Risk assessment exposure dose calculation

The reader should note tliat since many risk assessments have been conducted on the basis of fatal effects, there are also uncertainties on precisely what constitutes a fatal dose of thennal radiation, blast effect, or a toxic chemical. Where it is desired to estimate injuries as well as fatalities, tlie consequence calculation can be repeated using lower intensities of exposure leading to injury rather titan dcatli. In addition, if the adverse healtli effect (e.g. associated with a chemical release) is delayed, the cause may not be obvious. Tliis applies to both chronic and acute emissions and exposures. 

Assessments of risks associated with the use of chlorpyrifos insecticide products for workers have been made. The assessments are based on the results of field studies conducted in citrus groves, a Christmas tree farm, cauliflower and tomato fields, and greenhouses that utilized both passive dosimetry and biomonitoring techniques to determine exposure. The biomonitoring results likely provide the best estimate of absorbed dose of chlorpyrifos, and these have been compared to the acute and chronic no observed effect levels (NOELs) for chlorpyrifos. Standard margin-of-exposure (MOE) calculations using the geometric mean of the data are performed however, probability (Student s f-test) and distributional (Monte Carlo simulation) analyses are deemed to provide more realistic evaluations of exposure and risk to the exposed population. 

The purpose of this chapter is not to discuss the merits, or lack thereof, of using plasma cholinesterase inhibition as an adverse effect in quantitative risk assessments for chlorpyrifos or other organophosphate pesticides. A number of regulatory agencies consider the inhibition of plasma cholinesterase to be an indicator of exposure, not of toxicity. The U.S. Environmental Protection Agency, at this point, continues to use this effect as the basis for calculating the reference doses for chlorpyrifos, and it is thus used here for assessing risks. 

A total of 10,000 iterations or calculations of dose were performed as part of this simulation, and Figure 4 shows the resulting distribution of average daily doses of chlorpyrifos as determined by the Monte Carlo simulation. Common practice in exposure and risk assessment is to characterize the 50th percentile as a "typical" exposure and the 95th percentile as the "reasonable maximum" exposure.4 The distributional analysis for these calculated doses... 

The following example is based on a risk assessment of di(2-ethylhexyl) phthalate (DEHP) performed by Arthur D. Little. The experimental dose-response data upon which the extrapolation is based are presented in Table II. DEHP was shown to produce a statistically significant increase in hepatocellular carcinoma when added to the diet of laboratory mice (14). Equivalent human doses were calculated using the methods described earlier, and the response was then extrapolated downward using each of the three models selected. The results of this extrapolation are shown in Table III for a range of human exposure levels from ten micrograms to one hundred milligrams per day. The risk is expressed as the number of excess lifetime cancers expected per million exposed population. 

A third possibility consists of comparing the theoretically calculated lung cancer rate based on risk coefficients derived from miners with the actual cancer occurrence among non-miners, derived from Rn-d exposure assessment in dwellings and using appropriate exposure-dose conversion factors (Steinhausler et al.. 1983 ... 

The principal application of PBPK models is in the prediction of the target tissue dose of the toxic parent chemical or its reactive metabolite. Use of the target tissue dose of the toxic moiety of a chemical in risk assessment calculations provides a better basis of relating to the observed toxic effects than the external or exposure concentration of the parent chemical. Because PBPK models facilitate the prediction of target tissue dose for various exposure scenarios, routes, doses, and species, they can help reduce the uncertainty associated with the conventional extrapolation approaches. Direct application of modeling includes... 

Hazard identification is the process of collecting and evaluating information on the effects of an agent on animal or human health and well-being. In most cases, this involves a careful assessment of the adverse effects and what is the most sensitive population. The dose-response assessment involves evaluation of the relationship between dose and adverse effect. Typically, an effort is made to determine the lowest dose or exposure at which an effect is observed. A comparison is often made between animal data and any human data that might be available. Next is exposure assessment, in which an evaluation of the likely exposure to any given population is assessed. Important parameters include the dose, duration, frequency, and route of exposure. The final step is risk characterization, in which all the above information is synthesized and a judgment made on what is an acceptable level of human exposure. In the simplest terms, risk is the product of two factors hazard and exposure (i.e. hazard x exposure = risk). In real risk assessments, all hazards may not be known and exposure is often difficult to quantify precisely. As a result, the calculated risk may not accurately reflect the real risk. The accuracy of a risk assessment is no better than the data and assumptions upon which it is based. 

The widespread detection of phthalate metabolites in human urine has produced questions about public-health risks, especially with regard to antiandrogen effects that can influence male gonadal development (Gray et al. 2000 Parks et al. 2000). The extrapolation from urinary biomonitoring results to exposure and risk assessment has been facilitated by calculations that convert urinary metabolite concentrations to intake dose of the parent phthalate (Koo et al. 2002 Koch et al. 2003 Kohn et al. 2000 David 2000). The parent diester phthalates are rapidly and completely metabolized to the monoester metabolites, which are rapidly cleared by the kidney. Those features allow one to assume that the daily excretion rate of metabolite is equal to the daily intake rate of the parent chemical. Furthermore,... 

In summary, the PFOA risk assessment is a good example of biomonitoring-led risk assessment. There is no attempt to calculate exposure dose with pathway analysis, because the sources of human PFOA exposure are too uncertain. Instead, the biomonitoring data served as the sole source of human exposure information. Those data could be interpreted in a risk-assessment framework with the aid of PK mod-... 

The risk interpretation of biomonitoring results will tend to have additional uncertainties. That is because, in addition to the standard uncertainties encountered in risk assessment, there is the uncertainty of extrapolating from a blood or urinary concentration to an external dose. There will be variability both in the timing between sample draw and most recent exposure and in the relationship between blood concentration and dose. Those kinds of variability are compounded by uncertainty in the ability of a PK calculation or model to convert biomarker to dose accurately. For example, reliance on urinary biomarker results expressed per gram of urinary creatinine leads to an uncertain calculation of total chemical excretion per day because of the considerable variability in creatinine clearance per day. That complicates an otherwise simple approach to estimating dose. Furthermore, the conversion requires knowledge of fractional excretion via various pathways, which may not be present for a large sample of humans. The uncertainties created by these factors can be bounded via sensitivity and Monte... 

In general, calculation of the risk or dose from waste disposal in the numerator of the risk index in Equation 6.2 or 6.3 involves the risk assessment process discussed in Section 3.1.5.1. As summarized in Section 6.1.3, NCRP recommends that generic scenarios for exposure of hypothetical inadvertent intruders at waste disposal sites should be used in calculating risk or dose for purposes of waste classification. Implementation of models describing exposure scenarios for inadvertent intruders at waste disposal sites and their associated exposure pathways generally results in estimates of risk or dose per unit concentration of hazardous substances in waste. These results then are combined with the assumptions about allowable risk discussed in the previous section to obtain limits on concentrations of hazardous substances in exempt or low-hazard waste. 

For each generic exposure scenario to be used in classifying waste, and taking into account all relevant exposure pathways in each scenario, calculate the dose per unit concentration of each hazardous substance in the waste. These doses generally would be the highest values calculated over an assumed time frame for the risk assessment (see Section 6.4.5.3), taking into account the time-dependence of the concentrations of hazardous substances in the waste. For example, the quantity calculated for radionuclides would be the annual effective dose (sievert) per unit activity concentration (Bq nr3), and the quantity calculated for hazardous chemicals would be the dose (intake, mg kg 1 d-1) per unit concentration (kg m 3). 

In many respects, the foundations and framework of the proposed risk-based hazardous waste classification system and the recommended approaches to implementation are intended to be neutral in regard to the degree of conservatism in protecting public health. With respect to calculations of risk or dose in the numerator of the risk index, important examples include (1) the recommendation that best estimates (MLEs) of probability coefficients for stochastic responses should be used for all substances that cause stochastic responses in classifying waste, rather than upper bounds (UCLs) as normally used in risk assessments for chemicals that induce stochastic effects, and (2) the recommended approach to estimating threshold doses of substances that induce deterministic effects in humans based on lower confidence limits of benchmark doses obtained from studies in humans or animals. Similarly, NCRP believes that the allowable (negligible or acceptable) risks or doses in the denominator of the risk index should be consistent with values used in health protection of the public in other routine exposure situations. NCRP does not believe that the allowable risks or doses assumed for purposes of waste classification should include margins of safety that are not applied in other situations. 

The effects of genotoxic compounds are considered non-threshold. Thus, risk assessment for a given exposure is usually performed by a linear or sub-linear extrapolation from the high dose effects observed in animals to the lower human exposure. Since the outcome of the extrapolation depends on the model applied and extrapolation over different orders of magnitude is error prone, the European Food and Safety Authority (EFSA 2005) recommended to avoid this extrapolation and proposed the MOE approach. This approach uses the benchmark dose, or the T25 calculated from a carcinogenicity study and compares this with human exposure. A MOE of 10,000 and more is considered to be of minor concern. The advantage is that neither a debatable extrapolation from high to low doses needs to be performed nor are hypothetical cancer cases calculated. For details of the different approaches see, SCHER, SCCP, SCENIHR (2008). 

The route of exposure is another aspect of exposure in which health-relevance must be considered. In Section One of this book, there is a detailed discussion of exposure assessment methodologies, including the importance of identification of the most prevalent route of exposure (dermal, inhalation or oral) and the necessity of knowing the absorption of the pesticide to allow calculation of the absorbed dose for risk assessment. For epidemiological purposes, exposure-assessment smdies are usually limited to assessing contact exposure levels. Since dermal absorption is not known for many pesticides or complex mixtures, uptake through the dermal route can often not be estimated and contact exposure data are a poor proxy of internal exposure (absorbed dose) (Schneider et al., 1999). 

Probabilistic risk assessment methods are described herein for determining a popnlation s distribution of the dose from exposure and the combination of that exposnre characterization with appropriate toxicological information to form aggregate and cumulative risk assessments. An individual s dose from exposure is characterized as a set of chemical- and route-specific dose profiles over time. Toxic equivalence factors (TEFs) that reflect the toxic endpoint and exposure duration of concern are used to scale chemical- and route-specific doses to toxic equivalent doses (TEDs). The latter are combined in a temporally consistent manner to form a profile over time of the Total TED. For each individual, a Total MOE is calculated by dividing a toxicologically relevant benchmark dose (e.g. an EDio) by the individual s Total TED. The distribution of the Total MOE in a popnlation provides important information for risk management decisions. 

In addition, in vitro stndies can be nsed for semiqnantitative comparison of absorption of chemicals between species, between componnds within one species, and between different vehicles within one species. In this regard, it is important to realize that in vitro stndies give relative resnlts, i.e. that they should primarily be compared with results generated within the same test system. Various calculations can be made on the basis of in vitro data, dependent on the dose applied (infinite versus finite dose). The maximnm finx (derived from the linear part of the absorption versns time cnrve) may be nsed to semiqnantitatively compare absorption between species, componnds or vehicles, based on finite dose experiments. In this case, attention shonld be paid to the differences in maximnm flux values at relevant exposure levels. For example, if at 200 p.g/cm the flux through rat skin is ten times higher compared to hnman skin, bnt flnxes are comparable at the more relevant dose level of 20 p.g/cm, there shonld be no correction for differences in skin permeability in health risk assessment. 

This means that extrapolating from a high dose of a chemical to a very low dose to determine a threshold and calculate risk may not be appropriate (see Chapter 12), especially if a linear model is used which implies there is no safe dose (for example, for a carcinogen). If true this has profound implications for risk assessment, suggesting that we may sometimes have been more cautious than necessary Thus attempting to reduce exposure levels for chemicals excessively may be unnecessary and, worse, a waste of effort and money. 

For chemicals such as food additives, food contaminants, and industrial chemicals the threshold, that is the dose at which toxic effects become apparent, is determined from the dose-response graph and used in the risk assessment process. The threshold value is used, together with safety factors, to determine the acceptable daily intake (ADI) of a food additive, or the tolerable daily intake (TDI) of a food contaminant, or the threshold limit value (TLV in the USA, or maximum exposure limit (MEL) in the UK), for an industrial chemical (see box for calculation). For a drug, information about the dose in animals below which there are no adverse effects will be necessary before human volunteers can be exposed in clinical trials. More extensive safety evaluation is carried out for drugs than for... 

Risks associated with the ingestion of contaminated dust have been estimated by Butte and Heinzow [85] using the chronic oral reference dose available from the US-EPA Integrated Risk Assessment Information Service [156]. With a focus on small children (age 1-6 years, mean body weight 16 kg) and a daily intake of 100 mg house dust [24,83] tentative benchmarks for house dust were calculated. The assessment indicated for chlorpyrifos, DDT and diazinon that the tolerable exposure concentration in house dust might be exceeded in some samples and chlorpyrifos especially can be considered a potential hazard to householders. 

---