Soil Pollution Risks

               Are Soil Pollution Risks Established by Governments the Same as Actual Risks?

                                                                                                                                                                                         L. Reijnders
                                                                                                                                                                                         IBED, University of Amsterdam,
                                                                                                                                                                                         The Netherlands.


Though soil pollution policies in North America and the European Union increasingly use risk-based standards, the construction and application
of such standards are often deficient in taking account of actual risks. Standards refer to total concentrations of substances and not to the
biologically available amount. A number of countries neglect ‘background’ exposure and assumptions regarding routes of exposure to soil  
pollution can be very different and at variance with empirical data. Recent dose-effect studies are neglected in a number of cases. The application
of standards does not take account of the overall risk of soil pollution, but rather leads to the decision whether or not there is violation of at least
one standard for a specified (group of) substance(s). Standards for soil pollutants are often based on the assumption that only effect addition can
occur, whereas dose addition, antagonism and synergism and indirect effects may in fact apply. Several remedies for current shortcomings are

Keywords: soil pollution, risk

1. Introduction

There is an increasing use of risk-oriented policies to deal with the local effects of soil pollution. The risks that such policies deal with are: human
health risks and can also include ecotoxicological risks. These risks are expressed in terms of negative effects and chances between 0 and 1 that
such negative effects will occur. Examples of areas where risk-oriented policies are applied to soil pollution include the United States of America
[1,2], Canada [3] and countries in the European Union [4-10]. Historically, these risk oriented policies have followed the abandonment of policies
aimed at restoring soils to their original ‘clean’ state [e.g. 5].

Risk-based criteria or standards, developed in the framework of risk oriented policies, are applied to risks estimated with deterministic
methodologies, following the steps of hazard characterization, appraisal of exposure and risk characterization, while using exposure-risk relations
established beforehand. Risk-based criteria have been applied to decisions about soil remediation in the form of soil clean-up standards [2,9,10],
to the use of soils for specific purposes [11,12] and in the United States also to sediment management [13]. The risk-oriented policies considered
here [1-13], assume that background exposure to pollutants carries no risk and that a specified level of soil pollution carries a maximum tolerable
or maximum acceptable risk for organisms living locally. The latter is the main basis for standard setting.

In part, risk-oriented soil pollution legislation includes policy goals that are qualitative [10]. For instance, the primary UK legislation on
contaminated soil defines land as contaminated in need of risk management ‘if significant harm is being caused or there is a significant
possibility of such harm being caused’ [10]. Mostly, however policies have resulted in specific quantitative values for maximum tolerable or
acceptable soil pollution. The analysis of such values used in different industrialized countries has shown that there are very large differences,
roughly up to a factor 104 [14,15]. According to Provoost et al. [14,15], these differences to a large extent originate in different political choices (e.g.
including or excluding ecotoxicity) and in different assumptions as to the modeling of exposure to soil pollutants, including site related factors,
such as soil type and building constructions [14,15].

This paper deals with the way in which governments establish risk and with the question whether what governments establish reflects actual risk.

A part of the large divergence in standards for maximum tolerable or acceptable risk originate in between-country differences in political choices
(about what is tolerable or acceptable), soil types and building constructions. In principle, these differences do not lead to divergence between risk
as established by governments and actual risk.

However other factors may lead to such divergence. An example thereof is: the parameters used for modeling exposure to soil pollutants given
specific exposure routes. The differences in these parameters will not be considered here, as this subject has been extensively dealt with by
Provoost et al. [14,15]. Here the focus will be on another source for the divergence between risks as established by governments and actual risks.

In dealing with the risks of soil pollution, it would seem obvious to address the overall risk of the soil pollutants present to the extent that they are
biologically available for specific organisms, against the background of exposure from other sources. However actual practice is often different.

Firstly, in section 2 shortcomings in current governmental risk estimates will be discussed regarding exposure to a single pollutant focusing on
the presence of soil quality standards, exposures to all sources of the pollutant at hand, the account taken of recent dose-effect studies and
biological availability.

Secondly, in section 3 the matter of combination effects of soil pollutants will be considered.

In section 4, several remedies for current shortcomings will be proposed.

Section 5 summarizes the conclusions of this paper.

2. Risks related to one soil pollutant

In practice, there are several matters which are at variance with the proper establishment of actual risk related to one soil pollutant. These are: the
absence of standards for pollutants, neglect of background exposure, neglect of routes of exposure to soil pollution, neglect of available dose-
effect studies and neglect of biological availability. These will now be discussed in more detail.

2.1. Absence of quality standards

When data regarding soil pollutants are available, they should be compared with quality standards reflecting maximum tolerable risk of exposure.
However, such standards are not always in place. For instance, of the volatile organic carbon compounds detected in groundwater samples by the
US Geological Service, 21 were unregulated- with no standards in place [16]. Similarly Patterson et al. [17] found a variety of brominated ethenes
in Australian groundwater, all lacking standards. In the Netherlands there are no standards for heterocyclic polycyclic aromatics, though these can
be a substantial contributor to soil and sediment pollution risk [18].

When there is no standard, the government policies considered here [1-13] tend to neglect the soil pollutant involved. This may lead to a
divergence between risk as established by governments and actual risks.

2.2. Neglect of background exposure

For a proper estimate of soil pollution related risks, exposure to specific soil pollutants should be evaluated in combination with exposure to the
same substance that is not related to local soil contamination. Several countries, such as Canada, Germany, Spain and Belgium, do indeed
establish soil clean-up standards while considering background dietary and inhalatory exposure but others, e.g. Sweden, Norway and the
Netherlands, do not [9,14]. Neglecting background exposure or specific types of background exposure may have implications for risk estimates,
as will be elaborated in paragraph 2.4.

2.3 Neglect of routes of exposure to soil pollution

In evaluating exposure to soil pollutants, assumptions regarding exposure routes are important. In this respect difference between countries may
be noted. Soil clean-up standards for lead of Norway and Sweden differ in part because in Sweden the dominant exposure route is assumed to
be by drinking water and in Norway it is thought to be by drinking water and ingestion of soil [14]. These differences cannot be explained by
differences in habits between Swedes and Norwegians and at least one of these assumptions must be at variance with actual exposure patterns.

Inhalation of household dust and soil particles is not always taken into account in governmental decision making about risks of soil pollution. For
instance, in the Netherlands inhalation of soil particles has been neglected as an exposure route, but in e.g. Spain it is not [9]. Neglect of
inhalation would seem at variance with existing studies. Nawrot et al. [19] have studied the effects of cadmium pollution in soil (around former
thermal zinc plants) and found a significant increase in lung cancer risk correlated with cadmium exposure. They plausibly explain this in terms of
exposure of lung tissue to cadmium present in inhaled soil and household dust particles. Household dust particles have also been found to be
important in the exposure of children to pesticides in agricultural settings [20]. Studies of Laidlaw et al. [21,22] suggest that inhalation of soil
particles containing lead may be important in determining the body burden of lead in American cities. Increasing urban body burdens of lead have
been shown to be correlated with neurodevelopmental toxicity [23].

2.4 Neglect of available dose-effect studies

2.4.1 Dose-effect studies relevant to humans

As pointed out in the Introduction, it is assumed in soil pollution policy that background exposure represents no risk. This neglects a number of
epidemiological studies that have been done regarding background exposure, and more generally reflects deficiencies in the use of available
dose-effect studies in determining actual risk.

Akesson et al. [24] have analyzed the effects of low environmental cadmium exposure in an epidemiological study of Swedish women in the Lund
area , being 53-64 years of age, excluding women from areas with soils heavily polluted by cadmium. Akesson et al. [24] found associations
between the internal dose of cadmium and tubular and glomerular kidney effects, which may represent early signs of adverse effects. Women with
diabetes seemed to be at increased risk of experiencing such early signs. In view of these data it seems plausible that at a background exposure
that is common in Sweden, old women in the general population may be at risk for adverse cadmium effects [34] and that even a modest increase
in cadmium exposure due to polluted soil may lead to added risk. However, when establishing soil clean-up standards in Sweden this
background exposure has been neglected [14]. Nawrot et al. [25] have studied the relation between mortality and cadmium body burden in
Belgium. They obtained evidence that total mortality and non-cardiovascular mortality may be elevated at cadmium body burdens which can be
found among the population not living on soils that are currently considered to be a health risk.

Similarly there are now strong indications that the negative effects of lead on the neurophysiological and sexual development may well be found at
the level of background exposure common in Western European and US cities [26-31], though soil pollution policy, at least in European counties,
assumes that such background exposure is safe [14].
A study of women from the general population in France found that calcium pump activity in their newborns negatively correlated with hair mercury
levels, and this may well be responsible for subtle neurobehavioral deficits of children in the general population that also correlate with mercury
levels [32,33]. There is also evidence for negative cardiovascular effects of methylmercury in the adult general population of industrial countries
[33]. Again this is not reflected in risk estimates within the framework of soil pollution policies in European countries [14]. A study regarding
exposure of the general population in the Netherlands to PCBs and halogenated dioxins and benzofurans suggests reduced lung function,
retarded brain development and a negative haematological impact [34].

2.4.2 Ecotoxicological risks

Maximum acceptable or maximum tolerable ecotoxicological risks are usually derived from a limited number of studies concerning single species
under laboratory conditions. Laboratory conditions may be very different from actual conditions in the field, and thus findings in the field are often at
variance with laboratory studies [18,35]. In field studies it has been found that several factors which tend to be neglected in laboratory studies may
strongly impact toxic effects of soil pollutants. These include among others: density and adaptability of populations of affected organisms, the
presence of other environmental stress factors and the presence or absence of specific landscape elements such as buffer strips [18,35].

2.5 Biological availability

Biologically available pollutants determine risk [3]. Biological availability may vary strongly for different types of organisms [36]. Biological
availability of a compound in a specific soil is also dependent on physical, chemical and biological and spatial factors [3,35]. Examples of such
factors are pH, the amount and nature of organic and mineral compounds also present and the presence of organisms that can mobilize soil
pollutants [35, 37-39]. In practice, biological availability may be much at variance with total concentrations [40].

However standards reflecting potentially unacceptable risk tend to refer to total concentrations. In the case of elements (such as heavy metals),
moreover standards often do not refer to specific compounds though it may well be that the nature of the compound is a determinant of biological
availability. To the limited extent that biological availability is considered in site specific follow up studies, in-vitro tests are used that may give rise
to estimates that are at variance with in-vivo biological availability [3].

3. Combination effects

3.1 Limited accounting of combination effects

As to the overall risk of soil pollutants, the US Comprehensive Environmental Response, Compensation and Liability Act (1990) stipulates that
cumulative effects of the combination of substances present in soils should be considered. However actual standard setting practice has largely
focused on criteria relating to one element or compound. In some cases there are criteria for groups of compounds [4]. Such criteria limit the
amount (in g/kg soil) of groups of compounds but often do not address the possibility that the risk per unit of weight may be different for different
compounds. An exception to this are criteria for the presence of halogenated dioxins and benzofurans and planar biphenyls. The establishment of
risk in case of exposure to these compounds uses addition on the basis of equivalent toxicity [4]. This is a major improvement, though it has been
pointed out that this approach may still underestimate the risk of neurodevelopmental effects [41].

In the Netherlands, according to Van Zorge [4], the consideration of combination effects has led to introducing an additional safety factor. Indeed,
apart from a level of maximum tolerable or acceptable risk, a level of negligible risk has been defined which reflects the use of a safety factor for
combination effects. However in actual Dutch policy decisions, such as the decision to clean up soils, the legal basis usually necessitates a focus
on exceeding maximum tolerable risk levels for individual substances, when deciding whether soil pollution should be considered for remediation
[4]. Exceeding negligible risk standards is in the Netherlands not a basis for government intervention [42].

3.2 Importance of combination effects

Combination effects may be important in two respects. Firstly, coexisting soil contaminants may impact each others’ biological availability [43].
Secondly, exposure to a combination of pollutants may be associated with antagonistic, synergistic and additive interactions of these pollutants,
impacting their effect on organisms [44-47]. Some risks of pollutant mixtures can be predicted on the basis of existing knowledge. For instance
there is a fair chance that there will be dose additivity when effects are receptor mediated [48]. Also in case of narcotic effects, joint-mixture
ecotoxicological effects may be predicted [48]. If responses are dissimilar, response addition may be used [49]. A methodology to deal with the
ecotoxicity of mixtures giving rise to both dose-additive and response-additive effects has been proposed [48]. This two step model evaluates
mixture toxicity for the same mode of action with concentration additivity and the toxicity for different modes of action with response additivity. For
determining the severity of ecotoxicological effects in case of heavily polluted soils (in which legal maximum tolerable levels for one or more
substances are exceeded), a systematic approach to combination effects based on a mixture of concentration addition and response addition has
been proposed [41].

However, it should be noted that responses of ecosystems in the field may well diverge from estimates made on the basis of additivity, e.g. due to
mobility of organisms, density dependent effects or indirect effects such as secondary poisoning and trophic effects following from lowered
abundance of food sources [35,50]. Also synergism may occur [45,47]. Including such factors is not easy. More sophisticated modeling may still
lead to results that are widely of the mark in the real world [36].

4. Remedies for shortcomings

From sections 2 and 3 one may conclude that risks established within the framework of current risk oriented soil pollution policies tend to be at
variance with actual risks. This undermines the credibility of such policies and may be argued to be a good reason for abandoning risk oriented
policies or for correcting the shortcomings outlined in sections 2 and 3. Possible remedies for these shortcomings will be outlined here. It should
be realized that much scientific work will be required in order to make these remedies operational.

Remedies would seem possible which would allow for a significant improvement in risk estimates. Unregulated substances can get standards.
Standards may be regularly updated on the basis of new dose-effect studies. Risk estimates can include both background exposure and all
exposure routes for local soil pollution. Estimates of biological availability can be integrated in risk assessments and improved by better testing of
bioavailability or by in-vivo monitoring [3,51].

The deficiencies in taking account of combination effects in ecotoxicity, discussed in section 3, may be addressed by directly testing of ecotoxicity,
when the focus is on ecosystem functioning [52,53]. However it should be noted that small effects on the functioning of ecosystems may have
large effects over time [35]. This necessitates large numbers of replicate tests that may well be beyond routine practice [35].

In determining combination effects on human health, direct testing on humans is an ‘unethical option’. However biomarker-based monitoring of
some aspects of soil pollution relevant to humans may be an option. For instance Roos et al. [54] have applied a biomarker based test to original
and remediated soils that were contaminated by a variety of polycyclic aromatic hydrocarbons (PAH). They tested the expression profile of
cytochromes P 450 [54]. Xiao et al. [55] have measured genotoxic risk of soil contamination using an in-vitro assay with Salmonella. Though the
relation between such biomarker-based data gathered and the in-vivo risks awaits further elucidation, the application of tests based on
biomarkers for soil pollution is an interesting option in dealing with combination effects on humans.

Also, estimates of risk may be derived from biomarkers which may be monitored in people exposed to soil pollution. Such biomarkers have
emerged from epidemiological studies considering the combined effect of substances. An illustration thereof is the study by Lee et al. [56] which
found a graded association of the concentration of blood lead and urinary cadmium concentrations with oxidative stress related markers in the US
population. This suggests that oxidative stress may be useful as a biomarker for combination effects. It has furthermore been proposed to
evaluate effects of exposure to nitroarenes by measuring haemoglobin adducts [57], and of mixtures of volatile organochlorines by measuring
glutathione conjugative metabolites [58]. Bioassays based on aryl hydrocarbon (Ah) receptor mediated mechanisms have been proposed which
will allow a better alternative to the measurement of polyhalogenated aromatic hydrocarbons [41].

Another option is to estimate risks to human health by taking into account cumulative combination effects in line with established cause-effect
relations and research into the effects of actual combinations. It has been shown that risks of compounds with the same targets and the same
modes of action may be estimated on the basis of concentration addition, while including toxicity equivalence factors for the compounds involved
This has been shown to apply to receptor-mediated-and reactive mechanisms of toxicity, provided that no chemical reactions occur between the
components of the mixture considered [5,60]. Currently this approach is applied to halogenated dioxins, benzofurans and planar polybiphenyls,
though non-linear interactions are not completely absent in this category of compounds [61], and neurodevelopmental effects may be
underestimated, as pointed out before [41]. Extension of this approach is possible to e.g. polycyclic aromatics, including heterocyclic polycyclic
aromatics [18,62] organophosphates that inhibit the enzyme cholinesterase [44,63], compounds that bind to estrogen receptors [64-66],
carcinogens [67], a variety of petroleum products [68] and compounds that inhibit the MXR efflux pump [69].

5. Conclusion

The construction and application of risk-based standards for soil pollution are often deficient in taking account of actual risks. Standards refer to
total concentrations of substances and not to the biologically available amount. A number of countries neglect ‘background’ exposure and
assumptions regarding routes of exposure to soil pollution can be very different. Recent dose-effect studies are neglected in a number of cases.
The application of standards does not take account of the overall risk of soil pollution, but rather leads to the decision whether or not there is
violation of at least one standard for a specified (group of) substance(s). Excepting halogenated dioxins, benzofurans and planar biphenyls,
criteria for soil pollutants are based on the assumption that only effect addition can occur, whereas dose addition, antagonism and synergism and
indirect effects may in fact apply.

Several remedies to these shortcomings have been proposed. Regarding ecotoxicity direct testing would allow for a major improvement in risk
estimates. As to human health risks: including biological availability in risk estimates, more use of up to date knowledge about exposure routes,
dose-effect relations and combination effects, and biomonitoring of effects are options for improvement.


The comments of two anonymous reviewers are gratefully acknowledged.


1. Okrent D. 1999. On intergenerational equity and its clash with intragenerational equity and on the need for policies to guide the regulation of  
  disposal of wastes and other activities posing very long time risks. Risk Analysis 19: 877-901.

2. Belluck, D.A., Benjamin, S.L., Baveye, P., Sampson, J., Johnson, B. 2003. Widespread arsenic contamination of soils in residential areas and
  public spaces: an emerging regulatory or medical crisis? International Journal of Toxicology 22: 109-128.

3. Richardson, G.M., Bright, D.A., Dodd, M. 2006. Do current standards of practice in Canada measure what is relevant to human exposure at
  contaminated sites? II: oral bioaccessibility of contaminants in soil. Human and Ecological Risk Assessment 12: 606-618.

4. Van Zorge, J. A. 1996. Exposure to mixtures of chemical substances: is there a need for regulations? Food and Chemical Toxicology 34, 1033-

5. Swartjes, F.A. 1999. Risk-based assessment of soil and groundwater quality in the Netherlands: standards and remediation urgency. Risk
  Analysis 19:1235-1248.

6. Eikelboom, R.T., Ruwiel E., Goumans, J.J.M. 2001. The building materials decree: an example of a Dutch regulation based on the potential
  impact of materials on the environment. Waste Management 21: 295-302.

7. Crane, M. and Giddings, J.M. 2004. ‘Ecologically acceptable concentrations’ when assessing the environmental risks of pesticides under
  European Directive 91 414/EEC. Human and Ecological Risk Assessment 10: 733-747.

8. Nathanail, P., McCaffrey, C., Earl, N., Forster, N.D., Gillett, A.G., Ogden, R. 2005. A deterministic method for deriving site-specific human health
  assessment criteria for contaminants in soil. Human and Ecological Risk Assessment 11: 389-410.

9. Tarazona, J.V., Fernandez, M.D., Vega, M.M. 2005. Regulation of contaminated soils in Spain. Journal of Soil and Sediments 5:121-124.

10. Evans, J., Wood, G., Miller, A. 2006. The risk assessment-policy gap: An example from the UK contaminated land regime. Environment
   International 32: 1066-1071.

11. Huinink, J.T.M. 1998. Soil quality requirements of use in urban environments. Soil and Tillage Research 47: 157-162.

12. Urzelai, A., Vega, M., Angulo, E. 2000. Deriving ecological risk-based soil quality values in the Basque Country. Science of the Total
  Environment 247: 279-284.

13. Apitz, S.E. 2008. Is risk based, sustainable sediment management consistent with European policy. Journal of Soils and Sediments 8: 461-

14. Provoost, J., Cornelis, C., Swartjes, F. 2006. Comparison of soil clean-up standards fort race elements between countries: why do they differ?
  Journal of Soil and Sediments 6: 173-181.

15. Provoost, J., Reijnders, L., Swartjes, F., Bronders, J., Carlon, C., D’Allessandro, M., Cornelis, C. 2008. Parameters causing variation between
  soil screening values and the effect of harmonization. Journal of Soils and Sediments 8: 298-311.

16. Toccalino, P.L. and Norman, J.E. 2006. Health-based screening levels to evaluate US Geological Survey groundwater quality data. Risk
  Analysis 26: 1339-1348.

17. Patterson, M.M., Cohen, E., Prommer, H., Thomas, D.G., Rhodes, S., McKinley, A.I. 2007. Origin of mixed brominated ethane groundwater
  plume: contaminant degradation pathways and reactions. Environmental Science & Technology 41: 1352-138.

18. Leon Paumen, M. 2008. Invertebrate life cycle responses to PAC exposure. PhD thesis. Amsterdam: University of Amsterdam.

19. Nawrot, T., Plusquin, M., Hogervorst, J., Roels, H.A., Celis, H., Thijs, L., Vangronsveld, J., Van Hecke, E., Staessen, J. 2006. Environmental
  exposure to cadmium and risk of cancer: a prospective population-based study. The Lancet Oncology 7: 119-126.

20. Simcox, N.J., Fenske, R.A., Wolz, S.A., Lee, I.W., Kalman, D.A. 1995. Pesticides in household dust and soil: exposure pathways for children of
  agricultural families. Environmental Health Perspectives 103: 1126-1134.

21. Laidlaw, M.A.S., Mielke, H.W., Filippelli, G.M., Johnson, D.L., Gonzales, C.R. 2005. Seasonality and children’s blood lead levels: developing a
  predictive model using climate variables and blood data for Indianapolis, Indiana, Syracuse, New York and New Orleans Louisiana (USA).
  Environmental Health Perspectives 113: 793-800.

22. Laidlaw, M.A.S., Mielke, H.W., Filippelli, G.M., Johnson, D.L. 2006. Blood lead in children: Laidlaw et al. respond. Environmental Health
  Perspectives 114: A 19.

23. Rothenberg, S.J. and Rothenberg, J.C. 2005. Testing the dose-response specification in epidemiology: public health and policy
  consequences for lead. Environmental Health Perspectives 113: 1190-1195.

24. Akesson, A., Lundh, T., Vahter, M., Bjellerup, P., Lidfeldt, J., Nerbrand, C., Samisoe, G., Strömberg, U., Skerfving, S. 2005. Tubular and
  glomerular kidney effects in Swedish women with low environmental cadmium exposure. Environmental Health Perspectives 113: 1627-1631.

25. Nawrot, T.S., van Hecke, E., Thijs, L., Richart, T., Kuznetsova, T., Jin, Y., Vangronsveld, J., Roels, H.A., Staessen, J.A. 2008. Cadmium-related
  mortality and long-term secular trends in the cadmium body burden of an environmentally exposed population. Environmental Health
  Perspectives 116: 1620-1628.

26. Lamphear, R.P., Dietrich, K., Auinger, P., Cox, C. 2000. Cognitive defects associated with blood lead concentrations, 10 microgram/dl in United
  States children and adolescents. Public Health Reports 115: 521-529.

27. Wang, C.L., Chuang, H.J., Ho, C.K., Vang, C.Y. 2002. Relationship between blood lead concentrations and learning achievement among
  primary schoolchildren in Taiwan. Environmental Research A 89: 167-181.

28. Wu, T., Buck, G.M., Mendola, P. 2003. Blood levels and sexual maturation in US girls. Environmental Health Perspectives 111: 737-741.

29. Von Storch, H., Costa-Cabral, M., Hagner, C., Freser, F., Pacyna, Y., Pacyna, E., Kolb, S. 2003. Four decades of gasoline lead emissions and
  control policies in Europe: a retroactive assessment. Science of the Total Environment 311: 151-176.

30. Mielke, H.W., Gonzales, C.R., Powell, E., Jartun, M., Mielke Jr., P.W.2007. Nonlinear association between soil lead and blood lead of children in
  Metropolitan New Orleans, Louisiana: 2000-2005. Science of the Total Environment 388: 43-53.

31. Miranda, M.L., Kim, D., Overstreet-Galeano, M.A., Paul, C.J., Hull, A.P., Morgan, S.P. 2007. The relationship between early childhood blood lead
  levels and performance on end-of-grade tests. Environmental Health Perspectives 115: 1242-1247.

32. Huel, G., Sahuquillo, J., Debotte, G., Oury, J., Takser, L. 2008. Hair mercury negatively correlates with calcium pump activity in human term
  newborns and their mothers at delivery. Environmental Health Perspectives 116: 263-267.

33. Mergler, D., Anderson, H.A., Chan, L.H.M., Mahaffey, K.R., Murray, M., Sakamoto, M., Stern, H.A. 2007. Methylmercury exposure and health
  effects in humans: a worldwide concern. Ambio 36: 3-11.

34. ten Tusscher, G.W. 2002. Later childhood effects of perinatal exposure to background levels of dioxins in the Netherlands. PhD thesis.
  Amsterdam: University of Amsterdam.

35. Filser, J., Koehler, H., Ruf, A., Rombke, J., Prinzing, A., Schaefer, M. 2008. Ecological theory meets soil ecotoxicology: challenge and chance.
  Basic and Applied Ecology 9: 346-355.

36. Van Gestel, C.A.M. 2008. Physico-chemical; and biological parameters determine metal bioavailability in soils. Science of the Total
  Environment 406: 385-395.

37. Sokolik, G.A., Ovsiannikova, S.V., Ivanova, T.G., Leinova, S.L. 2004. Soil-plant transfer of plutonium and americium in contaminated regions of
  Belarus after the Chernobyl catastrophe.  Environment International 30: 939-947.

38. Zhao, C., Ren, J., Xue, C., Lin, E., 2005. Study on the relationship between soil selenium and plant selenium uptake. Plant and Soil 277: 197-

39. Liste, H., Prutz, I. 2006. Plant performance, dioxygenase-expressing rhizosphere bacteria, and biodegradation of weathered hydrocarbons in
  contaminated soil. Chemosphere 62: 1411-1420.

40. van der Geest, H, Leon Paumen, M. 2008. Dynamics of metal availability and toxicity in historically polluted floodplain sediments. Science of
  the Total Environment 406: 419-425.

41. Brouwer, A., Ahlborg, U.G., van den Berg, M., Birnbaum, L.S. 1995. Functional aspects of developmental toxicity of polyhalogenated
  hydrocarbons in experimental animals and human infants. European Journal of Pharmacology 293: 1-40.

42. Rutgers, M., Tuinstra, J., Spijker, J., Mesman, M., Wintersen, A., Posthuma, L. 2008. Risico’s voor het ecosysteem in stap twee van het
  saneringscriterium [Ecosystem risks in step 2 of the remediation criterion]. Bilthoven (the Netherlands): RIVM.

43. Li, J., Zhou, B., Liu, Y., Yang, Q., Cai, W. 2008. Influence of coexisting contaminants on bisphenol A sorption and desorption in soil. Journal of
  Hazardous Materials 151: 389-393.

44. Gennings, C., Carter Jr, W.H., Casey, M., Moser, V., Carchman, R., Simmons, J.E. 2004. Analysis of functional effects of five pesticides using a
  ray design. Environmental Toxicology and Pharmacology 14: 115-126.

45. Hayes, T.B, Case, P., Chui, S., Chung, D., Haeffele, C., Haston, K., Lee, M., Mai, V.P., Marjuoa, Y., Parker, J., Tsui, M. 2006. Pesticide mixtures,
  endocrine disruption and amphibian declines: are we underestimating the impact. Environmental Health Perspectives 114, Supplement 1: 40-

46. Perry, M.J., Venners, S.A., Barr, D.B., Xu, X. 2007. Environmental pyrethoid and organophosphorous insecticide exposures and sperm
  concentration. Reproductive Toxicology 23: 113-118.

47. Laetz, C.A., Baldwin, D.H., Clllier, T.K., Hebert, V., Stark, J.D., Scholz, N.L. 2009. The synergistic toxicity of pesticide mixtures: implications for
  risk assessment and the conservation of the endangered Pacific salmon. Environmental Health Perspectives 117: 348-353.

48. De Zwart, D. and Posthuma L. 2005. Complex mixture toxicity for single and multiple species: proposed methodologies. Environmental
  Toxicology and Chemistry 24: 2665-2676.

49. Altenburger, R., Walter, H., Grote, M. 2004. What contributes to the combined effect of a complex mixture? Environmental Science & Technology
  38: 6353-6362.

50. Filser, J., Wittmann, R., Lang A. 2000. Response types in Collembola towards copper in the microenvironment. Environmental Pollution 107:

51. Drexler, J.W. and Brattin, J.W. 2007. An in vitro procedure for estimation of lead relative bioavailability: with validation. Human and Ecological
  Risk Assessment 13: 383-401.

52. O’Halloran, K. 2006. Toxicological considerations of contaminants in the terrestrial environment for ecological risk assessment. Human and
  Ecological Risk Assessment 12: 74-83.

53. Römbke, J. 2006. Tools and techniques for the assessment of ecotoxicological impacts of contaminants in the terrestrial environment. Human
  and Ecological Risk Assessment 12: 84-101.

54. Roos, P.H., Tschirbs, S., Pfeifer, F., Welge, P., Hack, A., Wilhelm, M., Bolt, H.M. 2004. Risk potentials for humans of original and remediated
  PAH- contaminated soils: application of biomarkers of effect. Toxicology 205: 181-194.

55. Xiao, R., Wang, Z., Wang, C., Yu, G., Zhu, Y. 2006. Genotoxic risk identification of soil contamination at a major industrialized city in north east
  China by a combination of in vitro and in vivo bioassays. Environmental Science & Technology 40. 6170-6175.

56. Lee, D., Lim, J., Song, K., Boo, Y., Jacobs, D.R. Jr. 2006. Graded associations of blood lead and urinary cadmium concentrations with oxidative-
  stress-related markers in the US population: Results of the Third National Health and Nutrition Examination Survey. Environmental Health
  Perspectives 114: 350-354.

57. Neumann, H.G. 1996. Toxic equivalence factors, problems and limitations. Food and Chemical Toxicology 34: 1045-1051.

58. Dobrev, I.D., Andersen, M.E., Yang, R.S.H. 2002.  In silico toxicology: simulating interaction thresholds for human exposure to mixtures of
  trichloroethylene, tetrachloroethylene and 1,1,1 trichloroethane. Environmental Health Perspectives 110: 1031-1039.

59. Silva, E., Rajapakse, N., Kortenkamp, A. 2002. Something from ‘nothing’-eight weak estrogenic chemicals combined at concentrations below
  NOECs produce significant mixture effects. Environmental Science & Technology 36: 1751-1756.

60. Richter, M. and Escher, B. 2005. Mixture toxicity of reactive chemicals by using two bacterial growth assays as indicators of protein and DNA
  damage. Environmental Science & Technology 39: 8753-8761.

61. Wölfe, D. 1997. Interactions between 2,3,7,8 TCDD and PCs as tumor promoters: limitations of TEFs. Teratogenesis, Carcinogenesis and
  Mutagenesis 17: 217-224.

62. Reeves, W.R., Barhoumi, R., Burghardt, R.C., Lemke, S.J., Mayura, K., McDonald, T.J., Phillips, T.D., Donelly, K.C.2001. Evaluation of methods
  for predicting the toxicity of polyaromatic hydrocarbon mixtures. Environmental Science & Technology 35: 1630-1636.

63. El Masri, H., Mumtaz, M.M., Yushak, M.L. 2004. Application of physiologically-based pharmacokinetic modeling to investigate the toxicological;
  interaction between chlorpyrifos and parathion in rat. Environmental Toxicology and Pharmacology 16: 57-71.

64. Kortenkamp, A. and Altenburger, R. 1999. Approaches to assessing combination effects of estrogenic environmental pollutants. Science of the
  Total Environment 333: 131-140.

65. Sumpter, J.P. and Johnson, A.C. 2005. Lessons from endocrine disruption and their application to other issues concerning trace organics in
  the aquatic environment. Environmental Science & Technology 39: 4321-4332.

66. Brian, J.V, Harris, C.A., Scholze, M., Backhaus, T., Booy, P., Lamoree, M., Pojana, G., Jonkers, N., Runnalis, T., Bonfä, A., Marcomini, A.,
  Sumpter, J. 2005. Accurate prediction of the response of freshwater fish to a mixture of estrogenic chemical. Environmental Health Perspectives
  113: 721-728.

67. Marchant, C.A. 1996. Prediction of rodent carcinogenicity using the DEREK system for 30 chemicals currently being tested by the National
  Toxicology Program. Environmental Health Perspectives 104: 1065-1073.

68. Verhaar, H.J.M., Morroni, J.R., Reardon, K.F., Hays, S,M., Gaver, D.P., Carpenter, R.L., Yang, R.S.H. 1997. A proposed approach to study the
  toxicology of complex mixtures of petroleum products. Environmental Health Perspectives 195 (Supplement 1): 179-195.

69. Stevenson, C.N., Macmanus-Spencer, L.A., Luckenbach, T., Luthy, R.C., Epel, D. 2006. New perspectives on perfluorochemical ecotoxicology:
  inhibition and induction of an efflux transporter in the marine mussel Mytilus californianus. Environmental Science & Technology 40: 5580-5585.


Author’s Address:

              L. Reijnders
              University of Amsterdam,
              Nieuwe Achtergracht 166
              1018 WV Amsterdam
              The Netherlands
              Tel.: + 31-20-5256206
              Fax: + 31-20- 5257431