Reference Work Entry

Handbook of Sustainable Engineering

pp 137-161

Fundamental Toxicology Methods and Resources for Assessing Water-relatedContamination

  • Keith R. CooperAffiliated withDepartment of Biochemistry & Microbiology, Rutgers, The State University of NJ

Abstract

This chapter provides a brief overview of basic toxicological methods and approaches which can be used by engineers in the field to make a rapid environmental risk determination. In addition, Internet sites which deal with specific contaminants, standard operating procedures, and methods for assessing deleterious effects on organisms living and depending on ecosystem resources are provided.

Abstract

This chapter provides a brief overview of basic toxicological methods and approaches which can be used by engineers in the field to make a rapid environmental risk determination. In addition, Internet sites which deal with specific contaminants, standard operating procedures, and methods for assessing deleterious effects on organisms living and depending on ecosystem resources are provided.

1 Introduction

When talking about a systems engineering approach to solving these problems, it is critical that humans be included as an integral part of the World Ecosystem and not as a separate component acting independently. Sustainability of freshwater water supplies is a major issue for both developed and developing countries. Decreasing freshwater water supplies, due in part to climate change and increasing population densities (Fig. 10.1), are a major threat to stability around the globe (Brown 2003; Brown et al. 1998; Grumbine and Xu 2011). In certain parts of the world, scarcity of water is already having a major impact on the habitability of lands, which results in migration of humans and animals into water richer areas (Fig. 10.2). Water shortages will become more frequent due to increasing population growth, over pumping of aquifers/rivers, climate change, hydroelectric power, and shifts in food preferences (Wada et al. 2010). Water disputes will likely be the cause for major regional conflicts. Approximately seven billion people currently inhabit the earth and it is estimated to rise to nine billion by 2044. The current top five countries for total population include: China (1.3 billion), India (1.1 billion), the United States (300 million), Indonesia (240 million), and Brazil (200 million). Water sustainability is driven by having sufficient quantities of usable water to maintain human needs, crop irrigation, livestock, and ecologically important animals and plants. The density of a population also can result in threats to sustainability of local and regional resources (Fig. 10.1). Water usage is directly related to population density, agriculture, and ecosystem needs, and if demand outstrips supply then the ecosystem is not sustainable.
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Fig. 10.1

Worldwide population density map (Map Source: Center for International Earth Science Information Network) and water severity areas (arrows)

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Fig. 10.2

Pathways for contaminant movement (arrows) between the abiotic matrices comprising the ecosystem. Biotic factors can modify fate and transport of contaminants within an ecosystem by increasing or decreasing bioavailability

Water is essential for all life and both chemical and biological contaminants from both natural and anthropogenic sources can dramatically limit the suitability of water for human, animal, and plants. It is important to realize that local actions (eutrophication, overfishing, altered food types, industrialization, and fossil fuel usage) can impact water and water-related resources both locally and in some cases on a worldwide basis. The interconnectivity is due to the sharing of the air sheds, watershed, and oceans. The ramifications of economic prosperity, unfettered industrialization, and increased dependence on fossil fuels without proper environmental controls and regulations can result in both local and international water supplies becoming contaminated. The health and well-being of humans and all other species is inseparable from the health and well-being of the world ecosystems (Edwards 2005). Human beings have the responsibility to maintain the quality of water, air, and soil to enhance the well-being for all species (The Netherlands National Environmental Policy). For any successful sustainable program there are Four E’s: Ecological sustainability, Economic sustainability, Equitable resource allocation, and Education of the population. All of these concepts need to be included in any sustainable engineering approach that will be successful over the long term. If any of the Four E’s is not met, then long-term sustainability is unlikely. Therefore, it is essential that novel engineering approaches be developed to remove both biological and chemical contamination, as well as recycle water and limit natural resource damage.

This chapter provides a brief tutorial on basic toxicological concepts, tools, and information that are important for an engineer to be aware of when determining the suitability of a treatment process or the threats to a community or its ecosystem.

1.1 Toxicological Information Resources

This chapter provides a brief overview of both human and ecological toxicological approaches and standard sample collection methods. These topics are too expansive to be adequately covered in a single chapter; therefore, references to additional resources are provided throughout the text and within Table 10.1. The approaches described in this section are standard operating procedures (SOPs) used in toxicology assessments and will be useful in designing assessment protocols at any location. The sampling protocols are from a number of different government agencies and are listed in Table 10.1. Depending on the availability of resources, the SOPs can be modified to allow for some level of assessment with the understanding that the robustness of the results could be affected. The country in which an engineer is working may have SOPs that are available through the Environmental Ministry or comparable industries. It is important to realize that rules and regulations promulgated in highly developed countries for individual contaminants may not be appropriate or achievable for less developed countries since the relative risks from different pathogens or starvation may far out way the risks from an individual contaminant. There are a number of USEPA Web sites with specific compound assessments such as IRIS Toxicological Reviews http://​www.​epa.​gov/​iris/​, pesticides http://​www.​epa.​gov/​pesticides/​reregistration/​status.​htm, and Drinking Water Standards and Health Advisories http://​www.​epa.​gov/​ost/​. The values derived following these assessments are to protect the public based on very low risk numbers: 1 in a million or 1 in 100,000 and although desirable to be achieved may be impractical and not achievable in less developed countries.
Table 10.1

Listing of standard operating procedures (SOPs) and other resources used in toxicological assessments from various government agencies

Agency

Information available

 

WHO: IPCS

WHO works to establish the scientific basis for the safe use of chemicals, and to strengthen national capabilities and capacities for chemical safety. http://​www.​who.​int/​ipcs/​en/​

 

USEPA Emergency Response Team

SOPs for sampling air, water, sediment, and soil along with specific methods for analysis: general field sampling, sampling equipment decontamination, general air sampling, surface water sampling, groundwater well sampling, sediment sampling, and soil sampling. www.​ert.​org

 

USEPA: NCEA

The mission of NCEA is to provide guidance about how pollutants may impact our health and the environment. The compound of interest can be searched to see if a review has been carried out. http://​www.​epa.​gov/​ncea/​index.​htm

 

USEPA Duluth, Minnesota

ECOTOX database released in 2000 by the USEPA and managed by the USEPA Duluth laboratory has extensive ecological data both for terrestrial and aquatic species. http://​www.​epa.​gov/​ecotox/​ecotox_​home.​htm

 

US National Park Service

Environmental Contaminants Encyclopedia, Ed. Roy Irwin Discussing www.​nature.​nps.​gov/​hazardssafety/​toxic/​/

 

NOAA (USA) and UN

NOAA Status and Trends and Mussel Watch. http://​ccma.​nos.​noaa.​gov/​about/​coast/​nsandt/​welcome.​html and International Mussel Watch. http://​ccma.​nos.​noaa.​gov/​stressors/​pollution/​assessments/​as_​intl_​mw_​study.​html

 

Syracuse Research Corp. (USA)

Chemical, physical, and fate data on specific compounds PHYSPROP and BIOLOG for microbial degradation database http://​srcinc.​com/​what-we-do/​efdb.​aspx

 

Oak Ridge National Laboratory

Ecological Risk Analysis: Guidance, Tools and Applications (www.​esd.​ornl.​gov/​programs/​ecorisk/​ecorisk.​html)

 

ATSDR (USA)

Provides individual chemical Toxicological Profiles (www.​atsdr.​cdc.​gov/​)

 

ATSDR Agency for Toxic Substances and Disease Registry, ICPS InternationalProgramme on Chemical Safety, NCEA National Center for EnvironmentalAssessment, NOAA National Oceanographic and Atmospheric Administration,USEPA United States Environmental Protection Agency, WHO World HealthOrganization

2 Routes of Entry of Pollutants into Ecosystems

This is a simple overview of the fate and transport of chemicals of concern. It is important for any ecological based assessments to realize that chemicals of concern (COCs) can enter an ecosystem by many routes (transport through air, water, or soil/sediments) from point and nonpoint sources. The fate of the COC can be impacted by a number of physical parameters. Any town or city, whether small or large, is part of a larger water- and air shed, which has many different inputs. Therefore, it is essential that the surrounding lands adjacent to and within a drainage area be examined in relationship to their surrounding environs. Establishing a baseline level of chemicals present in the surrounding area and waters is important to be able to determine increases or decreases in a particular chemical. Because a number of chemicals occur naturally, it is important to understand the naturally occurring levels (Table 10.2). The flora and fauna that are present in surrounding areas are important in assessing the potential impact of specific COCs on the organisms and potentially humans. Because of the diversity of the plant and animals (ecological receptors), it is impossible to assess each species independently. It is vital, therefore, to identify representative or surrogate species that can be used to assess potential impacts on larger groups of organisms. Establishment of specific sites that can be monitored yearly over decades provides valuable chemical trend information. This is one of the underlying principles for historical and continued collection of Mussel Watch Data (oysters, mussels, and clams). This data set has existed since the mid-1980s and allows for assessing change in chemical occurrence. The use of a surrogate species may not always be the most sensitive organisms’ but they do allow for evaluation of a single species over multiple years as an indicator of improvement, remaining constant or deterioration of the organisms exposure and possible health. If such a species is identified, that specific species should be used as an indicator species for that location. If a sensitive species approach is taken, then protection of this species will protect other more resistant species. Biochemical and physiological parameters can be evaluated at the cellular and individual level which may or may not have effects at the population and community level, but can be correlated with COCs and habitat alterations.
Table 10.2

Element levels present in earth’s crust (parts per million, mg/Kg) (As reported in Mason B. Principles of Geochemistry, Third Edition)

Element

Crustal

average

Element

Crustal

average

Element

Crustal

average

 

H

1,400

Ge

1.5

Tb

0.9

 

Li

20

As

1.8

Dy

3.0

 

Be

2.8

Se

0.05

Ho

1.2

 

B

10

Br

2.5

Er

2.8

 

C

200

Rb

90

Tm

0.5

 

N

20

Sr

375

Yb

3.4

 

O

466,000

Y

33

Lu

0.5

 

F

625

Zr

165

Hf

3

 

Na

28,300

Nb

20

Ta

2

 

Mg

20,900

Mo

7

W

1.5

 

Al

81,300

Ru

0.01

Re

0.001

 

Si

277,200

Rh

0.005

Os

0.005

 

P

1,050

Pd

0.01

Ir

0.001

 

S

260

Ag

0.07

Pr

0.01

 

Cl

130

Cd

0.2

Au

0.004

 

K

25,900

In

0.1

Hg

0.08

 

Ca

36,300

Sn

0.2

Pb

13

 

Sc

22

Te

0.01

Bi

0.2

 

Ti

4,400

I

0.5

Th

7.2

 

V

135

Cs

3

U

1.8

 

Cr

100

Ba

425

   

Mn

950

La

30

   

Fe

50,000

Ce

60

   

Co

25

Pr

8.2

   

Ni

75

Nd

28

   

Cu

55

Sm

6.0

   

Zn

70

Eu

1.2

   

Ga

15

Gd

5.4

   

It must be realized that individual species can be impacted by nonchemical stressors that can have deleterious effects. Physical parameters such as pH, alkalinity, salinity, dissolved oxygen, water transparency, and temperature can result in the loss of keystone species. It should be recognized that physical parameters often limit distributions of organisms. Therefore, plants and animals living in an ecosystem can be impacted both by physical parameters and COCs. Because of this fact the EPA’s (1992) Framework for Ecological Risk Assessment includes “stressor response” which would include physical parameters as do later EPA documents (USEPA 199319951997a,b1998). The classic example of this is the effect of pH in freshwater systems and the loss of species dependent on pH tolerances. These physical parameters are easily measured and should be incorporated into any monitoring plan, since they have profound effects on COCs’ bioavailability and uptake (Rand et al. 1995; Schwarzenbach et al. 2003).

The selection of the animals and plants to be used in the risk assessment will be in some cases site specific. Generally, there will be both terrestrial and aquatic species and associated toxicity data used for determining which, if any, COCs pose a hazard. The selected species may also be used as indicator species that can be collected and analyzed for COCs. The sampling may be done to determine if COCs are reaching elevated levels. This will also allow for COC trends to be developed. Samples of water, soil, sediment, and air may be collected in concert or as independent samples to establish reference or baseline levels. In many locations, there is no plan in place. High volume air samples and semipermeable membrane water and soil samples could be added if grab samples would not be able to detect COCs (Huckins 2002; Petty et al. 2000; Sun et al. 2008).

2.1 Fate and Transport

Pollutants can enter an ecosystem through both human activities and from naturally occurring processes (weathering of rocks, volcanic activity, and others). While knowledge of sources will give you important information on the type of contaminants your environment might be exposed to, the concentration or dose of a compound will dictate the response observed in the wild (Crane and Newman 2000). As all toxicologists know “The Dose Makes the Poison” and once this level has been exceeded, toxicity will be observed (Di Giulio and Newman 2008; Eaton and Gilbert 2008). In all of the different matrices (surface and groundwater, sediments, and terrestrial soil), there are both point and nonpoint sources of contamination. Ecosystems are impacted from contamination contributed from atmospheric, terrestrial, and aquatic sources (Suter 19931998). Below is a brief description of sources of contaminants entering into the different media and illustrated in Fig. 10.2.

The conceptual diagram shown in Fig. 10.3 illustrates the requirement that the chemical can be bound or in a free state during exposure. Generally if the compound is sequestered on a particle or other organic material, it is not available to be taken up by an organism since the compound must be free to be able to be transported across a biological membrane. A number of factors, such as charge and size, as well as others, can determine whether a compound will be taken up by an organism. The free compound must diffuse or be transported across the membrane. Once inside the cell, the compound can be metabolized, stored, or excreted. If the compound reaches a target tissue where it causes toxicity at sufficient concentration then sublethal or lethal effects may be observed (Lehman-McKeeman 2008) (Fig. 10.4). The physical/chemical environments in which they exist can dramatically alter the chemical species which is available for accumulation (Spacie et al. 1995; Sprague 1985). These general principles apply to organic compounds as well as metallic compounds (USEPA 2007).
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Fig. 10.3

Conceptual diagram for evaluating bioavailability to organisms from a compound (suns) present in the environment in a bound (NOT BIOAVAILABLE) and free state (BIOAVAILABLE) for exposure, uptake across the biological membrane resulting in bioaccumulation within the organism, and subsequent internal transport and distribution

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Fig. 10.4

Dose-response curve with no observed effect level (NOEL). Lowest observed adverse effect level (NOAEL) and LC50 indicated on the graph

Atmospheric contamination can enter the terrestrial soil and surface waters through precipitation (wet deposition) or from suspended particulates (dry deposition) (Loganathan and Kannan 1994; Scheringer 2009). There can be a contribution from both the terrestrial soil compartment and the surface water compartment into the atmosphere via particle suspension and volatilization processes. The contamination of the terrestrial soil from atmospheric deposition or from crustal compounds (background) can flow into the surface waters and end up in suspension or depositing on the bottom as sediment. Groundwater can become contaminated from compounds leaching into the groundwater (Burgess et al. 2010). Surface waters can also contaminate groundwater when they are in direct contact (e.g., saltwater intrusion) or as the groundwater is being recharged. Groundwater contamination can reach surface waters in the basal flow of rivers and through seeps. The ability of a compound to move between these different compartments is determined by the physical/chemical properties of the individual compound. Biotic factors can result in modifying the amount of a COC that could be released from any of these compartments. Regional point sources (cities, superfund sites, and industrial activity) of contamination can result in contamination spreading into several of these compartments.

The movement of contaminants through the different matrices and ultimately into plants and animals (environmental fate) is determined by the physical characteristics of the compound (molecular weight, lipophilicity, chemical species) and that of the environmental compartment it is entering (Fig. 10.4). Bioavailability is an important concept that determines what levels of specific compounds will be present in the biota inhabiting contaminated and noncontaminated environments (Birak et al. 2001; Rand et al. 1995). Bioavailability is determined by many factors that determine how much of the compound or metal is in a free state that can be taken up by an organism or plant. Physical/chemical properties of the matrix into which a compound is associated and the properties of the surrounding medium will affect the level of compound that is free to be taken up by the organism. Factors such as pH, total organic carbon, and acid volatile sulfides can dramatically effect the uptake of a compound (DiToro et al. 1990). A full discussion of release from a specific matrix and the uptake mechanisms based on characteristics of a metal or organic compound are too complicated for this introductory chapter (Jorgensen et al. 1991; Sample and Suter 1994; Sample et al. 1996; Schwarzenbach et al. 2003; USEPA 2003).

A critical physical characteristic of a compound is polarity, which will determine the compound’s affinity for polar (water) and nonpolar matrixes (lipids and nonpolar materials). In the case of water, oxygen draws electrons away from the two hydrogen atoms, which creates a partial negative charge on the oxygen and a partial positive charge on the hydrogen atoms. Because of the polar nature of water, cations and anions (i.e., positively and negatively charged compounds) are attracted to water and are soluble. These include inorganic salts (alkali earth metals) and polar organic compounds containing polar groups (e.g., alcohols, acids, and amines). Compounds that are nonpolar generally will partition into nonpolar compartments that contain lipids (such as living organisms) and organic carbon (organic enriched particles). An important property of chemicals, which is used in estimating the potential to accumulate into a living organism is the octanol/water partitioning coefficient (K ow ). As the name implies, it is the concentration of a compound in octanol divided by the concentration of the compound in the water phase at equilibrium. This is a measure of the hydrophobicity of the compound. This value can be used to estimate the bioconcentration of compounds into organisms. In a similar fashion, compounds will partition between the different compartments described in Fig. 10.3 based on physical/chemical properties of the compound (e.g., Henry’s constant relating distribution of volatile compounds between air and water and soil) (Lyman et al. 1982; Schwarzenbach et al. 2003; Suter 1993).

The residence time of the parent compound and its metabolites are determined by the compound’s molecular stability (Lyman 1995). There are many methods by which a compound can be broken down in the environment (e.g., hydrolysis, auto-oxidation, photolysis, and enzymatic biotransformation). The rate of breakdown of a compound will determine the half-life in any particular medium (Newman and Unger 2003). Persistent organic pollutants (POPs) are generally recalcitrant to physical, chemical, and biochemical transformation (e.g., PFOA/PFOS, DDT/DDE, PCBs, and 2,3,7,8-sustituted dioxins). However, in some instance, metabolism (e.g., Benzo(a)pyrene, dieldrin) and photolysis (e.g., petroleum hydrocarbons) can result in activation of a compound to reactive chemical species that can result in cellular transformation leading to cancer, developmental effect, or acute toxicity.

Metals are nonbiodegradable and do not breakdown in a similar fashion to organic compounds in the environment (Faustman and Omenn 2008). Metals can exist as multiple converting species that are determined by the environmental chemistry of the medium they are located in. In the case of metal contaminants, they also are persistent and their ionization state will affect their residence time and transport within the environment (USEPA 2007). USEPA’s Framework for Metal Risk Assessment (http://​www.​epa.​gov/​osa/​metalsframework) is a document that discusses in great detail each of these issues and should be referred to when dealing with metal risk assessments (USEPA 2007). In the case of mercury, the metabolism to methyl mercury results in at least one bioavailable and toxic form of the metal. The presence of other metals such as selenium can antagonize the uptake of mercurial species (Dang and Wang 2011). Metals that are deposited onto the soil or into sediment can have very long residence times and may become mineralized and not biologically available. It is important to note that geological deposits, and the earth’s crust, are comprised of a number of metals that contribute to the general background concentrations for a specific region even without anthropogenic contributions. Background levels of metals in the earth’s crust are summarized in Table 10.2. Background metal concentrations can vary over several orders of magnitude based on the soil type, geography, and other factors. Additional sources of regional and state soil metal levels can be obtained from U.S. EPA ecological soil screening levels (EcoSSLs) document (USEPA 2003). Site-specific assessments will require gathering information on naturally occurring levels and the metals potential to impact flora, wildlife, and humans. A good case study of human activities and geochemistry is the arsenic levels present in Bangladesh that has contaminated groundwater supplies and is a major health issue (Burgess et al. 2010).

2.1.1 Atmospheric

The air shed that overlies a watershed can contribute large quantities of contaminants that are associated with particle deposition, gases, or as dissolved in rain or snow. The sources of contamination can be from both local and long distance transport. Incineration of waste and production of energy using coal and other sources can contribute to atmospheric contamination. Other sources include internal combustion engines, pesticide use on production agriculture, and volatile organic compounds that are used in refrigerants, solvents, and industrial processes. The magnitude of toxic chemical releases that may occur on a yearly basis into the atmosphere is often not appreciated, nor the resident time that these contaminants have in the atmosphere (Loganathan and Kannan 1994). From anthropogenic sources it is estimated that 6 billion tons of CO2, 100 million tons of SO x , 68 million tons of NO x , and 1.1 million tons of chlorofluorocarbons (CFCs) are released. Regional and local sources often contribute to higher local concentrations and corresponding higher concentrations entering surface water, soils, and biota.

2.1.2 Soil

Contamination of soil can occur through deliberate contamination from dumping of industrial and municipal wastes (metals, organic compound radioactive material), direct application of pesticides or herbicides, application of sewage sludge (heavy metals, nitrates, phosphates, detergents) and other soil amendments, and atmospheric deposition of long- and short-range contaminants (metals, sulfur dioxide, nitrogen oxides, dioxins, and furans and other organic compounds). The soil characteristics (organic content, buffering capacity) can affect the fate of the contaminants deposited on the soil and their impact on flora and fauna (Travis and Arms 1988). Erosion of soils is one of the major means by which contaminated soils can reach surface waters and also contribute to sediment contamination. Because watersheds represent large drainage basins, upstream soil contamination can impact downstream water and sediment quality, as it is deposited at the mouth of the rivers and bays or simply deposited in a slow moving portion of a stream. In many instances, soil acceptable levels of contamination are based on human and or terrestrial wildlife risk assessments and exposure is assessed differently for terrestrial pathways. But those same concentrations in soils, when deposited into aquatic systems, can cause deleterious effects to aquatic organisms. In some instances, studies indicate aquatic organisms are more sensitive to certain compounds in sediments, when the same level in soils would not elicit a response in terrestrial organisms.

For metals, it must be recognized that they naturally occur in the earth’s crust and represent nonanthropogenic background levels (Table 10.2). The values in this table allow for a quick reference benchmark for comparison to a soil sample. It should be emphasized that these are just general numbers and regional background levels should be consulted. Background levels can be directly compared to collected samples to determine if the collected samples are significantly higher and could indicate a local source. The actual concentrations would then have to be examined to see if the levels are at a toxic threshold level or higher. This result would determine what actions might be taken.

2.1.3 Surface Water

Surface waters receive anthropogenic (man-made) inputs from atmospheric deposition, non-point source run off from terrestrial sources (petroleum products from automobiles, road salt, etc.) and point source discharges (sewage treatment plants, industrial discharge, landfill leachate etc.). There are also natural sources of contaminants such as the earth’s crust and natural processes (petroleum seeps, volcanism) that release compounds and metals into water and into the atmosphere. Transport in water can occur either by being in solution or in suspension as a micelle, an electrically charged particle formed by an aggregate of molecule, or associated with particulate matter. Compounds and metals can adsorb onto particulate matter in suspension and some of these contaminant-enriched particles will eventually settle out in bottom sediment. In rivers, pollutants can be transported over great distances where they can enter into estuaries and ultimately the oceans. Contaminants associated with eroding soil can be a major source of chemicals entering into surface waters. Release of pollutants into moving surface water is followed by dilution and in some cases degradation. The point of entrance will normally represent the area where you will observe the highest concentration followed by a concentration gradient downstream. The stream flow and physical characteristics may result in depositional areas that should be sampled. In some instances, such as sewage outfalls, there may be a change in the biota upstream from the outfall when compared to biota present downstream from the outfall. The difference from upstream to downstream biota may reflect a contaminant or physicochemical parameters that impact the biota. In the case of small lakes and ponds, where there is no major outlet, chemicals may buildup in these waters. It is also important to understand that freshwater and salt water organisms may respond differently to the same compound, and because of this the no observed adverse effect level (NOAEL) levels are often different (Fig. 10.3). The surface waters and the associated particulate fractions are sinks for accumulating contaminants as they move from upland into estuaries and into the oceans. Surface waters can be contaminated from many different sources resulting in numerous anthropogenic (man-made) compounds being present.

2.1.4 Sediment

Particulates can be either suspended in the water column and or deposited along the bottoms of rivers, estuaries, and depositional areas (sediments) in the ocean. Sediment originates from the contribution of soil washing into the surface waters and from breakdown of organic matter in the receiving waters. Sediment can become contaminated from land sources, as well as from contaminants that accumulate in organisms and or adhere to the sediment from the water column (Fig. 10.4). The physical and chemical characteristics of the sediment will influence the amount of contamination. Sediments with high organic content and or charge will have elevated levels of organic and metals, respectively. The smaller the particle size, the greater the surface area to volume ratio and the higher the contaminant enrichment. Therefore, sediments with large grain size containing mainly sand will have little or no metals or organic contaminants associated with them. Sediment particle size also influences contaminant bioavailability both from physical/chemical characteristics and an organism’s size-specific particle selection. Bioavailability can be reduced in clay and silt sediments compared to coarser sediments. Thus when conducting sediment analyses, it is important to evaluate both grain size and total organic carbon (TOC) in order to assess your contaminant results effectively.

2.1.5 Ground Water

Groundwater, as defined, is water located below the surface of the soil and can be found in the soil pores (sometimes called pore water) or fractures in the geological deposited soils. An aquifer is a saturated zone where usable quantities of water can be withdrawn. Groundwater is recharged from rainwater leaching into the subsurface areas. Groundwater, unlike surface waters, flows very slowly and discharges into seeps or springs that feed into surface waters, sometimes over decades. Therefore, contamination that reached the aquifer decades ago may only now be being discharged into rivers and springs. In any one area, there can be several aquifers that are confined to different depths by geological formations. Generally, the shallow aquifers are at greatest risk of contamination. Contamination of groundwater can occur from pollutants seeping into the soil and leaching into the groundwater. Contamination can often occur from septic fields, underground storage tanks, direct injection into deep wells, or leaching from surface soil applications. The rate at which a contaminant leaches into an aquifer is determined both by the soil and chemical characteristics. A general rule is that sandy soil generally allows for easier penetration of contaminants than higher organic containing soils and clay. Aquifers that are below bedrock areas can become contaminated through cracks and fissures in the rock, well fracking, or contamination from abandoned wells. Contaminants entering the groundwater can include metals, volatile organic compounds, petroleum products, and leachable pesticides. Contaminant movement within the aquifer is determined in part by the polarity of the compound. Once contamination of an aquifer occurs, it is very difficult to remediate the contaminants. Degradation and mineralization of contaminants in aquifers occurs very slowly because of the lack of bacteria, light, and oxygen tension. Because of the slow recharge of aquifers, dilution of the contamination is minimal. Pump-and-treat (air stripping–volatiles, activated carbon–organics, and chelation–metals) and recharge measures are often used to remove contaminants, but these are expensive approaches.

3 Ecological Risk Assessment

The basic underlying principle of an ecological risk assessment is based on the USEPA Guidelines for Ecological Risk Assessment (USEPA 199219951997a). In any situation, it is necessary to follow a step-wise approach to evaluating potential hazards from chemicals of concern (COCs) to the terrestrial and aquatic wildlife and flora located at each location. The general approach has been adopted by USEPA (1998) where there is the Problem Formulation where integration of available information on potential COCs is carried out. The ANALYSIS PHASE attempts to obtain measurement for COCs (exposure analysis) or unusual abiotic conditions (stressor response profile) that exist. The final phase is the Risk Characterization (risk estimation). The final approach is to make scientific-based management recommendations that may include the following: (1) no action necessary COCs fall below any level of concern; (2) identify specific targeted baseline sampling for historical reference for any future incident; (3) specific targeted sampling approaches for confirming elevated COCs; and (4) remediation of contaminated sites or engineering modifications to reduce impact. The amount of available information on COCs present in the biota and the different environmental matrices (Problem Formulation and Analysis Phase) will affect the extent and uncertainty of the Risk Characterization that can be made and what management options are available. There is a need to follow up on any action that is taken to determine if the chemical effects are remaining the same, improving, or declining.

3.1 Screening Level for Ecological Risk Assessment

It must be realized that a “hazard” in toxicology is determined by the risk of a compound causing an adverse effect and actual exposure to that compound (Risk ×Exposure = Hazard). The Risk component incorporates a specified exposure dose under specific conditions. Therefore, there is no hazard if a COC has no adverse effect at a specific dose/concentration (risk component). If the organism does not come into contact with the chemical (exposure component), then there is no Hazard and the compound is not a COC. In either event, there will be no adverse effect. Although there are many approaches to estimating hazards that incorporate modeling, estimating the movement through fate and transport models and food webs, those methods are beyond the scope of this chapter. If the concentration of a COC in an organism is greater than the environmental media (soil, water, or sediment) it is living in, then the COC is said to bioconcentrate. When the organism’s COC burden is from both environmental media and food intake, then the COC is said to bioaccumulate. Biomagnification of a COC occurs when at each trophic level in the food web has a higher amount of COC than the trophic level below (Newman and Unger 2003).

Bioconcentration (BCF) and bioaccumulation (BAF) factors are two parameters which are important in both terrestrial and aquatic organisms for assessing the Hazard based on steady-state body burden. The steady-state tissue concentrations represent the maximum levels that are expected based on a particular level of exposure. The steady-state level is achieved by the accumulation, distribution, metabolism, and elimination of a compound within an organism (Fig. 10.3).
$$ \mathrm{BCF} = {\mathrm{concentration\ in\ organism/concentration\ in\ ambient\ medium}}$$
(10.1)
$$ \begin{array}{r}{\mathrm{BAF}} = {\mathrm{concentration\ in\ organism/concentration\ in\ food\ and}} \\ \qquad\qquad\mathrm{external\ media\ (ingested\ prey/water/soil/sediment)}\end{array} $$
(10.2)

In aquatic ecosystems where the main route of exposure is through the water column, then the BCF is an appropriate means to estimate body burdens. If food, abiotic media, and ambient water concentrations are important exposure routes, then (10.2) should be used to calculate the BAF. In the case of terrestrial organisms BCFs and BAFs can also be calculated, but require more detailed information (Sample and Suter 1994; Sample et al. 1996; Suter 1993). In the majority of cases, the BAF (10.2) is the appropriate value to be calculated for terrestrial wildlife.

It is known that most wildlife do not remain on a contaminated site or consume only contaminated food which would result in an area use factor (AUF) of less than 1. The conservative approach is to assume AUF an equal to 1. An AUF of 1 would assume that the organism only eats contaminated food or water for their entire life. While this assumption simplifies the assessment, due to the mobility and the diverse diets of most wildlife, it is likely to overestimate the actual exposure experienced. It should be remembered, however, that the purpose of the screening assessment is to identify potential risks and data gaps to be filled. As more data are collected, these assumptions can be modified to reflect the actual amount of time that a species spends within a specified area. A similar approach is used to determine appropriate land use designations of contaminated sites based on human contact time.

What is discussed below is an evaluation of contaminants’ ecological risk based on a simple two-tiered approach. This approach follows closely with that reported by Sample and Suter (1994). In the Hazard Quotient (HQ) approach, the concentrations of the contaminants in the environments are compared to the no observed adverse effects levels (NOAELs). These screening values are based on literature surveys for values that have been reported to cause no adverse effect (No Observed Adverse Effect Level – NOAEL or the Lowest Observed Adverse Effect Level – LOAEL).

The second approach incorporates the exposure of oral ingestion of contaminated media to reach an internal dose. This approach gives a benchmark that is either exceeded or not, based on known adverse effects in organisms compared to their determined dose. This approach is simple yet allows new information to be incorporated as new studies are carried out that better define the NOAEL or LOAEL values (Crane and Newman 2000). When these values are less than the toxicological benchmarks, the contaminant should be excluded from any further consideration (Hazard Quotient Approach). This assumes there are not contributions from other compounds to lower the internal dose necessary to cause toxicity. Although a much more elaborate evaluation is carried out for human risk assessment, the approach described below can be used as a starting point.

3.2 Hazard Quotient (HQ) Approach

The approach follows closely that proposed by Suter (1993). In the first tier, a screening assessment is performed where concentrations of contaminants in the environment are compared to a no observed adverse effect level (NOAEL)–based toxicological benchmarks. These benchmarks represent concentrations of chemicals (i.e., concentrations presumed to be nonhazardous to the biota) in environmental media (water, sediment, soil, food, etc.) that should not result in an adverse effect. One of the principal sources for the values that are used for this calculation is the NOAA Screening Quick Reference Tables (SQuiRT), which are available online at (response.​restoration.​noaa.​gov/​book_​shelf/​122_​squirt_​cards.​pdf) for reference. If site-specific values or more recent literature based information is available, then these values can be used for screening purposes. Within the NOAA SQuiRT tables are background levels for inorganic (p. 2) and organic compounds (pp. 5–8) in freshwater and marine sediments, fresh and saline surface water and ground water. These tables also give different effect levels: threshold effects levels (TEL), probable effects level (PEL), and effect range medium (ERM) which provide increasing levels of effect. This document also contains two tables summarizing specific extraction and detection methods for inorganic trace elements (p. 9) and organic compounds (p. 10) and a summary table listing appropriate containers for each analyte, preservation methods, maximum holding times, and required sample size (p. 11). Tables 10.3 and 10.4 are quick references for matrix and acceptable methods and holding times and storage parameters.
Table 10.3

Recommended analytical methods by matrix being analyzed

Analyte

Method number

Instrument

Matrix

 

PAHs

8270D

GC/MS

S, SD, X

 

PCBs

8270D

GC/MS

S, SD, X

 

Total petroleum hydrocarbon

8015B 418.1

GC/MS GC/FID

S, SD, X

 

Pest/PCB

8080/SW-846 608

GC/MS

S, SD, X W

 

Oil + Grease

5520A+F

IR and Fluorescence

W

 

Dioxins/Furans

8280B 8290A

GC/MS

S, SD, X

 

Metals

EPA-600/CFR40 7000B 7010

Flame AA

W

 
 

6010B SW-846

ICP

S, SD, X

 

VOA

8240/SW-846 624/CLP

GC/MS

S, SD W

 

Grain size

ASTM D422-63

NA

S, SD

 

Total organic carbon

SW 846-9060

NA

S, SD

 

S soil, SD sediment, W water, X tissue, NA not applicable

Table 10.4

Summary of collection and holding requirements based on analyte and specimen

Analytical parameter

Matrixa

Container type and volumeb

Preservation

Holding times

 

Pest/PCB

S

8 oz glass ∼ 500g

4C

7/40 days

 

TAL metals

S

8 oz glass ∼ 10–100g

4C

6 months

 

TAL metals

XE

8 oz glass ∼ 10–100g

≤ 0C

6 months

 

Pest/PCB

XE

8 oz glass ∼ 10–100g

≤ 0C

7/40 days

 

Pest/PCB

XM

Foil/Ziploc ∼ 500g

≤ 0C

7/40 days

 

TAL metals

XM

Foil/Ziploc ∼ 100g

≤ 0C

6 months

 

Pest/PCB

SD

8 oz glass ∼ 500g

4C

7/40 days

 

TAL metals

SD

4 oz glass ∼ 100g

4C

6 months

 

Pest/PCB

XF

Foil/Ziploc ∼ 500g

≤ 0C

7/40 days

 

TAL metals

XF

Foil/Ziploc ∼ 500g

≤ 0C

6 months

 

Pest/PCB

XI

Foil/Ziploc ∼ 100g

≤ 0C

7/40 days

 

TAL metals

XI

Foil/Ziploc ∼ 100g

≤ 0C

6 months

 

Pest/PCB

W

1 L amber ∼ 1,000 mL

4C

7/40 days

 

TAL metals

W

1 L poly ∼ 1,000 mL

4C

6 months

 

Lipids

X

Foil/Ziploc ∼ 1–10g

NA

7 days

 

Grain size

S,SD

32 oz glass ∼ 10–1,000g

NA

NA

 

TOC

S,SD

8 oz glass ∼ 1–10g

4C

28 days

 

aMatrix: Ssoil, W water, SD sediment, XE earthworm tissue, XF forage fish, XI invertebrate,X animal lipids, XM mammal

bThecontainer type is indicated but the amount needed is determined by the selectedanalytical technique and anticipated concentration level. The values listed in thetable are recommendations, and actual amounts should be discussed with theanalytical laboratory carrying out the analysis. The amounts collected can changethe detection limit for each analyte

In practice, when contaminant concentrations in food, water, soil, and sediment resources are less than the threshold toxicological benchmarks, the contaminants may be excluded from further consideration Buchman (1999). As stated above, this assumes no additive or synergistic effects from similar contaminants. The PEL and ERM values allow for comparison to the increasing likelihood of an effect on the biota. In the case of sediment values, the levels are provided both for freshwater and marine sediments for inorganic and organic compounds sampled by NOAA. Another source for freshwater sediment screening values is MacDonald et al. (2000). To determine which contaminants pose a risk, an HQ is calculated, where HQ = media concentration/benchmark. If the HQ ≥ 1, contaminant concentrations are sufficiently high that they may produce adverse effects. This basically means that the concentration observed in the environment is greater than or equal to a threshold or known effects level (i.e., PEL or ERM) and there is an increased likelihood of that compound may have an adverse effect on the organism and warrant further investigation. An HQ < 1 means there is no risk based on the laboratory derived values.

3.2.1 Example Calculation:

Values in ppb (ug/Kg) dry weight.

 

Threshold effects level (TEL)

Probable effects level (PEL)

Effects range medium (ERM)

Source

 

Arsenic

7,240

41,600

70,000

NOAA SQUIRT

 

Arsenic sample value sediment 40,000 ppb

40,000/7,240 = 5.5 TEL HQ > 1

40,000/41,600 = 0.96 PEL HQ < 1

40,000/70,000 = 0.57 ERM HQ < 1

  

Arsenic would be considered a chemical of concern since it was above the TEL value. If the PEL and ERM had HQ > 1, the detected concentration would indicate a probability of increasing toxic responses to benthic organisms. Scientific judgment needs to be exercised when multiple compounds of similar mechanisms of toxicity may be present in a sample or another factor may result in greater sensitivity of the organism to a contaminant.

Therefore, if the concentration of a contaminant exceeds a benchmark, HQ ≥ 1 that contaminant should be retained as a contaminant of concern (COC).

Conversely, if the HQ < 1, the contaminant would not be retained as a COC.

3.3 Estimating Body Burdens for Comparison to NOAELs for Wildlife

Calculations can be made to determine on a daily basis the amount of a contaminant that an organism is being exposed to from its food, water, or other ingestion exposure routes (e.g., sediment or soil). In this approach, both inhalation exposure and dermal adsorption are not included in the daily intake which can underestimate exposure, but normally the conservative assumptions made in estimating the other major exposure routes may negate the need to be concerned with other routes of exposure. Dermal exposure needs to be included when amphibians and reptiles are the organisms of concern. For humans, inhalation and dermal often play a major route of exposure and are included. The daily exposure estimations based on known or estimated levels of a contaminant can be calculated using standard calculations and based on literature-derived life history data (Jorgensen et al. 1991).

The total body burden can then be compared to LOAELs from literature studies or estimated from other reported endpoints (LOAEL, ED50 (dose resulting in 50% effect) or LD50 (dose resulting in 50% death). One approach to estimate a LOAEL from a NOAEL is to apply an uncertainty factor of 10 to the values or determine it from dose-response curves (Fig. 10.4).
$$\mathrm{LOAEL\ =\ NOAEL\ multiplied\ by\ an\ uncertainty\ factor\ of\ 10}$$
(10.3)

In the case where there is only LD50 data, either examine the slope of the line and extrapolate to the x-axis or apply a larger uncertainty factor (100 or 1,000) to estimate the value. The difficulty with extrapolating from these values is that the degree of uncertainty is greatly increased, but if no other value exists it may be the only alternative to estimate whether a contaminant potentially poses a risk.

The presentation of toxicity data on an mg/kg body weight (BW)/day basis allows comparisons to the calculated levels for a particular species of interest. If the levels that are calculated exceed the LOAEL, then the levels in these matrices would pose a hazard to that particular species and be considered a COC. When calculating these values, care must be taken to convert any dry weight concentrations to wet weight by applying the appropriate water content values for specific prey species or determined by the analytical laboratory at the time of analysis. The USEPA has developed a benchmark dose approach (BMD) to provide a more quantified method for determining critical points along the dose-response curve. A more detailed description and a computer model can be viewed and downloaded at http://​www.​epa.​gov/​ncea/​bmds.​htm.

4 Trophic Level Evaluation

Trophic level evaluation is not required for all compounds, because there are certain characteristics which preclude biomagnification up the food web. Compounds that are rapidly metabolized, are poorly absorbed from the gastrointestinal track or epidermis, or are rapidly eliminated generally need not be examined in a trophic model. In these cases, the compound can be compared to the LOAEL for determining if the concentrations are above a threshold of concern (as described in the previous section). In the case of trophic level evaluations, the organism being impacted is through direct contact with environmental matrices and through their food sources (bioaccumulation). Therefore, when calculating the daily intake of a compound, the sources of the contaminants include the following: water, sediment/soil, and food. Airborne exposure is not included in these calculations. In this way, representative organisms at different trophic levels can be selected and their daily intake calculated and compared to effects levels. By using this approach, the total daily intake can be calculated using both actual data from organisms collected on- or off-site or through estimated prey body burdens or matrix values. When the TDI (total daily intake of a compound) from all the sources is calculated, these values can be compared to the LOAEL value derived from the literature. When the TDI exceeds this benchmark effects value, the compound should be considered a COC. This is a highly conservative approach since it assumes that all of the food consumed is contaminated at a set level and that there is 100% uptake and retention. In the real world, it is unlikely that an organism will be exposed continuously to contaminated sediment/soil and prey at a constant level; there is never 100% uptake due to metabolism and elimination pathways. This conservative approach will help protect wildlife especially when tissue dose data for specific organisms is lacking.

In some instances where there is virtually no data on a chemical, the approach undertaken may have to be very simplistic. The Aquatic Food Chain Multiplying Factor (FCM) described below is such an approach (Sect. 4.2). If a large amount of data is available on actual matrix values (i.e., water, sediment, or soil contaminant concentrations) and lower trophic level prey dose, then a better estimate of the predator dose can be calculated using the estimated dose from all source inputs (Sect. 4). Because of the complexity of food webs even in the simplest of ecosystems, there is a need to select sentinel organisms that can represent large classes of organisms and also reflect the ecosystems being evaluated when determining compounds that have a potential to bioaccumulate.

4.1 Estimating Aquatic Organisms Body Burdens

If the BCF is not available for a specific compound, then the following relationship developed by Lyman et al. (1982) can be used based on the octanol-water partitioning equation (Eq. 10.4).
$$\mathrm{Log\,BCF\ =\ 0.76\ log\,P_{oct} - 0.23}$$
(10.4)
In a similar fashion, the BCF can be calculated from the water solubility of a compound (10.5) as reported by Lyman et al. (1982).
$$\mathrm{Log\,BCF = 2.791 - 0.564\ log\ Water\ Solubility\ (mg/L)}$$
(10.5)

The BCF values using (10.4) and (10.5) will allow for an estimated value to be calculated, but actual tissue and matrix concentrations are more accurate for BCF determinations.

4.2 Aquatic Food Chain Multiplying Factor (FCM)

The FCM is also a very simplistic approach where the concentration in a matrix is used to estimate the level in biota at different trophic levels. This approach does not differentiate between a bioconcentration factor (BCF) and a bioaccumulation factor (BAF). The differences between a BCF and BAF have been previously discussed in Sect. 3.1.

In order to estimate the tissue concentration in different trophic levels, the only information needed is the Log K ow for a chemical of interest (Sample et al. 1996). Although the FCM is a very simple approach, it is based on the assumption that the primary uptake route of these compounds is through passive diffusion, which is controlled by the lipophilicity of the compound. It is a conservative estimate since it assumes 100% uptake and does not take into account factors affecting absorption, metabolism, or elimination for different species. Estimates can be made for concentrations in zooplankton (trophic level 2), small fish (trophic level 3), and piscivorous fish and top predators (trophic level 4). Table 10.2 shows the multipliers for estimating the amount of a compound that will accumulate in each of the higher trophic level organisms based on the Log K ow . The FCM approach may be applicable for some metals if the organometallic forms (i.e., methyl mercury) are known to be biomagnified in higher trophic levels based on Log K ow (USEPA 1995).

The FCM approach is useful when there is a new compound that is identified and there is very little data on levels in higher trophic level organisms. By obtaining the Log K ow from the literature or even doing a simple partitioning experiment one can obtain this value. It must be realized that depending on the method from which it was derived, there can be differences in the actual K ow values; however, they are generally fairly similar. This value is then used to determine what multipliers (Table 10.5) for each trophic level. This is an approach that does not take into consideration any elimination pathways and will overestimate tissue levels when metabolism does play a major role. The estimated FCM tissue value can then be compared to the LOAEL value determined from literature values.
Table 10.5

Aquatic food chain multiplying factors (Taken from Sample et al. 1997 which was modified from USEPA 1993)

Log

Kow

Zooplankton

Sm. fish

Piscivorous

fish

Log

Kow

Zooplankton

Sm. fish

Piscivorous

fish

 

2

1

1.005

1

5.7

1

7.962

10.209

 

2.5

1

1.01

1.002

5.8

1

8.841

12.05

 

3

1

1.028

1.007

5.9

1

9.716

13.964

 

3.1

1

1.034

1.007

6.0

1

10.556

15.996

 

3.2

1

1.042

1.009

6.1

1

11.337

17.783

 

3.3

1

1.053

1.012

6.2

1

12.064

19.907

 

3.4

1

1.067

1.014

6.3

1

12.691

21.677

 

3.5

1

1.083

1.019

6.4

1

13.228

23.281

 

3.6

1

1.103

1.023

6.5

1

13.662

24.604

 

3.7

1

1.128

1.033

6.6

1

13.98

25.645

 

3.8

1

1.161

1.042

6.7

1

14.223

26.363

 

3.9

1

1.202

1.054

6.8

1

14.355

26.669

 

4

1

1.253

1.072

6.9

1

14.388

26.369

 

4.1

1

1.315

1.096

7.0

1

14.305

26.242

 

4.2

1

1.38

1.13

7.1

1

14.142

25.468

 

4.3

1

1.491

1.178

7.2

1

13.852

24.322

 

4.4

1

1.614

1.242

7.3

1

13.474

22.856

 

4.5

1

1.766

1.334

7.4

1

12.987

21.038

 

4.6

1

1.95

1.459

7.5

1

12.517

18.967

 

4.7

1

2.175

1.633

7.6

1

11.708

16.749

 

4.8

1

2.452

1.871

7.7

1

10.914

14.388

 

4.9

1

2.78

2.193

7.8

1

10.069

12.05

 

5.0

1

3.181

2.612

7.9

1

9.162

9.84

 

5.1

1

3.643

3.162

8.0

1

8.222

7.798

 

5.2

1

4.188

3.873

8.1

1

7.278

6.012

 

5.3

1

4.803

4.742

8.2

1

6.361

4.519

 

5.4

1

5.502

5.821

8.3

1

5.489

3.311

 

5.5

1

6.266

7.079

8.4

1

4.683

2.371

 

5.6

1

7.096

8.551

8.5

1

3.949

1.663

 

Trophic Level 2 = zooplankton; 3 = small fish; 4 = piscivorous fish, including top predators

For example, if a new compound has a Log K ow = 6. 0 (see Table 10.5 for multipliers for each trophic level and is detected at 10 mg/L in water, then it is assumed that zooplankton (level 2) will have 10 mg/Kg body levels (1 × 10 mg/Kg since 1 L equals 1 Kg). Trophic level 3 organisms would have 105.56 mg/Kg and level four would have 1,689 mg/Kg. Based on the estimated total dose these values can then be compared to LOAEL and or NOAEL values for representative species.

If for some reason there is no known K ow , then an alternative would be to find a compound with a similar structure and use that compounds’ K ow for the estimation. These estimates can also be used for calculating levels in higher trophic level organism as described. It is important to realize that the more assumptions that are made, the less reliable the calculated body burdens. That is why it is critical to not rely solely on these estimates without actual determined tissue levels from field samples.

As shown in Table 10.6 are some examples of calculated values for chemical BAF values for trophic level 3 and 4 organisms. These calculations are based on (10.4) and the FMC values presented in Table 10.5.
Table 10.6

Selected compounds for determining trophic level BAFs

Chemical

Log K ow

BCFa

Trophic

level 3b

FCM

Trophic

level 3

BAFa

Trophic

level 4

FCM

Trophic

level 4

BAF

 

Acetone

− 0. 24

0.39

1

0.39

1

0.39

 

Aldrin

6.5

51,286.14

13.66

7,000,000c

24.604

13,000,000c

 

Benzo(a)pyrene

6.11

25,917.91

11.34

294,000c

17.783

461,000c

 

Copper

 

290d

1

290

1

290

 

Mercury (Methyl Mercury)e

   

27,900.00

 

140,000.00

 

Seleniumf

   

2,600.00

 

6,800.00

 

aBCFcalculated using (Eq. 10.4)

bLevel 3 = small fish,level 4 = piscivorous fish, including top predators

c Rounded off values

dBAF calculated using Table 10.5

e Literature citation EPA 1995

f Peterson and Nebeker 1992

As shown in Table 10.6 with acetone, acetone would not be expected to bioaccumulate into higher trophic levels. Both aldrin and benzo(a)pyrene would be assumed to accumulate into the higher trophic level organisms. This model assumes that for most trace metals there is no bioaccumulation in higher trophic levels. Care should be exercised when there is evidence for specific metal bioaccumulation from prey species or where the metal may not be bioavailable such as with Cd (DiToro et al. 1992). In the case of organo-metals, there can be bioaccumulation in higher organisms, as seen with methyl mercury (USEPA 1997b). Selenium is also accumulated in part due to its similarity to calcium and using these transporters and sequestration areas within the cell. Using this method, the calculation of tissue concentrations can be used in estimating body burdens in these trophic level organisms for comparison to LOAELs to calculate a HQ. These values can also be used to estimate concentrations in prey species for terrestrial mammals (including man) and piscivorous birds.

5 Conclusion

It cannot be overemphasized that the approaches proposed in this chapter are only an initial approach to a very complex problem. In most environmental samples there are complex mixtures of compounds that can be assessed on an individual chemical basis, but there are likely complex chemical interactions that will influence the toxicity. These contaminants, although dealt with on an individual compound-by-compound basis in toxicology, does not preclude the possibility that chemical compounds can result in less than additive, additive, or greater than additive effects when an organism is exposed to multiple chemicals from an environmental mixture. The exception to this is when the chemicals act through a common mechanism and the literature supports an additive approach. This is the case for dioxin-like compounds, but in the majority of cases such information is not available (Van den Berg et al. 2008). The difficulty in dealing with complex mixtures is that there is no way short of testing the mixture to determine how the chemicals interact in the biological system. The reader should also appreciate that both chemical contamination and species have a spotty distribution in the environment that can effect the establishment of baseline values. It is important in any assessment to understand the activities that have historically or are currently ongoing (off-site and on-site) that could introduce chemicals into the exposure pathways. With the continued growth in world population in regional areas, there will be sustained pressure for water resources between competing groups.

6 Cross-References

Changing Energy Demand Behaviour:​ Potential of Demand-Side Management

Engineers and Community:​ How Sustainable Engineering Depends on Engineers’ Views of People

Ensuring Sustainability of Bioenergy in Practice

Groundwater Contamination:​ Role of Health Sciences in Tackling

Impact of New Technologies:​ How to Assess the Intended and Unintended Effects of New Technologies?​

Life Cycle Design and Life Cycle Strategy Planning

Materials Education for Sustainable Society

Supply Chain Management for Sustainability

Sustainable Earth System Engineering:​ Incentives and Perspectives

Sustainable Water Management in Response to Global Changes

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© Springer Science+Business Media Dordrecht 2013
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