Humic acid-rich DOM resulted in decreased free Pb ion concentration when compared to systems 

10 containing DOM with high concentrations of polysaccharides (Lamelas & Slaveykova, 2008). When the 

11 sequestering abilities of various components of DOM were compared, humic acid again was shown to be 

12 most efficient at reducing the Pb free ion concentration, followed by fulvic acid, alginic acid, 

13 polygalacturonic acid, succinoglycan, and xanthan (Lamelas et al., 2005). Lamelas et al. (2009)

14 considered the effect of humic acid on Pb(II) uptake by freshwater algae taking account of kinetics and 

15 cell wall speciation. The uptake flux was described by a Michaelis-Menten type equation. Comparison of 

16 Cu(II), Cd(II) and Pb(II) uptake by green freshwater algae, C. Kesslerii, in the presence of either citric 

17 acid or humic acid was made. The uptake fluxes, percentage adsorbed and percentage internalized for Cu 

18 and Cd were identical in the presence of either citric or humic acid. In contrast, however, there was a ten19 fold increase in the respective values for Pb. The increase in adsorbed Pb was attributed to the increase in 

20 adsorption sites from the adsorbed humic acid on the surface of the algae. Two hypotheses were 

21 considered to explain the increase in internalized Pb and the internalization flux: (1) direct interaction of 

22 Pb-humic acid complexes with the internalization sites, and (2) uptake of Pb(II) after dissociation from 

23 the Pb-humic acid complex. The authors favor the former hypothesis but no evidence is presented for the 

24 proposed ternary Pb-humic acid-internalized site complexes, nor is there an explanation as to why this 

25 behavior is not observed for Cd or Cu.

26 There is evidence, however, that DOC/DOM does not have the same effect on free Pb ion 

27 concentration in marine systems as in freshwater systems. No correlation was observed between DOM 

28 concentration or composition and Pb toxicity when examined using the marine invertebrate Paracentrotus 

29 lividus embryo-larval bioassay (Sanchez-Marin, Santos-Echeandia, et al., 2010). For marine invertebrates, 

30 the presence of humic acid increased both the uptake and toxicity of Pb, despite the fact that a larger 

31 fraction of Pb is complexed with humic acid (25 to 75%). Although the authors could not provide a 

32 precise explanation for this, they theorized that in marine environments, addition of humic acid could 

induce and enhance uptake of Pb via membrane Ca2+ 33 channels (Sanchez-Marin, Slaveykova, et al., 2010). 
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1 This mechanism was observed in the marine diatom Thalassiosira weissflogii, in that humic acids 

2 absorbed to cell surfaces increased metal uptake; however, water column Pb-humic acid associations did 

3 appear to reduce free Pb ion concentrations (Sanchez-Marin, Slaveykova, et al., 2010). Formation of a 

4 ternary complex that is better absorbed by biological membranes was another proposed mechanism that 

5 could describe the increased bioaccessibility to marine invertebrates of Pb bound to humic acid (Sánchez6 Marín et al., 2007). 

7 As little as 1 μmol of humic acid introduced into surface waters was sufficient to reduce Pb uptake 

8 by perennial ryegrass, Lolium perenne, grown in nutrient solution. This resulted from a decrease in the 

9 concentration of the free Pb fraction by several orders of magnitude following complexation with the OM. 

10 Pb content on the root surface was reduced to 8 μmol/g from 20 μmol/g following humic acid addition, 

11 and relative Pb absorption (absorption in the presence of humic acid divided by absorption in the absence 

12 of humic acid) was determined to be approximately 0.2 (Kalis et al., 2006). Conversely, humic acid may 

13 increase the bioaccessible Pb fraction for green algae through formation of a ternary complex that 

14 promotes algal uptake of the metal. Lamelas and Slaveykova (2007) found that aqueous Pb formed 

15 complexes with humic acid, which in turn would become adsorbed to Chlorella kesslerii algal surfaces, 

16 and that the presence of Pb sorbed to humic acid did not interfere with humic acid-algae complexation. 

17 The authors concluded that humic acids bound to algae acted as additional binding sites for Pb, thus 

18 increasing the concentrations associated with the algal fraction (Lamelas & Slaveykova, 2007).

19 Based on the above, the recent literature indicates the existence of a number of deviations from 

20 current models used to predict bioaccessibility of Pb. In marine aquatic systems, for instance, surface 

21 water DOM was found to increase (rather than decrease) uptake of Pb by fish gill structures, potentially 

22 through the alteration of membrane Ca-channel permeability. This phenomenon would not be accurately 

23 predicted by a BLM developed using data from freshwater organisms. Further, in both freshwater and 

24 marine environments, algal biosorption of labile Pb fraction was also increased by humic acid and DOM, 

25 likely through the formation of ternary complexes that increase Pb binding sites on the algal surface. 

26 Although it is unclear whether Pb in this form is available for toxic action on algae, it is likely to 

27 comprise a significant source of dietary Pb for primary consumers. Moreover, the attempted field 

28 verification of freshwater bioaccessibility models was conducted at sites with distinct point-sources of Pb 

29 contamination, and only one model (SHM) was found to adequately predict Pb bioaccessibility.

7.3.3. Bioavailability in Aquatic Systems

30 Bioavailability was defined in the 2006 Pb AQCD as “the proportion of a toxin that passes a 

31 physiological membrane (the plasma membrane in plants or the gut wall in animals) and reaches a target 

32 receptor (cytosol or blood).” In 2007, EPA took cases of bioactive adsorption into consideration and 

33 revised the definition of bioavailability as “the extent to which bioaccessible metals absorb onto, or into, 
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1 and across biological membranes of organisms, expressed as a fraction of the total amount of metal the 

2 organism is proximately exposed to (at the sorption surface) during a given time and under defined 

3 conditions” (Fairbrother et al., 2007). In brief, trace metals, and their complexes, must first diffuse from 

4 the external medium to the surface of the organism (mass transport). Metal complexes may dissociate and 

5 re-associate in the time that it takes to diffuse to the biological surface. To have an effect on the organism, 

6 metals must then react with a sensitive site on the biological membrane (adsorption/desorption processes), 

7 often but not necessarily followed by biological transport (internalization). Any of these processes may be 

8 the rate limiting step for the overall biouptake process. Internalization is, however, the key step in the 

9 overall biouptake process. Although the transport sites often have a high affinity for required metals they 

10 do not always have high selectivity and so a toxic metal may bind to the site of an essential metal with a 

similar ionic radius or co-ordination geometry, e.g., Pb2+, Cd2+ and Zn2+ are similar to Ca2+ 11 . At the 

12 molecular level, there are three major classes of transition metal transporter: P-type ATPases, zinc 

13 regulated transporter/iron-regulated transporter, and natural resistance associated macrophage proteins 

14 (Worms et al., 2006). Of these, natural resistance associated macrophage proteins have been shown to 

15 promote the uptake of various metals including Pb. This type of trace metal transport can be described by 

16 Michaelis-Menten uptake kinetics and equilibrium considerations.

17 According to the 2006 Pb AQCD, Pb adsorption, complexation, chelation, etc., are processes that 

18 alter its bioavailability to different aquatic species, and it was suggested that multiple exposure routes 

19 may be important in determining overall bioavailability of Pb. Given its low solubility in water, 

20 bioaccumulation of Pb by aquatic organisms may preferentially occur via exposure routes other than 

21 direct absorption from the water column, including ingestion of contaminated food and water, uptake 

22 from sediment pore waters, or incidental ingestion of sediment. If uptake and accumulation are 

23 sufficiently faster than depuration and excretion, Pb tissue levels may become sufficiently high to result in 

24 adverse effects (Luoma & Rainbow, 2005). Pb accumulation rates are controlled, in part, by metabolic 

25 rate. Other factors that influence bioavailability of Pb to organisms in aquatic systems are reviewed in 

26 Section 7.3.2. As summarized in the 2006 Pb AQCD, organisms exhibit three Pb accumulation strategies:

27 (1) accumulation of significant Pb concentrations with low rate of loss resulting in substantial 

28 accumulation; (2) balance between excretion and bioavailable metal in the environment; and (3) very low 

29 metal uptake rate without significant excretion, resulting in weak net accumulation (Rainbow, 1996). 

30 Uptake experiments with aquatic plants, invertebrates and vertebrates reviewed in the 2006 Pb AQCD 

31 showed increases in Pb uptake with increasing Pb in solution. The 2006 Pb AQCD findings included 

32 consideration of bioaccumulation in different trophic levels. Pb concentrations were found to be typically 

33 higher in algae and benthic organisms and lower in higher trophic-level consumers. 

34 In this section, recent information on bioavailability and uptake in algae, plants, invertebrates and 

35 vertebrates from marine and freshwater systems are reviewed with summary material from the 2006 Pb 

36 AQCD where appropriate. An overview of the BLM is presented as the most widely used method for 
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1 predicting both the bioaccessible and bioavailable fractions of Pb in the aquatic environment. This is 

2 followed by a discussion of bioavailability in algae, plants, invertebrates and vertebrates. As reviewed by 

3 Wang and Rainbow (2008), aquatic organisms exhibit distinct patterns of metal bioaccumulation. The 

4 authors suggest that the observed differences in accumulation, body burden, and elimination between 

5 species are due to metal biogeochemistry and physiological and biological responses of the organism. The 

6 studies presented below generally support the observations of Wang and Rainbow (2008) that closely 

7 related species can vary greatly in bioaccumulation of Pb and other non-essential metals. 

8 The bioaccumulation and toxicity of Pb to aquatic organisms are closely linked to the 

9 environmental fate of the metal under variable environmental conditions (Section 3.3) as they are highly 

10 dependent upon the relative proportion of free metal ions in the water column. However, information is 

11 lacking on the uptake of Pb through ingestion of Pb-sorbed particles or dietary exposure to biologically12 incorporated Pb. Such routes of exposure are not included in models such as the BLM that predict toxicity 

13 as a function of Pb concentration in the water column. This uncertainty may be greater for Pb than for 

14 other more soluble metals (such as Cu) as a greater proportion of the total mass of Pb in an aquatic 

15 ecosystem is likely to be bound to particulate matter. Therefore, estimating chronic toxicity of Pb to 

16 aquatic receptors may have greater uncertainty than predicting acute effects.

17 In addition to the biogeochemical effects that govern the environmental pool of accessible Pb, 

18 reactions of Pb with biological surfaces and membranes determines the bioavailability and uptake of the 

19 metal by aquatic organisms. The BLM predicts both the bioaccessible and bioavailable fraction of Pb in 

20 the aquatic environment, and can be used to estimate the importance of environmental variables such as 

21 DOC in limiting uptake by aquatic organisms (Alonso-Castro et al., 2009). The BLM integrates the 

22 binding affinities of various natural ligands in surface waters and the biological uptake rates of aquatic 

23 organisms to determine the site-specific toxicity of the bioavailable fraction. In the 2006 Pb AQCD, 

24 limitations of the use of BLM in developing air quality criteria were recognized including the focus of 

25 this model on acute endpoints and the absence of consideration of dietary uptake as a route of exposure. 

26 Atmospheric deposition of Pb to aquatic systems and subsequent effects on ecosystem receptors is likely 

27 characterized as a chronic, cumulative exposure rather than an acute exposure. Recommendations from 

28 the 2006 Pb AQCD included developing both chronic toxicity BLMs and BLMs that consider the dietary 

29 route of Pb uptake. This section reviews the literature from the past 5 years on applications of the BLM to 

30 predicting bioavailability of Pb to aquatic organisms. However, the primary focus of initial BLMs has 

31 been acute toxicity endpoints for fish and invertebrates following gill or cuticular uptake of metals. 

32 Di Toro et al. (2005) constructed BLMs for metals exposure in sediments, surface water, and 

33 sediment pore water to determine how to most accurately predict the toxicity of metals-contaminated 

34 sediments. Results from models were compared with literature-derived acute toxicity values for benthic 

35 and epibenthic invertebrates to establish the accuracy of the developed models. Although the models 

36 tended to overestimate the toxicity of aqueous and sediment-bound Pb in freshwater environments, it was 
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1 determined that the model significantly underestimated Pb toxicity to marine invertebrates (Di Toro et al., 

2 2005). This may be because pore water metal concentrations were not modeled. Consequently, these 

3 results may suggest that either 1) mobilization of Pb concentrations from sediments into pore water is 

4 greater in marine environments, or 2) marine invertebrates are significantly more susceptible to Pb 

5 exposures than are freshwater species.

6 A number of deviations from results predicted by Pb exposure models (such as the BLM) were 

7 documented by Ahlf et al. (2009). They highlighted that uptake of metals by sediment-dwelling bivalves 

8 was significantly greater than predicted, because bivalves accumulate Pb from multiple sources not 

9 included in the model, such as ingestion of algae, bacteria, and colloidal matter. Species-specific dietary 

10 assimilation of ingested particulate-bound metals is also likely to play a role in the toxicity of Pb to 

11 aquatic organisms, yet insufficient data are available to permit modeling of this additional factor (Ahlf et 

12 al., 2009). The authors outlined the need for additional data in developing bioavailability models for 

13 chronic metal exposures.

14 Similarly, although the presence of humic acid is considered to reduce the bioavailable fraction of 

15 metals in surface water, green algae uptake and biosorption of metals, including Pb, was actually 

16 increased by humic acid. The authors determined that humic acid bound to algal surfaces served to 

17 increase the total number of metal binding sites over those afforded solely by the algal surface (Lamelas 

18 & Slaveykova, 2007). This highlights the complexity of modeling chronic metals bioavailability through 

19 multiple exposure routes, as humic acid would decrease gill or cuticular uptake of metals from the water 

20 column, but could potentially enhance dietary exposure by increasing algal metal content. Slaveykova and 

21 Wilkinson (2005) also noted that humic acid is likely to interact with other biological membranes and 

22 alter their permeability to metals, especially in acidic environments. Further, they observed that increased 

23 surface water temperatures can not only increase membrane permeability but also change metabolic rates, 

24 both of which can enhance metals uptake and assimilation; however, this factor is not included in 

25 bioavailability models such as the BLM (Slaveykova & Wilkinson, 2005). Despite this, the authors noted 

26 that, in most cases, the BLM could predict acute metals toxicity with a reasonable degree of accuracy.

27 Veltman et al. (2010) proposed an integration of BLM and bioaccumulation models in order to 

28 more accurately predict metal uptake by fish and invertebrates. Although Pb was not the specific focus of 

29 the paper, calculated metal absorption efficiencies for marine fish species from both BLM and 

30 bioaccumulation models were determined to be highly comparable for Ag, Cd, Cu, and Zn. Authors also 

31 noted that affinity constants for Ca, Cd, Cu, Na, and Zn were highly similar across different aquatic 

32 species, including fish and invertebrates (Veltman et al., 2010). These findings suggest that the BLM can 

33 be integrated with bioaccumulation kinetics to account for both environmental chemical speciation and 

34 biological and physiological factors.
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7.3.3.1. Plants and Algae

1 Aquatic macrophytes and algae can accumulate Pb from either the water column or sediments, 

2 based on their specific microhabitats. For instance, rooted macrophytes may be more likely to accumulate 

3 Pb from sediment sources, while floating macrophytes or algae will take up Pb suspended or dissolved in 

4 the water column. However, significant species-dependent differences in bioaccumulation rates, as well as 

5 concentrations of sequestered metals within different parts of the plants (shoots versus roots), have also 

6 been observed and some authors have concluded that the plant species is a more important determinant of 

7 Pb uptake than is habitat type. Uptake and translocation studies of Pb in plants and algae reviewed in the 

8 2006 Pb AQCD indicated that all plants tend to sequester larger amounts of Pb in their roots than in their 

9 shoots. Recent studies on bioavailability of Pb to plants support the findings of the 2006 Pb AQCD and 

10 provide additional evidence for species-dependent differences in responses to Pb in water and sediments.

11 The microalgae Spirulina platensis was demonstrated to accumulate Pb from the water column, 

12 with 2.7, 6.9, 19, 45 and 145 μg Pb/mg accumulated at aqueous Pb concentrations of 5, 10, 30, 50, and 

13 100 μg Pb/L, following a 10-day incubation period (Arunakumara et al., 2008). Pb concentrations 

14 accumulated by algae appeared to decrease when culture time increased from 2 to 10 days. This may have 

15 occurred as a result of a gradual recovery of growth and an addition of biomass that would have reduced 

16 the concentration of Pb in algal tissue (known as “biodilution”). An aquatic moss, Fontinalis antipyretica, 

17 accumulated up to an average of 3 μmol Pb/g dry weight over a 7-day exposure to 100 μmol Pb, despite 

18 saturation of intracellular Pb concentrations after 5 days of exposure (Rau et al., 2007). Interestingly, 

19 experimentation with concurrent Cu and Pb exposure indicated that the presence of Cu increased the 

20 uptake of Pb by the green algae Chlamydomonas reinhardtii (Z. Z. Chen et al., 2010). The authors noted 

21 that, in the case of Cu-Pb binary exposures, uptake rates of Pb exhibited complex non-linear dynamics in 

22 other aquatic organisms as well.

23 Pb bioaccumulation studies conducted with five species of marine algae, (Tetraselmis chuii, 

24 Rhodomonas salina, Chaetoceros sp., Isochrysis galbana and Nannochloropsis gaditana) demonstrated 

25 that bioaccumulation rates varied with species. I. galbana accumulated the lowest concentrations of Pb 

26 (0.01 and 0.6 pg Pb/cell at water concentrations of 51 and 6,348 μg Pb/L), while Chaetoceros sp. was 

27 observed to be the most efficient Pb bioaccumulator, adsorbing 0.04 and 54 pg Pb/cell at 1.4 and 6,348 μg

28 Pb/L (Debelius et al., 2009).

29 When exposed to water concentrations of up to 100 μmol Pb, floating (non-rooted) coontail plants 

30 (Ceratophyllum demersum) accumulated an average Pb concentration of 1,748 mg Pb/kg after 7 days, 

31 although this was not significantly higher than levels accumulated in the first day of exposure (S. Mishra 

32 et al., 2006). Induction of the antioxidant system improved the tolerance of the aquatic plant Najas indica

33 for bioaccumulated Pb, allowing for increased biomass and the potential to accumulate additional Pb 

34 mass. High Pb accumulation (3,554 mg Pb/kg dry weight tissue following a 7-day exposure to 100 μmol 
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1 Pb) was considered to be a function of plant morphology; as a submerged, floating plant, N. indica

2 provides a large surface area for the absorption of Pb (Singh et al., 2010).

3 Given that atmospherically-derived Pb is likely to become sequestered in sediments, uptake by 

4 aquatic macrophytes is a significant route of Pb removal from sediments, and a potential route for Pb 

5 mobilization into the aquatic food web. The rooted aquatic macrophyte Eleocharis acicularis was 

6 determined to be a hyperaccumulator of Pb in an 11-month bioaccumulation experiment with mine 

7 tailings. When grown in sediments containing 1,930 mg Pb/kg, the maximum concentration of Pb in E.

8 acicularis was determined to be 1,120 mg Pb/kg dry weight. However, calculated BCF’s for Pb were all 

9 less than one, indicating that Pb uptake, although high, was less efficient than for other metals present (Ha 

10 et al., 2009).

11 Aquatic plants inhabiting a wetlands containing an average sediment Pb concentration of 99 

12 mg Pb/kg exhibited variable Pb tissue concentrations, but these do not appear to be related to macrophyte 

13 type (e.g., submerged, floating, emergent, etc.). Consequently, the authors concluded that uptake of Pb by 

14 aquatic plants appears to be dependent on species, at the exclusion of habitat or type. For instance, among 

15 the submerged plant species, Ceratophyllum demersum accumulated the greatest amount of Pb (22 μg/g 

16 dry weight), while Potamogeton malainus tissue contained the least amount of Pb, 2.4 μg/g dry weight 

17 (Bi et al., 2007). Tissues of the floating plants Azolla imbricata and Spirogyra communis were found to 

18 contain 12 and 20 mg Pb/kg dry weight, respectively, while emergent macrophytes Scirpus triqueter and 

19 Alternanthera philoxeroides accumulated 1.4 and 10 mg Pb/kg dry weight. Fritioff and Greger (2006)

20 determined that anywhere from 24–59% of the total Pb taken up by Potamogeton natans aquatic plants 

21 was sequestered in the cell wall fraction, depending on plant tissue and environmental Pb concentration. 

22 More importantly, no translocation of Pb was observed when plant tissues (leaf, stem, root) were exposed 

23 to Pb solutions separately (Fritioff & Greger, 2006). 

24 Dwivedi et al. (2008) reared nine different species of aquatic plants in a fly-ash contaminated 

25 medium containing approximately 7 mg Pb/kg dry weight. Not only did species exhibit different Pb 

26 accumulation efficiencies but they also compartmentalized sequestered Pb differently. The submerged 

27 macrophyte Hydrilla verticillata accumulated the greatest amount of Pb (approximately 180 mg Pb/kg 

28 dry weight tissue), but Pb was sequestered solely in the shoot tissue. In contrast, other plant species 

29 accumulated between 15 and 100 mg Pb/kg dry weight (Ranunculus scloralus and Marsilia quadrifolia) 

30 with the majority compartmentalizing the metal in root tissue, except for C. demersum and M. 

31 quadrifolia, which also utilized shoot tissue for Pb storage (Dwivedi et al., 2008). 

32 Pb concentrations in the root, leaf, and stem tissues of three aquatic plant species were found to 

33 correlate most closely with the concentration of the exchangeable Pb fraction (e.g., the fraction of Pb that 

34 is easily and freely leachable from the sediment). Authors noted that seasonal variations can alter the 

35 amount of Pb present in the exchangeable fraction, and that Pb was more likely than Cd or Cu to remain 
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1 tightly bound to sediments, and therefore the relationship between total sediment Pb and Pb in aquatic 

2 plant tissues was weaker (Ebrahimpour & Mushrifah, 2009). 

3 Lemna sp., a rooted macrophyte, incubated in a water extract of waste ash containing 19 μg Pb/L 

4 accumulated 3.5 mg Pb/kg dry weight over 7 days of exposure. Slight toxic effects, including suppression 

5 of growth, were observed over this exposure period, but this may have been a result of exposures to 

6 multiple metals in the water extract, including Cr, Mn, Cu, and Zn (Horvat et al., 2007). Lemna sp. was 

7 also demonstrated to be effective in the biosorption of Pb from solution, even in the presence of sediments 

8 (1 g per 700 mL water). Over 7 days of exposure to 5 and 10 mg Pb/L, plant biomass was found to 

9 contain an average of 2.9 and 6.6 mg Pb, respectively, versus 0.2 and 0.3 mg in sediment (Hurd & 

10 Sternberg, 2008).

11 Young Typha latifolia, another rooted macrophyte, were grown in 5 and 7.5 mg/L Pb-spiked 

12 sediment for 10 days to determine their value as metal accumulators. Within the exposure period, plants 

13 exposed to the lower concentration were able to remove 89% of Pb, while 84% of the Pb present in the 

14 higher treatment was taken up by T. latifolia. Pb concentrations measured in root and leaf tissue ranged 

15 from 1,365 to 4,867 mg Pb/kg and 272 to 927 mg Pb/kg, respectively, and were higher at the greater 

16 environmental Pb exposure (Alonso-Castro et al., 2009). 

17 Common reeds (Phragmites australis) grown in metal-impacted aquatic environments in Sicily, 

18 Italy, preferentially accumulated Pb in root and rhizome tissues (Bonanno & Lo Giudice, 2010). 

19 Environmental Pb concentrations in water and sediment averaged 0.4 μg Pb/L and 2.7 mg Pb/kg. These 

20 levels yielded root and rhizome concentrations of 17 and 15 mg Pb/kg, respectively, whereas stem and 

21 leaf Pb concentrations were lower (9.9 and 13 mg Pb/kg). These tissue concentrations were significantly 

22 correlated to both water and sediment concentrations (Bonanno & Lo Giudice, 2010). The roots of two 

23 salt marsh species, Sacorconia fructicosa and Spartina maritima significantly accumulated Pb, to 

24 maximum concentrations of 2,870 mg Pb/kg and 1,755 mg Pb/kg, respectively (Caetano et al., 2007). 

25 Roots had similar isotopic signature to sediments in vegetated zones indicating that Pb uptake by plants 

26 reflects the input in sediments. Conversely, the semi-aquatic plant Ammania baccifera, grown in mine 

27 tailings containing 35 to 78 mg Pb/kg, did not accumulate analytically detectable levels of Pb in either 

28 root or shoot tissues, despite the fact that other metals (Cu, Ni, Zn) were bioaccumulated (Das & Maiti, 

29 2007). This would indicate that at low/moderate environmental Pb concentrations, some plant species 

30 may not bioaccumulate significant (or measurable) levels of Pb.

31 The average concentration of Pb in the tissues of rooted aquatic macrophytes (Callitriche verna, P. 

32 natans, C. demersum, Polygonum amphibium, Veronica beccabunga) collected from two metals-polluted 

33 streams in Poland (average sediment concentration 38 to 58 mg Pb/kg) was less than 30 mg Pb/kg. Pb 

34 bioaccumulation in plants was significantly correlated with sediment Pb concentrations (Samecka35 Cymerman & Kempers, 2007). A similar significant correlation was established between reed sweet grass 
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1 root Pb concentration and sediment Pb concentrations, yielding BCFs ranging from 0.5 to 1.5, with an 

2 average BCF of 0.9 (Skorbiowicz, 2006).

3 Pb tissue concentrations of aquatic plants P. australis and Ludwigia prostrata collected from 

4 wetlands containing an average of 52 mg Pb/kg in surficial sediments were predominantly in root tissues, 

5 indicating poor translocation of Pb from roots. In the former, Pb decreased from an average of 37 mg

6 Pb/kg in roots to 17, 14, and 12 mg Pb/kg in rhizome, stem and leaf tissues, respectively, while L. 

7 prostrata Pb tissue concentrations decreased from 77 mg Pb/kg in fibrous root to 7 and 43 mg Pb/kg in 

8 stem and leaf tissues (H. J. Yang et al., 2008). The authors proposed that this diminished transfer ability 

9 explained the relatively low BCF’s for Pb uptake in these two species, when compared with those of other 

10 metals (Table 7-3). 

11 Despite no significant seasonal effect on surface water Pb concentrations, shining pondweed 

12 (Potamogeton lucens), a rooted aquatic macrophyte grown in an urbanized metal-contaminated lake in 

13 Turkey, exhibited seasonal alterations in Pb tissue concentrations. Average measured water Pb 

14 concentrations were 28 μg Pb/L in spring, 27 μg Pb/L in summer, and 30 μg Pb/L in autumn. Over this 

15 same time period, root tissue Pb concentrations significantly increased from 6 mg Pb/kg dry weight in 

16 spring, to 9 mg Pb/kg dry weight in summer, and to 10 mg/kg dry weight in autumn (Duman et al., 2006). 

17 No differences were detected in stem Pb concentrations between spring and summer (approximately 4 mg

18 Pb/kg dry weight), but stem Pb concentrations were found to be significantly higher in autumn (6 mg

19 Pb/kg dry weight). In the same system, P. australis plants accumulated the most Pb during winter: 103, 

20 23, and 21 mg Pb/kg dry weight in root, rhizome, and shoot tissue, respectively, in sediments containing 

21 13 mg Pb/kg dry weight. By contrast, Schoenoplectus lacustris accumulated maximum rhizome and stem 

22 Pb concentrations of 5.1 and 7.3 mg Pb/kg dry weight in winter, but sequestered the greatest amount of Pb 

23 in root tissues during the spring (30 mg Pb/kg dry weight) at a comparable sediment concentration, 18 mg 

24 Pb/kg dry weight (Duman et al., 2007). The authors suggest that this indicated that metal uptake was 

25 regulated differently between species.

26 Tree species that inhabit semi-aquatic environments have also been shown to absorb Pb from Pb27 contaminated sediments. Bald-cypress trees (Taxodium distichum) growing in sediments of a refinery28 impacted bayou in Louisiana accumulated significantly greater amounts of Pb than did trees of the same 

29 species growing in bankside soil, despite the lower Pb concentrations of sediments. Bankside soils 

30 contained greater than 2,700 mg Pb/kg versus concentrations of 10 to 424 mg Pb/kg in sediments, yet Pb 

31 concentrations in trees averaged 4.5 and 7.8 mg Pb/kg tissue, respectively (Devall et al., 2006). The 

32 authors theorized that Pb was more readily released from sediments and that soil dispersion to the swamp 

33 sediments provides additional, if periodic, loads of Pb into the system. 

34 BCFs for Pb in root tissue from mangrove tree species range between 0.09 and 2.9, depending on 

35 the species and the habitat, with an average BCF of 0.84. The average BCF for mangrove species leaf 

36 tissue was considerably less (0.11), as these species are poor translocators of Pb (MacFarlane et al., 2007). 
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1 In contrast, willow seedlings planted in Pb-contaminated sediment were more effective at removing Pb 

2 from the media than a diffusive gradient in thin film technique predicted (Jakl et al., 2009). The authors 

3 proposed that the plant’s active mobilization of nutrients from soil during growth also resulted in 

4 increased Pb uptake and sequestration.

5 Given that sediments are a significant sink for Pb entering aquatic systems, it is not surprising that 

6 rooted macrophytes bioaccumulate significant quantities of the metal. Although there are some 

7 similarities to Pb accumulation observed in terrestrial plants (e.g., preferential sequestration of the metal 

8 in root tissue), Pb appears to be more bioavailable in sediment than it is in soil. This may be a result of 

9 differences in plant physiology between aquatic and terrestrial plants (e.g., more rapid growth or more 

10 efficient assimilation of nutrients and ions from a water-saturated medium). While rooted macrophytes are 

11 likely to be chronic accumulators of Pb sequestered in sediments, aerial deposition of Pb into aquatic 

12 systems may result in pulsed inputs of labile Pb that would be available for uptake by floating 

13 macrophytes and algae. 

14 Reported values for BCF’s in aquatic plants from the 2006 Pb AQCD range from 840 to 20,000 (, 

15 Table AX7-2.3.1U.S. EPA, 2006). Duckweed (Lemna minor) had BCF values ranging from 840 to 3,560 

16 depending on the method of measurement. Additional BCF’s established for aquatic plants since the 2006 

17 Pb AQCD are summarized in Table 7-3 and include data on field-collected plants as well as BCF’s 

18 obtained from laboratory exposures.

Table 7-3. Bioconcentration factors for Pb in aquatic plants 

Species BCF Test conditions Reference

Typha latifolia 649 10 days, Pb nitrate-spiked water Alonso-Castro et al. (2009)

Spirulina platensis 1500 10 days, Pb nitrate-spiked water Arunakumara et al. (2008)

Ceratophyllum demersum 0.2 Field-collected plants Bi et al. (2007)

Spirogyra communis 0.2 Field-collected plants Bi et al. (2007)

Phragmites australis 6.4 Field-collected plants Bonanno and Lo Giudice (2010)

Taxodium distichum 0.02 Field-collected tissue Devall et al. (2006)

Hydrilla verticillata 26 Field-collected plants Dwivedi et al. (2008)

Eleocharis acicularis 0.8 11 mo, field-collected sediment Ha et al. (2009)

Lemna sp. 0.01 7 days, ash water extract Horvat et al. (2007)

Mangrove species 0.8 Field-collected tissue MacFarlane et al. (2007)

Glyceria aquatica 0.9 Field-collected tissue (roots) Skorbiowicz (2006)

Phragmites australis 0.7 Field-collected plants Yang et al. (2008)

Ludwigia prostrata 1.5 Field-collected plants Yang et al. (2008)

7.3.3.2. Invertebrates

19 Uptake and subsequent bioaccumulation of Pb in marine and freshwater invertebrates varies greatly 

20 between species and across taxa as previously characterized in the 2006 Pb AQCD. This section expands 

21 on the findings from the 2006 Pb AQCD on bioaccumulation and sequestration of Pb in aquatic 
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1 invertebrates. In the case of invertebrates, Pb can be bioaccumulated from multiple sources, including the 

2 water column, sediment, and dietary exposures, and factors such as proportion of bioavailable Pb, life 

3 stage, age, and metabolism can alter the accumulation rate. In this section, new information on Pb uptake 

4 from sediments by invertebrates will be considered, followed by a discussion on dietary and water routes 

5 of exposure and factors that influence species-specific Pb tissue concentrations such as invertebrate 

6 habitat and functional feeding group. 

7 The 2006 Pb AQCD summarized studies of uptake of Pb from sediment by aquatic invertebrates 

8 and noted that sediment pore water, rather than bulk sediment, is the primary route of exposure. However, 

9 a recent study suggests that in the midge, Chironomus riparius, total metal concentrations in bulk 

10 sediment are better predictors of metal accumulation than dissolved metal concentrations in sediment pore

11 water based on bioaccumulation studies using contaminated sediments from six different sites (Roulier et 

12 al., 2008). Vink (2009) studied six river systems and found that, for a range of metals, uptake by benthic 

13 organisms (the oligochaete, Limnodrilus (Family Tubificidae) and the midge, C. riparius) from the 

14 sediment pore water (as compared with surface water) was observed only occasionally, and solely for Pb. 

15 The physiological mechanisms of Pb uptake are still unclear but it is suggested that uptake and 

16 elimination of Pb obey different mechanisms than for other heavy metals. Additionally, Metian et al. 

17 (2009) showed that king scallop (Pecten maximus) exhibited low bioaccumulation efficiency of Pb from 

18 spiked sediment. 

19 The 2006 Pb AQCD recognized the potential importance of the dietary uptake pathway as a source 

20 of Pb exposure for invertebrates. Specifically, in a study with the freshwater amphipod Hyalella azteca, 

21 dietary exposure was found to contribute to the chronic toxicity of Pb, while acute toxicity was unaffected 

22 (J. M. Besser et al., 2004). Since the 2006 Pb AQCD, additional studies have considered the relative 

23 importance of water and dietary uptake of Pb in aquatic invertebrates. A stable isotope technique was used 

24 to simultaneously measure uptake of environmentally relevant concentrations of Pb (0.05 μmol Pb in the 

25 water column) by the freshwater cladoceran D. magna directly from water and through food, the green 

26 algae Pseudokirchneriella subcapitata. (Komjarova & Blust, 2009a). D. magna accumulated the metal 

27 from both sources, but the relative proportion of uptake from each source changed over the exposure 

28 period. After the first day of exposure, 12% of accumulated Pb was determined to have been absorbed 

29 from dietary (algal) sources, but this percentage decreased by day four of exposure to 4%. Pb absorbed 

30 from water exposure only resulted in Daphnia body burdens of approximately 300 μmol Pb/kg dry 

31 weight, and was similar to the amount absorbed by algae (Komjarova & Blust, 2009a).

32 Stable isotope analysis was to used measure uptake and elimination simultaneously in netspinning 

33 caddisfly larvae (Hydropsyche sp.) exposed to aqueous Pb concentrations of 0.2 to 0.6 μg Pb/L (Evans et 

34 al., 2006). The measured uptake constant for Pb in this study was 7.8 g/dry weight·day and the 

elimination rate constant of 0.15 d-1 35 for Pb-exposed larvae was similar in both presence and absence of 

36 the metal in the water. Caddisflies accumulated significant amounts of the metal over 18 days of 
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1 exposure. Measured tissue concentrations ranged from approximately 15 to 35 μg Pb/g. Hydropsychid Pb 

2 BCF’s ranged from 41 to 65, and averaged 54, indicating a relatively high accumulation rate when 

3 compared to other metals tested (average BCF of 17 for Cd, 7.7 for Cu, and 6.3 for Zn)(Evans et al., 

4 2006). In larvae of the mosquito, Culex quinquefasciatus, exposed to 100 μg Pb/L for seven days the BCF 

5 was 62 (Kitvatanachai et al., 2005).

6 In a comparison of dietary and waterborne exposure as sources of Pb to aquatic invertebrates, no 

7 correlation between Pb uptake and dietary exposure was observed in the amphipod H. azteca (Borgmann 

et al., 2007). Metian et al. (2009) investigated the uptake and bioaccumulation of 210 8 Pb in Chlamys varia

9 (variegated scallop) and king scallop to determine the major accumulation route (seawater or food) and 

10 then assess subsequent tissue distribution. Dietary Pb from phytoplankton in the diet was poorly 

assimilated (<20%) while more than 70% of Pb in seawater was retained in the tissues. In seawater, 210 11 Pb 

12 was accumulated more rapidly in C. varia than P. maximus and soft tissue distribution patterns differed 

13 between the species. C. varia accumulated Pb preferentially in the digestive gland (50%) while in P. 

14 maximus, Pb was equally distributed in the digestive gland, kidneys, gills, gonad, mantle, intestine, and 

adductor muscle with each tissue representing 12-30% of 210 15 Pb body load. An additional test with Pb16 spiked sediment with P. maximus showed low bioaccumulation efficiency of Pb from sediment. 

17 With the exception of the above-mentioned study with scallops (Metian et al., 2009) recent reports 

18 on Pb distribution generally supports the findings of the 2006 Pb AQCD that Pb is primarily sequestered 

19 in the gills, hepatopancreas, and muscle. Uptake of Pb by the crayfish (Cherax destructor) exposed to 

20 5,000 μg Pb/L for 21 days resulted in accumulation at the highest concentration in gill, followed by 

21 exoskeleton < mid-gut gland < muscle < hemolymph (Morris et al., 2005). The gills were the main sites 

22 of Pb accumulation in Pinctada fucata (pearl oyster) followed by mantle, in 72-hour exposures to 103.5 

23 μg Pb/L (Jing et al., 2007). Following a 10 day exposure to 2,500 μg Pb/L as Pb nitrate, accumulation of 

24 Pb was higher in gill than digestive gland of Mytilus edulis: after a 10 day depuration, Pb content was 

25 decreased in the gills and digestive gland of these mussels (Einsporn et al., 2009). In blue crabs, 

26 Callinectes sapidus, collected from a contaminated and a clean estuary in New Jersey, U.S., the 

27 hepatopancreas was found to be the primary organ for Pb uptake (Reichmuth et al., 2010). Body burden 

28 analysis following 96 hour exposure to 50, 100 and 500 μg Pb/L in the freshwater snail Biomphalaria 

29 glabrata indicated that bioaccumulation increased with increasing concentrations of Pb and the highest 

30 levels were detected in the digestive gland (Ansaldo et al., 2006). 

31 There is more information now on the cellular and subcellular distribution of Pb in invertebrates 

32 than there was at the time of writing the 2006 Pb AQCD. Specifically, localization of Pb at the 

33 ultrastructural level has been assessed in the marine mussel (M. edulis) through an antibody-based 

34 detection method (Einsporn et al., 2009; Einsporn & Koehler, 2008). Dissolved Pb was detected mainly 

35 within specific lysosomal structures in gill epithelial cells and digestive gland cells and was also localized 

36 in nuclei and mitochondria. Transport of Pb is thought to be via lysosomal granules associated with 
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1 hemocytes (Einsporn et al., 2009). In the digestive gland of the variegated scallop, Pb was also mainly 

2 bound to organelles, i.e., 66% of the total metal burden (Bustamante & Miramand, 2005). In the digestive 

3 gland of the cephalopod Sephia officinalis, (cuttlefish) most of the Pb was found in the organelles (62%) 

4 (Bustamante et al., 2006). In contrast, only 7% of Pb in the digestive gland of the octopus (Octopus 

5 vulgaris) was associated with the fraction containing nuclei, mitochondria, lysosome and microsomes: the 

6 majority of Pb in this species was found in cytosolic proteins (Raimundo et al., 2008).

7 Since the publication of the 2006 Pb AQCD, additional factors have been considered that may 

8 affect Pb uptake in aquatic organisms. Pb tissue concentrations fluctuated seasonally in mussels (Mytilus 

9 galloprovincialis) harvested near Istanbul, Turkey (Ozden, 2008). Tissue Pb concentrations were lowest 

10 during the summer months (average of 0.9 mg Pb/kg), followed by spring, autumn and winter (1.3, 1.4, 

11 and 1.6 mg Pb/kg, respectively). The authors speculated that the slight seasonal differences indicate that 

12 bioavailability of the metal may be related to seasonal changes in surface water or sediment chemistry. 

13 Additionally, alterations in growth over the year, as well as different rates of Ca uptake may have 

14 impacted Pb bioaccumulation rates. When the relationship between invertebrate habitat (epibenthic and 

15 benthic) and environmental Pb bioaccumulation was investigated, De Jonge et al. (2010) determined that 

16 different environmental fractions of Pb were responsible for invertebrate uptake and exposure. Pb uptake 

17 by benthic invertebrate taxa was not significantly correlated to AVS Pb levels, but rather to total sediment 

18 concentrations (De Jonge et al., 2009). Conversely, epibenthic invertebrate Pb body burdens were better 

19 correlated to AVS concentrations, rather than total Pb sediment concentrations (De Jonge et al., 2010). 

20 Reported BAF values for Pb in aquatic invertebrates from the 2006 Pb AQCD ranged from 499 to 

21 3,670 [See Table AX7-2.3.2 (U.S. EPA, 2006)]. Since the publication of the 2006 Pb AQCD, additional 

22 BAF values have been established for invertebrates in field studies which tend to be higher than BCF

23 values calculated in laboratory exposures (Casas et al., 2008; Gagnon & Fisher, 1997) (Table 7-4). A 

24 complicating factor in establishing BAF values is that laboratory studies usually assess uptake in water25 only or sediment only exposures while field studies take into account dietary sources of Pb as well as 

26 waterborne Pb resulting in BAF values that are frequently 100-1,000 times larger than BCF values for the 

27 same metal and species (DeForest et al., 2007). Mean Pb levels in both predatory and grazing 

28 zooplanktonic species in El Niagra reservoir, (in Aguascalientes, Mexico) were used to calculate BAF

29 values (Rubio-Franchini et al., 2008) to assess biomagnification of Pb. The BAF of the predatory rotifer 

30 Asplanchna brigthwellii (BAF 49,300) was up to four times higher than the grazing cladocerans Daphnia 

31 similis (BAF 9,022) and Moina micrura (BAF 8,046). Limpet (Patella sp.) from the Lebanese Coast had 

32 Pb BAF values ranging from 2,500 to 6,000 and in the same field study mussel (Brachidontes variabilis) 

33 Pb BAF values ranged from 7,500-8,000 (Nakhle et al., 2006).
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Table 7-4. Bioaccumulation factors for Pb in aquatic invertebrates 

Species BAF Test conditions Reference

Hydropsyche sp. (caddisfly 

larvae) 54 18 days, Pb laboratory exposure Evans et al. (2006)

Culex Quinquefasciatus

(mosquito) 62 7 days, Pb laboratory exposure Kitvatanachai et al. (2005)

Daphnia similis (zooplankton) 9,022 Field-collected Rubio-Franchini et al. (2008)

Moina micrura (zooplankton) 8,046 Field-collected Rubio-Franchini et al. (2008)

Asplanchna brigthwellii (rotifer) 49,344 Field-collected Rubio-Franchini et al. (2008)

Patella sp. (limpet) 6,000 Field-collected Nakhle et al.(2006)

Brachidontes variabilis (mussel) 8,000 Field-collected Nakhle et al.(2006)

1 Recently, several studies have attempted to establish biodynamic exposure assessments for various 

2 contaminants. In an in situ metal kinetics field study with the mussel M. galloprovincialis, simultaneous 

3 measurements of metal concentrations in water and suspended particles with mussel biometrics and 

4 physiological indices were conducted to establish uptake and excretion rates in the natural environment 

5 (Casas et al., 2008). A mean log of 4.3 of the metal concentration in mussels (ng Pb/kg wet flesh 

6 weight)/metal concentration in water (ng Pb/L) was determined for Pb for M. galloprovincialis in this 

7 study based on the rate constants of uptake and efflux in a series of transplantation experiments between 

8 contaminated and clean environments. Equilibrium concentrations of Pb in mussels leveled out at 

9 approximately 30 days with a concentration of 6.7 mg Pb/kg.

7.3.3.3. Vertebrates

10 Uptake of Pb by vertebrates considered here includes data from fish species as well as a limited 

11 amount of new information on amphibians and aquatic mammals. In fish, Pb is taken up from water via 

12 the gills and from food via ingestion. Amphibians and aquatic mammals are exposed to waterborne Pb 

13 primarily through dietary sources. In the 2006 Pb AQCD, dietary Pb was recognized as a potentially 

14 significant source of exposure to all vertebrates since Pb adsorbed to food, particulate matter and 

15 sediment can be taken up by aquatic organisms. 

16 Since the 2006 Pb AQCD, tissue accumulation of Pb via gill and dietary uptake has been further 

17 characterized in vertebrates, and new techniques such as the use of stable isotopes have been applied to 

18 further elucidate bioaccumulation of Pb. For example, patterns of uptake and subsequent excretion of Pb 

19 in fish as measured by isotopic ratios of Pb in each tissue can determine whether exposure was due to 

20 relatively long term sources (which favor accumulation in bone) or short term sources (which favors 

21 accumulation in liver) (Miller et al., 2005). New information since the 2006 Pb AQCD on uptake of Pb by 

22 fish from water is reviewed below, followed by studies on dietary uptake as a route of Pb exposure. Next, 

23 tissue accumulation patterns in fish species are reported with special consideration of the anterior intestine 

24 as a newly identified target of Pb from dietary exposures. New data on uptake studies in marine fishes are 
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1 presented followed by new evidence for additional Pb detoxification mechanisms in fish. Finally, new 

2 data on uptake and tissue distribution of Pb in amphibians and aquatic mammals are presented. 

Freshwater Fish

3 Pb uptake in freshwater fish is accomplished largely via direct uptake of dissolved Pb from the 

4 water column through gill surfaces and by ingestion of Pb-contaminated diets. According to the data 

5 presented in the 2006 Pb AQCD, accumulation rates of Pb are influenced by both environmental factors, 

6 such as water pH, DOC, and Ca concentrations, and by species-dependent factors, such as metabolism, 

7 sequestration, and elimination capacities. The effects of these variables on Pb bioaccumulation in fish are 

8 largely identical to the effects observed for invertebrates (discussed above). 

9 Since the publication of the 2006 Pb AQCD, multiple studies on uptake of Pb from water by 

10 fathead minnow have been conducted. Spokas et al. (2006) showed that Pb accumulates to the highest 

11 concentration in gill when compared to other tissues over a 24 day exposure. This pattern was also 

12 observed in larval fathead minnows exposed to 26 μg Pb/L for 10-30 days, where gill exhibited the 

13 highest Pb concentration compared to carcass, intestine, muscle and liver (Grosell et al., 2006a). In the 

14 larval minnows, Pb concentration in the intestine exhibited the highest initial accumulation of all tissues 

15 on day 3 but then decreased for the remainder of the experiment while concentrations in the other organs 

16 continued to increase. By day 30, gill tissue exhibited the highest Pb concentration (approximately 120 μg

17 Pb/g), followed by whole fish and carcass (whole fish minus gill, liver, muscle and intestine) Pb 

18 concentrations (approximately 70 to 80 μg Pb/g). However, in considering overall internal Pb body 

19 burden, nearly 80% was largely concentrated in the bone tissue, while gill contributed <5%.

20 In another study with fathead minnow, chronic (300 day) exposure to 120 μg Pb/L resulted in 

21 accumulation of approximately 200 nmol Pb/g tissue, although this number was decreased from initial 

22 body burdens of greater than 500 nmol Pb/g at test initiation (Mager et al., 2010). Tissue distribution at 

23 300 days was consistent with Grosell et al. (2006a) with highest concentration in gill, followed by kidney, 

24 anterior intestine, and carcass. Addition of humic acid and carbonate both independently reduced uptake 

25 of Pb in these fish over the exposure time period. Interestingly, fathead minnow eggs collected daily 

26 during 21 day breeding assays that followed the chronic exposure described above accumulated similar 

27 levels of Pb from the test solutions regardless of Pb concentration or water chemistry (e.g., addition of 

28 humic acid and carbonate) (Mager et al., 2010). Direct acute exposure from water rather than parental 

29 transfer accounted for the majority of the Pb accumulation in eggs. Similarly, exposure of fish to 157 nM 

30 Pb in base water for 150 days resulted in fathead minnow whole body concentrations of approximately 

31 150 nmol Pb/g tissue, with the most rapid accumulation rate occurring within the first 10 days of 

32 exposure, followed by an extended period of equilibrium (Mager et al., 2008). In this same study, fish 

were tested in two additional treatments: 177 nM Pb in hard water (Ca2+ 33 500 μM) or 187 nM Pb in humic 
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1 acid supplemented water (4 mg/L). While the addition of humic acid significantly reduced Pb 

2 bioaccumulation in minnows (to approximately 50 nmol Pb/g on a whole body basis), Ca sulfate did not 

3 alter uptake. Despite the fact that Ca-mediated Pb toxicity occurred in larval fathead minnow, there was 

4 no concurrent effect on whole body Pb accumulation. 

5 Uptake studies in other teleosts of Pb from freshwater have generally followed the pattern of 

6 uptake described above for fathead minnow. In the cichlid, Nile tilapia (Oreochromis niloticus) Pb 

7 accumulated significantly in gill (45.9 +34.4 μg/g dry weight at 10 μM, 57.4 +26.1 μg/g dry weight at 20 

8 μM) and liver (14.3 μg/g dry weight at 10 μM and 10.2 μg/g dry weight at 20 μM) during a 14-day 

9 exposure (Atli & Canli, 2008). In rainbow trout exposed to 100 μg Pb/L for 72 hours, the accumulation in 

10 tissues was gill>kidney>liver and this same pattern was observed in all concentrations tested (100-10,000 

11 μg Pb/L) (Suicmez et al., 2006). Sloman et al. (2005) investigated the uptake of Pb in dominant12 subordinate pairings of rainbow trout exposed to 46 μg/L or 325 μg/L Pb-nitrate for 48 hours. Significant 

13 Pb accumulation in gill, liver and kidney was only observed in the highest concentration. Pb accumulated 

14 preferentially in liver of subordinate trout when compared to dominant trout. Brown trout (Salmo trutta) 

15 exposed to aqueous Pb concentrations ranging from 15 to 46 μg Pb/L for 24 days accumulated 6 μg Pb/g 

16 dry weight in gill tissue and Pb concentrations in liver tissue reached 14 μg Pb/g dry weight. Interestingly, 

17 Pb in gill tissue peaked on day 11 and decreased thereafter, while liver Pb concentrations increased 

18 steadily over the exposure period, which may indicate translocation of Pb in brown trout from gill to liver 

19 (Heier et al., 2009).

20 Zebrafish (Danio rerio) Pb uptake rates from media containing 0.025 μmol Pb was significantly 

21 increased by neutral pH (versus a pH of 6 or 8) and by Ca concentrations of 0.5 mmol; uptake rate of Pb 

22 was increased from 10 L/kg·h to 35 L/kg·h by increasing pH from 6 to 7, and from 20 L/kg·h to 35 L/kg·h

23 by increasing Ca concentration from 0.1 mmol to 0.5 mmol (Komjarova & Blust, 2009c). This study also 

24 demonstrated that zebrafish gill tissue is the main uptake site for the metal, as Pb concentrations in these 

25 tissues were up to eight times as high as that in other tissues. 

26 The Eurasian silver crucian carp (Carrasius auratus) collected from a pond containing an average 

27 of 1,600 mg Pb/kg in the sediments exhibited increased Pb body burdens ranging from 12 to 68 mg Pb/kg 

28 dry weight (Khozhina & Sherriff, 2008). Pb was primarily sequestered in skin, gill, and bone tissues, but 

29 was also detected at elevated levels in muscle and liver tissues, as well as in eggs. Two fish species 

30 (Labeo rohita and Ctenopharyngodon idella) collected from the Upper Lake of Bhopal, India with 

31 average Pb concentration of 0.03 mg Pb/L in the water column contained elevated Pb tissue 

32 concentrations (Malik et al., 2010). However, while liver and kidney Pb concentrations were similar 

33 between the two species (1.5 and 1.1 μg Pb/g tissue and 1.3 and 1.0 μg Pb/g tissue for C. idella and L. 

34 rohita, respectively), they accumulated significantly different amounts of Pb in gill and muscle tissues. C. 

35 idella accumulated more than twice the Pb in these tissues (1.6 and 1.3 μg Pb/g) than did L. rohita (0.5 

36 and 0.4 μg Pb/g). 
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1 The studies reviewed above generally support the conclusions of the 2006 Pb AQCD that the gill is 

2 a major site of Pb uptake in fish and that there are species-dependent differences in the rate and pattern of 

3 Pb accumulation. As indicated in the 2006 Pb AQCD, exposure duration can be a factor in Pb uptake from 

4 water. In a 30 day exposure study, Nile tilapia fingerlings had a three-fold increase in Pb uptake at the gill 

5 on day 30 compared to Pb concentration in gill at day 10 and 20 (Kamaruzzaman et al., 2010). In addition 

6 to uptake at the gill, a time-dependent uptake of Pb into kidney in rainbow trout exposed to 570 μg Pb/L 

7 for 96 hours (Patel et al., 2006) was observed. Pb was accumulated preferentially in the posterior kidney 

8 compared to the anterior kidney. A similar pattern was observed by Alves and Wood (2006) in a dietary 

9 exposure. In catla (Catla catla) fingerlings, the accumulation pattern of Pb was kidney > liver > gill > 

10 brain > muscle in both 14 day and 60 day Pb exposures (Palaniappan et al., 2009). In multiple studies 

11 with fathead minnow at different exposure durations, tissue uptake patterns were similar at 30 days 

12 (Grosell et al., 2006a) and 300 days (Mager et al., 2010). In the larval minnows, Pb concentration in the 

13 intestine exhibited the highest initial accumulation of all tissues on day 3 but then decreased for the 

14 remainder of the experiment while concentrations in the other organs continued to increase (Grosell et al., 

15 2006a). By day 30, gill tissue exhibited the highest Pb concentration followed by whole fish and carcass 

16 (whole fish minus gill, liver, muscle and intestine). The most rapid rate of Pb accumulation in this species 

17 occurs within the first 10 days of exposure (Mager et al., 2008). African catfish (Clarias gariepinus) 

18 exposed to aqueous Pb concentrations of 50 to 1,000 μg Pb/L (as Pb nitrate) for 4 weeks accumulated 

19 significant amounts of Pb in heart (520-600 mg Pb/kg), liver (150-242 mg Pb/kg), and brain (120-230 mg

20 Pb/kg) tissues (Kudirat, 2008). Doubling the exposure time to 8 weeks increased sequestration of Pb in 

21 these tissues as well as in skin (125-137.5 mg Pb/kg) and ovaries (30-60 mg Pb/kg).

22 Since the publication of the 2006 Pb AQCD, several studies have focused on dietary uptake of Pb 

23 in teleosts. Alves et al. (2006) administered a diet of three concentrations of Pb (7, 77 and 520 μg Pb/g 

24 dry weight) to rainbow trout for 21 days. Doses were calculated to be 0.02 μg Pb/day (control), 3.7 

25 μg Pb/day (low concentration), 39.6 μg Pb/day (intermediate concentration) and 221.5 μg Pb/day (high 

26 concentration). Concentrations in the study were selected to represent environmentally relevant 

27 concentrations in prey. After 21 days exposure to the highest concentration, Pb accumulation was greatest 

28 in the intestine, followed by carcass, kidney and liver leading the authors to hypothesize that the intestine 

29 is the primary site of exposure in dietary uptake of Pb. All tissues, (gill, liver, kidney, intestine, carcass) 

30 sequestered Pb in a dose-dependent manner. The gills had the greatest concentration of Pb on day 7(8.0 

31 μg Pb/g tissue wet weight) and this accumulation decreased to 2.2 μg Pb/g tissue wet weight by the end of 

32 the experiment suggesting that the Pb was excreted or redistributed (Alves et al., 2006). Furthermore, 

33 with increasing dietary concentrations, the percentage of Pb retained in the fish decreased. Additionally, 

34 in this study red blood cells were identified as a reservoir for dietary Pb. Plasma did not accumulate 

35 significant Pb (0.012 μg Pb g wet weight in the high dose), however, Pb was elevated in blood cells (1.5 

36 μg Pb g wet weight in the high dose) (Alves et al., 2006).
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1 Additional studies have supported the anterior intestine as a target for Pb in fish. Nile tilapia 

2 exposed to dietary Pb for 60 days (100, 400, and 800 μg Pb/g dry weight) accumulated the greatest 

3 concentration of Pb in the intestine, followed by the stomach and then the liver (Dai, Du, et al., 2009). 

4 The amount of Pb in tissue increased with increasing dietary Pb concentration. In a 42 day chronic study 

5 of dietary uptake in rainbow trout, fish fed 50 or 500 μg Pb/g, accumulated Pb preferentially in anterior 

6 intestine (Alves & Wood, 2006). Pb accumulation in the gut was followed by bone, kidney, liver, spleen, 

7 gill, carcass, brain and white muscle (Alves & Wood, 2006). Ojo and Wood (2007) investigated the 

8 bioavailability of ingested Pb within different compartments of the rainbow trout gut using an in vitro gut 

9 sac technique. Although a significant increase in Pb uptake was observed in the mid-intestines, this was 

10 determined to be much lower than Pb uptake rates via gill surfaces. However, given that intestinal uptake 

11 rate for Pb did not significantly differ from those derived for essential metals (e.g., Cu, Zn, and Ni), this 

12 uptake route is likely to be significant when aqueous Pb concentrations are low and absorption via gill 

13 surfaces is negligible (Ojo & Wood, 2007). 

14 Following a chronic 63-day dietary exposure to Pb, male zebrafish had significantly increased Pb 

15 body burdens, but did not exhibit any significant impairment when compared with controls. Fish were fed 

16 diets consisting of field-collected Nereis diversicolor oligochaetes that contained 1.7 or 33 mg Pb/kg dry 

17 weight. This resulted in a daily Pb dose of either 0.1 or 0.4 mg Pb/kg (Boyle et al., 2010). At the end of 

18 the exposure period, tissue from male fish reared on the high-Pb diet contained approximately 0.6 mg 

19 Pb/kg wet weight, as compared with approximately 0.48 mg Pb/kg wet weight in the low-Pb dietary 

20 exposure group. Pb level was elevated in female fish fed the high-Pb diet, but not significantly so. 

21 Ciardullo et al. (2008) examined bioaccumulation of Pb in rainbow trout tissues following a 3-year 

22 chronic dietary exposure to the metal. Diet was determined to contain 0.19 μg Pb/g wet weight. Fish skin 

23 accumulated the greatest Pb concentrations (0.02 to 0.05 μg Pb/g wet weight), followed by kidney, gills, 

24 liver, and muscle. Pb accumulation in muscles (5 ng Pb/g) remained constant over all sampled growth 

25 stages (Ciardullo et al., 2008). The authors concluded that dietary Pb was poorly absorbed by rainbow 

26 trout. Comparison of dietary and water-borne exposures suggest that although accumulation of Pb can 

27 occur internally from dietary sources, toxicity does not correlate with dietary exposure, but does correlate 

28 with gill accumulation from waterborne exposure (Alves et al., 2006). Comparison of uptake rates across 

29 the gut and gill have shown that transporter pathways in the gill have a much higher affinity for Pb than 

30 do similar pathways in the gut (Ojo & Wood, 2007).

31 Since the 2006 Pb AQCD, several field studies have considered Pb uptake and bioaccumulation in 

32 fish as a tool for environmental assessment. Pb tissue concentrations were elevated in several species of 

33 fish exposed in the field to Pb from historical mining waste, and blood Pb concentrations were highly 

34 correlated with elevated tissue concentrations, suggesting that blood sampling may be a useful and 

35 potentially non-lethal monitoring technique (Brumbaugh et al., 2005). The Western Airborne 

36 Contaminants Assessment Project assessed concentrations of semi-volatile organic compounds and metals 
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1 in up to seven ecosystem components (air, snow, water, sediment, lichen, conifer needles and fish) in 

2 watersheds of eight core national parks during a multi-year project conducted from 2002-2007 (Landers et 

3 al., 2008). The goals of the study were to assess where these contaminants were accumulating in remote 

4 ecosystems in the Western U.S., identify ecological receptors for the pollutants, and to determine the 

5 source of the air masses most likely to have transported the contaminants to the parks. Results from this 

6 study are considered in in Chapter 3 of this ISA. 

Marine Fish

7 In comparison to freshwater fish, fewer studies have been conducted on Pb uptake in marine fish. 

8 Since marine fish drink seawater to maintain osmotic homeostasis, Pb can be taken up via gills and 

9 intestine (W. X. Wang & Rainbow, 2008). Pb was significantly accumulated in gill, liver, plasma, kidney, 

10 rectal gland, intestine, skin, muscle of a marine shark species, spotted dogfish (Scyliorhinus canicula)

11 exposed to 2,072 μg Pb/L for one week (De Boeck et al., 2010). Egg cases of the spotted dogfish exposed 

to 210 12 Pb in seawater for 21 days, accumulated radiolabeled Pb rapidly and the metal was subsequently 

13 detected in embryos indicating the permeability of shark eggs to Pb in coastal environments (Jeffree et al., 

14 2008). 

15 The 2006 Pb AQCD considered detoxification mechanisms in fish including mucus production and 

16 Pb removal by scales through chelation with keratin. Since the 2006 review, additional Pb detoxification 

17 mechanisms in marine fish have been further elucidated. Mummichog (Fundulus heteroclitus)

18 populations in metal-polluted salt marshes in New York exhibited different patterns of intracellular 

19 partitioning of Pb although body burden between sites was not significantly different (Goto & Wallace, 

20 2010). Mummichogs at more polluted sites stored a higher amount of Pb in metal rich granules as 

21 compared to other detoxifying cellular components such as heat-stable proteins, heat-denaturable proteins 

22 and organelles.

23 A study of Pb bioaccumulation in five marine fish species (Chloroscombrus chrysurus, Sardinella 

24 aurita, Ilisha africana, Galeoides decadactylus, Caranx latus) found that C. chrysurus was an especially 

25 strong bioaccumulator, yielding Pb concentrations of 6 to 10 mg Pb/kg (Gnandi et al., 2006). However, C. 

26 chrysurus metal content was not correlated to the Pb concentrations along the mine tailings gradient from 

27 which they were collected (8.5 and 9.0 μg Pb/L for minimum and maximum tissue concentrations, 

28 respectively). This lack of correlation was also observed for fish species that were considered to be 

29 weaker Pb bioaccumulators, indicating that diffuse, non-quantified sources of Pb (e.g., in sediments or in 

30 dietary sources) may be contributing to Pb uptake by marine fish.

31 This review of the recent literature indicates that the primary and most efficient mode of Pb 

32 absorption for freshwater fish is assimilation of labile Pb via gill surfaces; recent research indicates that 

33 chronic dietary Pb exposure may result in some Pb bioaccumulation although it is not the predominant 
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1 route of exposure. Nevertheless, if benthic invertebrates comprise a large portion of fish diets in 

2 chronically contaminated systems, assimilated Pb loads may be significant. This was demonstrated by 

3 Boyle et al. (2010), who showed that laboratory diets consisting of less than one third field-collected Pb4 contaminated invertebrates were sufficient to raise fish tissue Pb levels. However, data from field sites 

5 suggest that fish accumulation of Pb from dietary sources is highly variable and may be strongly 

6 dependent on the physiology of individual species and absorption capacities.

7 Reported BCF’s in fish species from the 2006 Pb AQCD were 42 for brook trout (Salvelinus 

8 fontinalis) and 45 for bluegill (Lepomis macrochirus). Since the 2006 Pb AQCD, additional BAF’s have 

9 been established in water-only exposures, from dietary exposures and from field-collected fish (Table 7-

10 5). 

Table 7-5. Bioaccumulation factors for Pb in fish 

Species BAF Test conditions Reference

Clarias gariepinus 800 56 days, Pb nitrate Kudirat (2008)

Ctenopharyngodon idella 44 Field-collected Malik et al. (2010)

Pimephales promelas 100 to 100,000 30 days, Pb nitrate Grosell et al. (2006a)

Carassius auratus 0.04 Field-collected Khozhina and Sherriff (2008)

Danio rerio 1.4 63 days, dietary exposure Boyle et al. (2010)

Amphibians

11 Since the 2006 Pb AQCD, there are a few new studies that consider uptake of Pb in amphibians. In 

12 a chronic study with tadpoles of the Northern Leopard frog (Rana pipiens), Pb tissue concentrations were 

13 evaluated following exposures to 3, 10, and 100 μg Pb/L from embryo to metamorphosis. The tadpole 

14 tissue concentrations ranged from 0.1 to 224.5 mg Pb/kg dry mass and were positively correlated to Pb 

15 concentrations in the water (T. H. Chen et al., 2006). Dose-dependent bioaccumulation of Pb was 

16 observed in the livers of tadpoles of the African clawed frog (Xenopus laevis) exposed to concentrations 

17 ranging from 0.001 to 30 mg Pb/L (2.91 to 114.5 Pb/g wet weight) for 12 days (Mouchet et al., 2007). Pb 

18 concentrations were measured in livers, bodies without liver and whole bodies in Southern leopard frog 

19 (Rana spenocephala) tadpoles exposed to Pb in sediment (45 to 7,580 mg Pb/kg dry weight) with 

20 corresponding pore water concentrations of 123 to 24,427 μg Pb/L from embryonic stage to 

21 metamorphosis (Sparling et al., 2006). There was 100% mortality at 3,940 mg Pb/kg and higher. In all 

22 body residues analyzed there was a significant positive correlation between Pb in sediment and Pb in 

23 sediment pore water. Concentrations of Pb in liver were similar to results with whole body and bodies 

24 without liver indicating that Pb is not preferentially sequestered in liver. 
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Mammals

1 Studies that consider uptake of Pb in aquatic mammals are limited. Kannan et al. (2006) in a 

2 comparison of trace element concentrations in livers of free-ranging sea otters (Enhydra lutris nereis) 

3 found dead along the California coast, detected Pb in all individuals sampled (N=80) in a range of 0.019 

4 to 1.06 μg Pb/g. The otters were classified by cause of death (infectious causes, non-infectious causes, 

5 those that died in an emaciated condition) and trace element patterns of tissue distribution were compared. 

6 Livers from emaciated otters had significantly elevated levels of Pb compared to non-diseased 

7 individuals. 

7.3.3.4. Food Web 

8 At the time of the publication of the 2006 Pb AQCD, trophic transfer of Pb through aquatic food 

9 chains was considered to be negligible (U.S. EPA, 2006). Measured concentrations of Pb in the tissues of 

10 aquatic organisms were found to be generally higher in algae and benthic organisms and lower in higher 

11 trophic-level consumers indicating that Pb was bioconcentrated but not biomagnified (Eisler, 2000; U.S. 

12 EPA, 2006). New literature since the 2006 Pb AQCD provides evidence of the potential for Pb to be 

13 transferred in aquatic food webs while other studies indicate Pb is decreased with increasing trophic level. 

14 This section incorporates recent literature on transfer of Pb through aquatic food chains including the 

15 application of stable isotope techniques to trace the accumulation and dilution of metals through 

16 producers and consumers.

17 Pb was transferred through at least one trophic level in El Niagra reservoir, Aguascalientes, 

18 Mexico, an ecosystem that lacks fishes (Rubio-Franchini et al., 2008). Pb was measured in sediment, 

19 water, and zooplankton samples of this freshwater system. BCF’s were calculated for predatory and 

20 grazing zooplanktonic species. The BCF of the rotifer A. brigthwellii (BCF 49,300) was up to four times 

21 higher than the grazing cladocerans D. similis (BCF 9,022) and M. micrura (BCF 8,046). According to 

22 the authors, since M. micrura are prey for A. brigthwellii this may explain the biomagnifications of Pb 

23 observed in the predatory rotifer and provides evidence that Pb biomagnifies at intermediate trophic 

24 levels. Partial evidence for biomagnification was observed in a subtropical lagoon in Mexico with 

25 increases of Pb concentration occurring in 14 of the 31 (45.2%) of trophic interactions considered 

26 (Ruelas-Inzunza & Páez-Osuna, 2008). The highest rate of transference of Pb as measured in muscle

27 tissue occurred between the prey species white shrimp (Litopenaeus vannimei) and mullet (Mugil 

28 cephalus) to pelican (Pelecanus occidentalis). 

29 The relative contribution of water and food as source of trace metals including Pb was investigated 

30 in the larvae of the alderfly Sialis velata (Croisetiere et al., 2006). Its prey, the midge (C. riparius) was 

31 reared in the laboratory and then exposed to trace elements in a metal-contaminated lake for one week 

32 prior to being fed to S. velata. During the one-week exposure period of C. riparius to the contaminated 
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1 water, five of six trace elements, including Pb, reached steady state within C. riparius. Alderfly larvae 

2 were held in the lab in uncontaminated lake water and feed one of the treated C. riparius per day for up to 

3 six days to measure Pb uptake via prey. A separate group of alderfly larvae were exposed directly to the 

4 contaminated lake water for six days and fed uncontaminated C. riparius while a third group was exposed 

5 to Pb via prey and water. Trace metal concentrations in S. velata that consumed contaminated C. riparius

6 increased significantly compared to S. velata in water-only exposures. Food was concluded to be the 

7 primary source of Pb (94%) to these organisms, not Pb in the water.

8 The trophic transfer of Pb from the sediment dwelling polychaete worm N. diversicolor to the

9 invertebrate polychaete predator Neris virens provides additional evidence for assimilation of Pb by a 

10 predator and the potential for further transport up the food chain (Rainbow et al., 2006). N. virens

11 significantly accumulated Pb from a diet of N. diversicolor and there was a significant inverse linear 

12 relationship between the trophic transfer coefficient and prey Pb concentration. In the same study, another 

13 predator, the decapod Palaemonetes varians, did not significantly accumulate Pb from N. diversicolor 

14 indicating that trophic transfer is dependent on species-specific differences in metal assimilation 

15 efficiencies and accumulation patterns. 

16 In a recent dietary metal study, field-collected invertebrates representing ecologically relevant 

17 sources of Pb were fed to zebrafish, to assess bioavailability of this metal via food. The polychaete worm 

18 N. diversicolor was collected from two sites; an estuary contaminated with Pb and a reference site with 

19 low metal concentrations (Boyle et al., 2010). Male zebrafish fed Pb-enriched N. diversicolor had 

20 significant increases in whole-body Pb burden when compared to zebrafish fed prey from the reference 

21 site, brine shrimp or flake food diets. There was a trend toward increased Pb levels in females under the 

22 same dietary regimen. In this study, deposit feeding invertebrates were shown to mobilize sediment23 bound metals in the food chain since zebrafish were exposed only to biologically incorporated metal.

24 The concentration of Pb in the tissues of various aquatic organisms was measured during the 

25 biomonitoring of mining-impacted stream systems in Missouri, U.S. Generally, Pb concentrations 

26 decreased with increasing trophic level: detritus contained 20 to 60 μg Pb/g dry weight, while periphyton 

27 and algae contained 1 to 30 μg Pb/g dry weight; invertebrates and fish collected from the same areas 

28 exhibited Pb tissue concentrations of 0.1 to 8 μg Pb/g dry weight (J. M. Besser et al., 2007). In addition, 

29 Pb concentrations in invertebrates (snails, crayfish, and other benthos) were negatively correlated with Pb 

30 concentrations in detritus, periphyton, and algae. Fish tissue concentrations, however, were consistently 

31 correlated only with detritus Pb concentrations (J. M. Besser et al., 2007).

32 Other studies have traced Pb in aquatic food webs and have found no evidence of biomagnification 

33 of Pb with increasing trophic level. Pb exposure at the base of the food web did not biomagnify in a 

34 simplified-four level marine food chain from Tetraselmis suecica (phytoplankton) to Artemia franciscana

35 (crustacean, brine shrimp) then L. vannamei (crustacean white shrimp) and finally to Haemulon scudderi

36 (fish, grunt) (Soto-Jiménez et al.). In the southeastern Gulf of California, Mexico, Pb was not positively 
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1 transferred (biomagnification factor <1) through primary producers (seston, detritus) and 14 consumer 

2 species in a lagoon food web (Jara-Marini et al., 2009). No biomagnification of Pb was detected from 

3 mesozooplankton to macrozooplankton in Bahia Blanca estuary, Argentina (Fernández Severini et al.). In 

4 a Brazilian coastal lagoon food chain, Pb was significantly higher in invertebrates than in fishes (Pereira 

5 et al., 2010). Watanabe et al. (2008) also observed decreasing Pb concentrations through a stream 

6 macroinvertebrate food web in Japan from producers to primary and secondary consumers.

7 Introduction of exotic species into an aquatic food web may alter Pb concentrations at higher tropic 

8 levels. In Lake Erie, the invasive round goby (Neogobius melanastomus) and the introduced zebra mussel

9 (Dreissena polymorpha) have created a new benthic pathway for transfer of Pb and other metals 

10 (Southward Hogan et al., 2007). The goby is a predator of the benthic zebra mussel, while the endemic 

11 smallmouth bass (Micropterus dolomieui) feed on goby. Since the introduction of goby into the lake, total 

12 Pb concentrations have decreased in bass. The authors attribute this decrease of Pb in bass to changes in 

13 food web structure, changes in prey contaminant burden or declines in sediment Pb concentrations. 

7.3.4. Biological Effects

14 This section focuses on the studies of biological effects of Pb on aquatic biota including algae,

15 aquatic plants, invertebrates, fish and other biota with an aquatic lifestage (e.g., amphibians) published 

16 since the 2006 Pb AQCD. Waterborne Pb is highly toxic to aquatic organisms with toxicity varying 

17 depending upon the species and lifestage tested, duration of exposure, the form of Pb tested, and water 

18 quality characteristics. The 2006 Pb AQCD noted that the physiological effects of Pb in aquatic organisms 

19 can occur at the biochemical, cellular, and tissue levels of organization and include inhibition of heme 

20 formation, adverse effects to blood chemistry, and decreases in enzyme levels. Functional growth 

21 responses resulting from Pb exposure include changes in growth patterns, gill binding affinities, and 

22 absorption rates. A review of the more recent literature corroborated these findings, and added 

23 information about induction of oxidative stress by Pb, alterations in chlorophyll, and changes in 

24 production and storage of carbohydrates and proteins. Since this document focuses on atmospheric 

25 sources of Pb to ecosystem receptors, areas of research not addressed here include literature related to 

26 exposure to Pb from shot or pellets. Biological effects of Pb on algae and plants are considered below, 

27 followed by information on effects on aquatic invertebrates and vertebrates. 

7.3.4.1. Plants and Algae

28 Effects of Pb on algae reported in the 2006 Pb AQCD included decreased growth, deformation and 

29 disintegration of algae cells, and blocking of the pathways that lead to pigment synthesis, thus affecting 

30 photosynthesis. Observations in additional algal species since the 2006 Pb AQCD support these findings. 

31 Pb exposure in microalgae species has been linked to several adverse effects, including disruption of 
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1 thylakoid structure and inhibition of growth in both Scenedesmus quadricauda and Anabeana flos-aquae 

2 (Arunakumara & Zhang, 2008). Arunakumara et al. (2008) determined the effect of aqueous Pb on the 

3 algal species S. platensis using solutions of Pb-nitrate. While low Pb exposures (5 μg Pb/mL) stimulated 

4 10-day algal growth, growth was inhibited at higher concentrations of 10, 30, 50, and 100 μg Pb/mL by 5, 

5 40, 49, and 78%, respectively. In addition to growth inhibition, algal chlorophyll a and b content were 

6 significantly diminished at the three highest Pb exposures (Arunakumara et al., 2008). Although no 

7 specific morphological abnormalities were linked to Pb exposure, filament breakage was observed in S. 

8 platensis at Pb concentrations >50 μg Pb/mL. The effect of Pb exposure on the structure and function of 

9 plant photosystem II was studied in giant duckweed, S. polyrrhiza (Ling & Hong, 2009). The Pb 

10 concentration of extracted photosystem II particles was found to increase with increasing environmental 

11 Pb concentration, and increased Pb concentration was shown to decrease emission peak intensity at 340 

12 nm, amino acid excitation peaks at 230 nm, tyrosine residues, and absorption intensities. This results in 

decreased efficiency of visible light absorption by affected plants. The authors theorized that Pb2+ 13 may 

replace either Mg2+ or Ca2+ 14 in chlorophyll or the oxygen-evolving center, inhibiting photosystem II 

15 function through an alteration of chlorophyll structure.

16 An increase in levels of antioxidant enzymes is commonly observed in aquatic plant, algae, and 

17 moss species exposed to Pb. An aquatic moss, F. antipyretica, exhibited increased SOD and ascorbate 

18 levels following a 2-day exposure to Pb-chloride solutions of concentrations of 1, 10, 100, and 1000 

19 μmol. When exposure duration was increased to 7 days, only SOD activity remained significantly 

20 increased by Pb exposure (Dazy et al., 2009). Bell-shaped concentration-response curves were commonly 

21 observed for the induction of antioxidant enzymes in F. antipyretica. The chlorophyll, carotenoid, and 

22 protein contents of the aquatic macrophyte Elodea canadensis were significantly reduced following Pb 

23 accumulation at exposures of 1, 10, and 100 mg Pb/L (Dogan et al., 2009). This, along with the induction 

24 of some antioxidant systems and the reduction of growth at the highest two exposures, indicated that 

25 exposure to the metal caused significant stress, and that toxicity increased with exposure. In addition, 

26 native Myriophyllum quitense exhibited elevated antioxidant enzyme activity (glutathione-S-transferase, 

27 glutathione reductase, peroxidase) following transplantation in anthropogenically polluted areas 

28 containing elevated Pb concentrations. These were correlated with sediment Pb concentrations in the 

29 range of 5 to 23 mg Pb/g dry weight (Nimptsch et al., 2005).

30 Toxicity and oxidative stress were also observed in coontail (C. demersum) rooted aquatic 

31 macrophytes following 7-day exposures to aqueous Pb (1 to 100 μmol), with increasing effects observed 

32 with greater exposure concentrations and times. Chlorosis and leaf fragmentation were evident following 

33 a 7-day exposure to the highest concentration, while induction of antioxidant enzymes (glutathione, 

34 superoxide dismutase, peroxidases, and catalase) was observed at lower exposure concentrations and 

35 times. However, as the duration and concentration of Pb exposure was increased, activities of these 

36 antioxidant enzymes decreased (S. Mishra et al., 2006).
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1 Sobrino et al. (2010) observed reductions in soluble starch stores and proteins with subsequent 

2 increases in free sugars and amino acids in Lemna gibba plants exposed to Pb (50 to 300 mg Pb/L); total 

3 phenols also increased with increasing Pb exposure. Authors noted that this species exhibited similar 

4 responses under extreme temperatures, drought, and disease (Sobrino et al., 2010). According to Odjegba 

5 and Fasidi (2006), exposure to 0.3 mmol of Pb for 21 days was sufficient to induce a gradual reduction of 

6 both chlorophyll and protein content in the macrophyte Eichhornia crassipes. Decreased proteins were 

7 theorized to be related to inefficient protein formation following disruption of nitrogen metabolism after 

8 Pb exposure (Odjegba & Fasidi, 2006). Foliar proline (which is thought to act as an antioxidant) 

9 concentrations were found to increase in a concentration-dependent manner as Pb concentrations increase 

10 from 0.1 to 5.0 mmol. 

11 Following 72-hour aqueous exposure to 41 μmol Pb-nitrate, phytochelatin and glutathione 

12 concentrations in the algae Scenedesmus vacuolatus were significantly increased over that of non-exposed 

13 algal cultures {F, 2006, 358857}. The 72-hour Pb exposure also significantly reduced S. vacuolatus

14 growth, and of all the metals tested (Cu, Zn, Ni, Pb, Ag, As, and Sb), Pb was determined to be the most 

15 toxic to the algae species. 

16 Pb exposure (as Pb-nitrate) caused oxidative damage, growth inhibition, and decreased biochemical 

17 parameters, including photosynthetic pigments, proteins, and monosaccharides, in Wolffia arrhiza plants. 

18 Fresh weight of plants was reduced following both 7- and 14-day exposures to Pb concentrations greater 

19 than 10 mmol, while chlorophyll a content was decreased at concentrations greater than 1 mmol Pb 

20 (Piotrowska et al., 2010).

21 Root elongation was significantly reduced in a number of wetland plant species (Beckmannia 

22 syzigachne, Juncus effusus, Oenanthe javanica, Cyperus flabelliformis, Cyperus malaccensis, and 

23 Neyraudia reynaudiana) following Pb exposures of 20 mg Pb/L (Deng et al., 2009). Further, while both 

24 Zn and Fe exposures exerted some selective pressure on plants, the authors did not observe the same with 

25 Pb, leading them to theorize that concentrations of bioavailable Pb were not present in high enough 

26 quantities to have such an effect. However, while Lemna sp. aquatic plants were determined to effectively 

27 sequester aqueous Pb, the plant growth rate was not significantly different from zero following exposures 

28 of 5 and 10 mg Pb/L, while exposure to 15 mg Pb/L was associated with notable plant mortality (Hurd & 

29 Sternberg, 2008). In fact, Paczkowska et al. (2007) observed that low Pb exposures (0.1 to 1.0 mmol for 9 

30 days) stimulated the growth of Lemna minor cultures, although there was concurrent evidence of 

31 chlorosis and induction of antioxidant enzymes. Additionally, Cd was found to be more toxic than Pb, 

32 although the authors determined that this resulted from poor uptake of Pb by L. minor (Paczkowska et al., 

33 2007)
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7.3.4.2. Invertebrates

1 Effects of Pb on aquatic invertebrates recognized in the 2006 Pb AQCD include adverse impacts on 

2 reproduction, growth, survival and metabolism. Pb was recognized to be more toxic in longer-term 

3 exposures than shorter-term exposures with chronic toxicity thresholds for reproduction in water fleas (D. 

4 magna) ranging as low as 30 μg Pb/L. As observed in terrestrial invertebrates, the antioxidant system, 

5 survival, growth and reproduction are affected by Pb in aquatic organisms. In aquatic invertebrates, Pb 

6 has also been shown to affect stress responses and osmoregulation. New evidence that supports previous 

7 findings of Pb on reproduction and growth in invertebrates are reviewed here as well as limited studies on 

8 behavioral effects. 

9 Recent literature strengthens the evidence indicating that Pb affects enzymes and antioxidant 

10 activity in aquatic invertebrates. Increased SOD activity was observed in mantles of pearl oyster but 

11 decreased with time although always remaining higher than in the control animals during 72-hour

12 exposures to 0.5 μM Pb (Jing et al., 2007). In contrast, activity of Se-dependent glutathione peroxidase 

13 was not changed with Pb exposure. SOD, catalase, and glutathione peroxidase were significantly reduced 

14 at environmentally relevant concentrations of Pb (2 μg Pb/L as measured in Bohai Bay, China) in the 

15 digestive gland of the bivalve Chlamys farreri (Y. Zhang et al., 2010). In contrast, Einsporn et al. (2009)

16 observed no change in catalase activity in the digestive gland and gill of blue mussel M. edulis following 

17 exposures to 2,500 μg Pb/L as Pb nitrate for 10 days and measured again following a 10 day depuration 

18 period. However, in this same species, glutathione-S-transferase activity was elevated in the gills after Pb 

19 exposure and remained active during depuration while no changes to glutathione-S-transferase activity 

20 were observed in the digestive gland. In black mussel (M. galloprovincialis) exposed 10 days to sublethal 

21 concentrations of Pb, fluctuations in SOD activity were observed over the length of the exposure 

22 (Vlahogianni & Valavanidis, 2007). Catalase activity was decreased in the mantle of these mussels but 

23 fluctuated in their gills, as compared with the control group. In the bivalve C. farreri exposed to Pb, there 

24 was induction of lipid peroxidation measured as MDA of 24% and a 37% reduction in 7-ethoxyresorufin25 o-deethylase (EROD) activity when compared to controls (Y. Zhang et al., 2010). In black mussel 

26 exposed for 10 days to sublethal concentrations of Pb, MDA levels were increased in mantle and gill 

27 (Vlahogianni & Valavanidis, 2007). 

28 Aminolevulinic acid dehydratase (ALAD) is a recognized biomarker of exposure across a wide 

29 range of taxa including bacteria (Korcan et al., 2007), invertebrates and vertebrates. Since the 2006 Pb 

30 AQCD, there are additional studies measuring changes in ALAD activity in field-collected bivalves and 

31 crustaceans. In the bivalve Chamelea gallina collected from the coast of Spain, ALAD inhibition was 

32 greater with higher concentrations of Pb measured in whole tissue (Kalman et al., 2008). In another study 

33 from Spain, ALAD activity was negatively correlated with total Pb concentration in seven marine 

34 bivalves (C. gallina, Mactra corallina, Donax trunculus, Cerastoderma edule, M. galloprovincialis, 
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1 Scrobicularia plana and Crassotrea angulata), however, the authors of this study indicated the need to 

2 consider species-dependent responses to Pb (Company et al., 2011). Pb content varied significantly 

3 among species and was related to habitat (sediment versus substrate) and feeding behavior. In red fingered 

4 marsh crab, Parasesarma erythodactyla, collected from sites along an estuarine lake in New South Wales, 

5 Australia, elevated glutathione peroxidase activity was correlated with individuals with higher metal body 

6 burdens (MacFarlane et al., 2006).

7 Studies of stress responses to Pb in invertebrates, conducted since the 2006 Pb AQCD, include 

8 induction of heat shock proteins and depletion of glycogen reserves. Induction of heat shock proteins in 

9 zebra mussel exposed to 500 μg Pb/L for 10 weeks exhibited a 12-fold higher induction rate as compared 

10 to control groups (Singer et al., 2005). Energetic reserves in the freshwater snail B. glabrata in the form 

11 of glycogen levels were significantly decreased by 20%, 57% and 78% in gonads compared to control 

12 animals following 96-hour exposures to 50, 100 and 500 μg Pb/L, respectively (Ansaldo et al., 2006). 

13 Decreases in glycogen levels were also observed in the pulmonary and digestive gland region at 50 and 

14 100 μg Pb/L treatment levels. Pb did not exacerbate the effects of sustained hypoxia in the crayfish (C. 

15 destructor) exposed to 5,000 μg Pb/L for 14 days while being subjected to decreasing oxygen levels in 

16 water (Morris et al., 2005). The crayfish appeared to cope with Pb by lowering metabolic rates in the 

17 presence of the metal. Activity of enzymes associated with the immune defense system in the mantle of 

18 pearl oyster were measured at 0, 24, 48 and 72 hour exposure to 104 μg Pb/L (Jing et al., 2007). Activity 

19 of AcPase, a lysosomal marker enzyme, was detected at 24 hours and decreased at subsequent time

20 points. Phenoloxidase activity was depressed compared with controls and remained significantly lower 

21 than control at 72 hours of exposure to Pb. 

22 The effect of Pb on the osmoregulatory response has been studied after the 2006 Pb AQCD. The 

23 combined effects of Pb and hyperosmotic stress on cell volume regulation was analyzed in vivo and in 

24 vitro in the freshwater red crab, Dilocarcinus pagei (Amado et al., 2006). Crabs held in either freshwater 

or brackish water lost 10% of their body weight after one day when exposed to 2,700 μg Pb2+ 25 /L. This 

26 weight loss was transient and was not observed during days 2-10 of the exposure. In vitro, muscle from 

27 red crabs exposed to hyperosmotic saline solution had increased ninhydrin-positive substances and 

28 muscle weight decreased in isosmotic conditions upon exposure to Pb indicating that this metal affects 

29 tissue volume regulation in crabs although the exact mechanism is unknown. 

30 Additional evidence of reproductive and developmental effects of Pb on aquatic invertebrates is 

31 available since the 2006 Pb AQCD. Sublethal concentrations of Pb negatively affected the total number of 

32 eggs, hatching success and embryonic survival of the freshwater snail B. glabrata exposed to 50, 100, or 

33 500 μg Pb/L (Ansaldo et al., 2009). Following exposure of adult snails for 96 hours, adults were removed 

34 and the eggs were left in the Pb solutions. The total number of eggs was significantly reduced at the 

35 highest concentration tested (500 μg Pb/L). Time to hatching was doubled and embryonic survival was 

36 significantly decreased at 50 and 100 μg Pb/L, while no embryos survived in the highest concentration. 
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1 Formation of tentacles and eyes was significantly impaired in embryos of the freshwater ramshorn snail 

2 Marisa cornuarietis at 15,000 μg Pb/L (Sawasdee & Köhler, 2010). Theegala et al. (2007) observed that 

3 the rate of reproduction was significantly impaired in Daphnia pulex at >500 μg Pb/L in 21 day 

4 exposures. Reproductive variables including average lifespan, rate of reproduction, generation time and 

5 rate of population increase were adversely affected in the rotifer Brachionus patulus under conditions of 

6 increasing turbidity and Pb concentration (Garcia-Garcia et al., 2007). 

7 In larvae of the mosquito, C. quinquefasciatus, exposed to 50 μg Pb/L , 100 μg Pb/L or 200 μg

8 Pb/L, Pb-nitrate exposure was found to significantly reduce hatching rate and egg-production at all 

9 concentrations and larval emergence rate at 200 μg Pb/L (Kitvatanachai et al., 2005). Larval emergence 

10 rates of 78% (F0), 86% (F1) and 86% (F2) were observed in the control group while emergence rates 

11 decreased in each generation 46% (F0), 26% (F1) and 58% (F2) in mosquitoes reared in a concentration 

12 of 200 μg Pb/L. The time to first emergence also increased slightly to 10 days in the Pb-exposed group as 

13 compared to the control group where emergence was first observed on day 9. In the F2 generation of 

14 parents exposed to 200 μg/L, the ratio of female to male offspring was 3.6:1.0. No effects were observed 

15 on oviposition preference of adult females, larval weight or larval deformation. 

16 Since the publication of the 2006 Pb AQCD, limited studies on marine invertebrates have indicated 

17 adverse effects of Pb on reproduction in saltwater environments. In a long term (approximately 60 days) 

18 sediment bioassay with the marine amphipod Elasmopus laevi, onset to reproduction was significantly 

19 delayed at 118 μg Pb/g compared to controls. In the higher concentrations, start of offspring production 

20 was delayed further; 4 days in 234 μg Pb/g and 8 days in 424 μg Pb/g (Ringenary et al., 2007). Fecundity 

21 was also reduced with increasing Pb concentration in sediment. Exposure of gametes to Pb prior to 

22 fertilization resulted in a decrease of the fertilization rates of the marine polychaete Hydroides elegans

23 (Gopalakrishnan et al., 2008). In sperm pretreated in 100 μg Pb/L filtered seawater for 20 minutes,

24 fertilization rate decreased by approximately 70% compared to controls. In a separate experiment, eggs 

25 were pretreated with Pb prior to addition of an untreated sperm suspension. The fertilization rate of eggs 

26 pretreated in 50 μg Pb/L filtered seawater decreased to 20% of the control. In another test with H. elegans

27 in which gametes were not pre-treated, but instead added directly to varying concentrations of Pb for 

28 fertilization, there appears to be a protective effect following fertilization due to the formation of the 

29 fertilization membrane during the first cell division that may prevent Pb from entering the oocytes 

30 (Gopalakrishnan et al., 2007).

31 The protective barrier against Pb toxicity formed by the egg structure in some invertebrates (e.g.,

32 Daphnia) was recognized in the 2006 Pb AQCD. Consideration of toxicity of Pb to embryos that develop 

33 surrounded by a protective egg shell has been expanded since the 2006 Pb AQCD. In a study with 

34 cuttlefish (S. officinalis) eggs, radioisotopes were used to assess the permeability of the egg to Pb at low 

exposure concentrations (210 35 Pb activity concentration corresponding to 512 μg/L Pb) (Lacoue-Labarthe et 

36 al., 2009). Retention and diffusion properties of the cuttlefish egg change throughout the development of 
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1 the embryo and since the eggs are fixed on substrata in shallow coastal waters they may be subject to 

acute and chronic Pb exposures. In the radiotracer experiments, 210 2 Pb was never detected in the internal 

3 compartments of the egg during the embryonic development stage, however concentrations associated 

4 with the eggshell increased throughout the 48 day exposure. These results are consistent with cuttlefish 

5 eggs collected from the field in which Pb was only detected in the eggshell and indicate the protective 

6 barrier provided by cuttlefish egg to Pb toxicity (Miramand et al., 2006). 

7 As noted in the 2006 Pb AQCD, Pb exposure negatively affects the growth of aquatic invertebrates. 

8 Some studies reviewed in the previous document suggested that juveniles do not discriminate between the 

9 uptake of essential and non-essential metals (Arai et al., 2002). In new literature, the freshwater 

10 pulomonate snail Lymnaea stagnalis has been identified as a species that is extremely sensitive to Pb 

11 exposure. Growth of juveniles was inhibited at EC20 <4 μg Pb/L. (Grosell & Brix, 2009; Grosell et al., 

2006b). In L. stagnalis exposed to 18.9μg/L Pb for 21 days, Ca2+ 12 influx was significantly inhibited and 

13 model estimates indicated 83% reduction in growth of newly hatched snails after 30 days at this exposure 

concentration (Grosell & Brix, 2009). The authors speculate that the high Ca2+ 14 demand of juvenile L. 

stagnalis for shell formation and interference of the Ca2+ 15 uptake pathway by Pb result in the susceptibility 

16 of this species.Wang et al., {, 2009, 533439} observed growth of embryos of the Asian Clam (Meretrix 

17 meretrix) was significantly reduced by Pb with an EC50 of 197 μg/L. In juvenile Catarina scallop, 

18 Argopecten ventricosus, exposed to Pb for 30 days, the EC50 for growth was 4,210 μg/L (A. S. Sobrino19 Figueroa et al., 2007). Rate of growth of the deposit feeding polychaete Capitella sp. decreased 

20 significantly with increasing concentrations of Pb associated with sediment (Horng et al., 2009). 

21 Aquatic invertebrate strategies for detoxifying Pb were reviewed in the 2006 Pb AQCD and include 

22 sequestration of Pb in lysosomal-vacuolar systems, excretion of Pb by some organisms, and deposition of 

23 Pb to molted exoskeleton. Molting of the exoskeleton can result in depuration of Pb from the body (see 

24 Knowlton et al., (1983) and Anderson et al., (1997) as cited in the 2006 Pb AQCD). New research has 

25 provided further evidence of depuration of Pb via molting in invertebrates. Mohapatra et al. (2009)

26 observed that Pb concentrations in body tissues were lower in the newly molted mud crabs (Scylla 

27 serrata) than in the pre-molt, hard-shelled crabs. Additionally, the carapace of hard shelled crabs have 

28 lower concentrations of Pb than the exuvium of the soft shell crabs, leading the authors to speculate that 

29 some of the metal might be excreted during the molting process. Bergey and Weis (2007) showed that 

30 differences in the proportion of Pb stored in exoskeleton and soft tissues changed during intermolt and 

31 immediate postmolt in two populations of fiddler crabs (Uca pugnax) collected from New Jersey. One 

32 population from a relatively clean estuary eliminated an average of 56% of Pb total body burden during 

33 molting while individuals from a site contaminated by metals eliminated an average of 76% of total Pb 

34 body burden via this route. Pb distribution within the body of crabs from the clean site shifted from 

35 exoskeleton to soft tissues prior to molting. The authors observed the opposite pattern of Pb distribution 

36 in fiddlers from the contaminated site where larger amounts of Pb were depurated in the exoskeleton.
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1 Behavioral responses of aquatic invertebrates to Pb reviewed in the 2006 Pb AQCD included 

2 avoidance. A limited number of new studies have considered additional behavioral endpoints. Valve 

3 closing speed was used as a measure of physiological alterations due to Pb exposure in the Catarina 

4 scallop (A. Sobrino-Figueroa & Caceres-Martinez, 2009). The average valve closing time increased from 

5 under one second in the control group to 3 to 12 seconds in juvenile scallops exposed to Pb (40 μg/L to 

6 400 μg/L) for 20 days. Damage to sensory cilia of the mantle was observed following microscopic 

7 examination of Pb-exposed individuals. Feeding rate of the blackworm L. variegatus was significantly 

8 suppressed by day 6 of a 10 day sublethal test in Pb-spiked sediments (Penttinen et al., 2008) as compared 

9 to feeding rates at the start of the experiment. However, this decrease of approximately 50% of the initial 

10 feeding rate was also observed in the controls; therefore it is likely caused by some other factor other than 

11 Pb exposure. 

12 Although Pb is known to cause mortality when invertebrates are exposed at sufficiently high 

13 concentrations, species that are tolerant of Pb may not exhibit significant mortality even at high 

14 concentrations of Pb. In a 10-day Pb-spiked sediment exposure (1,000 mg Pb/kg), 100% of individuals of 

15 the Australian estuarine bivalve Tellina deltoidalis survived (King et al., 2010). In the deposit feeding 

16 polychaete Capitella sp., exposure to varying concentrations of Pb associated with sediment up to 0.41 

17 μmol/g had no effect on survival (Horng et al., 2009). In freshwater habitats, odonates are highly tolerant 

18 of Pb with no significant differences in survival time of dragonfly larvae (Pachydiplax longipennis and 

19 Erythemis simplicicollis) exposed to concentrations as high as 185 mg Pb/L Pb (185,000 μg Pb/L) (Tollett 

20 et al., 2009). Other species are more susceptible to Pb in the environment and these responses are 

21 reviewed in Section 7.3.5.

7.3.4.3. Vertebrates

22 Biological effects of Pb on fish that have been studied since the 2006 Pb AQCD report are 

23 reviewed here, and limited new evidence of Pb effects on amphibians, and marine mammals are 

24 considered. As noted in the 2006 Pb AQCD, commonly observed effects of Pb on fish included inhibition 

25 of heme formation, alterations in brain receptors in fish, adverse effects on blood chemistry, and decreases 

26 in some enzyme activities. (U.S. EPA, 2006). Functional responses resulting from Pb exposure included 

27 increased production of mucus, changes in growth patterns, and gill binding affinities. According to Eisler 

28 (2000) and reviewed in the 2006 Pb AQCD, the general symptoms of Pb toxicity in fish include 

29 production of excess mucus, lordosis, anemia, darkening of the dorsal tail region, degeneration of the 

30 caudal fin, destruction of spinal neurons, ALAD inhibition, growth inhibition, renal pathology, 

31 reproductive effects, growth inhibition and mortality. More recent experimental data presented here 

32 expand and support these observations. As in terrestrial vertebrates, Pb has been shown to affect 

33 antioxidant and enzymatic activity in aquatic vertebrates and new evidence of this since the 2006 Pb 
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1 AQCD is reviewed in this section. This section also presents the limited new information available on the 

2 mechanism of Pb as a neurotoxicant in fish and effects of this metal on blood chemistry. Additional 

3 mechanisms of Pb toxicity have been elucidated in the gill and the renal system of fish since the 2006 Pb 

4 AQCD. Further supporting evidence of reproductive and growth effects of Pb on fish is discussed along 

5 with limited new information on behavioral effects of Pb. Finally, limited new information since the 2006 

6 Pb AQCD on physiological effects of Pb on amphibians and marine mammals is presented. 

Fish

7 In environmental assessments of metal-impacted habitats, ALAD is a recognized biomarker of Pb 

8 exposure (U.S. EPA, 2006). For example, lower ALAD activity has been significantly correlated with 

9 elevated blood Pb concentrations in wild caught fish from Pb-Zn mining areas although there are 

10 differences in species sensitivity (Schmitt et al., 2005; Schmitt et al., 2007). Suppression of ALAD 

11 activity in brown trout transplanted to a metal contaminated stream was linked to Pb accumulation on 

12 gills and in liver in a 23 day exposure (Heier et al., 2009). Costa et al. (2007) observed inhibition of 

13 ALAD in hepatocytes of the neotropical traira (Hoplias malabaricus) following dietary dosing of 21 μg 

14 Pb/g every 5 days for 70 days. Cytoskeletal and cytoplasmic disorganization were observed in 

15 histopathological examination of affected hepatocytes. In fathead minnow exposed to Pb in either control 

16 water (33 μg Pb/L), CaSO4 (37μg Pb/L) or (39 μg Pb/L) humic acid-supplemented water and 

17 subsequently analyzed by quantitative PCR analysis there were no significant changes in ALAD mRNA 

18 gene response leading the authors to speculate that water chemistry alone does not influence this gene 

19 response (Mager et al., 2008). 

20 Pb was shown to inhibit hepatic cytochrome P450 in carp (Cyprinus carpio), silver carp 

21 (Hypothalmichtys molitrix) and wels catfish (Silurus glanis) in a concentration-dependent manner from 0-

4.0 μg/mL (Pb2+ 22 ) (Henczova et al., 2008). The concentrations of Pb that resulted in 50% inhibition of 

23 EROD and 7-ethoxycoumarin-o-deethylase (ECOD) isoenzymes varied with the fish species. Silver carp 

24 was the least sensitive to the inhibitory effects of Pb (EROD 1.21, ECOD 1.52 μg Pb/L) while carp 

25 EROD activity was inhibited at 0.76 μg Pb/L. Interaction of Pb with cytochrome P450 was verified by 

26 spectral changes using Fourier Transform Infrared (FTIR) spectroscopy. Liver damage to African catfish 

27 exposed to Pb (50-1,000 μg Pb/L) for 4 or 8 weeks included hepatic vacuolar degeneration followed by 

28 necrosis of hepatocytes (Adeyemo, 2008b). The severity of observed histopathological effects in the liver 

29 was proportional to the duration of exposure and concentration of Pb.

30 Upregulation of antioxidant enzymes in fish is a well-recognized response to Pb exposure. Since 

31 the last review, additional studies demonstrating antioxidant activity as well as evidence for production of 

32 reactive oxygen species following Pb exposure are available. Silver crucian carp (Carassius auratus 

33 gibelio) injected with 10, 20 or 30 mg Pb/kg wet weight Pb-chloride showed a significant increase in the 
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1 rate of production of superoxide ion and hydrogen peroxide in liver (Ling & Hong, 2010). In the same 

2 fish, activities of liver SOD, catalase, ascorbate peroxidase, and glutathione peroxidase were significantly 

3 inhibited. Both glutathione and ascorbic acid levels decreased and malondialdehyde content increased 

4 with increasing Pb dosage, suggesting that lipid peroxidation was occurring and the liver was depleting 

5 antioxidants. In fathead minnow, three genes, glucose-6-phosphate dehydrogenase, glutathione-S6 transferase and ferritin were upregulated, in microarray analysis, during 30 day exposures to Pb in base 

water (33μg Pb/L), or (37μg Pb/L [hard]-water supplemented with 500 μM Ca2+ 7 ) or (39 μg Pb/L [DOC]-

8 water supplemented with 4 mg/L humic acid). However, no changes in whole body ion concentrations 

9 were observed (Mager et al., 2008). In the freshwater fish Nile tilapia, liver catalase, liver alkaline 

10 phosphatase, sodium and potassium-ATPase (NA, K-ATPase) and muscle Ca-ATPase activities were 

11 quantified in various tissues following a 14 day exposure to 5, 10, and 20 μM concentrations of Pb nitrate 

12 (Atli & Canli, 2007). Liver catalase activity significantly increased in the 5 and 20 μM concentrations 

13 while liver alkaline phosphatase activity was significantly increased only at the 20 μM concentration. No 

14 significant change in alkaline phosphatase activity was observed in intestine or serum. Ca-ATPase activity 

15 was significantly decreased in muscle. Na, K-ATPase was elevated in gill in the highest concentration of 

16 Pb while all concentrations resulted in significant decreases of this enzyme in intestine. In another study 

17 with O. niloticus, Pb had no effect on glutathione measured in liver, gill, intestine, muscle and blood and 

18 liver metallothionein levels following a 14 day exposure to 5, 10, and 20 μM concentrations of Pb nitrate 

19 (Atli & Canli, 2008). 

20 Metabolic enzyme activity in teleosts has also been measured following dietary exposures. Alves 

21 and Wood (2006) in a 42 day chronic dietary Pb study with 50 to 500 μg Pb/g found that gill Na, K22 ATPase activity was not affected in rainbow trout while increased Na, K-ATPase was observed in the 

23 anterior intestine. Metabolic activities measured in liver and kidney of Nile tilapia following 60 day 

24 dietary administration of 100, 400, and 800 μg Pb/g indicated that alanine transaminase, aspirate 

25 transaminase, and lactate dehydrogenase activities significantly decreased in kidney in a concentration26 dependent manner (Dai, Fu, et al., 2009) and increased in liver with increasing concentration of dietary 

27 Pb. In a subsequent study using the same exposure paradigm, the digestive enzymes amylase, trypsin and 

28 lipase in tilapia were inhibited by dietary Pb in a concentration-dependent manner (Dai, Du, et al., 2009). 

29 Lesions were also evident in histological sections from livers of Pb-exposed fish from this study and 

30 included irregular hepatocytes, cell hypertrophy, and vacuolation although no quantification of lesions by 

31 dose-group was presented. 

32 Data on the physiological effects of Pb on marine elasmobranchs are limited. De Boeck et al. 

33 (2010) exposed the spotted dogfish to 2,072 μg Pb/L for one week and measured metallothionein 

34 induction, and the electrolytes Na, K, Ca and Cl. No effects were observed in Pb-exposed fish in any of 

35 the physiological parameters measured in this study, however Pb was measured in all organs (De Boeck et 

36 al., 2010).
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1 Additional evidence of the neurotoxic effects of Pb on teleosts has become available since the 2006 

2 Pb AQCD. The mitogen-activated protein kinases (MAPK), extracellular signal-regulated kinase 

(ERK)1/2 and p38MAPK 3 were identified for the first time as possible molecular targets for Pb neurotoxicity 

in a teleost (Leal et al., 2006). The phosphorylation of ERK1/2 and p38MAPK 4 by Pb was determined in 

5 vitro and in vivo in the catfish (Rhamdia quelen). R. quelen exposed to 1,000 μg Pb/L acetate for two 

days showed a significant increase in phosphorylation of ERK1/2 and p38MAPK 6 in the nervous system. 

7 Incubation of cerebellar slices for 3 hours in 5 and 10μM Pb acetate also showed significant 

8 phosphorylation of MAPKs. The observed effects of Pb on the MAPK family of signaling proteins have 

9 implications for control of brain development, apoptosis and stress response. In the neotropical fish traira 

10 (Hoplias malabaricus) muscle cholinesterase was significantly inhibited after 14 dietary doses of 21 μg

11 Pb/g wet weight (Rabitto et al., 2005). Histopathological observations of brains of African catfish exposed 

12 to 500 μg Pb/L or 1,000 μg Pb/L Pb for 4 weeks included perivascular edema, focal areas of malacia, and 

13 diffuse areas of neuronal degeneration (Adeyemo, 2008b). 

14 Adverse effects of Pb on blood chemistry of fish were noted in the 2006 Pb AQCD and limited new 

15 literature since the last Pb review has considered effects on blood. In the African catfish, packed cell 

16 volume decreased with increasing concentration of Pb (25,000 to 200,000 μg Pb/L as Pb-nitrate) and 

17 platelet counts increased in a 96-hour exposure (Adeyemo, 2007). Red blood cell counts also decreased in 

18 some of the treatments when compared to controls, although the response was not dose-dependent and so 

19 may not have been caused by Pb exposure. 

20 The gill has long been recognized as a target of Pb in teleosts. Acute Pb toxicity at the fish gill 

21 primarily involves disruption of Ca homeostasis as previously characterized in the 2006 Pb AQCD 

22 (Rogers & Wood, 2004; Rogers & Wood, 2003). In addition to this mechanism, Pb was found to induce 

23 ionoregulatory toxicity at the gill of rainbow trout through a binding of Pb with Na-K, ATPase and rapid 

inhibition of carbonic anhydrase activity thus enabling noncompetitive inhibition of Na+ and Cl- 24 influx 

25 (Rogers et al., 2005). Alves et al. (2006) administered a diet of three concentrations of Pb (7, 77 and 520 

μg Pb/g dry weight) to rainbow trout for 21 days, and measured physiological parameters including Na+ 26

and Ca+ influx rate from water. Dietary Pb had no effect on brachial Na+ and Ca+ 27 rates except on day 8 

where Na+ 28 influx rates were significantly elevated. These studies suggest that Pb is intermediate between 

29 purely Ca antagonists such as Zn and Cd and disruptors of Na and Cl balance such as Ag and Cu. This 

30 finding has implications for BLM modeling since it suggests that both Ca and Na need to be considered 

31 as protective cations for Pb toxicity. Indeed, protection from Pb toxicity by both Na and Ca have been 

32 documented in freshwater fish (Komjarova & Blust, 2009b). 

33 Long-term exposures of Pb can adversely impact gill structure and function. Histopathological 

34 observations of gill tissue in the catfish (C. gariepinus) following an 8-week aqueous exposure to Pb 

35 nitrate revealed focal areas of epithelial hyperplasia and necrosis at the lower exposure concentrations (50

36 μg Pb/L and 100 μg Pb/L) (Adeyemo, 2008a). Hyperplasia of mucous cells and epithelial cells were 
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1 apparent in the tissue from fish exposed the highest concentrations of Pb in the study (500 μg Pb/L and 

2 1,000 μg Pb/L). In vitro incubation of gill tissue from fathead minnow with Pb concentrations of 2.5, 12.5 

3 and 25 mg Pb/L decreased the ratio of reduced glutathione to oxidized glutathione, indicating that lipid 

4 peroxidation at the gill likely contributes to Pb toxicity at low water hardness (Spokas et al., 2006). 

In addition to recent evidence of Pb interruption of Na+ and Cl- 5 at the gill (Rogers et al., 2005), Pb 

can interfere with the ionoregulation of Na+ and Cland reabsorption of Ca+, Mg2+ 6 , glucose, and water in 

7 the teleost kidney (Patel et al., 2006). Renal parameters including urine flow rate, glomerular filtration 

8 rate, urine pH, and ammonia excretion were monitored in a 96-hour exposure of rainbow trout to 1.2 mg

Pb/L as Pb nitrate. Rates of Na+ and Cl- 9 excretion decreased by 30% by 48 hours while Mg excretion 

10 increased two-to-three fold by 96 hours. Urine flow rate was not altered by Pb exposure, although urinary 

11 Pb excretion rate was significantly increased. After 24 hours of Pb exposure, the urine excretion rate of 

Ca+ 12 increased significantly by approximately 43% and remained elevated above the excretion rate in the 

13 control group for the duration of the exposure. Glomerular filtration rate significantly decreased only 

14 during the last 12 hours of the exposure. Ammonia excretion rate increased significantly at 48 hours as 

15 urine pH correspondingly decreased. At the end of the experiment glucose excretion was significantly 

16 greater in Pb-exposed fish. Although the exposures in this study approached the 96-hour LC50, 

17 nephrotoxic effects of Pb indicate the need to consider additional binding sites for this metal in the 

18 development of biotic ligand modeling (Patel et al., 2006). 

19 Limited new studies on reproductive effects of Pb in fish from oocyte formation to spawning are 

20 available. Decreased oocyte diameter and density in the toadfish (Tetractenos glaber) were associated 

21 with elevated levels of Pb in the gonad (Alquezar et al., 2006). The authors state this is suggestive of a 

22 reduction in egg size which ultimately may lead to a decline in female reproductive output. The effects of 

23 metals on embryonic stage of fish development in Cyprinus carpio and other species were reviewed in 

24 Jezierska et al. (2009) and included developmental abnormalities during organogenesis as well as 

25 embryonic and larval malformations. The authors concluded that the initial period of embryonic 

26 development, just after fertilization, and the period of hatching are the times at which developing embryos 

27 are most sensitive to metals. Reproductive performance of zebrafish as measured by incidence of 

28 spawning, numbers of eggs per breeding pair or hatch rate of embryos was unaffected following a 63 day 

29 diet of field-collected Pb-contaminated polychaetes that were representative of a daily dose of 0.3-0.48 g

30 Pb/kg·day (dry weight diet/wet weight fish) through food (Boyle et al., 2010). Mager et al. (2010)

31 conducted 21 day breeding exposures at the end of chronic 300 day toxicity testing with fathead minnow. 

32 Non-exposed breeders were switched to water containing Pb and Pb-exposed breeders were moved to 

33 control tanks and effects on egg hatchability and embryo Pb accumulation were assessed. Fish in the high 

34 Pb concentration (120 μg Pb/L) reduced total reproductive output, while a significant increase in average 

egg mass was observed in the high Pb HCO3

- 35 and DOC treatments as compared to egg mass size in 
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controls and in low HCO3

- 1 and DOC treatments with Pb. No significant differences were present between 

2 treatments in egg hatchability. 

3 Reproductive effects of Pb have also been observed at the cellular level, including alterations in 

4 gonadal tissue and hormone secretions that are associated with Pb-exposure. Histopathological 

5 observations of ovarian tissue in the African catfish following an 8-week aqueous exposure to Pb nitrate 

6 indicated necrosis of ovarian follicles at the lowest concentration tested (50 μg Pb/L) (Adeyemo, 2008a). 

7 Severe degeneration of ovarian follicles was observed in the highest concentrations of 500 μg Pb/L and 

8 1,000 μg Pb/L. Chaube et al. (2010) considered the effects of Pb on steroid levels through 12 and 24 hour 

9 in vitro exposures of post-vitellogenic ovaries from the catfish (Heteropneustes fossilis) to Pb-nitrate (0, 

10 001, 0.1, 1, 3, and 10 μg Pb/mL). Progesterone, 17-hydroxyprogesterone, 17, 20 beta11 dihydroxyprogesterone, corticosterone, 21-deoxycortisol and deoxycorticosterone were inhibited in a 

12 dose-dependent manner. Pb was stimulatory on the steroids estradiol-17-β, testosterone and cortisol at low 

13 concentrations, and inhibitory at higher concentrations. The disruption of steroid production and altered 

14 hormone secretion patterns observed at the low concentrations of Pb in this study are suggestive of the 

15 potential for impacts to fish reproduction (Chaube et al., 2010). 

16 Reduction of growth was noted as an adverse effect of Pb on fish in the 2006 Pb AQCD. No new 

17 evidence of growth effects in fish have been reported with the exception of Grosell et al. (2006a). In a 

series of exposures in which Ca+2 18 , DOC and pH were varied to assess effects on Pb toxicity to fathead 

19 minnows, Grosell et al. (2006a) observed a significant increase in growth in some groups exposed to 

20 higher concentrations, however, the increase in body mass was noted to have occurred in tanks with high 

21 mortality earlier in the exposure (Grosell et al., 2006a). No effects on growth rates were observed in 

22 rainbow trout administered a diet containing three concentrations of Pb (7, 77 and 520 μg Pb/g dry 

23 weight) for 21 days (Alves et al., 2006) or in Nile tilapia fed diets with 100, 400, or 800 μg/g Pb dry 

24 weight for 60 days (Dai, Du, et al., 2009). Growth and survival were not adversely affected in juvenile 

25 rainbow trout, fathead minnow and channel catfish (Ictalurus punctatus) fed a live diet of L. variegatus

26 contaminated with Pb (850-1,000 μg Pb/L·g dry mass for 30 days. (Erickson et al., 2010). In 30 day 

27 chronic tests in which a range of pH values (6.4, 7.5 and 8.3) were tested with low (25-32 μg Pb/L), 

28 intermediate (82-156 μg Pb/L) and high (297-453 μg Pb/L) concentrations of Pb, Mager et al. (2011) did 

29 not observe growth impairment in fathead minnows at environmentally relevant concentrations of Pb. 

30 Since the publication of the 2006 Pb AQCD, several studies integrating behavioral and 

31 physiological measures of Pb toxicity have been conducted on fish. The ornate wrasse (Thalassoma pavo) 

32 was exposed to sublethal (400 μg Pb/L) or a maximum acceptable toxicant concentration (1,600 μg Pb/L) 

33 dissolved in seawater for one week to assess the effects of Pb on feeding and motor activities (Giusi et al., 

34 2008). In the sublethal concentration group, hyperactivity was elevated 36% over controls. In the high 

35 concentration, a 70% increase in hyperactivity was observed and hyperventilation occurred in 56% of 

36 behavioral observations. Elevated expression of heat shock protein 70/90 orthologs was detected in the 
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1 hypothalamus and mesencephalic areas of the brains of Pb-treated fish. No changes in feeding activity 

2 were noted between non-treated and treated fish. 

3 Sloman et al. (2005) investigated the effect of Pb on hierarchical social interactions and the 

4 corresponding monoaminergic profiles in rainbow trout. Trout were allowed to establish dominant5 subordinate relationships for 24 hours, then were exposed to 46 μg Pb/L or 325 μg Pb/L (Pb-nitrate) for 

6 48 hours to assess effects on behavior and brain momoamines. In non-exposed fish, subordinate 

7 individuals had higher concentrations of circulating plasma cortisol and telencephalic 

8 5-hydroxyindoleacetic acid/5-hydroxytryptamine (serotonin) (5-HIAA/5-HT) ratios. In the high 

9 concentration of Pb, there was significant uptake of Pb into gill, kidney and liver when compared with the 

10 control group and dominant fish appeared to have elevated hypothalamic 5-HIAA/5HT ratios. Uptake of 

11 Pb into the liver was higher in subordinate fish when compared to the dominant fish. No significant 

12 differences were observed in cortisol levels or behavior after metal exposure. 

13 Mager et al. (2010) conducted prey capture assays with 10 day old fathead minnow larvae born 

14 from adult fish exposed to either 35 or 120 μg Pb/L for 300 days, then subsequently tested in a breeding 

15 assay for 21 days. The time interval between 1st and 5th ingestion of 10 prey items (Artemia nauplii) was 

16 used as a measure of behavior and motor function of offspring of Pb-exposed fish. Larvae were offered 10

17 Artemia and the number ingested within 5 minutes was scored. The number of larvae ingesting 5 Artemia 

18 decreased within the time period in offspring of Pb-exposed fish as compared to the control group, 

19 leading the authors to suggest this behavior is indicative of motor/behavioral impairment. 

Amphibians

20 Amphibians move between terrestrial and aquatic habitats and can therefore be exposed to Pb both 

21 on land and in water. The studies reviewed here are all aquatic or sediment exposures. Biological effects 

22 of Pb on amphibians in terrestrial exposure scenarios are reviewed in Sections 7.2.2.3 and 7.2.4.3. 

23 Amphibians lay their eggs in or around water making them susceptible to water-borne Pb during 

24 swimming, breeding and development. In the 2006 Pb AQCD amphibians were considered to be 

25 relatively tolerant to Pb. Observed responses to Pb exposure included decreased enzyme activity (e.g.,

26 ALAD reduction) and changes in behavior summarized in Table AX7-2.4.3 (U.S. EPA, 2006). Since the 

27 2006 Pb AQCD, studies conducted at environmentally relevant concentrations of Pb have indicated 

28 sublethal effects on tadpole endpoints including growth, deformity, and swimming ability. Genotoxic and 

29 enzymatic effects of Pb following chronic exposures have been assessed in laboratory bioassays.

30 Various sublethal endpoints (growth, deformity, swimming ability, metamorphosis) were evaluated 

31 in northern leopard frog (R. pipiens) tadpoles exposed to nominal concentrations of 3, 10, and 100 μg

32 Pb/L as Pb nitrate from embryonic stage to metamorphosis (T. H. Chen et al., 2006). In this chronic study, 

33 the concentrations represent the range of Pb found in surface freshwaters across the U.S. The lowest 
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1 concentration of 3 μg Pb/L approaches the EPA chronic criterion for Pb of 2.5 μg Pb/L at a hardness of 

2 100 mg/L or 4.5 μg Pb/L at a hardness of 170 mg/L (U.S. EPA, 2002). No effects were observed in the 

3 lowest concentration. In the 100 μg Pb/L treatment, tadpole growth rate was slower (Gosner stages 25-

4 30), 92% of tadpoles had lateral spinal curvature (compared with 6% in the control) and maximum 

5 swimming speed was significantly slower than the other treatment groups. In this study, Pb concentrations 

6 in the tissues of tadpoles were quantified and the authors reported that they were within the range of 

7 reported tissue concentrations from wild-caught populations. 

8 The effects of Pb-contaminated sediment on early growth and development were assessed in the 

9 southern leopard frog (Sparling et al., 2006). Tadpoles exposed to Pb in sediment (45, 75, 180, 540, 2,360, 

10 3,940, 5,520, and 7,580 mg Pb/kg dry weight) with corresponding sediment pore water concentrations of 

11 123, 227, 589, 1,833, 8,121, 13,579, 19,038 and 24,427 μg Pb/L from embryonic stage to metamorphosis 

12 exhibited sublethal responses to Pb in sediment at levels below 3,940 mg Pb/kg. There was 100% 

13 mortality in the 3,940, 5,520 and 7,580 mg Pb/kg exposures by day 5. The authors noted that the most 

14 profound effects of Pb on the tadpoles were on skeletal development. At 75 mg Pb/kg, subtle effects on 

15 skeletal formation such as clinomely and brachydactyly were observed. Skeletal malformations increased 

16 in severity at 540 mg Pb/kg and included clinodactyly, brachymely and spinal curvature and these effects 

17 persisted after metamorphosis. At the highest concentration with surviving tadpoles (2,360 mg Pb/kg) all 

18 individuals displayed severe skeletal malformations that impacted mobility. Other sublethal effects of Pb 

19 observed in this study were reduced rates of early growth of tadpoles at concentrations ≤ 540 mg Pb/kg 

20 and increased time to metamorphosis in the 2,360 mg Pb/kg (8,121 μg Pb/L sediment pore water) 

21 treatment. Conversely, no effects were observed on organogenesis in X. laevis embryos exposed to a 

22 range of Pb concentrations from 8,600 to 220,500 μg Pb/L using the Frog Embryo Teratogenesis Assay

23 (Gungordu et al., 2010).

24 Endpoints of oxidative damage were measured in testes of the black-spotted frog (Rana 

25 nigromaculata) treated with 100 μg Pb/L, 200 μg Pb/L, 400 μg Pb/L, 800 μg Pb/L or 1,600 μg Pb/L Pb26 nitrate by epidermal adsorption for 30 days (M. Z. Wang & Jia, 2009). All doses significantly increased 

27 MDA, a product of oxidative stress, and glutathione levels were elevated in all but the lowest treatment 

28 group. In the same study, damage to DNA assessed by DNA tail length showed effects at >200 μg Pb/L 

29 and DNA tail movement showed effects at >400 μg Pb/L. The authors concluded that the effects on 

30 endpoints of oxidative stress and DNA damage detected in testes indicated a possible reproductive effect 

31 of Pb to black-spotted frogs. 

32 The genotoxic potential of Pb to larvae of the toad (X. laevis) was assessed by determining the

33 number of micronucleated erythrocytes per thousand (MNE) following a 12 day exposure (Mouchet et al., 

34 2007). The lowest Pb concentrations with X. laevis (10 and 100 μg Pb/L) did not exhibit genotoxic effects 

35 while both 1,000 and 10,000 μg Pb/L significantly increased MNE to 14 and 202, respectively compared 

36 to the control (6 MNE). In another chronic genotoxic study, erythrocytic micronuclei and erythrocytic 
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1 nuclear abnormalities were significantly increased with increasing Pb concentrations (700 μg Pb/L , 1,400 

2 μg Pb/L , 14,000 μg Pb/L, 70,000 μg Pb/L) during 45, 60, and 75 day exposures of tadpoles Bufo raddei

3 (Y. M. Zhang et al., 2007). The authors noted that the erythrocytic micronuclei and erythrocytic nuclear 

4 abnormalities frequencies generally decreased with increasing exposure time and that this may be 

5 indicative of regulation of genotoxic factors by tadpoles. 

6 In a study with 4-day-old X. laevis tadpoles exposed to a range of concentrations of Pb from 25,500 

7 to 137,000 μg Pb/L for 24 hours, acetylcholinesterase was significantly inhibited in all treatments

8 (Gungordu et al., 2010). The authors suggest that the 35-60% inhibition of acetylcholinesterase is 

9 indicative of a neurotoxic effect. In the same study, glutathione-s-transferase activity significantly 

10 increased in a concentration-dependent manner. Alanine aminotransferase and aspartate aminotransferase 

11 activities were decreased, however, the degree of inhibition did not reflect Pb concentration. The 

12 concentrations used in this study were selected based on the LC50 of Frog Embryo Teratogenesis Assay 

13 results with X. laevis and are much higher than ambient levels of Pb.

Birds

14 In addition to effects on amphibians, consideration of toxicity of Pb to vertebrate embryos that 

15 develop surrounded by a protective egg shell has been expanded since the 2006 Pb AQCD. Pb treatment 

16 of mallard duck (Anas platyrhynchos), eggs by immersion in 100 μg Pb/L for 30 minutes on day 0 of 

17 development did not increase malformations or mortality of embryos (Kertész & Fáncsi, 2003). However, 

18 immersion of eggs in 2,900 μg Pb/L under the same experimental conditions resulted in increased rate of 

19 mortality and significant malformations including hemorrhages of the body, stunted growth, and absence 

20 of yolk sac circulatory system (Kertesz et al., 2006). The second study was conducted to emulate 

21 environmental levels of Pb following a dam failure in Hungary. 

Mammals

22 Although Pb continues to be detected in tissues of marine mammals in U.S. coastal waters (Bryan 

23 et al., 2007; Kannan et al., 2006; Stavros et al., 2007) few studies exist that consider biological effects 

24 associated with Pb-exposure. Pb effects on immune parameters including cell viability, apoptosis, 

25 lymphocyte proliferation, and phagocytosis were tested in vitro on phagocytes and lymphocytes isolated 

26 from the peripheral blood of bottlenose dolphin (Tursiops truncates) (Cámara Pellissó et al., 2008). No 

27 effects on viability of immune cells, apoptosis, or phagocytosis were observed in 72 hour exposure to 

28 concentrations of 1, 10, 20 and 50 mg Pb/L. Proliferative response of bottlenose dolphin leukocytes was 

29 significantly reduced at 50 mg Pb/L, albeit by only 10% in comparison to the control.
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7.3.5. Exposure and Response of Aquatic Species

1 To support the development of air quality criteria standards that are protective of aquatic 

2 ecosystems, threshold levels for Pb effects on aquatic populations must be evaluated. The Annex of the 

3 2006 Pb AQCD (U.S. EPA, 2006) summarized data on exposure-response functions for freshwater and 

4 marine invertebrates (Table AX7-2.4.1) and freshwater and marine fish (Table AX7-2.4.2). The recent 

5 exposure-response studies in this section expand on the findings from the 2006 Pb AQCD with 

6 information on newly-tested organisms (including microalgae, invertebrate, amphibian and fish species). 

7 A series of 72-hour Pb toxicity tests were conducted with five marine microalgae species (T. chuii, 

8 R. salina, Chaetoceros sp., I. galbana and N. gaditana) to determine the relative Pb sensitivities as 

9 measured by growth inhibition. The respective 72-hour EC50 values derived were 2,640, 900, 105, 1,340, 

10 and 740 μg Pb/L (Debelius et al., 2009). The authors noted that species cellular size, sorption capacity, or 

11 taxonomy did not explain differences in sensitivity to Pb, leaving the mechanism of response still open to 

12 question. Additionally, the aquatic freshwater microalgae Scenedesmus obliquus was significantly more 

13 susceptible to Pb exposure than Chlorella vulgaris algae, although these authors stated that both appeared 

14 to be very tolerant of the heavy metal. Laboratory 48-hour standard toxicity tests were performed with 

15 both of these species and respective EC50 values of 4,000 and 24,500 μg Pb/L were derived (Atici et al., 

16 2008). Experiments with the blue-green algae Spirulina platensis produced a LC50 value of 75.3 μg

17 Pb/mL (95% CI: 58.5, 97.0)(Arunakumara et al., 2008). 

18 In the 2006 Pb AQCD, adverse effects of Pb-exposure in amphipods (H. azteca) and water fleas (D. 

19 magna) were reported at concentrations as low as 0.45 μg Pb/L. Effective concentrations for aquatic 

20 invertebrates were found to range from 0.45 to 8,000 μg Pb/L. Since the publication of the 2006 Pb 

21 AQCD, recent studies have identified the freshwater snail L. stagnalis as a species that is extremely 

22 sensitive to Pb exposure (Grosell & Brix, 2009; Grosell et al., 2006b). Growth of juvenile L. stagnalis

23 was inhibited at an EC20 of < 4 μg Pb/L. In contrast, freshwater juvenile ramshorn snails M. cornuarietis 

24 were less sensitive to Pb with the same LOEC for hatching rate and LC50, calculated to be about 10,000 

25 μg Pb/L (Sawasdee & Köhler, 2010).

26 Additional studies on Pb effects in aquatic invertebrates published since the 2006 Pb AQCD have 

27 indicated differences in susceptibility of different lifestages of aquatic organisms to Pb. In a series of 

28 seawater and sediment exposures using adult and juvenile amphipods Melita plumulosa, juveniles were 

29 more sensitive to Pb than adults (King et al., 2006). In the seawater-only exposures, the 96-hour LC50 for 

30 adults was 3,000 μg Pb/L and 1,520 μg Pb/L for juveniles. Ten-day exposures of adults in seawater 

31 resulted in an LC50 of 1,270 μg Pb/L, an NOEC of 190 μg Pb/L and a LOEC of 390 μg Pb/L. In 

32 comparison, the LC50, NOEC, and LOEC value for the adults exposed in sediment was 3,560 μg Pb/L. 

33 Juvenile sediment tests results were LC50 1,980, NOEC 580 and LOEC 1,020 μg Pb/L. 
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1 In the freshwater mussel, Lampsilis siliquoidea (fatmucket) a Pb concentration response was 

2 observed in which newly transformed (5-day-old) juveniles were the most sensitive lifestage in a 96-hour 

3 toxicity test when compared to acute and chronic results with other lifestages (N. Wang et al., 2010). The 

4 96-hour EC50 values for the 5-day-old L. siliquoidea in two separate toxicity tests were 142 and 298 μg

5 Pb/L (mean EC50 220 μg Pb/L) in contrast to older juveniles (2 months old) with an EC50 >426 μg/L. The 

6 24-hour median effect concentration for glochidia (larvae) of L. siliquoidea in 48-hour acute toxicity tests 

7 was >299 μg/L. A 28 day exposure chronic value of 10 μg Pb/L was obtained from 2-month-old L. 

8 siliquoidea juveniles, and was the lowest genus mean chronic value ever reported for Pb (N. Wang et al., 

9 2010). A 96-hour test on newly transformed juveniles was also conducted on Lampsilis rafinesqueana 

10 (Neosho mucket), a mussel that is a candidate for the endangered species list. The EC50 for this species 

11 was 188 μg Pb/L. In contrast, a 24-hour LC50 of 4,500 μg Pb/L for adult black mussel (M. 

12 galloprovincialis) suggests that, in general, juvenile bivalves are more sensitive to Pb exposure than 

13 adults (Vlahogianni & Valavanidis, 2007). 

14 The acute toxicity of Pb to first-instar C. riparius larvae was tested in soft water, with hardness of 8 

15 mg/L as CaCO3. (Bechard et al., 2008). The 24-hour LC50 of 610 μg Pb/L for first instar C. riparius larvae 

16 was much lower than previous values reported for later instars in harder water. In a chronic test with 

17 Chironomus tentans, (8 day-old larvae exposed to Pb until emergence [approximately 27 days]), the 

18 NOEC was 109, and the LOEC was 497 μg Pb/L (Grosell et al., 2006b). The EC20 for reduced growth and 

19 emergence of the midge Chironomus dilutus was 28 μg Pb/L, observed in a 55-day exposure, while the 

20 same species had a 96-hour LC50 of 3,323 μg Pb/L (Mebane et al., 2008). The 24-hour LC50 for larvae of 

21 C. quinquefasciatus mosquitoes was 180 μg Pb/L (Kitvatanachai et al., 2005). A 48-hour LC50 of 5,200 

22 μg Pb/L was observed in water-only exposures of the blackworm Lumbriculus variegatus (Penttinen et 

23 al., 2008). 

24 Cladocerans are commonly tested aquatic organisms, with data from three species: D. magna, D.

25 pulex and Cerodaphnia dubia, representing approximately 70% of available metal toxicological literature 

26 on this group (L. C. Wong et al., 2009). Since the publication of the 2006 Pb AQCD, additional studies 

27 have generated acute toxicity values for other cladocerans. Median lethal concentrations for Moina 

28 micrura (LC50 690 μg Pb/L), Diaphanosoma birgei (LC50 3,160 μg Pb/L), and Alona rectangular (LC50

29 7,000 μg Pb/L) indicate differences in susceptibility to Pb in these freshwater species from Mexico 

30 (Garcia-Garcia et al., 2006). An acute study of Pb with D. pulex identified a 48-hour LC50 of 4,000 μg/L 

31 for this species (Theegala et al., 2007).

32 Exposure-response assays on other freshwater species have been conducted since the 2006 Pb 

33 AQCD. In the mayfly Baetis tricaudatus, the 96-hour LC50 was 664 μg Pb/L (Mebane et al., 2008). An 

34 EC20 value of 66 μg Pb/L was derived for B. tricaudatus by quantifying the reduction in the number of 

35 molts over a 10-day exposure to Pb (Mebane et al., 2008). For rotifer Brachionus patulus neonates, the 

36 24-hour LC50 was 6,150 μg Pb/L (Garcia-Garcia et al., 2007). In a 48-hour toxicity test with the rotifer 
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1 Brachionus calyciflorus, an NOEC (194 μg Pb/L), a LOEC (284 μg Pb/L), and an EC20 of 125 μg Pb/L 

2 was established for this species (Grosell et al., 2006b). 

3 Since the publication of the 2006 Pb AQCD, Pb toxicity to larval stages of marine species has been 

4 assessed at sublethal and lethal concentrations. The effective concentrations at which Pb resulted in 50% 

5 of abnormal embryogenesis of the Asian clam (M. meretrix) was 297 μg Pb/L. The 96-hour LC50 for 

6 larvae of the same species was 353 μg Pb/L {Wang, 2009, 533439}. In comparison, juvenile Catarina 

7 scallop (A. ventricosus) had a LC50 of 830 μg Pb/L in a 96-hour exposure (A. S. Sobrino-Figueroa et al., 

8 2007). Morphological deformities were observed in 50% of veliger larvae of blacklip abalone (Haliotis 

9 rubra) at 4,100 μg Pb/L following a 48-hour exposure to Pb, suggesting this species is not as sensitive to 

10 Pb as other marine invertebrate larvae (Gorski & Nugegoda, 2006).

11 In the marine polychaete H. elegans, EC50 values of gametes, embryos, larvae (blastula to 

12 trochophore and larval settlement), and adults, exhibited dose-responses to Pb that reflected the 

13 differential sensitivity of various lifestages of this organism (Gopalakrishnan et al., 2008). The EC50

14 values for sperm and egg toxicity were 380 and 690 μg Pb/L respectively. Larval settlement measured as 

15 the metal concentration causing 50% reduction in attachment was most sensitive to Pb with an EC50 of 

16 100 μg Pb/L, while the EC50 for abnormal development of embryos was 1,130 μg Pb/L. The LC50 values 

17 for adult worms in 24-hour and 96-hour tests were 25,017 and 946 μg Pb/L, respectively. Manzo et al.

18 (2010) established a LOEC of 500 μg Pb/L and a maximum effect at 3,000 μg Pb/L in an embryotoxicity 

19 assay with sea urchin P. lividus. The EC50 for developmental defects in this species was 1,150 μg Pb/L 

20 with a NOEL of 250 μg Pb/L.

21 There have been only a few new exposure-response studies in amphibians since the 2006 Pb 

22 AQCD. Southern leopard frog tadpoles exposed to Pb in sediment (45 to 7,580 mg/kg dry weight) with 

23 corresponding sediment pore water concentrations from 123 to 24,427 μg Pb/L from embryonic stage to 

24 metamorphosis exhibited concentration-dependent effects on survival (Sparling et al., 2006). The LC50

25 value for Pb in sediment was 3,738 mg/kg, which corresponds to 12,539 μg Pb/L in sediment pore water. 

26 In the same study, concentration-dependent effects on skeletal development were observed. The 40 day27 EC50 for deformed spinal columns in the tadpoles was 1,958 mg Pb/kg (corresponding to 6,734 μg Pb/L 

28 sediment pore water) and the 60 day-EC50 was 579 mg Pb/kg (corresponding to 1,968 μg Pb/L sediment 

29 pore water) (Sparling et al., 2006). A 96-hour LC50 of 96,100 μg Pb/L was determined for X. laevis

30 embryos exposed to a range of Pb concentrations from 8,600 to 220,500 μg Pb/L using the Frog Embryo 

31 Teratogenesis Assay (Gungordu et al., 2010).

32 In the studies reviewed for the 2006 Pb AQCD, freshwater fish demonstrated adverse effects at 

33 concentrations ranging from 10 to >5,400 μg Pb/L, generally depending on water quality parameters (e.g.,

34 pH, hardness, salinity)(U.S. EPA, 2006). Pb tended to be more toxic in longer-term exposures and 

35 correlated to Pb-uptake in tissues. Table AX7-2.4.2 of the 2006 Pb AQCD summarizes effects of Pb to 

36 freshwater and marine fish. At the time of the 2006 Pb AQCD, there was a lack of exposure-response data 
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1 in marine fish. No new exposure-response studies in marine fish have been conducted since the previous 

2 Pb review. 

3 A series of studies published since the 2006 Pb AQCD have been conducted and have further 

4 elucidated the influence of water chemistry parameters on Pb uptake and toxicity in fathead minnow 

5 resulting in additional dose-response data for this species. Grosell et al. (2006b) conducted a series of 30-

day exposures with larval fathead minnow in which varying concentrations ofCa2+ 6 (as CaSO4) and DOC 

7 were tested. The effects of reduced pH (6.7) and increased pH (8.1) compared to a control pH of 7.4 on 

8 Pb toxicity were also assessed in this study. DOC, CaSO4 and pH influenced Pb toxicity considerably 

9 over the range of water parameters tested. The 30-day LC50 for low hardness (19 mg CaSO4/L) in basic 

10 test water was 39 μg dissolved Pb/L and the highest LC50 value (obtained from the protection from 

11 increased concentrations of DOC and CaSO4) was 1,903 μg dissolved Pb/L (Grosell et al., 2006a). 

12 Mager et al. (2010) conducted 300-day chronic toxicity tests at 35 and 120 μg Pb/L with fathead 

13 minnow under conditions of varied DOC and alkalinity to assess the effects of these water quality 

14 parameters on fish growth and Pb-uptake. In additional tests with fathead minnow, Mager et al. (2011)

conducted both 96-hour acute and 30-day chronic tests to further characterize Ca2+ 15 , DOC, pH, and 

alkalinity values on Pb toxicity. Increased Ca2+ 16 , DOC and NaHCO3 concentration afforded protection to 

17 minnows in acute studies. The role of pH in Pb toxicity is complex and likely involves Pb speciation and 

competitive interaction of H+ with Pb2+ 18 (Mager et al., 2011). 

19 In the 2006 Pb AQCD, fish size was recognized as an important variable in determining the adverse 

20 effects of Pb. Acute (96-hour) and chronic (60-day) early-lifestage test exposures were conducted with 

21 rainbow trout to develop ACR’s for this species (Mebane et al., 2008). Two early-lifestage chronic tests 

22 were conducted, the first with an exposure range of 12-384 μg Pb/L (69 days) at 20 mg CaCO3/L water 

23 hardness and the second with an exposure range of 8 to 124 μg Pb/L (62 days) and a water hardness of 29 

24 mg CaCO3/L. In the 69-day test, the following chronic values were observed for survival: NOEC=24 

25 μg/L, maximum acceptable toxicant concentration=36 μg/L, EC10=26 μg/L, EC20=34 μg/L, and LC50=55 

26 μg/L. Results from the 62-day test, with fish length as the endpoint, were NOEC=8 μg/L, MATC=12 

27 μg/L, EC10=7μg/L, EC20=102 μg/L and LC50=120 μg/L. In acute tests run concurrently with the chronic 

28 tests, 96-hour LC50 values were 120 and 150 μg/L, respectively. Data from this study resulted in ACR’s 

29 for trout lower than previously reported. The low ACR values were due to the acute tests which produced 

30 LC50 values that were 10 to 25 times lower than earlier studies with trout (Mebane et al., 2008). The 

31 authors speculated that the lower LC50 values were due to the age of the fish used in the study (two to four 

32 week old fry) and that testing with larger and older fish may not be protective of more sensitive lifestages.
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7.3.6. Community and Ecosystem Effects

1 As discussed in the 2006 Pb AQCD, exposure to Pb is likely to have adverse impacts in aquatic 

2 environments via effects at several levels of ecological organization (individuals, populations, 

3 communities, or ecosystems). These adverse effects resulting from toxicity of Pb would be evidenced by 

4 changes in species composition and richness, in ecosystem function, and in energy flow. The 2006 Pb 

5 AQCD concluded that, in general, there was insufficient information available for single materials in 

6 controlled studies to permit evaluation of specific impacts on higher levels of organization (beyond the 

7 individual organism). Furthermore, Pb rarely occurs as a sole contaminant in natural systems making the 

8 effects of Pb difficult to ascertain. New information on effects of Pb at the population, community and 

9 ecosystem level is reviewed below. 

10 In laboratory studies, Pb exposure has been demonstrated to alter predator-prey interactions, as 

11 well as feeding and avoidance behaviors. Additionally, field studies have linked Pb contamination to 

12 reduced primary productivity and respiration, and to altered energy flow and nutrient cycling. However, 

13 because of the complexity inherent in defining such effects, there are relatively few available population, 

14 community, or ecosystem level studies that conclusively relate Pb exposure to aquatic ecosystem effects. 

15 In addition, most of the available work is related to point-source Pb contamination, with very few studies 

16 considering the effects of diffuse Pb pollution.

17 Both plant species and type of habitat were determined to be factors affecting the rate of Pb 

18 accumulation from contaminated sediments. While the rooted aquatic plant E. canadensis was observed 

19 to accumulate the highest concentrations of Pb, the authors concluded that submerged macrophytes 

20 (versus emergent plants) as a group were the most likely to accumulate Pb and other heavy metals 

21 (Kurilenko & Osmolovskaya, 2006). This would suggest that certain types of aquatic plants, such as 

22 rooted and submerged species, may be more susceptible to aerially-deposited Pb contamination, resulting 

23 in shifts in plant community composition as a result of Pb pollution.

24 Alteration of macrophyte community composition was demonstrated in the presence of elevated 

25 surface water Pb concentrations at three lake sites impacted by mining effluents. A total of 11 species of 

26 macrophytes were collected. Study sites 2 and 3 exhibited similar dissolved Pb concentrations (78 to 92 

27 μg Pb/L, depending on season) and contained six and eight unique macrophyte species, respectively (V. 

28 K. Mishra et al., 2008). The site with the highest Pb concentrations (103 to 118 μg Pb/L) had the lowest 

29 number of resident macrophyte species; these included E. crassipes, L. minor, Azolla pinnata and 

30 Spirodela polyrrhiza. Based on analysis of plant tissue Pb concentrations, the authors theorized that 

31 certain species may be more able to develop Pb tolerant eco-types that can survive at higher Pb 

32 concentrations (V. K. Mishra et al., 2008). 

33 Exposure to three levels of sediment Pb contamination (322, 1,225, and 1,465 μg Pb/g dry weight) 

34 had variable effects on different species within an aquatic nematode community (Mahmoudi et al., 2007). 
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1 Abundance, taxa richness, and species dominance indices were altered at all Pb exposures when 

2 compared with unexposed communities. Further, while the species Oncholaimellus mediterraneus

3 dominated control communities (14% of total abundance), communities exposed to low and medium Pb 

4 concentrations were dominated by Oncholaimus campylocercoides (36%) and Marylynnia stekhoveni

5 (32%), and O. campylocercoides (42%) and Chromadorina metulata (14%), respectively. Communities 

6 exposed to the highest Pb sediment concentrations were dominated by Spirinia gerlachi (41%) and 

7 Hypodontolaimus colesi (29%). Given this, the authors concluded that exposure to Pb significantly 

8 reduced nematode diversity and resulted in profound restructuring of the community structure.

9 The faunal composition of seagrass beds in a Spanish coastal saltwater lagoon was found to be 

10 impacted by Pb in sediment, plants, and biofilm (Marín-Guirao et al., 2005). Sediment Pb concentrations 

11 ranged from approximately 100 to 5,000 mg Pb/kg and corresponding biofilm concentrations were 500 to 

12 1,600 mg Pb/kg, with leaf concentrations up to 300 mg Pb/kg. Although multiple community indices 

13 (abundance, Shannon-Wiener diversity, Simpson dominance index) did not vary from site to site, 

14 multivariate analysis and similarity analysis indicated significant differences in macroinvertebrate 

15 communities between sites with different sediment, biofilm, and leaf Pb concentrations. Differences were 

16 largely attributable to three amphipod species (Microdeutopus sp., Siphonoecetes sabatieri, Gammarus 

17 sp.). This indicates that, although seagrass abundance and biomass were unaffected by Pb exposure, 

18 organisms inhabiting these plants still may be adversely impacted.

19 In certain freshwater habitats, exposure to Pb has been shown to result in significant alterations of 

20 invertebrate communities. Macroinvertebrate community structure in mine-influenced streams was 

21 determined to be significantly correlated to Pb sediment pore water concentrations. Multiple invertebrate 

22 community indices, including Ephemeroptera, Plecoptera, Trichoptera (EPT) taxa richness, Missouri 

23 biotic index, and Shannon-Wiener diversity index, were integrated into a macroinvertebrate biotic 

24 condition score (Poulton et al., 2010). These scores were determined to be significantly lower at sample 

25 sites downstream from mining sites where Pb pore water and bulk sediment concentrations were elevated.

26 Rhea et al. (2006) examined the effects of multiple heavy metals in the Boulder River, MT, U.S., 

27 watershed biofilm on resident macroinvertebrate assemblages and community structure, and determined 

28 that, among all the metals, biofilm Pb concentrations exerted the greatest influence on the 

29 macroinvertebrate community indices. Pb biofilm concentrations were significantly correlated with 

30 reduced EPT taxa richness, reduced EPT abundance, and an increase in Diptera species abundance. 

31 Interestingly, Pb concentrations in invertebrate tissues were correlated to an increase in Hydropsychidae 

32 caddisfly abundance, but this may have resulted from the intrinsically high variability in tissue Pb 

33 concentrations. The authors concluded that Pb-containing biofilm represented a significant dietary 

34 exposure for impacted macroinvertebrate species, thus altering invertebrate community metrics (Rhea et 

35 al., 2006).
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1 Kominkova and Nabelkova (2005) examined ecological risks associated with metal contamination 

2 (including Pb) in small urban streams. Although surface water Pb concentrations in monitored streams 

3 were determined to be very low, concentrations of the metal in sediment were high enough to pose a risk 

4 to the benthic community (e.g., 34 to 101 mg Pb/kg). These risks were observed to be linked to benthic 

5 invertebrate functional feeding group, with collector-gatherer species exhibiting larger body burdens of 

6 heavy metals than other groups (Kominkova & Nabelkova, 2005). In contrast, benthic predators and 

7 collector-filterers accumulated significantly lower metals concentrations. Consequently, it is likely that 

8 sediment-bound Pb contamination would differentially affect members of the benthic invertebrate 

9 community, potentially altering ecosystems dynamics.

10 Invertebrate functional feeding group may also affect invertebrate Pb body burdens in those 

11 systems where Pb bioconcentration occurs. The predaceous zooplanktonic rotifer, A. brigthwellii collected 

12 from a Pb-impacted reservoir in Mexico, contained 384 ng Pb/mg and exhibited a water-to-tissue BCF of 

13 49,344. The authors theorized that Pb biomagification may have been observed in this case because the 

14 cladoceran M. micrura is both a known Pb accumulator and a favorite prey item of the rotifer (Rubio15 Franchini et al., 2008). They showed that M. micrura had twice the Pb body burden of D. similis, another 

16 grazing cladoceran species present in the reservoir. These two species exhibited average Pb tissue 

17 concentrations of 57 and 98 ng Pb/mg, respectively, with respective water column BCFs of 9,022 and 

18 8,046. Conversely, an examination of the simultaneous uptake of dissolved Pb by the algae P. subcapitata

19 and the cladoceran D. magna suggests that the dietary exposure route for the water column filter-feeder is 

20 minor. Although Pb accumulated in the algal food source, uptake directly from the water column was 

21 determined to be the primary route of exposure for D. magna (Komjarova & Blust, 2009c). 

22 For many invertebrate species, sediment Pb concentrations may be the most important driver in 

23 determining Pb uptake. For instance, while Hg and Cd body burdens in lentic invertebrates were affected 

24 by lake ecological processes (e.g., eutrophication), a similar effect was not observed for Pb concentrations 

25 in crayfish tissue, despite a high variability between sites. Although this may be a result of differing 

26 bioaccumulation tendencies, the authors suggested that other factors, including the potential for sediment 

27 exposures, may be responsible for Pb uptake in lentic invertebrates (Larsson et al., 2007).

28 A risk assessment conducted in southern Florida freshwater canals determined that the 90th 

29 percentile of bulk Pb sediment concentrations in systems was 105 mg Pb/kg, which was predicted to 

30 result in a sediment pore water concentration of 2.6 μg Pb/L. This estimated pore water concentration was 

31 contrasted with acute 10th percentile toxic concentrations derived from a series of species sensitivity 

32 distributions: 8.7 μg Pb/L for arthropods, 223 μg Pb/L for fish species, and 116 μg Pb/L for all species 

33 (Rand & Schuler, 2009). Although the predicted sediment pore water Pb concentration was below the 

34 derived acute toxicity hazardous concentration for 10% of species (HC10) values, it was considered 

35 possible that chronic exposure to such concentrations could impact some arthropod populations. A 

36 chronic species sensitivity distribution constructed with Pb NOEC values for all aquatic species produced 
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1 a chronic 10th percentile Pb NOEC value of 1.6 μg/L, a further indication that some aquatic species 

2 (likely arthropod species) may be impacted by the Pb contamination in southern Florida canals (Rand & 

3 Schuler, 2009).

4 Caetano et al. (2007) investigated the mobility of Pb in salt marshes using total content and stable 

5 isotope signature. They found that roots had similar isotopic signature to sediments in vegetated zones 

6 indicating that Pb uptake by plants reflects the input in sediments. At one site, there was a high 

7 anthropogenic Pb content while at the other natural mineralogical sources dominated. The roots of S. 

8 fructicosa and S. maritima significantly accumulated Pb, having maximum concentrations of 2,870 mg

9 Pb/kg and 1,755 mg Pb/kg, respectively, indicating that below-ground biomass played an important role 

10 in the biogeochemical cycling of Pb.

11 In addition to the ecological effects discussed above, there is additional evidence that Pb exposure 

12 could alter bacterial infection (and potentially disease transmission) in certain fish species. Following 96-

13 hour exposures to 4,000 μg Pb/L, bacterial density in Channa punctatus fish was observed to be 

14 significantly altered when compared to non-exposed fish. Bacteria population densities in fish spleen, 

15 gills, liver, kidneys and muscle tissues were higher following Pb exposure, with bacterial abundance in 

16 the gills too numerous to quantify (Pathak & Gopal, 2009). In addition, bacteria inhabiting Pb-exposed 

17 fish were more likely to exhibit antibacterial resistance than colonies isolated from non-exposed fish. 

18 Although the mechanism remains unknown, this study suggests that Pb exposure may increase the 

19 likelihood of infection in fish, potentially affecting fish abundance and recruitment.

20 In summary, despite the fact that alterations of macrophyte communities may be highly visible 

21 effects of increased sediment Pb concentrations, several recently published papers propose that ecological 

22 impacts on invertebrate communities are also significant, and can occur at environmental Pb 

23 concentrations lower than those required to impact plant communities. High sediment Pb concentrations 

24 were linked to shifts in amphipod communities inhabiting plant structures, and potentially to alterations in 

25 ecosystem nutrient processing through selective pressures on certain invertebrate functional feeding 

26 groups (e.g., greater bioaccumulation and toxic effects in collector-gatherers versus predators or filter27 feeders). Increased sediment pore water Pb concentrations were demonstrated to likely be of greater 

28 importance to invertebrate communities, as well. Interestingly, recent research also suggests that Pb 

29 exposure can alter bacterial infestations in fish, increasing both microbial density and resilience, and 

30 potentially increasing the likelihood of serious disease outbreak. 

7.3.7. Critical Loads in Aquatic Systems

31 Since the publication of the 2006 Pb AQCD there is no new significant information on critical 

32 loads of Pb in aquatic systems. Refer to Section 7.3.6 of the 2006 Pb AQCD for a discussion of critical 

33 loads of Pb in aquatic systems. 
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7.3.8. Characterization of Sensitivity and Vulnerability

1 Data from the literature indicate that exposure to Pb may affect survival, reproduction, growth, 

2 metabolism, and development in a wide range of aquatic species. Often, species differences in 

3 metabolism, sequestration, and elimination rates control relative sensitivity and vulnerability of exposed 

4 organisms. Diet and lifestage at the time of exposure also contribute significantly to the determination of 

5 sensitive and vulnerable populations and communities. Further, environmental conditions in addition to 

6 those discussed as affecting bioavailability may also alter Pb toxicity. The 2006 Pb AQCD reviewed the 

7 effects of genetics, age, and body size on Pb toxicity. While genetics appears to be a significant 

8 determinant of Pb sensitivity, effects of age and body size are complicated by environmental factors that 

9 alter metabolic rates of aquatic organisms. A review of the more recent literature corroborated these 

10 findings, and identified seasonally-affected physiological changes and life-stage as other important 

11 determinants of differential sensitivity to Pb.

7.3.8.1. Seasonally-Affected Physiological Changes

12 A study by Duman et al. (2006) identified species and seasonal effects of Pb uptake in aquatic 

13 plants. P. australis accumulated higher root Pb concentrations than S. lacustris. Additionally, the P. 

14 australis Pb accumulation factor was significantly higher during the winter versus other seasons, while 

15 the Pb accumulation factor for S. lacustris was greatest in spring and autumn. The Pb accumulation factor 

16 for a third species, P. lucens, was greatest in autumn (Duman et al., 2006). Most significantly, these 

17 changes in bioaccumulation were not linked with biomass increases, indicating that species-dependent 

18 seasonal physiological changes may control Pb uptake in aquatic macrophytes (Duman et al., 2007). 

19 Significant interspecies differences in Pb uptake were observed for plants representing the same genus 

20 (Sargassum), indicating that uptake of Pb by aquatic plants also may be governed by highly species21 dependent factors (Jothinayagi & Anbazhagan, 2009). 

22 Couture et al. (2010) investigated seasonal and decadal variations in Pb sources to mussels (M. 

23 edulis) from the French Atlantic shoreline. Pb concentrations in the mussels were 5-66 times higher than 

the natural background value for the north Atlantic. The 206Pb/207 24 Pb signature indicated that the 

25 bioaccumulated Pb was anthropogenic in origin. The signature was not, however, the same as that emitted 

26 in western Europe as a result of leaded gasoline combustion, although that was a major emission source to 

27 the atmosphere during a large part of the study period (1985-2005). Instead, it was most similar to that of 

28 Pb released into the environment from wastewater treatment plants, municipal waste incinerators and 

29 industries such as metal refineries and smelters. Thus continental runoff rather than atmospheric 

30 deposition was identified as the main source of Pb to the French coastal area. The strong seasonal 

variations in 206Pb/208 31 Pb were used to conclude that resuspension of Pb triggered by high river runoff 

32 events was a key factor affecting bioaccumulation of Pb in M. edulis. In another biota monitoring study, 
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1 Pearce and Mann (2006) investigated variations in concentrations of trace metals in the U.K. including Pb 

2 in the shells of pod razor shell (Ensis siliqua). Pb concentration varied from 3.06-36.2 mg Pb/kg and 

3 showed a regional relationship to known sources, e.g., former metal mining areas such as Cardigan Bay, 

4 Anglesey, and industrial activity in Liverpool Bay. Seasonal variations were also found for Pb in both 

5 Cardigan Bay and Liverpool Bay, relating to increased winter fluxes of Pb (and other metals) into the 

6 marine environment. Heier et al. (2009) established the speciation of Pb in water draining from a shooting 

7 range in Norway and looked at the time dependent accumulation in brown trout. They found that high 

8 molecular weight (>10 kDaltons) cationic Pb species correlated with high flow episodes and 

9 accumulation of Pb on gills and in the liver. Thus, high flow episodes can remobilize metals from a 

10 catchment and induce stress to aquatic organisms. 

7.3.8.2. Increased Nutrient Uptake

11 Singh et al. (2010) proposed that metal-resistant plants have the capacity to not only up-regulate

12 antioxidant synthesis, but also have the ability to increase nutrient consumption and uptake to support 

13 metal sequestration and detoxification via production of antioxidants (Singh et al., 2010). Therefore, it is 

14 likely that such plant species would be significantly less susceptible to Pb exposure than those species 

15 without those abilities.

7.3.8.3. Temperature and pH

16 Water temperature also appears to affect the toxicity of Pb to aquatic organisms, with higher 

17 temperatures leading to greater responses. Pb toxicity to crayfish increased 7 to 10% when the water 

18 temperature was increased by 4°C, and by 14% when the temperature increased by 7°C. The authors 

19 determined that the increased toxicity was a result of the negative impact of Pb on crayfish respiration, 

20 which was exacerbated by the lower dissolved oxygen concentrations at higher water temperatures (M. A. 

21 Q. Khan et al., 2006). The sequestration ability of L. minor macrophytes was similarly impacted by 

22 increased surface water temperature; plants absorbed a maximum Pb concentration of 8.6 mg /g at 30°C, 

23 while uptake at 15°C was only 0.3 mg/g (Uysal & Taner, 2009). Decreased pH was also demonstrated to 

24 increase the uptake of environmental Pb in aquatic plants (Uysal & Taner, 2009; C. Wang, X. Yan, et al., 

25 2010). Additionally, Birceanu et al. (2008) determined that fish (specifically rainbow trout) were more 

26 susceptible to Pb toxicity in acidic, soft waters characteristic of sensitive regions in Canada and 

27 Scandinavia. Hence, fish species endemic to such systems may be more at risk from Pb contamination 

28 than fish species in other habitats. 
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7.3.8.4. Life Stage

1 A comparison of C.riparius Pb LC50 values derived from toxicity tests with different instars 

2 indicates a significant effect of lifestage on Pb sensitivity for aquatic invertebrates. Bechard et al. (2008)

3 calculated a first instar C. riparius 24-hour LC50 value of 613 μg Pb/L, and contrasted this value with the 

4 24-hour and 48-hour LC50s derived using later instar larvae—350,000 and 200,000 μg Pb/L, respectively. 

5 This disparity would suggest that seasonal co-occurrence of aquatic Pb contamination and sensitive early 

6 instars could have significant population-level impacts (Bechard et al., 2008). Similarly, Wang et al. 

7 (2010) demonstrated that the newly transformed juvenile mussels, L. siliquoidea and L. rafinesqueana, at 

8 5 days old were more sensitive to Pb exposure than were glochidia or two to six month- old juveniles, 

9 suggesting that Pb exposure at particularly sensitive lifestages could have a significant influence on 

10 population viability (N. Wang et al., 2010).

11 Pb exposures also differentially affected life-stages of the marine polychaete H. elegans. Pb water 

12 concentrations of 100 μg Pb/L and greater significantly affected fertilization and embryonic development, 

13 but the greatest effects were exhibited by 24-hour-old larvae (Gopalakrishnan et al., 2007). The authors 

14 suggested that timing of Pb exposure may have different impacts on marine polychaete populations, if life 

15 cycles are off-set (Gopalakrishnan et al., 2007). Further, given that the adult lifestage is sedentary, 

16 reduction of the mobile early lifestage as a result of Pb exposures may disproportionally affect sessile 

17 polychaetes. For instance, larval settlement was significantly reduced at Pb exposures of 50 μg Pb/L and 

18 greater (Gopalakrishnan et al., 2008).

7.3.8.5. Species Sensitivity

19 Both inter- and intra-specific difference in Pb uptake and bioaccumulation may occur in 

20 macroinvertebrates of the same functional-feeding group. Cid et al. (2010) reported significant differences 

21 in Pb bioaccumulation between field collected Ephoron virgo mayflies and Hydropysche sp, caddisflies, 

22 with only the mayfly exhibiting increased Pb tissue concentrations when collected from Pb-contaminated 

23 sites; the caddisfly Pb tissue concentrations were similar between reference and Pb-contaminated areas. 

24 The authors also examined the lifestage specific accumulation of Pb for E. virgo mayflies, and although 

25 there was no statistical difference in Pb tissue concentrations between different lifestages, Pb 

26 bioaccumulation did change as mayflies aged (Cid et al., 2010). Data from 20 years of monitoring of 

27 contaminant levels in filter-feeding mussels Mytilus sp. and oysters Crassotrea virginica in coastal areas 

28 of the U.S. through the National Oceanic and Atmospheric Administration (NOAA) Mussel Watch 

29 program indicate that Pb is on average three times higher in mussels than in oysters (Kimbrough et al., 

30 2008).

31 Species-specific Ca requirements have also been shown to affect the vulnerability of aquatic 

32 organisms to Pb. The snail, L. stagnalis, exhibits an unusually high Ca demand due to CaCO3 formation 
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1 required for shell production and growth, and exposure to Pb prevents the uptake of needed Ca, leading to 

2 toxicity. Consequently, aquatic species that require high assimilation rates of environmental Ca for 

3 homeostasis are likely to be more sensitive to Pb contamination (Grosell & Brix, 2009). Grosell and 

4 colleagues also noted that reduced snail growth following chronic Pb exposure was likely a result of 

5 reduced Ca uptake (Grosell et al., 2006b).

6 There is some indication that molting may comprise an additional sequestration and excretion 

7 pathway for aquatic animals exposed to Pb (Bergey & Weis, 2007; Mohapatra et al., 2009; Soto-Jiménez 

8 et al.; Tollett et al., 2009). Crab species U. pugnax (Bergey & Weis, 2007) and Scylla serrata (Mohapatra 

9 et al., 2009), white shrimp L. vannamei (Soto-Jiménez et al.) as well as Libellulidae dragonfly nymphs 

10 (Tollett et al., 2009) have been shown to preferentially sequester Pb in exoskeleton tissue, where it is later 

11 shed along with the hardened exterior tissue. Consequently, aquatic arthropod species and those species 

12 that shed their exoskeleton more frequently may be able to tolerate higher environmental Pb 

13 concentrations than non-arthropods or slow-growing molting species, as this pathway allows them to 

14 effectively lower Pb body burdens.

15 Some tolerant species of fish (e.g., mummichog) have the ability to sequester accumulated Pb in 

16 metal-rich granules or heat-stable proteins (Goto & Wallace, 2010). Fish with such abilities are more 

17 likely to thrive in Pb-contaminated environments than other species. In contrast, the effect of Pb exposure 

18 on fish bacterial loads demonstrated by Pathak and Gopal (2009) suggest that infected fish populations 

19 may be more at risk to the toxic effects of Pb than healthier species. Aqueous Pb was demonstrated to 

20 both increase bacteria density in several fish organs and to improve the likelihood of antibacterial 

21 resistance (Pathak & Gopal, 2009).

22 Tolerance to prolonged Pb exposure may develop in aquatic invertebrates and fish. Multi23 generational exposure to low levels of Pb appears to confer some degree of metal tolerance in 

24 invertebrates such as Chironomus plumosus larvae; consequently, previous population Pb exposures may 

25 decrease species’susceptibility to Pb contamination (Vedamanikam & Shazilli, 2008). However, the 

26 authors noted that metal tolerant larvae were significantly smaller than larvae reared under clean 

27 conditions, and that transference of Pb-tolerant C. plumosus larvae to clean systems resulted in a 

28 subsequent loss of tolerance. Evidence of acclimation to elevated Pb in fathead minnow was suggested in 

29 the variations in ionoregulatory parameters that were measured on day 10 and 30 in fish exposed to 115 

μg Pb/L for 30 days. At the end of the experiment, whole body Ca2+ was elevated while Na+ and K+ 30

31 recovered from elevated levels at 30 days (Grosell et al., 2006a). 

32 A series of species sensitivity distributions constructed by Brix et al. (2005) indicated that 

33 sensitivity to Pb was greatest in crustacean species, followed by coldwater fish, and warmwater fish and 

34 aquatic insects, which exhibited a similar sensitivity. Further, analysis of both acute and chronic 

35 mesocosm data sets indicated that Pb-contaminated systems exhibited diminished species diversity and 

36 taxa richness following both types of exposure (Brix et al., 2005). Wong et al. (2009) constructed Pb 
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1 species sensitivity distributions for both cladoceran and copepod species. A comparison of the two curves 

2 indicated that cladoceran species, as a group, were more sensitive to the toxic effects of Pb than were 

3 copepods, with respective hazardous concentration values for 5% of the species (HC5) values of 35 and 

4 77 μg Pb/L. This difference in sensitivities would indicate that cladoceran species are more likely to be 

5 impacted at lower environmental Pb concentrations than copepods, potentially altering community 

6 structures or ecosystem functions (L. C. Wong et al., 2009).

7.3.8.6. Ecosystem Vulnerability

7 Relative vulnerability of different aquatic ecosystems to effects of Pb can be inferred from the 

8 information discussed above on species sensitivity and the influence of water quality variables on the 

9 bioavailability and toxicity of Pb. It is, however, difficult to categorically state that certain plant, 

10 invertebrate or vertebrate communities are more vulnerable to Pb than others, since toxicity is dependent 

11 on many variables and data from field studies are complicated by co-occurrence of other metals and 

12 alterations of pH, such as in mining areas. Aquatic ecosystems with low pH and low DOM are likely to be 

13 the most sensitive to the effects of atmospherically-deposited Pb. Examples of such systems are acidic, 

14 soft waters such as sensitive regions in Canada and Scandinavia (Birceanu et al., 2008). In the U.S., 

15 aquatic systems that may be more susceptible to effects of Pb include habitats that are acidified due to 

16 atmospheric deposition of pollutants, runoff from mining activities or lakes and streams with naturally

17 occurring organic acids. Hence, fish and invertebrate species endemic to such systems may be more at 

18 risk from Pb contamination than corresponding species in other habitats.

19

7.3.9. Ecosystem Services

20 There are limited publications at this time that address Pb impacts to ecosystem services associated 

21 with aquatic systems and most of the available literature is for estuaries, salt marsh and freshwater 

22 wetlands rather than lakes and streams. The evidence reviewed in the ISA illustrates that Pb can affect the 

23 ecological effects in each of the four main categories of ecosystem services (Section 7.1.2) as defined by 

24 Hassan et al. (2005). These effects are sorted into ecosystem services categories and summarized here:

25  Supporting: food for higher trophic levels, biodiversity 

26  Provisioning: clean drinking water, contamination of food by heavy metals, decline in health 

27 of fish and other aquatic species

28  Regulating: water quality

29  Cultural: ecosystem and cultural heritage values related to ecosystem integrity and 

30 biodiversity, wildlife and bird watching, fishing 
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1 A few recent studies consider the impact of Pb and other heavy metals on ecosystem services 

2 provided by estuaries (Smith et al., 2009) and salt marsh {Bromberg Gedan, 2009, 672706}. These 

3 systems are natural sinks for metals and other contaminants. They provide habitat and breeding areas for 

4 both terrestrial and marine wildlife and are locations for bird watching. In a Monte-Carlo modeling 

5 approach designed to assess the degree of risk of Pb and Hg to wading birds in estuarine habitats in the 

6 U.K., the authors found a high probability that Pb may pose an ecologically relevant risk to dunlin, 

7 Calidris alpina (Smith et al., 2009). However, the authors noted that a major source of uncertainty in this 

8 study was the NOAEL values for Pb. Pb can be toxic to salt marsh plant species and decaying plant 

9 detritus may result in resuspension of Pb into the aquatic food chain {Bromberg Gedan, 2009, 672706}.

10 Ecological services provided by freshwater wetlands are similar to those provided by estuaries and 

11 are sinks for atmospheric Pb as well as Pb from terrestrial runoff (Landre et al., 2010; Watmough & 

12 Dillon, 2007). Several studies have addressed the response of natural wetlands to Pb (Gambrell, 1994; 

13 Odum, 2000). Recent reviews of pollution control by constructed wetlands (Mander & Mitsch, 2009), 

14 removal of metals by constructed wetlands (Marchand et al., 2010) and phytoremediation of metals by 

15 wetland plants (Rai, 2008) indicate that these systems can remove Pb from the aquatic environment. The 

16 use of plants as a tool for immobilization of Pb and other metals from the environment is not limited to 

17 wetland species. Recent advances in the phytoremediation of metals are reviewed in Dickinson et al. 

18 (2009). 

19 The impact of Pb on ecological services provided by specific components of aquatic systems has 

20 been considered in a limited number of studies. Aquatic fauna can take up and bioaccumulate metals. If 

21 the bioaccumulating species is a food source, the uptake of metals may make it toxic or more dangerous 

22 for people or other wildlife to consume. For example, oysters and mussels bioaccumulate Pb from 

23 anthropogenic sources, including atmospheric deposition, and are a food source that is widely consumed 

24 by humans and wildlife (Couture et al., 2010). Their capacity to bioaccumulate Pb makes them good 

25 bioindicators of environmental contamination and they have been used as monitors of coastal pollutants 

26 by the NOAA Mussel Watch program since 1986. The conclusions of a recent assessment report are that 

27 the highest concentrations of Pb are found in mussels and oysters near urban and industrial centers and 

28 the only region that exceeded the Food and Drug Administration action level for Pb in clams, oysters and 

29 mussels of 1.7 mg Pb/kg wet weight was Lake Michigan, where maximum concentrations were 1.9 mg

30 Pb/kg wet weight in the years 2004-2005 (Kimbrough et al., 2008). Although bioaccumulation may 

31 render aquatic fauna toxic to consumers, bioaccumulation is a way to sequester the metals and remove 

32 them from waters and soils. Sequestration for this purpose is an ecosystem service that has been 

33 quantified. For example, the total ecological services value of a constructed intertidal oyster (Crassostrea 

34 sp.) reef in improving water quality and sequestering metals including Pb was calculated in the Yangtze 

35 River estuary to be about $500,000 per year (Quan et al., 2009). Other aquatic organisms have been 
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1 considered for their role in remediation of Pb in the environment. Theegala et al. (2007) discuss the high 

2 uptake rate of Pb by D. pulex as the basis for a possible Daphnia-based remediation for aquatic systems.

7.3.10. Summary of Aquatic Effects

3 This summary of the effects of Pb on aquatic ecosystems covers information from the publication 

4 of the 2006 Pb AQCD to present. Refer to Section 7.4: Causality determinations for Pb in Terrestrial and 

5 Aquatic Systems for a synthesis of all evidence dating back to the 1977 AQCD considered to determine 

6 causality.

7.3.10.1. Biogeochemistry and Chemical Effects

7 Once the atmospherically-derived Pb enters surface waters its fate and bioavailability are 

8 influenced by Ca concentration, pH, alkalinity, and total suspended solids, and DOC, including humic 

9 acids. Once in sediments, Pb bioavailability may be influenced by the presence of sulfides and Fe and Mn 

10 oxides, physical disturbance, the presence of other metals, biofilm and organisms. In many, but not all 

11 aquatic organisms, Pb dissolved in the water can be the primary exposure route to gills or other biotic 

12 ligands. A more detailed understanding of the biogeochemistry of Pb in aquatic systems (both the water 

13 column and sediments) is critical to accurately predicting toxic effects of Pb to aquatic organisms. As 

14 recognized in the 2006 Pb AQCD and further supported in this review, chronic exposures to Pb may also 

15 include dietary uptake, and there is an increasing body of evidence showing that differences in uptake and 

16 elimination of Pb vary with species. Currently available models for predicting bioavailability focus on 

17 acute toxicity and do not consider all possible routes of uptake. They are therefore of limited applicability, 

18 especially when considering species-dependent differences in uptake and bioaccumulation of Pb. 

7.3.10.2. Bioavailability 

19 There is evidence over several decades of research previously reviewed in Pb AQCDs and in recent 

20 studies reviewed in this ISA that Pb bioaccumulates in plants, invertebrates and vertebrates in aquatic 

21 systems, just as it does in terrestrial systems. According to the 2006 Pb AQCD, and further supported in 

22 this review, Pb adsorption, complexation, chelation, etc., are processes that alter bioavailability to aquatic 

23 biota. Given the low solubility of Pb in water, bioaccumulation of Pb by aquatic organisms may 

24 preferentially occur via exposure routes other than direct absorption from the water column, including 

25 ingestion of contaminated food and water, uptake from sediment pore waters, or incidental ingestion of 

26 sediment. 

27 As reviewed by Wang and Rainbow (2008) and supported by additional studies reviewed in this 

28 ISA, there are considerable differences between species in the amount of Pb taken up from the 

29 environment and in the levels of Pb retained in the organism. The bioaccumulation and toxicity of Pb to 
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1 aquatic organisms are closely linked to the environmental fate of the metal under variable environmental 

2 conditions (Section 3.3) as they are highly dependent upon the proportion of free metal ions in the water 

3 column.

4 Recent studies on bioavailability of Pb in aquatic plants and algae support the findings of previous 

5 Pb AQCDs that all plants tend to sequester larger amounts of Pb in their roots than in their shoots and 

6 provide additional evidence for species differences in compartmentalization of sequestered Pb and 

7 responses to Pb in water and sediments. Given that atmospherically-derived Pb is likely to become 

8 sequestered in sediments, uptake by aquatic plants is a significant route of Pb removal from sediments, 

9 and a potential route for Pb mobilization into the aquatic food web. Although there are some similarities 

10 to Pb accumulation observed in terrestrial plants (e.g., preferential sequestration of the metal in root 

11 tissue), Pb appears to be more bioavailable in sediment than it is in soil. Trees that inhabit semi-aquatic 

12 environments have also been shown to absorb Pb from contaminated sediments. 

13 In the case of invertebrates, Pb can be bioaccumulated from multiple sources, including the water 

14 column, sediment, and dietary exposures, and factors such as amount of bioavailable Pb, lifestage, age, 

15 and metabolism can alter the accumulation rate. Additional studies have considered the relative 

16 importance of water versus dietary uptake of Pb in aquatic invertebrates. Use of stable isotopes has

17 enabled simultaneous measurement of uptake and elimination in several aquatic species to assess the 

18 relative importance of water versus dietary uptake. In uptake studies of various invertebrates, Pb was 

19 mainly found in the gills and digestive gland/hepatopancreas.

20 There is more information now on the cellular and subcellular distribution of Pb in invertebrates 

21 than there was at the time of writing the 2006 Pb AQCD. Specifically, localization of Pb at the 

22 ultrastructural level has been assessed in the marine mussel (M. edulis), scallop and cuttlefish and was 

23 found to be bound principally to organelles (Einsporn et al., 2009; Einsporn & Koehler, 2008). Since the 

24 2006 Pb AQCD, BCF’s have been measured for several species in field studies and these BCF’s tend to be 

25 higher than calculated BCF’s from laboratory exposures (Table 7-4).

26 Tissue accumulation of Pb via gill and dietary uptake has been further characterized in aquatic 

27 vertebrates and stable isotope techniques have been applied to further elucidate bioaccumulation of Pb in 

28 this ISA. The conclusions of the 2006 Pb AQCD that the gill is a major site of Pb uptake in fish and that 

29 there are species differences in the of Pb accumulation and distribution of Pb within the organism are 

30 supported in this review. In general, the accumulation of Pb in fish tissues is observed to be 

31 gill>kidney>liver. The anterior intestine has been newly identified as a site of uptake of Pb through 

32 dietary exposure studies (Alves et al., 2006). Additional detoxification strategies for Pb have been 

33 elucidated since the 2006 Pb AQCD. Mummichogs at more polluted sites stored a higher amount of Pb in 

34 metal rich granules as compared to other detoxifying cellular components such as heat-stable proteins, 

35 heat-denaturable proteins and organelles (Goto & Wallace, 2010).
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1 There are few new studies on Pb uptake by amphibians and mammals. At the time of the 

2 publication of the 2006 Pb AQCD, trophic transfer of Pb through aquatic food chains was considered to 

3 be negligible. Measured concentrations of Pb in the tissues of aquatic organisms were generally higher in 

4 algae and benthic organisms than in higher trophic-level consumers indicating that Pb was 

5 bioconcentrated but not biomagnified (Eisler, 2000; U.S. EPA, 2006). Some studies published since the 

6 2006 Pb AQCD support the potential for Pb to be transferred in aquatic food webs, while other studies 

7 indicate that Pb concentration decreases with increasing trophic level (biodilution). 

7.3.10.3. Biological Effects

8 Evidence in this review further supports the findings of the previous Pb AQCDs that waterborne Pb 

9 is highly toxic to aquatic organisms, with toxicity varying with species and lifestage, duration of 

10 exposure, form of Pb, and water quality characteristics.

11 Effects of Pb on algae reported in the 2006 Pb AQCD included decreased growth, deformation and 

12 disintegration of algae cells, and blocking of the pathways that lead to pigment synthesis, thus affecting 

13 photosynthesis. Observations in additional algal species since the 2006 Pb AQCD support these findings. 

14 Effects on plants supported by additional evidence in this review and evidence from previous reviews 

15 include oxidative damage, decreased photosynthesis and reduced growth. The mechanism of Pb toxicity 

16 in plants is likely mediated by damage to photosystem II through alteration of chlorophyll structure. 

17 Elevated levels of antioxidant enzymes are commonly observed in aquatic plant, algae, and moss species 

18 exposed to Pb. 

19 As observed in terrestrial invertebrates, upregulation of antioxidant enzymes is a common 

20 biomarker of Pb exposure in aquatic invertebrates. Since the 2006 Pb AQCD, there is additional evidence 

21 for Pb effects on antioxidant enzymes, lipid peroxidation, stress response and osmoregulation. Studies of 

22 reproductive and developmental effects of Pb in aquatic invertebrates in this review provide further 

23 support for findings in the 2006 Pb AQCD. These new studies include reproductive endpoints for rotifers 

24 and freshwater snails as well as multigenerational effects of Pb in mosquito larvae. Growth effects are 

25 observed at lower concentrations in some aquatic invertebrates since the 2006 Pb AQCD, including 

26 juveniles of the freshwater snail L. stagnalis where growth is affected at <4 μg Pb/L (Grosell et al., 

27 2006b). Behavioral effects of Pb in aquatic invertebrates reviewed in this ISA include decreased valve 

28 closing speed in scallops and slower feeding rate in blackworms.

29 Additional mechanisms of Pb toxicity have been elucidated in the gill and the renal system of fish 

30 since the 2006 Pb AQCD. Further supporting evidence of reproductive, behavioral, growth effects and 

31 effects on blood parameters have become available since the 2006 Pb AQCD. The mitogen-activated 

protein kinases, ERK1/2 and p38MAPK 32 were identified for the first time as possible molecular targets for 

33 Pb neurotoxicity in a teleost (Leal et al., 2006). Pb toxicity at the fish gill primarily involves disruption of 
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1 Ca homeostasis as previously characterized in the 2006 Pb AQCD (Rogers & Wood, 2004; Rogers & 

2 Wood, 2003). In addition to this mechanism, Pb was found to induce ionoregulatory toxicity at the gill of 

3 rainbow trout through a binding of Pb with Na-K, ATPase and rapid inhibition of carbonic anhydrase 

activity thus enabling noncompetitive inhibition of Na+ and Cl- 4 influx. 

5 In the 2006 Pb AQCD amphibians were considered to be relatively tolerant to Pb. Observed 

6 responses to Pb exposure included decreased enzyme activity (e.g., ALAD reduction) and changes in 

7 behavior summarized in Table AX7-2.4.3 of the 2006 Pb AQCD (U.S. EPA, 2006). Since the 2006 Pb 

8 AQCD, studies conducted at concentrations approaching environmental levels of Pb have indicated 

9 sublethal effects on tadpole endpoints including growth, deformity, and swimming ability.

7.3.10.4. Exposure and Response

10 Concentration-response data on plants, invertebrates and vertebrates is consistent with findings in 

11 previous reviews of species differences in sensitivity to Pb in aquatic systems. Growth in plants continues 

12 to be an endpoint adversely affected by Pb exposure. The lowest EC50 for growth observed in marine 

13 microalgae and freshwater microalgae was in the range of 100 μg/L.

14 In the 2006 Pb AQCD, concentrations at which effects were observed in aquatic invertebrates 

15 ranged from 5 to 8,000 μg Pb/L. Several studies in this review have provided evidence of effects at lower 

16 concentrations. Among the most sensitive species, growth of juvenile freshwater snails L. stagnalis was 

17 inhibited at an EC20 of <4 μg Pb/L. (Grosell & Brix, 2009; Grosell et al., 2006b). A chronic value of 10 

18 μg Pb/L obtained in 28-day exposures of 2-month-old L. siliquoidea juveniles was the lowest genus mean 

19 chronic value ever reported for Pb (N. Wang et al., 2010). 

20 In the 2006 Pb AQCD, adverse effects were found in freshwater fish at concentrations ranging from 

21 10 to >5,400 μg Pb/L, generally depending on water quality variables (e.g., pH, hardness, salinity).

22 Additional testing of Pb toxicity under conditions of varied alkalinity, DOC, and pH has been conducted 

23 since the last review. However, adverse effects in fish observed in recent studies fall within the range of 

24 concentrations observed in the previous AQCD. 

7.3.10.5. Community and Ecosystem Effects 

25 Since the publication of the 2006 Pb AQCD, additional evidence for community and ecosystem 

26 level effects of Pb have been observed primarily in microcosm studies or field studies with other metals 

27 present. One ecological effect reported in previous Pb AQCDs is a shift in community composition in Pb28 impacted habitats towards more Pb-tolerant species. New studies in this ISA provide evidence in 

29 additional habitats for community composition shifts associated with Pb. Alteration of aquatic plant 

30 community composition was demonstrated in the presence of elevated surface water Pb concentrations at 

31 three lake sites impacted by mining effluents. Lakes with the highest levels of Pb had the lowest number 
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1 of aquatic plant species when compared to sites with lower Pb concentrations. In an aquatic macrophyte 

2 community, both plant species and type of habitat were determined to be factors affecting the rate of Pb 

3 accumulation from contaminated sediments. While the rooted macrophyte E. canadensis was observed to 

4 accumulate the highest concentrations of Pb, the authors concluded that submerged macrophytes (versus 

5 emergent plants) as a group were the most likely to accumulate Pb and other heavy metals (Kurilenko & 

6 Osmolovskaya, 2006). This would suggest that certain types of aquatic plants, such as rooted and 

7 submerged species, may be most susceptible to atmospherically-deposited Pb, resulting in shifts in plant 

8 community composition as a result of Pb pollution.

9 Despite the fact that alterations of macrophyte communities may be highly visible effects of 

10 increased sediment Pb concentrations, several recently published papers propose that ecological impacts 

11 on invertebrate communities are also significant, and can occur at environmental Pb concentrations lower 

12 than those required to impact plant communities. High sediment Pb concentrations were linked to shifts in 

13 amphipod communities inhabiting plant structures, and potentially to alterations in ecosystem nutrient 

14 processing through selective pressures on certain invertebrate functional feeding groups.

15 Sensitive species may become locally extinct from habitats where Pb toxicity is greater. Birceanu et 

16 al. (2008) determined that fish, specifically rainbow trout, were more susceptible to Pb toxicity in acidic, 

17 soft waters characteristic of sensitive regions in Canada and Scandinavia. Hence, fish species endemic to 

18 such systems may be more at risk from Pb contamination than fish species in other habitats. A series of 

19 species sensitivity distributions constructed by Brix et al. (2005) indicated that sensitivity to Pb was 

20 greatest in crustacean species, followed by coldwater fish, and warmwater fish and aquatic insects, which 

21 exhibited a similar sensitivity. 

7.3.10.6. Critical Loads, Sensitivity and Vulnerability

22 Since the 2006 Pb AQCD there is no new significant information on critical loads of Pb in aquatic 

23 systems. Recent studies have identified seasonally-affected physiological changes and life-stage as 

24 important determinants of differential sensitivity to Pb in aquatic organisms. These factors are in addition 

25 to species differences in metabolism, sequestration, and elimination rates, diet, lifestage, genetics, age, 

26 and body size that were considered in the 2006 Pb AQCD. Although evidence is available to support Pb 

27 impacts to supporting, provisioning, regulating and cultural ecosystem services, there is insufficient data 

28 available to adequately quantify these adverse effects.
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7.4. Causality Determinations for Lead in 

Terrestrial and Aquatic Systems

1 This section presents key conclusions regarding causality determinations for welfare effects of Pb

2 (Table 7-6). Evidence considered in establishing causality was drawn from the 1977 (U.S. EPA, 1977), 

3 1986 (U.S. EPA, 1986) and 2006 Pb AQCD (U.S. EPA, 2006) for Pb where appropriate as well as the 

4 current ISA. EPA’s framework for causality described in Chapter 1 was applied and the causal 

5 determinations are highlighted. In this ISA, effects determined to be causal at the species level contribute 

6 to the body of evidence for causal effects at the community and ecosystem scale. Some of the effects of 

7 Pb observed in terrestrial and aquatic organisms are also considered in the chapters of the ISA that 

8 evaluate evidence for human health effects associated with Pb exposure. 

Table 7-6. Summary of Pb causal determinations for plants, invertebrates and vertebrates

Effect Causality Determination 

Bioaccumulation as it Affects Ecosystem Services-All organisms Causal 

Mortality-Plants Inadequate

Mortality- Invertebrates and Vertebrates Causal 

Growth-Plants Causal

Growth-Invertebrates Causal

Growth-Vertebrates Suggestive

Physiological Stress-All organisms Causal

Hematological Effects-Invertebrates and Vertebrates Causal

Development and Reproduction-Invertebrates and Vertebrates Causal 

Development and Reproduction-Plants Inadequate

Neurobehavior-Invertebrates and Vertebrates Causal 

Community and Ecosystem Level Effects Causal

7.4.1. Bioaccumulation of Lead in Terrestrial and Aquatic 

Biota as it Affects Ecosystem Services

9 Pb deposited on the surface of, or taken up by organisms has the potential to alter the services 

10 provided by terrestrial and aquatic biota to humans. Ecosystem services are the benefits people obtain 

11 from ecosystems. They include supporting, provisioning, regulating and cultural services that are vital for 

12 the functioning of the biosphere and provide the basis for the delivery of tangible benefits to human 

13 society. There is compelling evidence over several decades of research that Pb bioaccumulates in plants, 

14 invertebrates and vertebrates in terrestrial and aquatic systems. Generally, there are considerable 

15 differences between species in the amount of Pb taken up from the environment and in the amounts of Pb 
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1 retained in the organism. In order for Pb to reach a biological receptor, the metal must first cross the 

2 membranes of organisms to the target organ or site of storage. This process varies between plants, 

3 invertebrates and vertebrates, and furthermore, uptake and sequestration are at times similar in unrelated

4 species, while substantially different between related ones. Uptake of Pb from environmental media is 

5 dependent upon the bioaccessibility of Pb (reviewed in Chapter 3) which is influenced by many factors 

6 including, but not limited to, temperature, pH, presence of humic acid and dissolved organic matter, 

7 presence of other metals, and speciation of Pb. 

8 Terrestrial plants accumulate Pb via direct stomatal uptake into foliage and incorporation of 

9 atmospherically-deposited Pb from soil into root tissue, followed by variable translocation to other 

10 tissues. Near smelters and other point-sources of Pb, leaf uptake may in some cases be greater than root 

11 uptake, otherwise root uptake is predominant. Translocation of soil Pb to shoots and leaves is limited in 

12 most species as plants sequester a large portion of Pb in root tissue. There are considerable species13 dependent differences in rate of uptake and translocation of Pb to other parts of the plant. Uptake and 

14 sequestration of Pb primarily in roots of terrestrial plants is also observed in wetland species and algae. 

15 Rooted aquatic plants take up Pb primary from sediments while floating aquatic plants take up Pb from 

16 water. 

17 Uptake of atmospherically deposited Pb from soil is the primary exposure route in terrestrial plants 

18 and invertebrates. Bioaccessibility of Pb to soil-dwelling organisms is influenced by soil factors including 

19 soil type, amount of OM, pH, and CEC and there are considerable differences in uptake among species. 

20 Species exhibit different accumulation efficiencies and compartmentalize sequestered Pb differently. 

21 There is limited evidence of contributions to total Pb body burden via dietary exposures in primary and 

22 secondary consumers. 

23 Aquatic organisms can uptake and bioaccumulate Pb from the water column, sediments or via 

24 dietary exposure. However, as in terrestrial organisms, uptake and subsequent bioaccumulation of Pb in 

25 aquatic plants, invertebrates and vertebrates varies greatly between species and across taxa. Invertebrates 

26 may also sequester Pb in the exoskeleton, which is subsequently shed. Pb in aquatic invertebrates is 

27 primarily sequestered in the gill and hepatopancreas. Uptake of Pb in fish is well characterized and occurs 

28 primarily via direct uptake of dissolved Pb from the water column through gill surfaces and by ingestion 

29 of Pb-contaminated diets. Pb in these organisms is primarily sequestered in the gill. In dietary exposures 

30 in fish, Pb also targets the anterior intestine indicating the importance of water-only versus dietary uptake 

31 exposures.

32 Pb is bioaccumulated in plants, invertebrates and vertebrates inhabiting terrestrial and aquatic 

33 systems that receive Pb from atmospheric deposition. This represents a potential route for Pb mobilization 

34 into the food web or into food products. For example, Pb bioaccumulation in leaves and roots of an edible 

35 plant may represent an adverse impact to the provisioning of food, an essential ecosystem service. Recent 

36 research has suggested that dietary Pb (i.e., Pb adsorbed to sediment, particulate matter, and food) may 
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1 contribute to exposure and toxicity in primary and secondary order consumers (including humans). 

2 Although there is no consistent evidence of trophic magnification there is substantial evidence of trophic 

3 transfer. It is through consumption of Pb-exposed prey or Pb-contaminated food that atmospherically 

4 deposited Pb reaches species that may have very little direct exposure to it. Overall, based on the

5 consistency of findings across taxa, the evidence is sufficient to conclude that there is a causal 

6 relationship between Pb exposures and bioaccumulation of Pb that affects ecosystem services

7 associated with terrestrial and aquatic biota.

7.4.2. Mortality

8 The relationship between Pb exposure and mortality has been well demonstrated in terrestrial and 

9 aquatic species as presented in Sections 7.2.5 and 7.3.5 of this ISA and in the previous Pb AQCDs. 

10 Toxicological studies have established LC50 values for some species of plants, invertebrates and 

11 vertebrates. From the LC50 data on Pb in this review and previous Pb AQCDs a wide range of sensitivity 

12 to Pb is evident across taxa. However, the LC50 is usually much higher than current environmental levels 

13 of Pb, even though physiological dysfunction that adversely impacts the fitness of an organism often 

14 occurs well below concentrations that result in mortality. 

15 Pb is generally not phytotoxic to plants at concentrations found in the environment away from 

16 point-sources, probably due to the fact that plants often sequester large amounts of Pb in roots, and that 

17 translocation to other parts of the plant is limited. 

18 Invertebrates are generally more sensitive to Pb exposure than other taxa, with survival adversely 

19 impacted in a few species at concentrations occurring near point-sources or at concentrations that 

20 approach ambient levels. These impacted species may include candidate or endangered species. The 

21 freshwater mussel Lampsilis rafinesqueana (Neosho mucket), is a candidate for the endangered species 

22 list. The EC50 for foot movement (a measure of viability) for newly transformed juveniles of this species 

23 was 188 μg Pb/L. (N. Wang et al., 2010). However, tolerance to Pb varies substantially among 

24 invertebrate species. As reviewed in the 2006 Pb AQCD, the LC50 values for soil nematodes vary from 

25 10-1,550 mg Pb/kg dry weight dependent upon soil OM content and soil pH (U.S. EPA, 2006). In 

26 earthworms, 14 and 28 day LC50 values typically fall in the range of 2,400-5,800 mg Pb/kg depending 

27 upon the species tested. Toxicity of Pb to aquatic invertebrates is highly dependent on water quality 

parameters such as pH, DOC and Ca2+ 28 . For example, 48 hour LC50 values for C. dubia range from 280 to 

29 >2,700 μg Pb/L when tested at varying pH levels (U.S. EPA, 2006). Other invertebrates such as odonates

30 may be tolerant of Pb concentrations that greatly exceed environmental levels. Some invertebrates are 

31 able to detoxify Pb such as through sequestration of Pb in the exoskeleton which is subsequently molted. 

32 Early experiments from Pb mining areas indicated local extinction of fish from streams. Mortality 

33 in fish is dependent on the sensitivity of the species tested and on water quality parameters. Higher 
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1 toxicity tends to occur in acidic waters where more free-Pb ion is available for uptake. The interaction 

2 between water quality parameters and Pb toxicity may result in a range of concentrations that cause 

3 equivalent toxicity. For example, 96-hr LC50 values in fathead minnow range from 810->5,400 μg Pb/L in 

4 varying pH and hardness (U.S. EPA, 2006). Increased mortality is also a function of age of the fish. Thirty 

5 day LC50 values for larval fathead minnows ranged from 39 to 1,903 μg Pb/L in varying concentrations of 

6 DOC, CaSO4 and pH (Grosell et al., 2006b). In a recent study of rainbow trout fry at 2 to 4 weeks post7 swim up, the 96-hour LC50 was 120 μg Pb/L at a hardness of 29 mg/L as CaCO3, a value much lower than 

8 in testing with older fish (Mebane et al., 2008). 

9 In terrestrial avian and mammalian species, toxicity is observed in laboratory studies over a wide 

10 range of doses (<1 to >1,000 mg Pb/kg body weight-day) as reviewed for the development of Eco-SSL’s 

11 {U.S. EPA, 2005 #19390}. Mortality associated with Pb exposure in vertebrates is supported by the 

12 consistently positive associations between Pb exposure and mortality observed in human epidemiologic 

13 studies (Section 5.3.1). 

14 The evidence is inadequate to conclude that there is a causal relationship between Pb 

15 exposure and mortality in plants.

16 The evidence is sufficient to conclude that there is a causal relationship between Pb exposures 

17 and mortality in sensitive terrestrial and aquatic animal taxa. 

7.4.3. Growth Effects

18 Evidence for Pb effects on growth is strongest in plants with limited information on invertebrates 

19 and vertebrates. There is evidence over several decades of research that Pb inhibits photosynthesis and 

20 respiration in plants all of which reduce the growth of the plant (U.S. EPA, 1977, 1986, 2006). Many 

21 laboratory toxicity studies report effects on plants; however, there are few field studies. Specifically, Pb 

22 has been shown to affect photosystem II with the hypothesized mechanism being that Pb may replace 

23 either Mg or Ca in chlorophyll, altering the pigment structure and decreasing the efficiency of visible 

24 light absorption in exposed plants. Decreases in chlorophyll a and b content have been observed in 

25 various algal and plant species. The lowest 72-hour EC50 for growth inhibition reported for algae was 105 

26 μg Pb/L in Chaetoceros sp. Most primary producers experience EC50 values for growth in the range of 

27 1,000 to 100,000 μg Pb/L (U.S. EPA, 2006). 

28 In previous AQCDs, growth effects of Pb have been reported in fish (growth inhibition), birds 

29 (changes in juvenile weight gain), and frogs (delayed metamorphosis, smaller larvae). Growth effects 

30 observed in invertebrates and vertebrates underscore the importance of lifestage to overall Pb 

31 susceptibility. In general, juvenile organisms are more sensitive than adults. Several studies since the 2006 

32 Pb AQCD have demonstrated adverse effects of Pb on growth at lower concentrations than in previous 

33 literature. Among the animal taxa tested, aquatic invertebrates were the most sensitive to the effect of Pb, 
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1 with adverse effects being reported as low as 4 μg Pb/L. Growth of juvenile freshwater snails L. stagnalis

2 was inhibited at EC20 <4 μg Pb/L (Grosell & Brix, 2009; Grosell et al., 2006b). In the freshwater mussel, 

3 fatmucket (L. siliquoidea) juveniles were the most sensitive lifestage (N. Wang et al., 2010). A chronic 

4 value of 10 μg Pb/L in a 28 day exposure of 2-month-old fatmucket juveniles was the lowest genus mean 

5 chronic value ever reported for Pb. Evidence is sufficient to conclude that there is a causal relationship 

6 between Pb exposures and growth effects in plants and invertebrates. Evidence is suggestive of a 

7 causal relationship between Pb exposures and growth effects in vertebrates.

7.4.4. Physiological Stress

8 In this review and in previous Pb AQCDs there is consistent and coherent evidence that Pb induces 

9 oxidative stress in plants, invertebrates, and vertebrates. This is consistent with evidence in humans and 

10 experimental animal studies for oxidative stress development, due in many instances to the antagonism of 

11 normal metal ion functions (Section 5.2.4). This oxidative stress is characterized by increased reactive 

12 oxygen species and membrane and lipid peroxidation that can promote tissue damage, cytotoxicity, and 

13 dysfunction. Increased reactive oxygen species are often followed by a compensatory and protective 

14 upregulation in antioxidant enzymes, such that this observation is indicative of oxidative stress 

15 conditions. Additionally, continuous reactive oxygen species production may overwhelm this defensive 

16 process leading to further oxidative stress and injury.

17 Building on this strong body of evidence presented previously, recent studies provide consistent 

18 and coherent evidence of upregulation of antioxidant enzymes and increased lipid peroxidation associated 

19 with Pb exposure in many species of plants, invertebrates and vertebrates. In plants, increases of 

20 antioxidant enzymes with Pb exposure occur in algae, aquatic mosses, floating and rooted aquatic 

21 macrophytes, and terrestrial species. There is considerable evidence for antioxidant activity in 

22 invertebrates, including gastropods, mussels, and crustaceans, in response to Pb exposure. Markers of 

23 oxidative damage are also observed in fish and amphibians. Across all biota, there are differences in the 

24 induction of antioxidant enzymes that appear to be species-dependent.

25 Additional stress responses to Pb exposure observed in terrestrial and aquatic invertebrates 

26 including elevated heat shock proteins, osmotic stress, lowered metabolism and decreased glycogen levels 

27 have been reported since the 2006 Pb AQCD. Heat shock protein induction by Pb exposure has been 

28 observed in zebra mussels and mites. Elevated expression of heat shock protein orthologs were reported 

29 for the first time in the hypothalamic and mesencephalic brain regions of Pb-treated fish (Giusi et al., 

30 2008). Crayfish exhibit a Pb-induced hypometabolism under conditions of environmental hypoxia in the 

31 presence of the metal (Morris et al., 2005). Tissue volume regulation is adversely affected in freshwater 

32 crabs (Amado et al., 2006). Glycogen levels in the freshwater snail B. glabrata were significantly 
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1 decreased at near environmentally relevant concentrations of Pb (50 μg/L and higher) (Ansaldo et al., 

2 2006). 

3 Upregulation of antioxidant enzymes and increased lipid peroxidation are considered to be reliable 

4 biomarkers of stress, and suggest that Pb exposure induces a stress response in those organisms which

5 may increase susceptibility to other stressors and reduce individual fitness. Evidence is sufficient to 

6 conclude that there is a causal relationship between Pb exposures and physiological stress in plants, 

7 invertebrates and vertebrates. 

7.4.5. Hematological Effects

8 Hematological responses are commonly reported effects of Pb exposure in invertebrates and 

9 vertebrates in both aquatic and terrestrial systems. In environmental assessments of metal-impacted 

10 habitats, ALAD is a recognized biomarker of Pb exposure (U.S. EPA, 2006). ALAD activity is negatively 

11 correlated with total Pb concentration in bivalves. Lower ALAD activity has been correlated with 

12 elevated blood Pb levels in fish and mammals as well. In the 1986 AQCD, decreases in red blood cell 

13 ALAD activity following Pb exposure were well documented in birds and mammals. Further evidence 

14 from the 2006 Pb AQCD and this review of Pb effects on ALAD enzymatic activity, including effects in 

15 bacteria, amphibians and additional field and laboratory studies on fish, suggest this enzyme is an 

16 indicator for Pb exposure across a wide range of taxa. Limited evidence of Pb effects on other blood 

17 parameters including altered serum profiles and changes in white blood cell counts in fish and amphibians 

18 support the finding of the hematological system as a target for Pb in natural systems. This evidence is 

19 strongly coherent with observations from human epidemiologic and animal toxicology studies (Section 

20 5.7). There is consistent toxicological and epidemiologic evidence that exposure to Pb induces adverse 

21 effects on hematological endpoints, including altered heme synthesis mediated through decreased ALAD 

22 and ferrochelatase activities, decreased red blood cell survival and function, and increased red blood cell 

23 oxidative stress. Taken together, the overall weight of epidemiologic and toxicological evidence is 

24 sufficient to conclude that a causal relationship exists between exposure to Pb and hematological effects 

25 in humans and laboratory animals (Section 5.7). Based on observations in both terrestrial and aquatic 

26 organisms and additionally supported by toxicological and epidemiological findings in laboratory animals 

27 and humans evidence is sufficient to conclude that there is a causal relationship between Pb exposures 

28 and hematological effects in invertebrates and vertebrates. 

7.4.6. Developmental and Reproductive Effects

29 Evaluation of the literature on Pb effects in aquatic and terrestrial species indicates that exposure to 

30 Pb is associated with reproductive effects. Various endpoints have been measured in aquatic and terrestrial 

31 organisms to assess the effect of Pb on development, fecundity and hormone homeostasis. Evidence in 
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1 this review and the previous Pb AQCDs from invertebrate and vertebrate studies indicate that Pb is 

2 adversely affecting reproductive performance in multiple species. However, there are typically only 

3 limited studies available from different taxa. 

4 In plants, few studies are available that specifically address reproductive effects of Pb exposure. 

5 One study with grass pea showed Pb exposure increased pollen sterility.

6 Several studies with invertebrates provide evidence for adverse impacts to embryonic development, 

7 specifically in snails, clams and rotifers. In addition to affecting the embryo, Pb can alter developmental 

8 timing, sperm morphology and hormone homeostasis. In fruit flies, Pb exposure increased time to 

9 pupation and decreased pre-adult development. Sperm morphology was altered in earthworms exposed to 

10 Pb-contaminated soils. Pb may also disrupt hormonal homeostasis in invertebrates as studies with moths 

11 have suggested. 

12 Reproductive effects have also been observed in multi-generational studies. Larval settlement rate, 

13 and rate of population increase was adversely impacted in rotifers. Rotifers have decreased fertilization 

14 rate associated with Pb exposure that appeared to be due to decreased viability of sperm and eggs. 

15 Evidence of multi-generational toxicity of Pb is also present in terrestrial invertebrates, specifically 

16 springtails, mosquitoes, carabid beetles and nematodes where decreased fecundity in progeny of Pb17 exposed individuals was observed. 

18 In aquatic vertebrates there is evidence for reproductive and developmental effects of Pb. Pb19 exposure in frogs has been demonstrated to delay metamorphosis, decrease larval size and produce subtle 

20 skeletal malformations. Previous Pb AQCDs have reported developmental effects in fish, specifically 

21 spinal deformities in larvae. In the 2006 Pb AQCD, decreased spermatocyte development in rainbow trout 

22 was observed at 10 μg Pb/L and, in fathead minnow testicular damage occurred at 500 μg Pb/L. In fish, 

23 there is new evidence of Pb in this ISA on effects on steroid profiles and additional reproductive 

24 parameters. Reproduction in fathead minnows was affected in breeding exposures following 300 day 

25 chronic toxicity testing. However, reproductive performance was unaffected in zebrafish exposed to Pb26 contaminated prey. Additional reproductive parameters in fish observed to be impacted by Pb include 

27 decreased oocyte diameter and density in toadfish associated with elevated Pb levels in gonad.

28 In terrestrial vertebrates, evidence from Chapter 7 of this ISA and in previous Pb AQCDs indicates 

29 an association between observed adverse reproductive effects and Pb exposure. Reproductive effects 

30 observed in birds near areas of Pb-contamination or where Pb tissue concentration has been correlated 

31 with effects include declines in clutch size, number of young hatched, number of young fledged, 

32 decreased fertility, and decreased eggshell thickness. Decreased testis weight was observed in lizards. In 

33 mammals, few studies in the field have addressed Pb specifically, due to most available data in wild or 

34 grazing animals being from near smelters, where animals are co-exposed to other metals. 

35 In Chapter 5 evidence from mammals demonstrates a consistency of adverse effects of Pb on sperm 

36 and the onset of puberty in males and females with strong evidence from both toxicology and 
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1 epidemiology studies. Other reproductive endpoints including spontaneous abortions, pre-term birth, 

2 embryo development, placental development, low birth weight, subfecundity, hormonal changes, and 

3 teratology were also affected, but less consistently. New toxicological data support trans-generational 

4 effects, a finding that is also an area of emerging interest in ecology. The evidence presented in Section 

5 5.8 is sufficient to conclude that there is a causal relationship between Pb exposure and reproductive 

6 effects.

7 Adverse effects of Pb on reproduction in invertebrates and vertebrates indicate that Pb is likely 

8 adversely affecting fecundity of Pb-exposed organisms in both aquatic and terrestrial habitats, and 

9 therefore the evidence is sufficient to conclude that there is a causal relationship between Pb exposures

10 and reproductive effects in terrestrial and aquatic invertebrates and vertebrates. In plants, the 

11 evidence is inadequate to conclude a causal relationship between Pb exposures and plant 

12 reproductive effects.

7.4.7. Neurobehavioral Effects

13 Evidence from laboratory studies and limited data from field studies reviewed in Chapter 7 have 

14 shown adverse effects of Pb on neurological endpoints in both aquatic and terrestrial animal taxa. These 

15 include changes in behaviors that may decrease the overall fitness of the organism. There is also evidence 

16 from both invertebrate and vertebrate studies that Pb adversely affects behaviors that may decrease the 

17 ability of an organism to escape predators or capture prey.

18 Central nervous system effects in fish recognized in previous Pb AQCDs include effects on spinal 

neurons and brain receptors. New evidence from this review identifies the MAPKs ERK1/2 and p38MAPK 19

20 as possible molecular targets for Pb neurotoxicity in catfish (Leal et al., 2006). Additionally, there is new 

21 evidence for neurotoxic action of Pb in invertebrates with exposure to Pb observed to cause changes in 

22 the morphology of GABA motor neurons in nematodes (C. elegans) (Du & Wang, 2009).

23 Decreased food consumption of Pb-contaminated diet has been demonstrated in some invertebrates 

24 (snails) and vertebrates (lizards, pigs). Pb may also decrease the ability of an organism to capture prey or 

25 escape predation. For example, Pb exposure has been demonstrated to adversely affect prey capture 

26 ability of certain fungal and fish species. In limited studies available on snails, tadpoles, scallops,

27 hatchling turtles and fish there is evidence that Pb may affect the ability to escape or avoid predation. The 

28 motility of nematodes was adversely affected in Pb-contaminated soils (D. Y. Wang & Xing, 2008). In a 

29 laboratory study, Pb-exposed gull chicks exhibited abnormal behaviors such as decreased walking, erratic 

30 behavioral thermoregulation and food begging that could make them more vulnerable in the wild (Burger 

31 & Gochfeld, 2005). Lizards exposed to Pb through diet in the laboratory exhibited abnormal coloration 

32 and posturing behaviors. Other behavioral effects affected by Pb exposure include increased hyperactivity 

33 in fish and hypoxia-like behavior in frogs. 
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1 These findings are coherent with findings from studies in laboratory animals as described in

2 Section 5.3 that show that Pb induces changes in learning, memory, attention and motor skills. Pb induced 

3 new behaviors including hyperactivity and mood disorders. Within the sensory organs, the visual and 

4 auditory systems are affected by Pb exposure. Changes in structure and function of neurons and 

5 supporting cells in the brain are detailed including effects on the blood brain barrier. Mechanisms 

6 including the displacement of physiological cations, oxidative stress and changes in neurotransmitters and 

7 receptors are detailed. Based on evidence from several cohort and cross-sectional studies in diverse 

8 populations, the overall weight of the available evidence provides clear and consistent evidence of 

9 association between blood Pb concentrations and decrements in neurodevelopmental outcomes in young 

10 children (Section 5.1). In addition to the consistency of findings in children, the evidence is strengthened 

11 by the coherence of findings with toxicological studies and by coherence of association of blood Pb with 

12 a spectrum of related endpoints including IQ, verbal and reading skills, motor coordination, mood and 

13 attention problems, and behavioral problems. The evidence presented in the health chapter is sufficient to 

14 conclude that there is a causal relationship between Pb exposure and neurobehavioral effects (Section 

15 5.3). These data from laboratory toxicology studies, especially neurobehavioral findings and structural 

16 changes are highly coherent with data from ecological studies. Overall, the evidence from aquatic and 

17 terrestrial systems is sufficient to conclude that there is a causal relationship between Pb exposures and 

18 neurobehavioral effects in invertebrates and vertebrates.

7.4.8. Other Physiological Effects

19 In addition to the above mentioned physiological effects of Pb on organisms in terrestrial and 

20 aquatic systems for which there is sufficient evidence to infer causality, there are a few recent studies that 

21 can be linked to effects observed in humans for which there is insufficient evidence across taxa. Pb 

22 exposure has been demonstrated to result in changes to DNA structure and chromosomal alterations in 

23 plants, gastropods, mussels and fish. DNA damage, chromosomal damage and aberrations, and 

24 micronucleus formation are also observed in humans and laboratory animals exposed to Pb (Section 

25 5.10). Additional new evidence in this review indicates that Pb can interfere with renal function in fish, 

specifically with ionoregulation of Na and Cl and reabsorption Ca2+, Mg2+ 26 glucose and water (Patel et al., 

27 2006). In humans and laboratory animals, Pb is a recognized nephrotoxicant and is considered to be 

28 causal of kidney damage (Section 5.5). 

7.4.9. Community and Ecosystem Level Effects 

29 Uptake of Pb into aquatic and terrestrial organisms and subsequent effects on survival, 

30 reproduction, growth, behavior and other physiological variables at the species scale are likely to result in

31 effects at the population, community and ecosystem scale. The effects may include alteration of predator-
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1 prey dynamics, species richness, species composition, and biodiversity. There are few field studies 

2 available that directly consider effects of Pb on these measures of ecosystem health. Ecosystem-level 

3 studies are complicated by the confounding of Pb exposure with other factors such as trace metals and 

4 acidic deposition. In natural systems, Pb is often found co-existing with other stressors, and observed 

5 effects may be due to cumulative toxicity.

6 Most direct evidence of community and ecosystem level effects is from near point sources. For 

7 terrestrial systems there are several decades of research on impacts to natural ecosystems near smelters, 

8 mines, and other industrial sources of Pb where Pb levels are elevated. Those impacts include decreases in 

9 species diversity and changes in floral and faunal community composition. For aquatic systems, the 

10 literature focuses on evaluating ecological stress from Pb originating from urban and mining effluents 

11 rather than atmospheric deposition. In laboratory studies and simulated ecosystems, where it is possible to 

12 isolate the effect of Pb, this metal has been shown to alter competitive behavior of species, predator-prey 

13 interactions and contaminant avoidance. These dynamics may change species abundance and community 

14 structure at higher levels of ecological organization. Effects of Pb on mortality, growth, physiological 

15 stress, blood, neurobehavior and developmental and reproductive endpoints at the individual level are 

16 expected to have ecosystem-level consequences, and thus provide consistency and plausibility for 

17 causality in ecosystem-level effects.

18 Avoidance response to Pb exposure varies widely in different species and this could affect 

19 community composition. For example, frogs and toads lack avoidance response while snails and fish 

20 avoid higher concentrations of Pb (U.S. EPA, 2006). New evidence since the 2006 Pb AQCD indicates 

21 that some species of worms avoid Pb-contaminated soils (Langdon et al., 2005)

22 In terrestrial ecosystems, most studies show decreases in microorganism abundance, diversity, and 

23 function with increasing soil Pb concentration. Specifically, shifts in nematode communities, bacterial 

24 species, and fungal diversity have been observed. Furthermore, presence of arbuscular mycorrhizal fungi 

25 may protect plants growing in Pb-contaminated soils. Increased plant diversity was shown to ameliorate 

26 effects of Pb contamination on a microbial community. 

27 In aquatic ecosystems there are numerous field studies on reductions of species abundance, 

28 richness or diversity particularly in benthic macroinvertebrate communities coexisting with other metals. 

29 For example, in the 2006 Pb AQCD, the Coeur d’Alene River watershed in Idaho, U.S. was used as a case 

30 study for Pb effects at the population and community level. Significant negative correlations were 

31 observed between Pb in water column and total taxa richness and EPT taxa richness in the river. In a 

32 simulated aquatic microcosm a reduction in abundance and richness of protozoan species was observed 

33 with increasing Pb concentration from 50 to 1,000 μg Pb/L (Fernandez-Leborans & Antonio-Garcia, 

34 1988)

35 Since the last Pb AQCD, there is further evidence for effects of Pb in sediment-associated 

36 communities. Exposure to three levels of sediment Pb contamination (322, 1,225, and 1,465 μg Pb/g dry 
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1 weight) in a microcosm experiment significantly reduced nematode diversity and resulted in profound 

2 restructuring of the community structure (Mahmoudi et al., 2007). Sediment-bound Pb contamination 

3 appears to differentially affect members of the benthic invertebrate community, potentially altering 

4 ecosystems dynamics in small urban streams (Kominkova & Nabelkova, 2005). Although surface water 

5 Pb concentrations in monitored streams were determined to be very low, concentrations of the metal in 

6 sediment were high enough to pose a risk to the benthic community (e.g., 34 to 101 mg Pb/kg). These 

7 risks were observed to be linked to benthic invertebrate functional feeding group, with collector-gatherer 

8 species exhibiting larger body burdens of heavy metals than benthic predators and collector-filterers.

9 In a new study conducted since the 2006 Pb AQCD, changes to aquatic plant community 

10 composition have been observed in the presence of elevated surface water Pb concentrations at three lake 

11 sites impacted by mining effluents. The site with highest Pb concentration (103-118 μg Pb/L) had lowest 

12 number of aquatic plant species when compared to sites with lower Pb concentrations (78-92 μg Pb/L) (V. 

13 K. Mishra et al., 2008). This shift toward more Pb-tolerant species is also observed in terrestrial plant 

14 communities near smelter sites (U.S. EPA, 1986, 2006). Certain types of plants such as rooted and 

15 submerged aquatic plants may be more susceptible to aerially-deposited Pb resulting in shifts in Pb 

16 community composition. High Pb sediment concentrations are linked to shifts in amphipod communities 

17 inhabiting plant structures. 

18 In many cases it is difficult to characterize the nature and magnitude of effects and to quantify 

19 relationships between ambient concentrations of Pb and ecosystem response due to existence of multiple 

20 stressors in natural systems. However, the evidence for Pb effects at higher levels of ecological 

21 organization is sufficient to conclude that there is a causal relationship between Pb exposures and the 

22 alteration of species richness, species composition and biodiversity in terrestrial and aquatic 

23 ecosystems. 
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