From toxicity tests

In the case of surface water, the LOQ must not exceed a concentration which has an impact on nontarget organisms deemed to be unacceptable according to the requirements of Annex VI. At present, no harmonized limits for surface water exist. Therefore, provisions in Annex VI of Directive 91/414/EEC will be used to calculate guidance limits for analytical methods for surface water. In SANCO/825/00 the limits given in Table 6 are established [the relevant concentrations (the lowest will always be taken into consideration) depend on the species as indicated and can be taken from toxicity tests]. 

As protozoa and nematodes live in pore water in the soil, most of the methods are adapted from toxicity tests designed for aquatic samples. Among the protozoa the tests with cihates Tetrahymena pyriformis, Tetrahymena thermophiia, Colpoda cucullus, Colpoda inflata, Colpoda steinii, Paramecium caudatum, and Paramecium aurelia have been developed [ 102,112-117]. It is the opinion of some authors that the sensitivity of infusorians is higher than that of microorganisms [115,116]. 

Appraising the toxic potential of biologically available contaminants in sediment should include three compartments the whole sediment (with standardized direct contact assays when these are available), the porewater, and the elutriate (aqueous extract). Additional hazard information can also be obtained from toxicity testing conducted on organic extracts using methanol or acetone. 

Additional information is also requested from toxicity testing in birds (Octa and Deca), although it is also questioned whether such testing would actually eliminate uncertainties considering the inherent difficulties in achieving sufficiently high exposures in the available tests. 

Apart from toxicity tests involving the use of live animals, there are other ways of evaluating the toxic properties of chemicals that stem from an understanding of their mode of action. For example, the Ames test aids in the identification of substances that act as carcinogens or mutagens in mammals. Also, the study of the relationship between structure and toxicity (i.e., quantitative structure-activity relationships or QSARs) can provide support for the identification of toxic substances. These approaches will become more viable as molecular mechanisms of toxicity become better known, and they can lead to an understanding of the molecular characteristics that cause a chemical to interact adversely with cellular macromolecules. 

Distinguishing between toxicity tests and bioassays is not only of theoretical relevance. From toxicity tests, a clear dose-response graph could be drawn and effect concentrations (ECio, EC50 and EC90) as well as LOEC and NOEC could be calculated. The evaluation of bioassays should preferably be limited to the expression of inhibition values of the original sample and the calculation of dilution rates (G-values) at which a specific rate of inhibition, most commonly 10%, is reached. For example, the G-values derived from a wastewater treatment process should provide an estimation basis for the treatment plant operator regarding the possible inhibition effect of a certain wastewater stream (for methods see DIN-standards listed in Table 4.2). 

Raw toxico-kinetic values, those derived directly from toxicity test data, were converted to the more familiar bioconcentration-based kinetic values by using the relationship between the internal toxicant concentrations associated with the completion of toxicity and bioconcentration tests. Simple kinetics information is based on the estimated internal dose. Given that the internal dose for acutely toxic narcotics appears to be constant while the internal dose for bioconcentration of those same chemicals increases with increasing log then it is hypothesized that the difference in the internal concentration endpoints, which is explained by Kg, can be used to convert kinetic data between the two test procedures. 

The cumulative dose-effect curve (Figure 3.1b) and the frequency-response histogram (Figure 3.1a) offer two distinct approaches to analyzing data from toxicity tests. Nevertheless, the two graphs are fully interconvertible. The cumulative... 

Rejection of the null hypothesis provides prima facie statistical evidence of increased disease incidence in an at-risk population however, it does not constitute conclusive proof. One or more of several other lines of evidence are needed to buttress the statistical finding (a) evidence that the at-risk population has been exposed to the toxic chemical (b) evidence from toxicity testing in animals or from accidental human exposures that the chemical produces the disease symptoms observed in the at-risk population (c) evidence that the molecular mechanism of action is consistent with the observed disease symptoms and/or (d) evidence that the incidence of chemical disease conforms to a dose-effect relationship. 

Several important characteristics can be gleaned from toxicity testing and used to assess human health risk (Chapter 8) and ecological risk (Chapter 9). 

When a federal or state agency develops chemical safety standards, these are almost always lower than the threshold values generated from the dose-reponse relationship. The reason for this is purely pragmatic, as the results from toxicity testing are generally... 

Assessment tools used in ecotoxicology range from toxicity tests (link the response of damaged biological systems to a particular substance), chemical analyses of substances (proof of the presence of a toxicant), and field surveys (characteristics of the damaged ecosystem). These three tools help in the assessment of risk. Indirect effects of toxicants are more difficult to predict. For example, an organism s response to a toxicant depends on the presence of other organisms in the community and also on the environmental conditions. Most of the relationships between environmental conditions and toxicity have yet to be established. 

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