What are the four steps that describe what the body does with an environmental hazard?

Background

Environmental risk assessments require determination of risks for physiologically diverse individuals who are exposed to several air and water pollutants. In risk assessment evaluation, for the estimation of the threshold dose/concentration, a safe human dose/concentration (SHD/SHC) can be calculated. Different national and health agencies have different terminologies for the SHD/SHC. For example, the US Environmental Protection Agency (EPA) refers to this safe dose as a reference dose (RfD) or reference concentration (RfC) in the form of a dose in food or water, amount of chemical in contact with the skin, or concentration of chemical in air. The US Food and Drug Administration uses the term allowable daily intake. The World Health Organisation uses the term acceptable daily intake. Because a chemical may produce more than one toxic effect, the first step for these assessments is to identify the adverse effect that occurs at the lowest dose. The second step is identifying a threshold dose or the dose below which no deleterious effect is expected to occur. The threshold dose is referred to as the lowest observed adverse effect level (LOAEL), that is, lowest dose tested that produced an adverse effect. However, for risk assessment purposes, the no observed adverse effect level (NOAEL), defined as the highest exposure level at which no statistically or biologically significant increases are seen in the frequency or severity of adverse effect between the exposed population and its appropriate control population, is desired. Alternative to the NOAEL is a benchmark dose (BMD) or benchmark dose lower limit (BMDL). According to, for example, US EPA, deriving a BMDL involves selecting a predetermined change in the response rate of an adverse effect (called the benchmark response, generally in the range of 1–10% depending on the power of a toxicity study) and the BMDL is a statistical lower confidence limit on the dose that produces the selected response.

The US EPA has derived toxicity factors or values (RfDs, RfCs, and cancer slope factors) for many of the most toxic chemicals such as pollutants in air, food, or water. Values for a number of these chemicals are available from the EPA's online Integrated Risk Information System. These profiles often contain minimal risk levels, but are frequently based on different critical studies or derived with different uncertainty factors (UFs).

RfD/RfC as BMD is quantitative dose–response assessment of noncancer toxicity for ingested or inhaled chemicals. Noncancer toxicity refers to adverse health effects other than cancer and gene mutations. These effects include those on the tissue or organ where the chemical enters the body, such as the gastrointestinal tract, respiratory tract, or skin, and also effects that follow absorption and distribution of the toxicant to a site remote to its entry point. An acute RfD or RfC is an estimate of a continuous ingestion or inhalation exposure for an acute duration (24 h or less) while chronic RfD or RfC is an estimate of a daily ingestion of or contact with a pollutant or a continuous inhalation exposure for a chronic duration (up to a lifetime) to the human population (including sensitive subgroups). RfD or RfC is the ingestion dose (or amount in contact with the skin) or air exposure concentration above which chronic exposure could cause disease.

The RfC methodology deviates from the RfD approach by substituting the no-effects atmospheric concentration for the no-effects inhalation dose. RfC was expanded to account for the dynamics of the respiratory system and needs dosimetric adjustments to account for the species-specific relationships of exposure concentrations to deposited and delivered doses. The physicochemical characteristics of the inhaled agent can determine its interaction with the respiratory tract and disposition.

In acute exposure, an exposure time correction for adjustment of the concentration to account for different exposure periods was needed. Because the effects of acute exposure to airborne contaminants depends on momentary concentrations than on integrated exposures, the use of the total dose over a period of time is not always appropriate for acute exposures.

The RfD or RfC derivation begins with the identification of a NOAEL, LOAEL, or BMDL, as the point of departure (POD). The POD is determined for the specified adverse effect in animal experiments or in human epidemiological or occupational studies to human equivalent doses or concentrations. The RfD or RfC is an estimate that is derived from the POD for the critical effect by consistent application of safety factor or UFs. The UFs are applied to account for recognized uncertainties in the extrapolations from the experimental data conditions to an estimate appropriate to humans.

The uncertainties of available data are because of different effects in the same tissue, different end points in some studies, and different species that are used in various studies. One of several, generally 3- to 10-fold, factors is used in deriving the oral/dermal RfD or inhalation RfC from experimental data. Modifying factor may also be applied when scientific uncertainties such as statistically minimal or inadequate sample size or poor exposure characterization is not addressed by the standard UFs.

The RfD or RfC is defined as

RfDorRfC=NOAEL(orLOAEL orBMDL)÷(UF×MF)

The RfD and the RfC are reported in milligrams per kilogram per day and milligram per cubic meter, respectively.

2.2 Persistence, bioaccumulation potential, and toxicity of pharmaceuticals

The absorption and interaction capacities of pharmaceuticals with living organism makes them a threat for the whole ecosystem. Biologically active compounds existed in pharmaceuticals may lead to their adverse effects on the wildlife (so-called nontarget organisms). Pharmaceutical occurrence in environment is a reason for scientific challenges to prevent the increase of negative impacts and provide the good quality standards (Sangion and Gramatica, 2016).

Consequently, the Environmental Risk Assessment (ERA) for human and veterinary pharmaceuticals was accepted by the European legislation with two EC Directives 2004/27/EC and 2004/28/EC laws. Today the ERA is conducted based on the EMA guidelines (Sangion and Gramatica, 2016).

An ERA refers to acute toxic risk occurred in the aquatic environment. This risk calculated based on the ratio between the predicted environmental concentration (PEC) of the compounds, and the highest predicted no-effect concentration (PNEC) of these compounds. A PEC: PNEC ratio <0.1 is considered insignificant risk; 0.1–1, low risk; 1–10, moderate risk; and >10, high risk. The environmental hazard of a substance is classified by the following characteristics (Deblonde and Hartemann, 2013):

persistence-ability to resist degradation in the aquatic environment;

bioaccumulation–accumulation in adipose tissue of aquatic organisms; and

toxicity-potential to poison aquatic organism.

Each of these features is assigned a numerical value (0–3). The total of these numerical values shows the persistence, bioaccumulation, and toxicity index (PBT index) for the substance, which ranges from zero to nine. The higher PBT index of a pharmaceutical is related to its greater hazard in the environment. Pharmaceuticals with a PBT index of 9 are antifungals (ketoconazole, miconazole, terbinafine), antiinfectives (ofloxacin, efavirenz), antineoplastic agents (dasatinib, docetaxel, tamoxifen, and megestrol) and drugs for the nervous system (propofol, bromocriptine, clozapine, citalopram, etc.) (Deblonde and Hartemann, 2013).

Bioconcentration is the ability of an organism to accumulate a chemical from the ambient environment. The bioconcentration factor (BCF) is the ratio of the concentration of the chemical accumulated in the organism to its concentration in the ambient environment. A range BCF values for the bioaccumulation of pharmaceuticals have been reported. Table 2.1 show BCF ranges related to several pharmaceuticals in fish prey.

Table 2.1. BCF ranges of environmentally relevant pharmaceuticals in fish prey (Zenker et al., 2014).

BCF, Bioconcentration factor.

With permission.

According to the OECD the main criteria for accumulation is a logKow>3, meaning there is a tendency for accumulation (Rahman et al., 2012). The authors Howard and Muir shown 92 out of 275 pharmaceuticals detected in the environment have the potential bioaccumulative properties (Howard and Muir, 2011). Some ionophore antibiotics used as veterinary drugs were detected in sediments at higher concentrations than in water (Kim and Carlson, 2006). Large amounts of gemfibrozil, IBU, and DIC were also found to be bound to sewage sludge (Yu and Wu, 2012).

Biomarkers

In environmental risk assessment, identifying changes in ecological systems in relation to the contamination level is a crucial step. However, ecological indicators, as structural metrics, although can detect the degradation of surface waters they are not reliable indicators of impairments caused by contaminants. Therefore, it is necessary to develop biological measures that may serve as descriptors of cause–effect, diagnose sublethal physiological effect and low level of ecological impairment, and may inform about further degradation or improvement of benthic communities.

Biochemical and cellular responses can detect early disruption of organisms homeostasis due to chemical stressors. They are expected to be sound indicators for assessing the impact of anthropogenic contaminants in freshwater ecosystems and studying cause–effect relationships in laboratory and field studies. Numerous biomarkers, with variable specificity and sensitivity, have been used for several decades in terrestrial and aquatic invertebrates. A lot of attention has been given to molecular and biochemical targets as enzymatic activities and proteins involved in defense metabolism of invertebrates (gluthatione-S-transferase, metallothioneins, stress, multixenobiotic resistance proteins), linked to detoxification of reactive oxygen species (such as catalase, superoxide dismutase, glutathione peroxidase), and in markers of oxidative tissue damage (lipid peroxidation). More specific biomarkers are also widely used to assess the impact of chemicals mechanisms of action in invertebrates, as by example, cholinesterasic activity or vitellogenin (Vg) and vitellin-like protein. Markers of genotoxicity and immunotoxicity have been also developed in marine and freshwater macroinvertebates.

The implementation of molecular end points to provide information on chemical impacts is now particularly relevant in regard to the rapid development of proteomic and genomic approaches, offering many opportunities to identify and use molecular responses highly specific and directly related to a type of contaminants. Invertebrates ecotoxicology should benefit from this increasing availability of relevant molecular tools for toxicity assessment, which has been still limited up to now in nonvertebrates species.

Abstract

The terrestrial environmental risk assessment of chemical substances covers the prediction and evaluation of the likelihood and magnitude of the potential adverse effects; these substances may produce on terrestrial ecosystems and soil-based biological communities. Traditionally, the assessment has been based on two parallel evaluations, one for the soil and the other for aboveground compartment. The methodology for the soil compartment was mostly developed through the adaptation of that developed for the aquatic environment. For the above compartment, specific methods and conceptual models have been developed. A particular emphasis has been dedicated to the assessment and prediction of the potential adverse effects of agrochemicals and on tools for assessing and managing soil pollution including site-specific assessment related to waste and industrial soil contamination.

5 Limitations of conventional ERA for the case of pharmaceuticals

5.5.8 Environmental Risk Assessment (ERA)

ERA can be defined as the process of assigning magnitudes and probabilities to the adverse effects of human activities. The process involves identification of hazards (e.g., the release of toxic chemicals to the environment) by quantification of the relationship between an activity associated with an emission to the environment and its impacts. The entire ecological hierarchy is considered in this context, implying that the impacts on the cellular level, organism level, population level, ecosystem level and the entire ecosphere should be considered.

The application of environmental risk assessment is rooted in the recognition that:

The cost of elimination of all environmental effects is impossibly high,

The decision in practical environmental management must always be made on basis of incomplete information.

Humans use about 100,000 chemicals in such quantities that they may threaten the environment. However, we only know a small fraction about what we must know about these chemicals to be able to make a proper and complete environmental risk assessment of them.

ERA is a process complementary to environmental impact assessment, EIA (Jørgensen, 1991), with the latter used to assess the impacts of a human activity. EIA is predictive, comparative and concerned with all possible effects on the environment, including secondary and tertiary (indirect) effects, while ERA attempts to assess the probability of a given (defined) adverse effect resulting from a considered human activity.

Both ERA and EIA use models to find the expected environmental concentration (EEC), which is translated into impacts for EIA, and to risks of specific effects for ERA.

Uncertainty plays an important role in risk assessment. Risk is the probability that a specified harmful effect will occur or, in the case of a graded effect, the relationship between the magnitude of the effect and its probability of occurrence.

Risk assessment has emphasized risks to human health, ignoring ecological effects to a certain extent. However, it has now been acknowledged that some chemicals with little or no risk to human health can cause severe effects to aquatic and other organisms. Examples include chlorine, ammonia and certain pesticides. An up-to-date risk assessment, therefore, comprises consideration of the entire ecological hierarchy, which is the ecologist's perspective in terms of level of organization. Organisms interact directly with the environment, and they are exposed to toxic chemicals. The reproducing population is the smallest meaningful level in the ecological sense. However, populations do not exist in vacuum, but rather within a community of other organisms of which the population is a part. The community occupies a physical environment, with which it forms an ecosystem.

Moreover, both the various adverse effects and the ecological hierarchy have different scale in time and space, which must be included in a proper environmental risk assessment (Fig. 5.18). For example, oil spills occur at a spatial scale similar to those of populations, but they are briefer events than population processes. Thus, a risk assessment of a oil spill requires considerations of reproduction and recolonization that typically occur on a longer time scale than the spill, and that determine the magnitude of the population responses and its significance to natural population variance safety factors. Uncertainties have three basic causes:

The inherent randomness of the world (i.e., stochasticity),

Errors in execution of assessment,

Imperfect or incomplete knowledge.

The inherent randomness refers to the uncertainty that can be described and estimated, but cannot be reduced because it is a characteristic of the system. The meteorological factors (e.g., rainfall, temperature, wind) are effectively stochastic at levels of interest for risk assessment. Many biological processes, such as colonization, reproduction and mortality, also must be described stochastically.

Human errors are inevitably attributes of all human activities. This type of uncertainty includes incorrect measurements, data recording errors, computational errors, etc.

Lack of knowledge results in undefined uncertainty that cannot be described or quantified. It is a result of practical constraints on our ability to accurately describe, count, measure or quantify everything relevant to an estimate of risks. Prominent examples are an inability to test all toxicological responses of all species exposed to a pollutant, and simplifications needed in a model to predict the expected environmental concentration.

The most important feature distinguishing risk assessment from impact assessment is the emphasis in risk assessment on characterizing and quantifying uncertainty. Thus, it is of particular interest in risk assessment to be able to analyze and estimate the analyzable uncertainty. They include natural stochasticity, parameter errors and model errors. Statistical methods may provide direct estimates of uncertainties, and are widely used in model development (e.g., see Jørgensen, 1994).

The use of statistics to quantify uncertainty is complicated in practice by the need to consider errors in both the dependent and independent variables, and to combine errors when multiple extrapolations should be made. Monte Carlo analysis is often used to overcome these difficulties (e.g., see Bartell et al., 1984).

Model errors include inappropriate selection or aggregation of variables, incorrect functional forms and incorrect boundaries. The uncertainty associated with model errors are usually assessed by field measurements utilized for calibration and validation of the model (Jørgensen, 1994).

Risk assessment of chemicals can be divided into nine steps (Fig. 5.19). The nine steps correspond to questions the risk assessment attempts to answer in order to quantify the risk associated with the use of a chemical, as follows:

Step 1. Which hazards are associated with the application of the chemical? Answering this question involves gathering data on the types of hazards, including possible environmental damages and human health effects. The health effects include congenital, neurological, mutagenic and cancerogenic effects. It may also include characterization of the behavior of the chemical within the body (interactions with organs, cells or genetic material). What is the possible environmental damage, including lethal effects and sublethal effects on growth and reproduction of various populations?

A variety of toxicity tests has been devised to attempt to quantify the potential danger posed by chemicals. Some recommended tests involve experiments with subsets of natural systems (i.e., microcosms) or with entire ecosystems. The majority of the testing of new chemicals for possible effects, however, has been confined to laboratory studies on a limited number of test species. Results from these laboratory assays provide useful information for quantifying the relative toxicity of different chemicals. They are used to forecast effects in natural systems, although their justifications have been seriously questioned.

Step 2. What is the relation between dose and responses of the type defined above in Step 1? It implies knowledge of no effect concentration (NEC), LDx-, LCy- and ECz-values, where x, y and z express a probability of harm. The answer can be found by laboratory examinations or estimation methods. Based on these answers, a most probable level of no effect, NEL·, is assessed.

Data needed for Steps 1 and 2 above can be obtained directly from scientific libraries. They also are increasingly being found via online data searches in bibliographic and factual databases. Data gaps should be filled with estimated data (e.g., see Jørgensen et al., 1997).

Step 3. Which uncertainty (safety) factors reflect the degree of uncertainty that must be taken into account when experimental laboratory data or empirical estimations methods are extrapolated to real situations? Safety factors ranging from 10 to 10,000 are typically used. If adequate knowledge about a chemical is available, a safety factor of 10 may be applied. On the other hand, if the available information has a high uncertainty, a safety factor of 10,000 may be appropriate. Safety factors ranging between 100–1000 are most frequently applied. The result of NELx, the safety factor, is called the predicted non-effect level, PNEL. The complexity of environmental risk assessment is often simplified by deriving predicted no effect concentration, PNEC, for different environmental compartments (e.g., water, soil, air, biota, sediment).

Step 4. What are the sources and quantities of emissions? The answer to this question requires a complete knowledge of the production and use of the chemical compounds of concern, including an assessment of how much of the chemical is wasted to the environment by production and use? The chemical also may be a waste product, making it very difficult to determine the amounts involved. The very toxic dioxins, for example, are waste products of the incineration of organic waste.

Step 5. What is (are) the actual exposure concentration(s)? The answer to this question is called the predicted environmental concentration, PEC. Measuring environmental concentrations can assess exposure. It also may be predicted with a model when the emissions are known. The use of models is necessary in most cases, either because we are considering a new chemical, or because the assessment of environmental concentrations require a very high number of measurements to determine the variations in concentrations over time and space. Further, it always provides an additional degree of certainty to compare model results with measurements, implying that it is always advisable both to develop a model and to make at least a few measurements of concentrations in the ecosystem components, where the highest concentrations are expected. Most models will require an input of parameters which describe the properties of the chemicals and the organisms, and which also will require an extensive application of handbooks and a wide range of estimation methods.

The development of an environmental, ecotoxicological model, therefore, requires an extensive knowledge of the physical-chemical-biological properties of the considered chemical compound(s). The selection of a proper model is discussed in Section 5.4.

Step 6. What is the ratio PEC/PNEC? This ratio is often called the risk quotient. It should not be considered an absolute assessment of risks, but rather a relative ranking of risks. The ratio is usually found for a wide range of ecosystems (e.g., aquatic ecosystems, terrestrial ecosystems, ground water).

Step 7. How will one classify the risk? The valuation of risks are made in order to decide on risk reductions (Step 9). Two risk levels are defined:

The upper limit (maximum permissible level, MPL), and

The lower limit (negligible level, NL). It may also be defined as a percent of the MPL. The two risk limits create three zones: A black, unacceptable, high risk zone > MPL, a gray, medium risk level, and a white, low risk level < NL.

Step 8. What is the relation between risk and benefit? This analysis involves examining socioeconomic, political and technical factors, which is beyond the topic of this volume. The cost-benefit analysis is also difficult, because the costs and benefits are often of a different order.

Step 9. How can the risk be reduced to an acceptable level? The answer to this question requires a thorough technical, economic and legislative investigation before it can be given. Assessment of alternatives is often an important aspect in risk reduction.

Steps 1–3 and 5 require knowledge of the properties of the chemical compounds of concern, which again implies an extensive literature search and/or selection of the best feasible estimation procedure.

Abstract

Site-specific environmental risk assessments estimate the kind, likelihood, and magnitude of the environmental effects associated with a particular stress, e.g., an anthropogenic activity, on one or several interconnected ecosystems within a defined geographical area. There are many options and alternatives to be considered when conducting a site-specific assessment, which should be defined in the problem formulation. The main elements to be considered include the regulatory or nonregulatory nature of the assessment; the identification of the stress source(s) and the limits and coverage of the studied ecosystem(s); the overall aim of the assessment, e.g., predictive, diagnostic, or restorative; and the expected outcome, e.g., just the identification/confirmation of risk or also the selection/implementation of risk control and management measures.

8.2 Environmental Risk Assessment

8.2.1 Overview of Environmental Risk Assessment

A brief introduction to the concepts of ERA is given in this section to introduce readers to the concepts and ideas behind the application of ecotoxicological models to assess an environmental risk.

Treatment of industrial wastewater, solid waste, and smoke is very expensive. Consequently, the industries attempt to change their products and production methods in a more environmentally friendly direction to reduce the treatment costs. Therefore, industries need to know how much the different chemicals, components, and processes are polluting our environment. In other words: What is the environmental risk of using a specific material or chemical compared with other alternatives? If industries can reduce their pollution just by switching to another chemical or process, then they will reduce their environmental costs and improve their green image. An assessment of the environmental risk associated with the use of a specific chemical and a specific process enables industries to make the right selection of materials, chemicals, and processes to benefit the economy of the enterprise and the quality of the environment.

Similarly, society needs to know the environmental risks of all chemicals to phase out the most environmentally threatening chemicals and set standards for the use of all other chemicals. The standards should ensure there is no serious risk in using the chemicals, provided that the standards are followed carefully. Modern abatement of pollution includes ERA, which is defined as the process of assigning magnitudes and probabilities to the adverse effects of human activities. This process involves identification of hazards such as the release of toxic chemicals to the environment by quantifying the relationship between an activity associated with an emission to the environment and its effects. The entire ecological hierarchy is considered in this context including the effects on the cellular (biochemical) level, the organism level, the population level, the ecosystem level, and the entire ecosphere.

The application of ERA is rooted in the recognition that:

1.

The elimination cost of all environmental effects is impossibly high.

2.

Practical environmental management decisions must always be made on the basis of incomplete information.

We use about 100,000 chemicals in amounts that might threaten the environment, but we know only about 1% of what is necessary to make a proper and complete ERA of these chemicals. Section 8.5 is a short introduction to available estimation methods to apply if information about properties of chemical compounds is unavailable in the literature. A list of relevant properties and how they impact the environment is also given.

ERA is in the same family as environmental impact assessment (EIA), which attempts to assess the impact of a human activity. EIA is predictive, comparative, and concerned with all possible effects on the environment, including secondary and tertiary (indirect) effects, whereas ERA attempts to assess the probability of a given (defined) adverse effect as a result of human activity.

Both ERA and EIA use models to find the expected environmental concentration (EEC), which is translated into impacts for EIA and to risks of specific effects for ERA. Development of ecotoxicological models for assessing environmental risks is detailed in the following section. An overview of ecotoxicological models is given in Jørgensen et al. (1995).

Legislation and regulation of domestic and industrial chemicals for the protection of the environment have been implemented in Europe and North America for decades. Both regions distinguish between existing chemicals and introduction of new substances. For existing chemicals, the EU requires a risk assessment to humans and the environment according to a priority setting. An informal priority setting (IPS) is used for selecting chemicals among the 100,000 listed in “The European Inventory of Existing Commercial Chemical Substances.” The purpose of the IPS is to select chemicals for detailed risk assessment from among the EEC high production volume compounds, that is, >1000 t/y (about 2500 chemicals). Data necessary for the IPS and an initial hazard assessment are called Hedset and cover issues such as environmental exposure, environmental effects, exposure to humans, and human health effects.

At the UNCED meeting on the Environment and Sustainable Development in Rio de Janeiro in 1992, it was decided to create an Intergovernmental Forum on Chemical Safety (IGFCS, Chapter 19 of Agenda 21). Its primary task is to stimulate and coordinate global harmonization in the field of chemical safety and covers the following principal themes: assessment of chemical risks, global harmonization of classification and labeling, information exchange, risk reduction programs, and capacity building in chemical management.

Uncertainty plays an important role in risk assessment (Suter, 1993). Risk is the probability that a specified harmful effect will occur or, in the case of a graded effect, the relationship between the magnitude of the effect and its probability of occurrence.

Risk assessment has traditionally emphasized risks to human health over the concerns of ecological effects. However, some chemicals such as chlorine, ammonia, and certain pesticides — which have no risk or only a small amount of risk to human health — cause severe effects on ecosystems such as aquatic organisms. An up-to-date risk assessment is comprised of considerations of the entire ecological hierarchy, which is the ecologist's worldview in terms of levels of organization. Organisms interact directly with the environment, so they can be exposed to toxic chemicals. The species-sensitivity distribution is therefore more ecologically credible (Calow, 1998). A reproducing population is the smallest meaningful level ecologically. However, populations do not exist in a vacuum; they require a community of other organisms of which the population is a part. The community occupies a physical environment with which it forms an ecosystem.

Moreover, both the various adverse effects and the ecological hierarchy have different scales in time and space, which must be included in a proper ERA (Figure 8.1). For example, oil spills occur at a spatial scale similar to those of populations, but they are briefer than population processes. Therefore, a risk assessment of an oil spill requires the consideration of reproduction and recolonization on a longer time scale to determine the magnitude of the population response and its significance to natural population variance.

Figure 8.1. The spatial and time scale for various hazards (hexagons, italic) and for the various levels of the ecological hierarchy (circles, non-italic).

8.2.3 Step-by-Step Guide for Ecological Risk Assessment

Chemical risk assessment is divided into nine steps shown in Figure 8.2. The nine steps correspond to questions that the risk assessment attempts to answer when quantifying the risk associated with the use of a chemical.

Figure 8.2. The presented procedure in nine steps to assess the risk of chemical compounds. Steps 1–3 require extensive use of ecotoxicological handbooks and ecotoxicological estimation methods to assess the toxicological properties of the chemical compounds considered, while step 5 requires the selection of a proper ecotoxicological model.

Step 1:

Which hazards are associated with the application of the chemical? This involves gathering data on the types of hazards such as possible environmental damage and human health effects. The health effects include congenital, neurological, mutagenic, endocrine disruption (e.g., estrogen), and carcinogenic effects. It may also include characterization of the behavior of the chemical within the body (interactions with organs, cells, or genetic material). The possible environmental damage including lethal effects and sub-lethal effects on growth and reproduction of various populations is considered in this step.

As an attempt to quantify the potential danger posed by chemicals, a variety of toxicity tests have been devised. Some of the recommended tests involve experiments with subsets of natural systems, such as microcosms, or with entire ecosystems. The majority of testing new chemicals for possible effects has, however, been confined to studies in the laboratory on a limited number of test species. Results from these laboratory assays provide useful information for quantification of the relative toxicity of different chemicals. They are used to forecast effects in natural systems, although their justification has been seriously questioned (Cairns et al. 1987).

Step 2:

What is the relation between dose and responses of the type defined in step 1? It implies knowledge of NEC and LDx values (dose that is lethal to x% of the organisms considered), LCy values (concentration lethal to y% of the organisms considered), and ECz values (concentration giving the indicated effect to z% of the considered organisms) where x, y, and z express a probability of harm. The answer can be found by laboratory examination or we may use estimation methods. Based upon these answers, a most probable level of no effect (NEL) is assessed. Data needed for steps 1 and 2 are obtained directly from scientific libraries, but are increasingly found via online data searches in bibliographic and factual databases. Data gaps should be filled with estimated data. It is very difficult to completely know about a chemical's effect on all levels from cells to ecosystem as some effects are associated with very small concentrations (the estrogen effect). Therefore it is far from sufficient to know NEC, LDx-, LCy-, and ECz-values.

Step 3:

Which uncertainty (safety) factors reflect the amount of uncertainty that must be taken into account when experimental laboratory data or empirical estimation methods are extrapolated to real situations? Usually, safety factors of 10–1000 are used. The choice was discussed earlier and is usually in accordance with Table 8.1. If good knowledge about the chemical is available, then a safety factor of 10 may be applied. If, on the other hand, it is estimated that the available information has a very high uncertainty, then a safety factor of 10,000 may be recommended. Most frequently, safety factors of 50–100 are applied. NEL times the safety factor is the predicted noneffect level (PNEL). The complexity of ERA is often simplified by deriving the predicted no-effect concentration (PNEC) for different environmental components (water, soil, air, biotas, and sediment).

Step 4:

What are the sources and quantities of emissions? The answer requires thorough knowledge of the production and use of the considered chemical compounds, including an assessment of how much of the chemical is wasted in the environment by production and use. The chemical may also be a waste product, which makes it very difficult to determine the amounts involved; for instance, the very toxic dioxins are waste products from incineration of organic waste.

Step 5:

What is (are) the actual exposure concentration(s)? The answer to this question is the PEC. Exposure can be assessed by measuring environmental concentrations. It may also be predicted by a model when the emissions are known. The use of models is necessary in most cases either because we are considering a new chemical, or because the assessment of environmental concentrations requires a very large number of measurements to determine the variations in concentrations. Furthermore, it provides an additional certainty to compare model results with measurements, which implies that it is always recommended both to develop a model and to make at least a few measurements of concentrations in the ecosystem components when and where it is expected that the highest concentration will occur. Most models demand an input of parameters, describing the properties of the chemicals and the organisms, which also requires extensive application of handbooks and a wide range of estimation methods. The development of an environmental, ecotoxicological model requires extensive knowledge of the physical-chemical-biological properties of the chemical compound(s) considered. The selection of a proper model is discussed in this chapter and in Chapter 2.

Step 6:

What is the ratio PEC/PNEC? This ratio is often called the risk quotient. It should not be considered an absolute assessment of risk but rather a relative ranking of risks. The ratio is usually found for a wide range of ecosystems such as aquatic and terrestrials well as ground water. Steps 1–6 shown in Figure 8.3 agree with Figure 8.2 and the information given in the previous six steps.

Figure 8.3. Steps 1–6 are shown in more detail for practical applications. The result of these steps leads naturally to the assessment of the risk quotient.

Step 7:

How will you classify the risk? Risk valuation decides on risk reductions (step 9). Two risk levels are defined: (1) the upper limit, that is, the maximum permissible level (MPL); and (2) the lower limit, that is, the negligible level, NL. It may also be defined as a percentage of MPL, for instance, 1% or 10% of MPL. The two risk limits create three zones: a black, unacceptable, high risk zone >MPL; a gray, medium risk level; and a white, low risk level Step 8:

What is the relation between risk and benefit? This analysis involves examination of socioeconomic, political, and technical factors, which are beyond the scope of this volume. The cost-benefit analysis is difficult because the costs and benefits are often of a different order and dimension.

Step 9:

How can the risk be reduced to an acceptable level? The answer to this question requires deep technical, economic, and legislative investigation. Assessment of alternatives is often an important aspect in risk reduction.

Steps 1, 2, 3, and 5 require knowledge of the properties of the focal chemical compounds, which again implies an extensive literature search and/or selection of the best feasible estimation procedure. In addition to “Beilstein,” (http://www.reaxys.com/info/) it is recommended to have on hand the following handbooks of environmental properties of chemicals and methods for estimation of these properties in case literature values are not available:

Jørgensen, S. E., Nielsen, S. N., and Jørgensen, L. A. (1991). Handbook of Ecological Parameters and Ecotoxicology, Elsevier, Amsterdam. Year 2000 published as a CD called Ecotox. It contains three times the amount of parameters then the 1991 book edition. See also Chapter 2 for further details about Ecotox.

Howard, P. H. et al. (1991). Handbook of Environmental Degradation Rates. Lewis Publishers, New York.

Verschueren, K. (2007). Several editions have been published, the latest in 2007. Handbook of Environmental Data on Organic Chemicals. Van Nostrand Reinhold, New York.

Mackay, D., Shiu, W. Y., and Ma, K. C. (1991, 1992). Illustrated Handbook of Physical-Chemical Properties and Environmental Fate for Organic Chemicals. Volume I. Mono-aromatic Hydrocarbons. Chloro-benzenes and PCBs, 1991. Volume II. Polynuclear Aromatic Hydrocarbons, Polychlorinated Dioxins, and Dibenzofurans, 1992.Volume III. Volatile Organic Chemicals, 1992. Lewis Publishers, New York.

Jørgensen, S. E., Halling-Sørensen, B., and Mahler, H. (1997). Handbook of Estimation Methods in Environmental Chemistry and Ecotoxicology. Lewis Publishers, Boca Raton, FL.

Steps 1–3 are sometimes denoted as effect assessment or effect analysis, and steps 4–5 are exposure assessment or effect analysis. Steps 1–6 may be called risk identification, while ERA encompasses all 9 steps presented in Figure 8.2. Step 9 is very demanding, as several possible steps in reduction of the risk should be considered, including treatment methods, cleaner technology, and substitutes to replace the examined chemical.

8.2.4 Risk Assessment of medicinal and veterinarian chemicals

In North America, Japan, and the EU, medicinal products are considered similar to other chemical products because there is no difference between a medicinal product and other chemical products. In the EU, technical directives for human medicinal products do not include any reference to ecotoxicology and the assessment of their potential risk. However, a detailed technical draft guideline issued in 1994 indicated that the applied approach for veterinary medicine would also apply to human medicinal products. Presumably, ERA will be applied to all medicinal products in the near future when sufficient experience with veterinary medicinal products has been achieved. Veterinary medicinal products, on the other hand, are released in larger amounts to the environment as manure. It is also possible to perform an ERA where the human population is in focus. Ten steps corresponding to Figure 8.3 are shown in Figure 8.4, which is not significantly different from Figure 8.3. The principles for the two types of ERA are the same. Figure 8.4 uses the nonadverse effect level (NAEL) and nonobserved adverse effect level (NOAEL) to replace the PNEC, and the PEC is replaced by the tolerable daily intake (TDI).

Figure 8.4. ERA for human exposure. It leads to a margin of safety that corresponds to the risk quotient in Figures 8.2 and 8.3.

This type of ERA is of particular interest to veterinary medicine that may contaminate food products for human consumption. For instance, the use of antibiotics in pig feed has attracted a lot of attention, as they may be found as residue in pig meat or may contaminate the environment though the application of manure as natural fertilizer.

Selection of a proper ecotoxicological model is the first step in the development of an environmental exposure model, as required in step 5. It will be discussed in more detail in the next section.

Assessing Effects

For assessment purposes, ERAs commonly use the concept of a screening index (SI) which is simply the ratio of the estimated dose rate (to an individual biota) to the reference dose rate, viz.

[2]SI=Estimated dose rateReference dose rate

This comparison assumes that the numerator and the denominator of the SI are based on a common assessment of dose relevant to the endpoint of interest (e.g., mortality, reproductive capacity).

When the estimated SI is below 1, it is considered that an effect to a (population of) biota is unlikely. When an SI is estimated to be greater than 1, an effect may be possible and further detailed evaluations are carried out to investigate whether the SI being greater than 1 is an artifact of undue conservative assumptions and the nature of the calculation, or whether an actual effect might be possible. At this time, the reference dose rates described in the previous section provide an appropriate basis for assessment.

The interpretation of SIs also needs to consider aspects such as the spatial and temporal distribution of radiation, the natural cycles of populations, and background radiation levels.

Arctic caribou provide one example of a situation where natural background can result in quite high doses to biota. Caribou and reindeer subsist on lichens during the winter months but must graze over a wide area to obtain enough nourishment. Lichens do not have a root system but are highly efficient at collecting and retaining nutrient material deposited on their surfaces. Furthermore, lichens do not turn over annually but integrate all deposited materials, including radionuclides, over several decades. Hence this food chain integrates radioactivity not only over time, but also over spatial area. Estimated radiation doses to various tissues of caribou taken from areas with natural radiation in Nunavat, Canada, are over the reference dose rate. Therefore, it can be important to consider natural background in ERA calculations.

It is evident that there is a wide range of complex factors to account for in the extrapolation of dose estimation from an individual level to the population level. When applying the SI approach to nonhuman biota at the individual level, great caution is therefore necessary about the interpretation of the predicted outcomes (as in most cases they would be overly conservative at the population level). A reliable interpretation also requires knowledge on the extrapolation from the individual level to the population level of both nonhuman biota dose rates and the related potential effects.

How is an Environmental Risk Assessment Performed?

Although the goal of any ERA is generally the same (to identify the hazard, estimate the risk posed, and consider steps to reduce risk if necessary), there are a number of different frameworks used by different groups to conduct an ERA. For instance, if the hazard you’re evaluating includes economic implications (e.g., reducing noise pollution around airports by limiting use of certain types of aircraft), you’d need to include that in your risk assessment, while if you’re estimating the risk posed by a new pesticide to salmon in a river, you’d be basing your assessment on a different set of measures. Regardless of the metrics used, most ERAs involve several common phases: planning and scoping, problem formulation, and risk assessment analysis and characterization (Fig. 1.12.1).

Planning and scoping: Risk assessors will typically start by addressing a set of basic questions. Who or what is at risk? What is the environmental stressor of concern? What is the source of this stress agent? How is exposure to this stress agent propagated across environments and populations? What are the impacts of exposure to this stress agent? How long does it take for impacts to appear?

Problem formulation: The objective of the problem formulation phase is to gather all of the details needed to outline what the ERA should address. This helps to define the ecological entity you want to assess risk for and come up with an exact plan to assess this risk. This includes specifying whether the goal is to assess risk for a particular species, a functional group of species, a specific habitat, or geographic location of concern. It also needs to lay out all of the specific questions or concerns that need to be included in the assessment.

For example, suppose a chemical plant is considering the level of waste treatment necessary to minimize the risks posed by the production of a new compound, Chemical X, that will be made at the facility. Regulators design an ERA focused on the river into which the plant effluent is discharged. Specific items to be considered in this problem formulation phase might include the following: what area downriver below the discharge might be affected, and how does the area of impact differ during low flow or high flow conditions? How much of the chemical might reach a public water supply intake located 1 km below the discharge? Which species of aquatic life might be affected? Are there other substances present in the river that might interact with Chemical X?

Model development: Before estimating the extent and likelihood of risk, it’s helpful to develop a model that diagrams the relationships and pathways between the source of the hazard and the receptors at risk of harm. Such a model helps identify the key pieces of data or information necessary to complete the assessment.

In our chemical facility example, the hazard is the chemical, the source is the discharge pipe entering the river, the pathway is movement of the chemical from the waste discharge downstream in the river, and the receptors would be fish, other aquatic organisms, or the municipal water supply of concern. This means we will need information for things such as the flow of water through the system, the solubility of Chemical X, its bioaccumulation in the species of interest, etc.

Risk assessment: Here, the actual risk to the receptors is estimated for the key components identified in the model development. Calculations are used to determine the levels of exposure that will lead to harmful effects, the expected duration or lag in the timing of this risk, which plants and animals are most at risk, and what degree of exposure is likely to have harmful effects.

In our example, mathematical models could be used to calculate how much of Chemical X would be released into the ecosystem and accumulated in receptors over a given period of time. If data are available that show whether or not the chemical will be toxic to the species in the river, they can be used to estimate the risk of adverse effects. If not, site-specific laboratory tests may be needed to establish the risk.

Field studies, theoretical models, or exploration of existing research could also be used to estimate the likely concentration of Chemical X entering the drinking water intake under different flow conditions and how these levels compare to federal drinking water standards. This could then be translated into the risk of harmful impacts on populations that drink the water (Fig. 1.12.2).

An important part of the assessment process, in addition to estimating the possible risks, is to estimate the likelihood of occurrence. If, for instance, Chemical X is moderately toxic to aquatic life, but is extremely volatile, the probability of effects on aquatic life in the river would be low because the chemical would leave the water column soon after discharge. If Chemical X is persistent, the likelihood of adverse effects would be much higher.

Typically, there will be more than one receptor to evaluate, so multiple lines of evidence will be generated. In our example, one line would be the toxicity of Chemical X to aquatic life, and a second would be human health impacts related to drinking water. If all lines of evidence lead to the same conclusion (e.g., high risk or no risk), it increases the level of confidence in the results of the assessment. Conflicting lines of evidence, of course, can make the assessment more challenging.

Once various risks are estimated, an additional factor should be considered: level of certainty in the assessments. For example, perhaps the data on the toxicity of Chemical X are very limited, or the presence of other chemicals in the water raises the challenging prospect of synergistic effects. These would likely increase the level of uncertainty and lower the degree of confidence in the ERA.

Risk management: If an ERA identifies a significant risk, the next step is to develop a management approach to reduce or eliminate the risk. In our example, if the ERA determines that Chemical X poses a risk to aquatic life in the river, more effective removal technologies at the plant might be required, or, failing that, the plant might decide not to produce Chemical X. If the public is exposed to a risk, communication about the risk and proposed steps to manage it is also an important component of the process.

An Example

Chapman et al. (2010) performed an ERA to determine the risk posed by metal contaminants in soils around two lighthouses, one in Nova Scotia and the other in New Brunswick, Canada. Because of peeling paint, levels of lead and other trace metals in soils around the two lighthouses were reported to be high.

The scientists developed five lines of evidence to estimate the risk associated with exposure to the lead contaminated soils:

a.

Comparison of metal levels in those soils to Canadian soil quality guidelines;

b.

Comparison of metal levels in soil leachate (water passed through soil samples) to water quality guidelines;

c.

Growth of wheat shoots in soil samples (Fig. 1.12.3);

Figure 1.12.3. Wheat shoot growth weeks 1–6 presented as average lengths in centimeters (4 replicates) in samples from Nova Scotia. Miracle-Gro’s “Seed Starting Soil” and Aurenz-Spezial-Blumenerde (ASB) Green World’s “Garden Soil, Original Grower Mix” were used as control soils. Samples NS0–NS35 were collected in a transect out from the lighthouse, with NS0 collected directly in the structure’s drip line and NS35 collected at a distance of 35 m from the lighthouse (Chapman et al., 2010).

d.

Growth of lettuce shoots in soil samples; and

e.

Survival of the invertebrate Daphnia magna in leachate collected from soil samples. Note that in all cases, controls were used in the tests.

While the lines of evidence from (a) and (b) above indicated higher risk (total metal levels in soils and leachates often exceeded soil quality guidelines), lines of evidence from the biological exposures (c–e) were much weaker, suggesting that the risk posed by the soils to the environment could be less than indicated by simple comparisons to available guidelines. The authors concluded that some of the metals in the soil were tightly bound and therefore biologically unavailable, resulting in reduced impacts on the test species. Note that if the scientists had simply compared metal levels in the soils to available guidelines, they would have come to a different conclusion regarding the level of risk.