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Environmental Exposure Assessment
Environmental Change see Global Climate Change and Environmental Toxicology
Environmental Exposure Assessment A Di Guardo, University of Insubria, COMO, Italy Ó 2014 Elsevier Inc. All rights reserved.
Definition While exposure to human beings is generally evaluated by considering a variety of ways in which exposure could take place, such as at the workplace (occupational exposure), from use of consumer products (consumer exposure), and indirectly via the environment, the environmental exposure of the organisms in the ecosystems is generally assessed by evaluating the concentration of a specific chemical in the main environmental media (air, soil, water, sediment) and in biota as food source for other organisms (predators) by secondary poisoning. Exposure is therefore derived by considering that the organism is living in or interacting with such media, in lieu of considering the internal concentrations as resulting from the exposure. This makes the study apparently simple, but potentially complicated if the variety of environments (and ecosystems) in which the huge multiplicity of organisms live and interact is considered. As an example, biomagnification (accumulation of a chemical through the food chain) can occur for certain types of molecules, and its intensity depends on the type and length of the food chain, which could be considerably different in different ecosystems. Environmental exposure assessment (EEA) is also one component of the phases of risk assessment and in particular it is usually compared with the effect assessment in order to characterize the ecological risk in ecological risk assessment (ERA) (Figure 1). The concentrations reached in one specified compartment or environmental medium (air, water, soil, sediment) are compared with their corresponding concentrations which are considered as thresholds (no effect concentrations) for not causing any observable adverse effect to the ecosystem. This is usually performed by comparing a predicted environmental concentration (PEC), generally produced using environmental fate models (EFMs) and referred to a particular phase or compartment, with a predicted no effect concentration (PNEC), generally calculated starting from toxicological data such as acute toxicity (LD50 or Exposure assessment
Effect assessment
Risk characterization Figure 1
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Exposure assessment in environmental risk assessment.
LC50) or measured no observable effect concentration (or levels) (no-observed-effect concentration (NOEC) or no-observed-effect level (NOEL)) and referred to the same phase or compartment. When the ratio (also called risk quotient) is higher than 1, risk may occur and measures should be taken.
Ecosystem Complexity and Exposure EEA is a complicated task because it deals with the evaluation of the exposure of the ecosystems, which are many and involve a high degree of diversity. Therefore, EEA should balance the desire of completeness and overall protection with the need to simplify and streamline the processes. The first point in complexity regards the biodiversity of organisms: for example, the high number and variety of species of organisms. More than 1.5 million species have been classified (mostly insects, spermatophytes, molluscs, fungi), which are adapted to living in a multiplicity of environments, at different spatial scales. At global level, this is evident observing the very different conditions which determine the main terrestrial biomes at the different latitudinal regions, from average annual air temperature (and its annual, seasonal, daily variation) to rainfall abundance and pattern, together with the variability of the abiotic substrate. Spatial variability can be therefore observed at the different continental, regional, local, and even at much smaller scale (e.g., a single lake, river, or environment characterized by peculiar characteristics). When observing temporal scales, the wide variety of life cycles of the organisms is rather important in considering the effect of the duration of the exposure: as an example, generation times can be less than a day for bacteria to years for plants and animals. The interaction of such spatial and temporal issues with organism life cycles determines the complexity of the exposure which cannot be in principle obtained knowing only the external exposure (or the concentration in the environmental media or compartment where the organism lives). The internal dose, which is the real amount of the chemical reaching the organism and capable of determining the effects, is dependent on the interaction between the organism and a number of factors, such as the exposure time, the homogeneity of the chemical distribution in the environmental phase, its bioavailability, a number of abiotic factors, and a number of biotic factors (surface to volume ratio, feeding and growth rates, life history, behavior, etc.). Exposure time can be very different because of the organism life cycle and duration of each phase (juvenile or larval stages vs adult stage) or because the chemical can be emitted with different temporal patterns: for example, the
Encyclopedia of Toxicology, Volume 2
http://dx.doi.org/10.1016/B978-0-12-386454-3.00553-4
Environmental Exposure Assessment
discharge of chemicals into the environment can be constant, determining steady state or time invariant conditions, such as the effluents of a wastewater treatment plant in normal operating conditions (without a flood event) or can be rapidly changing, such as the pesticide pulses in surface water determined, for example, by rainfall over treated fields and consequent runoff. Also chemical distribution in the environmental media can be rather inhomogeneous, depending in part on the poor mixing of certain phases (soil, sediment vs air or water) and on the movement of chemical in the compartment: for example, cracks and macropores in soil could facilitate the vertical movement of waterborne chemicals, while in zones without cracks or macropores chemical moves more slowly. Bioavailability of the chemical is also important in soil, sediment, or water environments: the bioavailable fraction of a chemical in an environmental medium (e.g., water) can be defined as the fraction of the chemical which is available for organism uptake and therefore capable of determining an effect. Such fraction (truly dissolved) could be a small quantity compared to the total (bulk) concentration present (and measured) in water. The bioavailable fraction of a chemical can be influenced by the sorption of the chemical onto the organic carbon in suspended solids and to dissolved organic carbon, or, by other environmental conditions, such as pH, salinity, and cation exchange capacity. Abiotic factors include climatic conditions (such as temperature, insulation, wind speed, and rainfall) or substrate characteristics such as texture and structure in soil. Their variations depend on both spatial and temporal issues and may greatly influence the exposure of the organisms. Among the biotic factors, surface to volume ratio of the organisms is important when the chemical must diffuse through a surface as a cuticle; and the higher surface to volume ratio for small organisms will make diffusion much faster and therefore may enhance the exposure of small and not mobile organisms. Feeding and growth rates are specific parameters regulating how an organism would grow according to the availability of food, its intake, and the efficiency of its transformation in new body structures. These two parameters are very variable among the species and of course could regulate the intake of chemical through the diet and/or its accumulation or dilution in the body. Life histories and life cycles are indeed important in the exposure because many organisms can change type of exposure medium (aquatic vs terrestrial) or type of food during their development through the larval and adult stages. The sensitivity to the chemical could also change during the evolution through these stages, so that the same exposure concentration could produce very different effects in different larval stages of the same species. Last but not least, organism behavior can influence the exposure because the organism may be rather mobile through the environment, or migrate, hibernate, or possess resistant structures, which could reduce the exposure.
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are probably a factor of 100 less, so in the order of tens of thousands. In order to characterize their identity, their properties in order to evaluate their fate and behavior as well as the toxicological profile, many initiatives were devised to collect and classify information on new and existing chemicals, such as Toxic Substances Control Act (TSCA) in the United States, Canadian Environmental Protection Act (CEPA), and more recently Registration, Evaluation, Authorisation and Restriction of Chemicals (REACH) in European Union. A number of priority lists were created in the past to handle substances for which undesired and initially unrecognized effects were found: from chemicals capable of altering pH of rain water, generating acid rain (such as sulfate oxides and nitrogen oxides) to the so-called persistent organic pollutants (POPs) and persistent, bioaccumulative, and toxic (PBT) chemicals. POPs and PBTs are organic chemicals that, according to the Stockholm Convention, a global agreement adopted in 2001 to ban or regulate the use of a number of POP chemicals, possess a particular combination of physical and chemical properties such that, once released into the environment, they remain intact for exceptionally long periods of time (many years); become widely distributed throughout the environment as a result of natural processes involving soil, water, and, most notably, air; accumulate in the fatty tissue of living organisms including humans, and are found at higher concentrations at higher levels in the food chain; and are toxic to both humans and wildlife. The chemical families initially 12 and 9 were later included. Among those families some organochlorine pesticides can be found, such as aldrin, chlordane, DDT, dieldrin, endrin, heptachlor, hexachlorobenzene, mirex, toxaphene; industrial chemicals such as polychlorinated biphenyls (PCBs); and by-products, such as polychlorinated dibenzo-p-dioxins and polychlorinated dibenzofurans (dioxins and furans). More recently, a number of chemicals recognized were being capable of interfering with the endocrine system in animals and they are generally known as endocrine disruptors or environmental hormone disruptors. Some of the POPs listed above are considered endocrine disruptor chemicals. Other chemicals of environmental relevance are pesticides, biologically active substances which are deliberately emitted into the environment at relevant rates. They are designed to eliminate animal pests and weeds in order to protect crops but could escape from treated fields and reach surface and groundwaters as well as not treated fields and natural areas due to drift and air transport of their vapor phase. Among the most recently recognized environmental chemicals, there are many human and veterinary pharmaceutical substances, which, because of their characteristics, may travel through the wastewater treatment plants with no or little reduction in amount and end up in surface waters or in soils through the sewage sludge, when used as biosolids to restore soil fertility.
Assessment of Concentrations in Environmental Media Environmental Chemicals Registered chemical substances are nowadays about 60 millions, as recorded in the Chemical Abstracts Service (CAS) database. Among those, chemicals in commerce produced at a rate which could make them significant for the environment
Environmental exposure can be assessed by evaluating the multimedia path of a chemical in order to establish the prevailing concentrations in the compartment by means of measurements or modeling approaches. Both approaches have advantages and disadvantages and they are often both required
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to obtain an adequate level of insight into the understanding of environmental exposure.
Monitoring Environmental monitoring generally provides data on average concentrations in environmental media (air, water, soil, sediment). Peak concentrations are obtained when the measurement is performed at the point of discharge (air, water). While these data are important to estimate the order of magnitude of emissions, they generally do not allow to portrait a complete picture of the concentrations in the environment, because the chemical can undergo a series of transformations and transfer among media before reaching the point of measurement. Monitoring implies a number of activities in order to capture concentrations in the environmental compartments: from the preparation of a statistically sound sampling scheme to the selection of a sampling method, transport, storage, analytical and laboratory requirements for the analysis, as well as data quality and reporting issues. Samples of air, biota, water, soil, and sediment can be taken by employing several techniques: for example, air can be collected in a cartridge using a pump and a flowmeter to obtain the volume of air measured, water with sampling bottles or automatic sampling devices, soil with a core sampler, and sediment with a grab sampler such as a dredge. Passive sampling techniques can also be used to record concentrations in a certain phase such as air with shielded polyurethane foam (PUF) devices, or water with semipermeable membrane devices (SPMDs) or solid-phase micro extraction (SPME) techniques. Artificial passive samplers have advantages over the ‘natural’ passive samplers (such as leaves) in a way that they can be standardized and cleaned up before sampling in order to obtain comparable initial conditions. Passive samplers can also be used to sample a phase (e.g., air) for relatively long time (weeks, months) and therefore detecting chemicals present at very low concentrations. However, their sampling rate depends on a number of environmental conditions (such as temperature, and wind speed, for some air samplers) and specific physical–chemical properties; therefore, they generally provide order of magnitude estimates. Current passive devices also have the disadvantage of generally providing average concentrations of the sampling period and missing the information on peak concentrations. Planning a monitoring program raises a series of questions, such as those related to which parameters and chemical should be included; and where and when they should be measured. Additionally, it may be inconsistent and produce no results without prior knowledge of the environmental fate of a chemical. Data obtained without a properly selected sampling and statistically sound sampling scheme may produce a poor representation of the spatial and temporal characteristics of the contamination. This means that they can seldom catch the variability of concentration change in a territory in time and the relative spatial gradient. Sometimes, when the amount of spatial data is sufficient, geostatistical techniques can be used to reconstruct a spatial trend (e.g., in soil) but they are usually limited to static representations of contamination at a certain point in time. Finally, environmental monitoring is an a posteriori approach, in other terms the contamination needs to be in place to be measured, while it would of course be
preferable to act before a contamination (and a potential damage) has occurred. Finally, monitoring data can be used to gather a picture of the contamination of a certain compartment or a certain ecosystem and later used to ‘calibrate’ or ‘validate’ EFMs.
Modeling Modeling is an a priori approach and represents the only possibility of generating information for predictive purposes and for estimating potential exposure before a new chemical is used. On the other hand, models cannot be calibrated and validated without experimental data. In order to avoid mistakes when applying predictive models, the proper model for the specific environmental situation must be carefully selected. EFMs are nowadays among the most used tools to evaluate the fate of chemicals in the environment, due to their general predictive nature and the substantial ease of implementation. EFMs can be physical dispersion models, used to calculate concentrations in a phase (air, water, soil) in the proximity of a point source or partitioning or compartmental models, also known as multimedia fate models (MFMs). Among them, the fugacity-based models, developed by Don Mackay and coworkers, can be found. They are also called ‘Mackay’s type’ models. Fugacity is an equilibrium criterion which can be used to define the partition of nonpolar chemicals between phases (instead of concentration as in other models). Other equilibrium criteria were adopted later to deal with polar and/or dissociated chemicals: the aquivalence first and, more recently, the concept of activity. Many MFMs are called box models because they divide the environment in well-mixed boxes or compartments which can exchange chemicals among them; chemicals can enter the environment through direct emission or passively carried by advective fluxes in air and water (Figure 2). Chemicals could also disappear from the modeled environment due to advection or degradation and can move through the environment due to compartment exchange which in turn depends on substance and environmental characteristics. MFMs generally require data on chemical properties, environment conditions, and rates or quantities of chemical discharges and produce a picture of the environmental fate of
Air
Soil
Water Sediment
Figure 2 Multimedia environment, with the four main compartments (air, water, soil, sediment). Arrows represent exchange fluxes between compartments. Emission of chemical, advection, and degradation not depicted.
Environmental Exposure Assessment
a chemical, often in the form of a mass balance. More specifically the data required are the following: l
physical–chemical properties and half-lives; chemical emissions and temporal/spatial patterns; l environmental compartment characteristics; and l mass transfer coefficients between compartments. l
The relevant physical–chemical properties for chemicals (water solubility, vapor pressure, octanol/water partition coefficient (Kow), other solids/water partitioning coefficients, octanol/air partitioning coefficient, Henry’s Law Constant or air/water partitioning coefficient, pKa, etc.) and the half-lives in the various environmental compartments (air, particles, water, soil, sediment, suspended solids, vegetation, animals) are the basic parameters needed to simulate the behavior of nonpolar organic chemicals in the environment. These basic parameters are used in the models to calculate partitioning parameters of the selected chemical between the phases of the compartments: vapor pressure and water solubility are used to calculate Henry’s Law Constant and air/water partitioning coefficient (Kaw), Kow is used to calculate the bioconcentration factor in aquatic biota, partitioning to organic carbon (Koc) and to solid phases in suspended solids, soil, and sediment (Kp), and, together with Kaw, the octanol/air partitioning coefficient, which can be used to calculate the partitioning in particulate matter in air and plant leaves. Half-lives are sought for the main compartments and used to calculate the degradation losses. Chemical emissions are key factors for the calculations of the fate of chemicals because they determine the amount introduced in the modeling environment. In steady state approaches (e.g., European Uniform System for Evaluation of Substances (EUSES) or the fugacity model, equilibrium criterion (EQC) model) they can be annual emissions (kg year1), while in dynamic approaches a temporal pattern should be specified (e.g., 2 kg ha1 for the first 2 h). Very often the temporal pattern is ignored or unknown. This is generally true also for the spatial pattern. Emission can be estimated by evaluating monitoring data, using estimation techniques based on population or employee density or using existing national and international emission databases such as the US Environmental Protection Agency (US EPA), Toxics Release Inventory (TRI) Program or the European Pollutant Release and Transfer Register (EPTR). The environmental compartment characteristics regard their composition and their temporal and spatial variabilities. Models require sizes and volumes as well as volume fractions of solids, air, water, etc. in subcompartments for each box. In most of the regulatory approaches, these are fixed parameters characterizing ‘average’ characteristics. Environmental temperature is usually considered among these parameters, as well as organic carbon fraction, fraction of air/ water/solids/lipids, soil texture, etc. Compartments can be discretized vertically or horizontally: in the vertical discretization each compartment can be made as one box or layered (e.g., soil, air, water, sediments), while horizontal compartments can be described with regionally averaged properties (organic carbon, texture, hydrological data, etc.) vs distributed (typical/ spatially explicit) properties (often supported by a geographic information system (GIS)).
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Mass transfer coefficients are specific intercompartmental velocities used in box models to predict the rate of transfer among phases. They are generally implemented in nonequilibrium models and are referred to fixed environmental conditions. Environmental models require calibration/verification/ validation procedures to assure their adequacy and datasets of environmental/measured chemical concentrations in order to proceed in such direction.
Environmental Fate Models Many EFMs have been developed in the past 30 years. Most are steady state models (in which chemical discharge is constant in time), such as the EQC and ChemCAN models or the EUSES model. Some were later developed as unsteady state or dynamic models (historically in terms of chemical discharge, now also in terms of environmental scenario changes). The steady state models are better suited to situations in which chemical emission does not vary significantly during a certain period of time (e.g., sewage treatment plant (STP) discharges). They are generally adopted for the simple mass balance equations obtained in such situations and because they do not oblige the user to provide a time-varying emission profile (at hour, day, week, month, year level). Dynamic models are better suited to handle episodic discharges, such as pulse discharges (e.g., pesticides) or mass movement of chemical caused by some environmental phenomena (such as a runoff event triggered by rainfall). Models can be built with different types of spatial resolution, presenting large compartments (regional) or smaller or site-specific descriptions (local), sometimes nested (a local model into a regional one), such as in the EUSES model, for which global, continental, regional, and local scenarios are present. Other approaches include different models (e.g., EQC, ChemCAN, SoilFug) which are used to calculate the fate of chemicals from regional to site-specific (local) sizes in a tiered approach. Some models were developed to predict the fate of agrochemicals applied to soil at specified times, some of those are suited to predict the fate of pesticides leaching toward groundwater (e.g., MACRO, PEARL, PRZM models), while others are used to predict surface and subsurface runoff of chemicals reaching surface waters such as streams and rivers. When the chemical has the potential for long-range transport, regional or continental scales are not adequate and global models were developed to handle such situations. Most of these models are created linking regional models which simulate a number of latitudinal sectors connected via air and water. In another approach, existing meteorological models are connected to MFM to better handle the transport in the troposphere. Recently, many modeling efforts are devoted to introduce more ecological realism in predictive approach of exposure, including a higher degree of temporal and spatial dynamics, in order to follow the variability of the fate of a chemical.
Bioaccumulation Models While EFMs are generally employed to obtain PECs in the main environmental compartments, they fall short in predicting accumulation in biota, especially when bioaccumulation is
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Environmental Exposure Assessment
important. Therefore, bioaccumulation models are predictive tools designed to evaluate the concentrations that a chemical could reach in an organism (aquatic or terrestrial) from the concentrations in environmental media (e.g., air, water, sediment, soil). The knowledge of the concentration reached in the organism or in certain tissue or organ is important to interpret the toxicological response of organisms. These levels usually depend on the interaction between the organism and the environmental media directly, but sometimes (when the chemical possesses certain physicochemical characteristics and adequate metabolic half-life, such as PBTs) a trophic enrichment could be observed.
Aquatic Organism Models For the aquatic organisms three main phenomena can be cited: bioconcentration (process in which the chemical concentration in an aquatic organism exceeds that in water as a result of exposure to waterborne chemical, usually where the chemical is taken up only from water via the respiratory surfaces such as gills and/or the skin), biomagnification (process in which the chemical concentration in an aquatic organism exceeds that in the organism’s diet, due to dietary absorption), and bioaccumulation (which sums up all possible routes of exposure). For terrestrial ecosystems similar concepts are adopted for terrestrial animals, while plants are generally considered in terms of accumulation from the chemical present in soil or in the air compartments. For the aquatic organisms, initial approaches were simple correlations (Kow based) mainly to predict bioconcentration in fish. Then more complex models appeared and several processes, detailing uptake and clearance of chemicals, were introduced. These were gill ventilation, food uptake, egestion, and metabolism, as well as reproductive and growth dilution losses. The organism was considered a single compartment. A full mass balance could be compiled, usually in steady state formulations. Mechanistic explanations were found to describe the process of bioaccumulation, separating its two components, bioconcentration and biomagnification. Role of food vs gill uptake was shown, especially for POPs, so that the importance of the biomagnification processes in determining the uptake of POPs in the trophic chains was shown. Bioavailability of chemicals in the environmental media (such as air, water, sediment, soil) and food was then investigated, and it was shown to influence the fraction available for uptake. The mechanism of uptake and release in aquatic air–breathing organism was later understood, showing that water-respiring organisms were more efficient in clearing chemical than airrespiring organisms. Therefore, air-breathing warm-blooded animals, which require more food, found were being subject to greater bioaccumulation and biomagnification from food. Other models were later developed for other aquatic organisms, such as benthic invertebrates, birds, and up to marine mammals. Some of the models were physiologically based toxicokinetic (PBTK) models, similar to those developed for pharmacokinetic studies in toxicology. These single-species models were then sometimes integrated into more complex food web models, allowing to reconstruct chemical accumulation through food chains or webs. The increase in complexity required a consequent increase in species-specific parameters (which are seldom
available) forcing the models to be based on ‘generic’ physiological parameters (such as egestion or metabolic rates) and relying on the available specific parameters (such as lipid fraction or ventilation rate).
Terrestrial Organism Models While most of the efforts were put in aquatic food web modeling, a consistent lack of approaches exists for terrestrial animals, both for single organism model and food webs. As an example, models can be found for earthworms, or some species of birds (such as seagull) and are generally scarce for other terrestrial species. Recent models are available for agricultural food chains, such as employed to investigate the uptake process in cows. When dealing with uptake in plant biomass, it has to be considered that vegetation uptake was mostly investigated in the past 25 years, and several models are now available to predict root uptake and translocation to the aerial part and in the air to leaf path, especially for nonpolar chemicals. While root uptake is generally proportional to Kow, translocation is mostly limited to low-range Kow chemicals (in the vicinity of log Kow ¼ 2) and therefore uptake of more hydrophobic chemicals in leaves is generally driven by the air to leaf partitioning. An exception was found for some species of the Cucurbitaceae family, which were capable of accumulating dioxins in the aerial parts from contaminated soil via root uptake. The accumulation of chemicals from air to leaves in canopies of existing forests was recently developed to quantify the so-called forest filter effect which depicts the forest leaf’s capability to sequester chemical from air and transfer it to the soil environment and therefore to influence the soil ecosystem. Forest filter was measured in a variety of forest types (from mixed broadleaf woods to pure conifer woods) and parameters were devised to describe the uptake from air (in which Koa is the main chemical descriptor of the uptake). The role of some ecological parameters was also investigated, showing the relationship with leaf area index (LAI) (m2 of leaves m2 of soil) and specific leaf area (SLA) (m2 kg1). LAI gives an indication of the amount of foliar surface per soil unit area and is a measure of leaf stratification or foliar density, while SLA represents the amount of surface for unit of weight of leaf. SLA is a characteristic of each species, even if it is influenced by the canopy density, in other terms from LAI. The ratio of LAI to SLA (and its variation in time) gives the amount of leaf biomass per square meter of soil, parameter which can be used in the accumulation models.
See also: Risk Assessment, Ecological; Environmental Processes; PBT (Persistent, Bioaccumulative, and Toxic) Chemicals; Persistent Organic Pollutants; Pesticides; Predicted No Effect Concentration (PNEC); REACH; Toxic Substances Control Act.
Further Reading Kelly, B.C., Gobas, F.A.P.C., 2003. An arctic terrestrial food-chain bioaccumulation model for persistent organic pollutants. Environ. Sci. Technol. 37, 2966–2974. Kuemmerer, K. (Ed.), 2004. Pharmaceuticals in the Environment, Sources, Fate, Effects and Risks. Springer, Heidelberg.
Environmental Exposure Assessment
Mackay, D., 2001. Multimedia Environmental Models: The Fugacity Approach, second ed. Lewis Publishers, Boca Raton. Mackay, D., Di Guardo, A., Paterson, S., Kicsi, G., Cowan, C.E., 1996. Assessing the fate of new and existing chemicals: a five stage process. Environ. Toxicol. Chem. 15, 1618–1626. Mackay, D., Fraser, A., 2000. Bioaccumulation of persistent organic chemicals: mechanisms and models. Environ. Pollut. 110, 375–391. Nizzetto, L., Cassani, C., Di Guardo, A., 2006. Deposition of PCBs in mountains: the forest filter effect of different forest ecosystem types. Ecotoxicol. Environ. Safety 63, 75–83. Nizzetto, L., Jarvis, A., Brivio, P.A., Jones, K.C., Di Guardo, A., 2008. Seasonality of air-forest canopy POP exchange. Environ. Sci. Technol. 42, 8778–8783. Shoeib, M., Harner, T., 2002. Characterization and comparison of three passive air samplers for persistent organic pollutants. Environ. Sci. Technol. 36, 4142–4151. Suter II, G. W, 2007. Ecological Risk Assessment, second ed. CRC Press, Boca Raton. TGD, 2003. Technical Guidance Document on Risk Assessment, Part II, European Communities. Institute for Health and Consumer Protection, European Chemicals Bureau.
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Thibodeaux, L.J., Mackay, D., 2011. Handbook of Chemical Mass Transport in the Environment. CRC Press, Boca Raton. Trapp, S., Franco, A., Mackay, D., 2010. Activity-based concept for transport and partitioning of ionizing organics. Environ. Sci. Technol. 44, 6123–6129. van Leeuwen, C.J., Vermeire, T.G. (Eds.), 2007. Risk Assessment of Chemicals: An Introduction, second ed. Springer, Dordrecht.
Relevant Websites http://www.epa.gov/ – Environmental Protection Agency. http://echa.europa.eu/ – European Chemicals Agency. http://prtr.ec.europa.eu/ – European Pollutant Release and Transfer Register. http://ec.europa.eu/environment/chemicals/reach/reach_intro.htm – Registration, Evaluation, Authorisation and Restriction of Chemicals EU site. http://chm.pops.int/default.aspx – Stockholm convention. http://www.epa.gov/tri/ – Toxics Release Inventory.
Environmental Exposure Assessment A Di Guardo, University of Insubria, COMO, Italy Ó 2014 Elsevier Inc. All rights reserved.
Definition While exposure to human beings is generally evaluated by considering a variety of ways in which exposure could take place, such as at the workplace (occupational exposure), from use of consumer products (consumer exposure), and indirectly via the environment, the environmental exposure of the organisms in the ecosystems is generally assessed by evaluating the concentration of a specific chemical in the main environmental media (air, soil, water, sediment) and in biota as food source for other organisms (predators) by secondary poisoning. Exposure is therefore derived by considering that the organism is living in or interacting with such media, in lieu of considering the internal concentrations as resulting from the exposure. This makes the study apparently simple, but potentially complicated if the variety of environments (and ecosystems) in which the huge multiplicity of organisms live and interact is considered. As an example, biomagnification (accumulation of a chemical through the food chain) can occur for certain types of molecules, and its intensity depends on the type and length of the food chain, which could be considerably different in different ecosystems. Environmental exposure assessment (EEA) is also one component of the phases of risk assessment and in particular it is usually compared with the effect assessment in order to characterize the ecological risk in ecological risk assessment (ERA) (Figure 1). The concentrations reached in one specified compartment or environmental medium (air, water, soil, sediment) are compared with their corresponding concentrations which are considered as thresholds (no effect concentrations) for not causing any observable adverse effect to the ecosystem. This is usually performed by comparing a predicted environmental concentration (PEC), generally produced using environmental fate models (EFMs) and referred to a particular phase or compartment, with a predicted no effect concentration (PNEC), generally calculated starting from toxicological data such as acute toxicity (LD50 or Exposure assessment
Effect assessment
Risk characterization Figure 1
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Exposure assessment in environmental risk assessment.
LC50) or measured no observable effect concentration (or levels) (no-observed-effect concentration (NOEC) or no-observed-effect level (NOEL)) and referred to the same phase or compartment. When the ratio (also called risk quotient) is higher than 1, risk may occur and measures should be taken.
Ecosystem Complexity and Exposure EEA is a complicated task because it deals with the evaluation of the exposure of the ecosystems, which are many and involve a high degree of diversity. Therefore, EEA should balance the desire of completeness and overall protection with the need to simplify and streamline the processes. The first point in complexity regards the biodiversity of organisms: for example, the high number and variety of species of organisms. More than 1.5 million species have been classified (mostly insects, spermatophytes, molluscs, fungi), which are adapted to living in a multiplicity of environments, at different spatial scales. At global level, this is evident observing the very different conditions which determine the main terrestrial biomes at the different latitudinal regions, from average annual air temperature (and its annual, seasonal, daily variation) to rainfall abundance and pattern, together with the variability of the abiotic substrate. Spatial variability can be therefore observed at the different continental, regional, local, and even at much smaller scale (e.g., a single lake, river, or environment characterized by peculiar characteristics). When observing temporal scales, the wide variety of life cycles of the organisms is rather important in considering the effect of the duration of the exposure: as an example, generation times can be less than a day for bacteria to years for plants and animals. The interaction of such spatial and temporal issues with organism life cycles determines the complexity of the exposure which cannot be in principle obtained knowing only the external exposure (or the concentration in the environmental media or compartment where the organism lives). The internal dose, which is the real amount of the chemical reaching the organism and capable of determining the effects, is dependent on the interaction between the organism and a number of factors, such as the exposure time, the homogeneity of the chemical distribution in the environmental phase, its bioavailability, a number of abiotic factors, and a number of biotic factors (surface to volume ratio, feeding and growth rates, life history, behavior, etc.). Exposure time can be very different because of the organism life cycle and duration of each phase (juvenile or larval stages vs adult stage) or because the chemical can be emitted with different temporal patterns: for example, the
Encyclopedia of Toxicology, Volume 2
http://dx.doi.org/10.1016/B978-0-12-386454-3.00553-4
Environmental Exposure Assessment
discharge of chemicals into the environment can be constant, determining steady state or time invariant conditions, such as the effluents of a wastewater treatment plant in normal operating conditions (without a flood event) or can be rapidly changing, such as the pesticide pulses in surface water determined, for example, by rainfall over treated fields and consequent runoff. Also chemical distribution in the environmental media can be rather inhomogeneous, depending in part on the poor mixing of certain phases (soil, sediment vs air or water) and on the movement of chemical in the compartment: for example, cracks and macropores in soil could facilitate the vertical movement of waterborne chemicals, while in zones without cracks or macropores chemical moves more slowly. Bioavailability of the chemical is also important in soil, sediment, or water environments: the bioavailable fraction of a chemical in an environmental medium (e.g., water) can be defined as the fraction of the chemical which is available for organism uptake and therefore capable of determining an effect. Such fraction (truly dissolved) could be a small quantity compared to the total (bulk) concentration present (and measured) in water. The bioavailable fraction of a chemical can be influenced by the sorption of the chemical onto the organic carbon in suspended solids and to dissolved organic carbon, or, by other environmental conditions, such as pH, salinity, and cation exchange capacity. Abiotic factors include climatic conditions (such as temperature, insulation, wind speed, and rainfall) or substrate characteristics such as texture and structure in soil. Their variations depend on both spatial and temporal issues and may greatly influence the exposure of the organisms. Among the biotic factors, surface to volume ratio of the organisms is important when the chemical must diffuse through a surface as a cuticle; and the higher surface to volume ratio for small organisms will make diffusion much faster and therefore may enhance the exposure of small and not mobile organisms. Feeding and growth rates are specific parameters regulating how an organism would grow according to the availability of food, its intake, and the efficiency of its transformation in new body structures. These two parameters are very variable among the species and of course could regulate the intake of chemical through the diet and/or its accumulation or dilution in the body. Life histories and life cycles are indeed important in the exposure because many organisms can change type of exposure medium (aquatic vs terrestrial) or type of food during their development through the larval and adult stages. The sensitivity to the chemical could also change during the evolution through these stages, so that the same exposure concentration could produce very different effects in different larval stages of the same species. Last but not least, organism behavior can influence the exposure because the organism may be rather mobile through the environment, or migrate, hibernate, or possess resistant structures, which could reduce the exposure.
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are probably a factor of 100 less, so in the order of tens of thousands. In order to characterize their identity, their properties in order to evaluate their fate and behavior as well as the toxicological profile, many initiatives were devised to collect and classify information on new and existing chemicals, such as Toxic Substances Control Act (TSCA) in the United States, Canadian Environmental Protection Act (CEPA), and more recently Registration, Evaluation, Authorisation and Restriction of Chemicals (REACH) in European Union. A number of priority lists were created in the past to handle substances for which undesired and initially unrecognized effects were found: from chemicals capable of altering pH of rain water, generating acid rain (such as sulfate oxides and nitrogen oxides) to the so-called persistent organic pollutants (POPs) and persistent, bioaccumulative, and toxic (PBT) chemicals. POPs and PBTs are organic chemicals that, according to the Stockholm Convention, a global agreement adopted in 2001 to ban or regulate the use of a number of POP chemicals, possess a particular combination of physical and chemical properties such that, once released into the environment, they remain intact for exceptionally long periods of time (many years); become widely distributed throughout the environment as a result of natural processes involving soil, water, and, most notably, air; accumulate in the fatty tissue of living organisms including humans, and are found at higher concentrations at higher levels in the food chain; and are toxic to both humans and wildlife. The chemical families initially 12 and 9 were later included. Among those families some organochlorine pesticides can be found, such as aldrin, chlordane, DDT, dieldrin, endrin, heptachlor, hexachlorobenzene, mirex, toxaphene; industrial chemicals such as polychlorinated biphenyls (PCBs); and by-products, such as polychlorinated dibenzo-p-dioxins and polychlorinated dibenzofurans (dioxins and furans). More recently, a number of chemicals recognized were being capable of interfering with the endocrine system in animals and they are generally known as endocrine disruptors or environmental hormone disruptors. Some of the POPs listed above are considered endocrine disruptor chemicals. Other chemicals of environmental relevance are pesticides, biologically active substances which are deliberately emitted into the environment at relevant rates. They are designed to eliminate animal pests and weeds in order to protect crops but could escape from treated fields and reach surface and groundwaters as well as not treated fields and natural areas due to drift and air transport of their vapor phase. Among the most recently recognized environmental chemicals, there are many human and veterinary pharmaceutical substances, which, because of their characteristics, may travel through the wastewater treatment plants with no or little reduction in amount and end up in surface waters or in soils through the sewage sludge, when used as biosolids to restore soil fertility.
Assessment of Concentrations in Environmental Media Environmental Chemicals Registered chemical substances are nowadays about 60 millions, as recorded in the Chemical Abstracts Service (CAS) database. Among those, chemicals in commerce produced at a rate which could make them significant for the environment
Environmental exposure can be assessed by evaluating the multimedia path of a chemical in order to establish the prevailing concentrations in the compartment by means of measurements or modeling approaches. Both approaches have advantages and disadvantages and they are often both required
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to obtain an adequate level of insight into the understanding of environmental exposure.
Monitoring Environmental monitoring generally provides data on average concentrations in environmental media (air, water, soil, sediment). Peak concentrations are obtained when the measurement is performed at the point of discharge (air, water). While these data are important to estimate the order of magnitude of emissions, they generally do not allow to portrait a complete picture of the concentrations in the environment, because the chemical can undergo a series of transformations and transfer among media before reaching the point of measurement. Monitoring implies a number of activities in order to capture concentrations in the environmental compartments: from the preparation of a statistically sound sampling scheme to the selection of a sampling method, transport, storage, analytical and laboratory requirements for the analysis, as well as data quality and reporting issues. Samples of air, biota, water, soil, and sediment can be taken by employing several techniques: for example, air can be collected in a cartridge using a pump and a flowmeter to obtain the volume of air measured, water with sampling bottles or automatic sampling devices, soil with a core sampler, and sediment with a grab sampler such as a dredge. Passive sampling techniques can also be used to record concentrations in a certain phase such as air with shielded polyurethane foam (PUF) devices, or water with semipermeable membrane devices (SPMDs) or solid-phase micro extraction (SPME) techniques. Artificial passive samplers have advantages over the ‘natural’ passive samplers (such as leaves) in a way that they can be standardized and cleaned up before sampling in order to obtain comparable initial conditions. Passive samplers can also be used to sample a phase (e.g., air) for relatively long time (weeks, months) and therefore detecting chemicals present at very low concentrations. However, their sampling rate depends on a number of environmental conditions (such as temperature, and wind speed, for some air samplers) and specific physical–chemical properties; therefore, they generally provide order of magnitude estimates. Current passive devices also have the disadvantage of generally providing average concentrations of the sampling period and missing the information on peak concentrations. Planning a monitoring program raises a series of questions, such as those related to which parameters and chemical should be included; and where and when they should be measured. Additionally, it may be inconsistent and produce no results without prior knowledge of the environmental fate of a chemical. Data obtained without a properly selected sampling and statistically sound sampling scheme may produce a poor representation of the spatial and temporal characteristics of the contamination. This means that they can seldom catch the variability of concentration change in a territory in time and the relative spatial gradient. Sometimes, when the amount of spatial data is sufficient, geostatistical techniques can be used to reconstruct a spatial trend (e.g., in soil) but they are usually limited to static representations of contamination at a certain point in time. Finally, environmental monitoring is an a posteriori approach, in other terms the contamination needs to be in place to be measured, while it would of course be
preferable to act before a contamination (and a potential damage) has occurred. Finally, monitoring data can be used to gather a picture of the contamination of a certain compartment or a certain ecosystem and later used to ‘calibrate’ or ‘validate’ EFMs.
Modeling Modeling is an a priori approach and represents the only possibility of generating information for predictive purposes and for estimating potential exposure before a new chemical is used. On the other hand, models cannot be calibrated and validated without experimental data. In order to avoid mistakes when applying predictive models, the proper model for the specific environmental situation must be carefully selected. EFMs are nowadays among the most used tools to evaluate the fate of chemicals in the environment, due to their general predictive nature and the substantial ease of implementation. EFMs can be physical dispersion models, used to calculate concentrations in a phase (air, water, soil) in the proximity of a point source or partitioning or compartmental models, also known as multimedia fate models (MFMs). Among them, the fugacity-based models, developed by Don Mackay and coworkers, can be found. They are also called ‘Mackay’s type’ models. Fugacity is an equilibrium criterion which can be used to define the partition of nonpolar chemicals between phases (instead of concentration as in other models). Other equilibrium criteria were adopted later to deal with polar and/or dissociated chemicals: the aquivalence first and, more recently, the concept of activity. Many MFMs are called box models because they divide the environment in well-mixed boxes or compartments which can exchange chemicals among them; chemicals can enter the environment through direct emission or passively carried by advective fluxes in air and water (Figure 2). Chemicals could also disappear from the modeled environment due to advection or degradation and can move through the environment due to compartment exchange which in turn depends on substance and environmental characteristics. MFMs generally require data on chemical properties, environment conditions, and rates or quantities of chemical discharges and produce a picture of the environmental fate of
Air
Soil
Water Sediment
Figure 2 Multimedia environment, with the four main compartments (air, water, soil, sediment). Arrows represent exchange fluxes between compartments. Emission of chemical, advection, and degradation not depicted.
Environmental Exposure Assessment
a chemical, often in the form of a mass balance. More specifically the data required are the following: l
physical–chemical properties and half-lives; chemical emissions and temporal/spatial patterns; l environmental compartment characteristics; and l mass transfer coefficients between compartments. l
The relevant physical–chemical properties for chemicals (water solubility, vapor pressure, octanol/water partition coefficient (Kow), other solids/water partitioning coefficients, octanol/air partitioning coefficient, Henry’s Law Constant or air/water partitioning coefficient, pKa, etc.) and the half-lives in the various environmental compartments (air, particles, water, soil, sediment, suspended solids, vegetation, animals) are the basic parameters needed to simulate the behavior of nonpolar organic chemicals in the environment. These basic parameters are used in the models to calculate partitioning parameters of the selected chemical between the phases of the compartments: vapor pressure and water solubility are used to calculate Henry’s Law Constant and air/water partitioning coefficient (Kaw), Kow is used to calculate the bioconcentration factor in aquatic biota, partitioning to organic carbon (Koc) and to solid phases in suspended solids, soil, and sediment (Kp), and, together with Kaw, the octanol/air partitioning coefficient, which can be used to calculate the partitioning in particulate matter in air and plant leaves. Half-lives are sought for the main compartments and used to calculate the degradation losses. Chemical emissions are key factors for the calculations of the fate of chemicals because they determine the amount introduced in the modeling environment. In steady state approaches (e.g., European Uniform System for Evaluation of Substances (EUSES) or the fugacity model, equilibrium criterion (EQC) model) they can be annual emissions (kg year1), while in dynamic approaches a temporal pattern should be specified (e.g., 2 kg ha1 for the first 2 h). Very often the temporal pattern is ignored or unknown. This is generally true also for the spatial pattern. Emission can be estimated by evaluating monitoring data, using estimation techniques based on population or employee density or using existing national and international emission databases such as the US Environmental Protection Agency (US EPA), Toxics Release Inventory (TRI) Program or the European Pollutant Release and Transfer Register (EPTR). The environmental compartment characteristics regard their composition and their temporal and spatial variabilities. Models require sizes and volumes as well as volume fractions of solids, air, water, etc. in subcompartments for each box. In most of the regulatory approaches, these are fixed parameters characterizing ‘average’ characteristics. Environmental temperature is usually considered among these parameters, as well as organic carbon fraction, fraction of air/ water/solids/lipids, soil texture, etc. Compartments can be discretized vertically or horizontally: in the vertical discretization each compartment can be made as one box or layered (e.g., soil, air, water, sediments), while horizontal compartments can be described with regionally averaged properties (organic carbon, texture, hydrological data, etc.) vs distributed (typical/ spatially explicit) properties (often supported by a geographic information system (GIS)).
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Mass transfer coefficients are specific intercompartmental velocities used in box models to predict the rate of transfer among phases. They are generally implemented in nonequilibrium models and are referred to fixed environmental conditions. Environmental models require calibration/verification/ validation procedures to assure their adequacy and datasets of environmental/measured chemical concentrations in order to proceed in such direction.
Environmental Fate Models Many EFMs have been developed in the past 30 years. Most are steady state models (in which chemical discharge is constant in time), such as the EQC and ChemCAN models or the EUSES model. Some were later developed as unsteady state or dynamic models (historically in terms of chemical discharge, now also in terms of environmental scenario changes). The steady state models are better suited to situations in which chemical emission does not vary significantly during a certain period of time (e.g., sewage treatment plant (STP) discharges). They are generally adopted for the simple mass balance equations obtained in such situations and because they do not oblige the user to provide a time-varying emission profile (at hour, day, week, month, year level). Dynamic models are better suited to handle episodic discharges, such as pulse discharges (e.g., pesticides) or mass movement of chemical caused by some environmental phenomena (such as a runoff event triggered by rainfall). Models can be built with different types of spatial resolution, presenting large compartments (regional) or smaller or site-specific descriptions (local), sometimes nested (a local model into a regional one), such as in the EUSES model, for which global, continental, regional, and local scenarios are present. Other approaches include different models (e.g., EQC, ChemCAN, SoilFug) which are used to calculate the fate of chemicals from regional to site-specific (local) sizes in a tiered approach. Some models were developed to predict the fate of agrochemicals applied to soil at specified times, some of those are suited to predict the fate of pesticides leaching toward groundwater (e.g., MACRO, PEARL, PRZM models), while others are used to predict surface and subsurface runoff of chemicals reaching surface waters such as streams and rivers. When the chemical has the potential for long-range transport, regional or continental scales are not adequate and global models were developed to handle such situations. Most of these models are created linking regional models which simulate a number of latitudinal sectors connected via air and water. In another approach, existing meteorological models are connected to MFM to better handle the transport in the troposphere. Recently, many modeling efforts are devoted to introduce more ecological realism in predictive approach of exposure, including a higher degree of temporal and spatial dynamics, in order to follow the variability of the fate of a chemical.
Bioaccumulation Models While EFMs are generally employed to obtain PECs in the main environmental compartments, they fall short in predicting accumulation in biota, especially when bioaccumulation is
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important. Therefore, bioaccumulation models are predictive tools designed to evaluate the concentrations that a chemical could reach in an organism (aquatic or terrestrial) from the concentrations in environmental media (e.g., air, water, sediment, soil). The knowledge of the concentration reached in the organism or in certain tissue or organ is important to interpret the toxicological response of organisms. These levels usually depend on the interaction between the organism and the environmental media directly, but sometimes (when the chemical possesses certain physicochemical characteristics and adequate metabolic half-life, such as PBTs) a trophic enrichment could be observed.
Aquatic Organism Models For the aquatic organisms three main phenomena can be cited: bioconcentration (process in which the chemical concentration in an aquatic organism exceeds that in water as a result of exposure to waterborne chemical, usually where the chemical is taken up only from water via the respiratory surfaces such as gills and/or the skin), biomagnification (process in which the chemical concentration in an aquatic organism exceeds that in the organism’s diet, due to dietary absorption), and bioaccumulation (which sums up all possible routes of exposure). For terrestrial ecosystems similar concepts are adopted for terrestrial animals, while plants are generally considered in terms of accumulation from the chemical present in soil or in the air compartments. For the aquatic organisms, initial approaches were simple correlations (Kow based) mainly to predict bioconcentration in fish. Then more complex models appeared and several processes, detailing uptake and clearance of chemicals, were introduced. These were gill ventilation, food uptake, egestion, and metabolism, as well as reproductive and growth dilution losses. The organism was considered a single compartment. A full mass balance could be compiled, usually in steady state formulations. Mechanistic explanations were found to describe the process of bioaccumulation, separating its two components, bioconcentration and biomagnification. Role of food vs gill uptake was shown, especially for POPs, so that the importance of the biomagnification processes in determining the uptake of POPs in the trophic chains was shown. Bioavailability of chemicals in the environmental media (such as air, water, sediment, soil) and food was then investigated, and it was shown to influence the fraction available for uptake. The mechanism of uptake and release in aquatic air–breathing organism was later understood, showing that water-respiring organisms were more efficient in clearing chemical than airrespiring organisms. Therefore, air-breathing warm-blooded animals, which require more food, found were being subject to greater bioaccumulation and biomagnification from food. Other models were later developed for other aquatic organisms, such as benthic invertebrates, birds, and up to marine mammals. Some of the models were physiologically based toxicokinetic (PBTK) models, similar to those developed for pharmacokinetic studies in toxicology. These single-species models were then sometimes integrated into more complex food web models, allowing to reconstruct chemical accumulation through food chains or webs. The increase in complexity required a consequent increase in species-specific parameters (which are seldom
available) forcing the models to be based on ‘generic’ physiological parameters (such as egestion or metabolic rates) and relying on the available specific parameters (such as lipid fraction or ventilation rate).
Terrestrial Organism Models While most of the efforts were put in aquatic food web modeling, a consistent lack of approaches exists for terrestrial animals, both for single organism model and food webs. As an example, models can be found for earthworms, or some species of birds (such as seagull) and are generally scarce for other terrestrial species. Recent models are available for agricultural food chains, such as employed to investigate the uptake process in cows. When dealing with uptake in plant biomass, it has to be considered that vegetation uptake was mostly investigated in the past 25 years, and several models are now available to predict root uptake and translocation to the aerial part and in the air to leaf path, especially for nonpolar chemicals. While root uptake is generally proportional to Kow, translocation is mostly limited to low-range Kow chemicals (in the vicinity of log Kow ¼ 2) and therefore uptake of more hydrophobic chemicals in leaves is generally driven by the air to leaf partitioning. An exception was found for some species of the Cucurbitaceae family, which were capable of accumulating dioxins in the aerial parts from contaminated soil via root uptake. The accumulation of chemicals from air to leaves in canopies of existing forests was recently developed to quantify the so-called forest filter effect which depicts the forest leaf’s capability to sequester chemical from air and transfer it to the soil environment and therefore to influence the soil ecosystem. Forest filter was measured in a variety of forest types (from mixed broadleaf woods to pure conifer woods) and parameters were devised to describe the uptake from air (in which Koa is the main chemical descriptor of the uptake). The role of some ecological parameters was also investigated, showing the relationship with leaf area index (LAI) (m2 of leaves m2 of soil) and specific leaf area (SLA) (m2 kg1). LAI gives an indication of the amount of foliar surface per soil unit area and is a measure of leaf stratification or foliar density, while SLA represents the amount of surface for unit of weight of leaf. SLA is a characteristic of each species, even if it is influenced by the canopy density, in other terms from LAI. The ratio of LAI to SLA (and its variation in time) gives the amount of leaf biomass per square meter of soil, parameter which can be used in the accumulation models.
See also: Risk Assessment, Ecological; Environmental Processes; PBT (Persistent, Bioaccumulative, and Toxic) Chemicals; Persistent Organic Pollutants; Pesticides; Predicted No Effect Concentration (PNEC); REACH; Toxic Substances Control Act.
Further Reading Kelly, B.C., Gobas, F.A.P.C., 2003. An arctic terrestrial food-chain bioaccumulation model for persistent organic pollutants. Environ. Sci. Technol. 37, 2966–2974. Kuemmerer, K. (Ed.), 2004. Pharmaceuticals in the Environment, Sources, Fate, Effects and Risks. Springer, Heidelberg.
Environmental Exposure Assessment
Mackay, D., 2001. Multimedia Environmental Models: The Fugacity Approach, second ed. Lewis Publishers, Boca Raton. Mackay, D., Di Guardo, A., Paterson, S., Kicsi, G., Cowan, C.E., 1996. Assessing the fate of new and existing chemicals: a five stage process. Environ. Toxicol. Chem. 15, 1618–1626. Mackay, D., Fraser, A., 2000. Bioaccumulation of persistent organic chemicals: mechanisms and models. Environ. Pollut. 110, 375–391. Nizzetto, L., Cassani, C., Di Guardo, A., 2006. Deposition of PCBs in mountains: the forest filter effect of different forest ecosystem types. Ecotoxicol. Environ. Safety 63, 75–83. Nizzetto, L., Jarvis, A., Brivio, P.A., Jones, K.C., Di Guardo, A., 2008. Seasonality of air-forest canopy POP exchange. Environ. Sci. Technol. 42, 8778–8783. Shoeib, M., Harner, T., 2002. Characterization and comparison of three passive air samplers for persistent organic pollutants. Environ. Sci. Technol. 36, 4142–4151. Suter II, G. W, 2007. Ecological Risk Assessment, second ed. CRC Press, Boca Raton. TGD, 2003. Technical Guidance Document on Risk Assessment, Part II, European Communities. Institute for Health and Consumer Protection, European Chemicals Bureau.
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Thibodeaux, L.J., Mackay, D., 2011. Handbook of Chemical Mass Transport in the Environment. CRC Press, Boca Raton. Trapp, S., Franco, A., Mackay, D., 2010. Activity-based concept for transport and partitioning of ionizing organics. Environ. Sci. Technol. 44, 6123–6129. van Leeuwen, C.J., Vermeire, T.G. (Eds.), 2007. Risk Assessment of Chemicals: An Introduction, second ed. Springer, Dordrecht.
Relevant Websites http://www.epa.gov/ – Environmental Protection Agency. http://echa.europa.eu/ – European Chemicals Agency. http://prtr.ec.europa.eu/ – European Pollutant Release and Transfer Register. http://ec.europa.eu/environment/chemicals/reach/reach_intro.htm – Registration, Evaluation, Authorisation and Restriction of Chemicals EU site. http://chm.pops.int/default.aspx – Stockholm convention. http://www.epa.gov/tri/ – Toxics Release Inventory.