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Integrated Environmental Assessment and Management — Volume 6, Number 2—pp. 308–317 ß 2010 SETAC

Learned Discourses: Timely Scientific Opinions

Learned Discourses: Timely Scientific Opinions

Timely Scientific Opinions Intent. The intent of Learned Discourses is to provide a forum for open discussion. These articles reflect the professional opinions of the authors regarding scientific issues. They do not represent SETAC positions or policies. And, although they are subject to editorial review for clarity, consistency, and brevity, these articles are not peer reviewed. The Learned Discourses date from 1996 in the North America SETAC News and, when that publication was replaced by the SETAC Globe, continued there through 2005. The continued success of Learned Discourses depends on our contributors. We encourage timely submissions that will inform and stimulate discussion. We expect that many of the articles will address controversial topics, and promise to give dissenting opinions a chance to be heard. Rules. All submissions must be succinct: no longer than 1000 words, no more than 6 references, and at most one table or figure. Reference format must follow the journal requirement found on the Internet at http:// www.setacjournals.org. Topics must fall within IEAM’s sphere of interest. Submissions. All manuscripts should be sent via email as Word attachments to Peter M Chapman ([email protected]).

Learned Discourses Editor Peter M. Chapman Golder Associates Ltd. 500-4260 Still Creek Drive Burnaby, BC V5C 6C6, Canada [email protected]

SETAC’s Learned Discourses appearing in the first 7 volumes of the SETAC Globe Newsletter (1999–2005) are available to members online at http://communities.setac.net. Members can log in with last name and SETAC member number to access the Learned Discourse Archive.

BROADENING OUR VIEW ON CHEMICAL DIVERSITY IN THE BALTIC SEA Harri T. Kankaanpa ä ¨* Finnish Environment Institute, Marine Research Centre, Helsinki, Finland *[email protected] DOI: 10.1002/ieam.38

At a glance, the Baltic Sea may seem peripheral and only locally interesting. In reality, many of its characteristics are unique and deserve closer examination, not least because they contribute to the health of this vulnerable marine system. The biodiversity of the geologically young Baltic Sea is low compared to the oceans. Primary production and thus biosynthesis of chemical components is sporadically high.

In a Nutshell. . . Environmental Chemistry Broadening Our View on Chemical Diversity in the Baltic Sea, by Harri Kankaanpa¨ä. Naturally produced molecules truly are contaminants; their increasing contribution is causing a shift in the chemical balance of the Baltic Sea. Ecotoxicology Predicting Hormesis in Mixtures, by Nina Cedergreen. There are several possible mechanisms for biphasic doseresponse patterns; for now, predictions must rely on empirical data. The Nanococktail—Ecotoxicological Effects of Engineered Nanoparticles in Chemical Mixtures, by Nanna Hartmann and Anders Baun. The potential biological effects of engineered nanoparticles interacting with other environmental contaminants need to be addressed experimentally. Is There a Distinct Tropical Ecotoxicology?, by Eduardo da Silva and Amadeus Soares. Key differences include biological processes and human management/interest in contaminant and non-contaminant stressors. Epigenetics: An Emerging Field in Environmental Toxicology, by Juliette Legler. That epigenetic alterations due to chemical exposure in 1 generation may have effects on subsequent non-exposed generations has far-reaching implications for (ecological) risk assessment. Environmental Management Cradle to Cradle: Old Wine or New Spirits?, by Jose´ Potting and Carolien Kroeze. It would be a terrible loss to forfeit (again) the energy and enthusiasm for change once embodied in Life-Cycle Assessment. DOI: 10.1002/ieam.44

As a result, the internal inputs of bioactive substances are high in relation to those from external sources. Further complexity arises from an incomplete understanding of biotransformation and biodegradation, particularly in relation to the natural constituents. Contemporary techniques (e.g., DNA pyrosequencing) enable high-speed genetic analyses that provide information on biodiversity in marine systems. The situation regarding chemical diversity is quite different: no comprehensive surveys to characterize marine chemical diversity have been conducted. Ideally, techniques such as field-deployed on-line mass chromatographic or spectrometric instruments to reveal the quantity of different atoms and molecules swarming in seawater would be required. This would also generate estimates of chemical diversity over time, which could be complemented with toxicity data. So far, this has not been

Learned Discourses—Integr Environ Assess Manag 6, 2010

Figure 1. Man-made organohalogen compounds (sum of seven ICES chlorobiphenyl congeners, DDT, DDE, DDD, a- and g-HCH plus HCB; circles), and cyanobacteria-produced peptide toxins (gray squares; Karjalainen et al. 2008) in herring from eastern Gulf of Finland during 1985–2007. Organohalogen concentrations are based on frequent HELCOM monitoring; peptide toxins only on sporadic surveys.

possible, and we have had to get by with limited resolution and scattered information on marine chemistry. The situation with man-made contaminants has improved over the past decades. Purification of biota from PCBs and pesticides (Figure 1) is statistically significant. Dioxins are still at unacceptable levels in fish, but the Swedish guillemot egg data show a reduction to 1/3 of peak concentrations from the early 1970s. Emissions of organochlorine compounds from pulp mills were already cut back dramatically in the mid1990s. Oil /PAH concentrations in subsurface water have also declined to approximately 1/5 of levels in the late 1970s. Trace elements in herring occur at below European Union (EU) alert levels. Radioactive 137Cs in seawater is currently at only 3% of its peak activity value after the 1986 Chernobyl accident. In addition to these man-made contaminants, we need to examine the question of natural versus man-made. Two groups (or perhaps genera) of molecules serve to highlight the pool of natural organic compounds: natural organohalogens (reported since the 1970s, reviewed by Gribble 2004) and phytoplankton-produced phycotoxins (since the 1980s). The structures, toxicities, and occurrence of phycotoxins are well established, although their ecological consequences at the community level are not. In contrast, these characteristics are not well established for naturally produced organohalogens. Both groups may also possess beneficial antimicrobial or anticancer properties. In the highly eutrophied Baltic Sea organohalogens likely arise from primary production. The sedimentation rate of natural organohalogens in the Gulf of Finland was estimated at 20–56 t/a in 1997. Spring bloom diatoms of genera Achnanthes, Diatoma, Skeletonema, and Chaetoceros, and cyanobacteria Nodularia, Aphanizomenon sp., and Anabaena were associated with production of uncharacterized organohalogens (Kankaanpaä¨ 1997). The red alga Ceramium tenuicorne (Ku¨tz.) and especially cyanobacteria have been identified as probable sources of, for example, polybrominated dibenzo-p-dioxins (PBDDs; brominated analogues of the extremely toxic dioxins; Malmva¨rn et al. 2005). Recently, Haglund et al. (2007) reported that these biologically-derived PBDDs bioaccumulate in the Baltic Sea.

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To understand the generation of natural organohalogens, it is appropriate to look at the underlying mechanisms. The following are needed for biological production of organohalogens: 1) organic substrate; 2) hydrogen peroxide; 3) halides; 4) protons; and 5) haloperoxidases (or equivalent) catalysing the reaction. Because oxidation of chloride (Eh ¼ 1.36 V) and bromide (Eh ¼ 1.07 V) cannot be accomplished using oxygen (Eh ¼ þ0.77 V) as an electron acceptor, reduction of hydrogen peroxide (Eh ¼ þ1.77 V) takes place instead. The current quantity of chloride in the Baltic Sea water is approximately 86 billion tons. H2O2 occurs in marine waters and can be produced photochemically and biologically (oxidative stress); Herut et al. (1998) reported H2O2 at 20–80 nM in surface water in the southern Baltic Sea. Protons are always present, and haloperoxidases are available in marine organisms such as algae and bacteria. Bioactive peptides (nodularin-R and microcystin-LR) are currently the most abundant and important group of phycotoxins for the Baltic Sea, and these are produced by the cyanobacteria Nodularia spumigena and Anabaena sp. respectively. Two other phycotoxin groups, paralytic shellfish toxins (alkaloids, from Alexandrium sp.; Kremp et al. 2009) and diarrhetic shellfish toxins (polyethers and polyether macrolides, produced by Dinophysis sp.), occur at least locally in the Baltic Sea. The nonribosomal biosynthesis of peptide toxins underlines the importance of enzyme-catalyzed synthesis in the marine environment. The different contaminants have distinct temporal maxima. For instance, spring bloom (March–May) seems to generate most of the organohalogens, whereas peptide phycotoxins occupy the water column during July–August and polyether and polyether macrolide toxins in August– September. In general, the winter period is the low season for both man-made and natural organic contaminants, except for oil-derived polyaromatic compounds. Sedimentation rates for peptide phycotoxins and natural organohalogens are 1–2 orders of magnitude higher than those for PCB and DDT metabolites. We need to face the fact that naturally-produced molecules truly are contaminants, and their increasing contribution is causing a shift in chemical balance. Yet uncharacterized molecules may also contribute to the chemical crossfire experienced by organisms. Natural contaminants all integrate into total biological stress. In the harsh conditions of anoxic environments, O2 depletion, H2S, CH4, NH3, and CO2 naturally dictate biological responses. The extent to which combined external radiation from natural primordial 40K and Chernobyl-derived 137Cs affect the benthic surroundings remains unresolved. Local chemical diversity is a challenge to monitoring guidelines. It is paramount that regional differences in the composition of biologically active compounds be considered. The EU Marine Strategy Framework Directive (MSFD; Directive 2008/56/EC) urges member countries to prepare an initial assessment of the current environmental status of the waters concerned by 2012. The Directive further advises countries to determine good environmental status on the basis of descriptors, including pesticides and hydrocarbons. Accumulating evidence indicates that these chemical descriptors are inadequate for the Baltic Sea and that the MSFD for this region needs to be revised to include naturally-produced bioactive compounds. Re-examination of Baltic Sea chemical monitoring under the Helsinki Commission COMBINE program is also necessary. Furthermore, we need to continue

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exploring marine systems for their chemical diversity, arising from compounds, both known and yet uncharacterized.

REFERENCES EC. 2008. Directive 2008/56/EC of the European Parliament and of the Council of 17 June 2008 establishing a framework for community action in the field of marine environmental policy (Marine Strategy Framework Directive). Official Journal of the European Union, L164: 19–40. Gribble G. 2004. Amazing organohalogens. Am Sci 92: 342–349. Haglund P, Malmva¨rn A, Bergek S, Bignert A, Kautsky L, Nakano T, Wiberg K, Asplund L. 2007. Brominated dibenzo-p-dioxins: A new class of marine toxins? Environ Sci Technol 41: 3069–3074. Herut B, Shoham-Frider E, Kress N, Fiedler U, Angel DL. 1998. Hydrogen peroxide production rates in clean and polluted coastal marine waters of the Mediterranean. Red and Baltic Seas. Mar Pollut Bull 36: 994–1003. Kankaanpa¨ä H. 1997. Sedimentation, distribution, sources and properties of organic halogen material in the Gulf of Finland. Monographs of the Boreal Environment Research (6) [Dissertation] Univ. Helsinki. Karjalainen M, Pa¨äkko ¨nen JP, Peltonen H, Sipia¨ V, Valtonen T, Viitasalo M. 2008. Nodularin concentrations in Baltic Sea zooplankton and fish during a cyanobacterial bloom. Mar Biol 155: 483–491. Kremp A, Lindholm T, Drebler N, Erler K, Gerdts G, Eirtovaara S, Leskinen E. 2009. Bloom forming Alexandrium ostenfeldii (Dinophyceae) in shallow waters of the ˚ land Archipelago. Northern Baltic Sea. Harmful Algae 8: 318–328. A Malmva¨rn A, Eriksson U, Kautsky L, Asplund L. 2005. Cyanobacteria from the Baltic Sea: A producer of hydroxylated polybrominated diphenyl ethers (HO-PBDEs), methoxylated polybrominated diphenyl ethers (MeO-PBDEs) and polybrominated dibenzo-p-dioxin (PBDD)? Organohalogen Compd 67: 1380–1383.

Integr Environ Assess Manag 6, 2010—PM Chapman, Editor

biphasicity is often the rule rather than the exception, with metabolites in the human hormone system being a good example (Ohlsson et al. 2009). Hence, any morphological traits induced by biphasically responding metabolites, such as, for example, hormones, are likely to respond in a similar manner. Because we rarely know the mechanism behind the individual biphasic dose–response curves, predicting the size and dose range of hormesis in mixtures must, for now, be based on empirical data rather than on any predefined models. We did so in a study on the response of root length in lettuce and area-specific growth rate in duckweed exposed to mixtures of natural compounds or pesticides giving biphasic concentration–response curves of varying degrees (Belz et al. 2008). We asked 3 questions in the study: 1) Can the concentration range of the hormetic growth stimulation be predicted in binary mixtures of 2 hormetic compounds; 2) Can the size of the hormetic growth stimulation be predicted in these studies; and 3) Do we need to use biphasic dose–response models for adverse effect risk assessment purposes? Regarding the first question, all data showed that the concentration of both the maximal growth stimulation and the concentration where the response decreased relative to control values followed the same pattern as the EC50, when pictured in an isobologram. An isobolgram shows the combination of 2 chemicals giving a certain effect at various mixture ratios. Hence, if the mixture effects at EC50 levels

PREDICTING HORMESIS IN MIXTURES Nina Cedergreen* University of Copenhagen, Frederiksberg, Denmark *[email protected] DOI: 10.1002/ieam.41

‘‘Hormesis’’ is a term used to describe the phenomenon of biphasic dose–response or concentration–response relationships. These relationships are characterized by an initial increase in the measured response, followed by a decrease as a function of increasing dose or concentration of a chemical or other stressor. The common term ‘‘hormesis’’ can be misleading, because it insinuates a common mechanism behind the biphasic dose–response relationships. A closer look at the mechanisms behind different biphasic dose–response curves, however, suggests that this may rarely be the case! For most biphasic dose–response curves the physiological mechanism causing the response increase is not known. There are some general causes of biphasic dose–response curves: 1) deficient controls, where the deficiency is alleviated by the chemical; 2) parasites or diseases that are more susceptible to the test chemical than their host where the response is measured; 3) release of density-dependent pressure on a population, which will give better growth conditions for the individuals; 4) trade-off between traits, for example increased brood size at the expense of brood fitness; and 5) a dosedependent mode of action, where small doses are stimulatory, as, for example, estrogens are to mammalian breast cancer cells, while higher doses are cytotoxic. There are likely many more causes of biphasic dose–response curves that we are still not aware of. For those working with metabolomics,

Figure 1. Selected concentration–response relationships for the inhibition of root length of Lactuca sativa. The figures, adapted from Belz et al. (2008), illustrate how both the dose range and the size of the maximal response is approximately proportional to the mixture ratio of 50:50%, both for mixtures between 2 hormetic compounds (a) and between a hormetic and a nonhormetic compound (b).


could be described with the model of Concentration Addition or an antagonistic isobole model, the concentrations defining the hormetic peak could be as well. These results correspond with an example of synergy and hormesis described by Flood et al. (1985). The size of the hormetic peak was well described by a linear model depending on the mixture ratio of 2 hormetic compounds or of a hormetic and a nonhormetic compound, as illustrated in Figure 1 for the 50:50% mixture ratio. This, however, only applied for the lettuce data, whereas the size of the hormetic peak seemed more random for the duckweed data, which was probably a result of less accurately defined curves. The last question concerned the issue of using biphasic models rather than monotonous dose–response models when determining effect values used for risk assessment purposes. This is a matter of great controversy. Our results showed that, as long as risk assessment is based on an adverse effect endpoint (we used EC50 and EC20), there were no significant differences between the estimated values depending on the type of model used, neither for the compounds tested alone nor for the mixtures. This, however, does not exclude the fact that traditional risk assessments are based on adverse effects below which ‘‘no-effect’’ is expected. And this is obviously not the case for biphasic dose–response curves, where a response increase is observed below the defined No Observable Adverse Effect Dose or Concentration. The presented example is one where we do not know the physiological mechanism behind the biphasic response of the chemicals. We, therefore, do not know how general our observations are when extrapolating to other systems where biphasic dose–response patterns are observed. Results of recent and yet unpublished work, investigating mixtures of chemicals inducing biphasic dose–response curves on specific metabolites, also follow the patterns described above. However, until we know more about the underlying mechanisms behind biphasic dose–response curves, we will depend on empirical data for predicting hormesis in mixtures.

REFERENCES Belz RG, Cedergreen N, Sørensen H. 2008. Hormesis in mixtures - can it be predicted? Sci Tot Environ 404:77–87. Flood JF, Smith GE, Cherkin A. 1985. Memory enhancement - Supra-additive effect of subcutaneous cholinergic drug-combinations in mice. Psychopharmacology 86:61–67. Ohlsson A, Ullera˚s E, Oskarsson A. 2009. A biphasic effect of the fungicide prochloraz on aldosterone, but not cortisol, secretion in human adrenal H295R cells - underlying mechanisms. Toxicol Lett 91:174–180.

THE NANO COCKTAIL: ECOTOXICOLOGICAL EFFECTS OF ENGINEERED NANOPARTICLES IN CHEMICAL MIXTURES Nanna B. Hartmann* and Anders Baun Technical University of Denmark, Lyngby, Denmark *[email protected] DOI: 10.1002/ieam.39

Around 2003, the first concerns related to potential environmental exposure and related effects of engineered nanoparticles (ENPs) were raised in the scientific literature.

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Today it is evident that the widespread and diverse use of products containing ENPs, as well as the expected increase in the number of such products, will lead to emissions to the environment. Parallel to escalating industrial production, increasing focus has been placed upon their ecotoxicological effects. As is the case for traditional chemicals, ENPs will be part of a complex mixture, rather than existing as single contaminants, when present in the environment. However, only a limited number of studies, exist on the effects of ENPs in mixtures with other contaminants. Mixture interactions between ENPs and other chemical compounds may influence bioavailability and, hence, the effects, of the individual compounds. Seen from the point of view of the organism, these interactions can be beneficial as well as negative, but are in both cases difficult to predict at our present state of knowledge. Interactions between ENPs and other chemicals will take place already at the stage of particle production, for example, in the processes of synthesis and functionalization. When the particles are used in industrial applications, medical, and consumer products, they again come into contact with chemical compounds such as surfactants and preservatives. Finally, the different routes of disposal also lead to interactions with environmental contaminants present in, for instance, wastewater streams, which may include both metals and organics. Thus, unintentional mixing of ENPs with other compounds is inevitable before, during, and after their intended use. Furthermore, some ENPs are also produced with coatings or doped with metallic or nonmetallic compounds. This is done, for example, to shift the activation wavelengths of the photocatalyst TiO2 from ultraviolet to visible wavelengths (Zaleska 2008). A relatively inert and low-toxicity particle thereby becomes a carrier of potentially toxic compounds. Combined effects of traditional chemicals in a mixture do not necessarily require a direct physical interaction between the individual compounds. This is the case when 2 compounds are competing for the same binding site, thereby leading to a less than additive effect, or when 1 compound is promoting the uptake of another, leading to synergistic effects. In the case of ENPs interacting with other chemical compounds, other interaction scenarios need to be included due to the presence of a solid particle phase. This issue is particularly relevant because ENPs have certain properties that may be favorable for sorption and interactions with other contaminants: a large surface area relative to mass, and a size possibly small enough for particles to enter into organisms, organs, and cells. It is evident that there are several possible scenarios for the interactions between ENPs, environmental contaminants, and aquatic organisms. Although some of these scenarios are so far purely theoretical, others have already been documented in toxicity studies involving algae, crustaceans and fish. Below we briefly comment on the scenarios outlined in Figure 1.

Nanoparticles not interacting with environmental contaminants In some cases the physicochemical properties of the ENPs and coexisting contaminants may not allow for interactions (corresponding to Scenarios 1 and 2). We have previously investigated changes in toxicity of 4 chemical compounds (atrazine, methyl parathion, phenanthrene, and pentachlorophenol) in the presence of C60 nanoparticles (Baun et al.

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Figure 1. Different scenarios of nanoparticles (NP) acting as modifying factors on the effect of environmental contaminants (EC) toward an organism (in this case algae). Scenario 1: No interaction between NP and EC. The effect of the EC is unchanged. The NP has no effect on the organism. Scenario 2: No interaction between NP and EC. Effect—and possibly uptake—of both independently. Scenario 3: Interaction between NP, EC, and organism. Increased bioavailability of the EC as a result of increased local EC concentration (or at least the sorbed EC is bioavailable). Scenario 4: Interaction between NP, EC, and organism. Increased bioavailability as a result of membrane rupture (caused directly or indirectly by NP) leading to increased uptake and/or effect of EC. Scenario 5: Interaction between NP and EC. EC uptake facilitated by NP (trojan horse) and increased EC body burden. Scenario 6: Release of free ions from NP, leading to competition with EC (in this case a metal) for binding sites. This will result in reduced EC uptake and effect. However, the overall effect on the organism may be unchanged, reduced, or decreased depending on the specific EC and NP. Scenario 7: Interaction (sorption) of the EC onto NP. The bioavailability (and effect) of the EC is reduced. Possible inherent effect of the NP on the organism.

2008). For atrazine and methyl parathion, we observed a limited sorption to the C60 aggregates, and the addition of C60 did not affect the toxicity of these 2 compounds (Scenario 1).

Nanoparticles increasing bioavailability/toxicity of environmental contaminants In the same study, toxicity of phenanthrene to both green algae and daphnids was found to increase in the presence of C60, despite 85% sorption of phenanthrene to C60aggregates. This indicates that phenanthrene sorbed to C60 aggregates is available to the organisms (Scenario 4) (Baun et al. 2008). High sorption of PAHs to C60 has also been observed by others, indicating that the presence of C60 can influence the environmental fate and biological exposure of PAHs (e.g., Hu et al. 2008). In a recent study, we found that cadmium sorption to TiO2 nanoparticles resulted in reduced concentrations of dissolved cadmium species. However, the negative effects on algal growth rates were higher than could be explained by dissolved cadmium concentrations. The reason for this is not yet known, but it may be a result of either bioavailability of sorbed cadmium, local high concentrations (Scenario 3), the TiO2 nanoparticles affecting cell permeability leading increased cadmium bioavailability (Scenario 4), or a carrier effect of TiO2, transporting cadmium into the algal cells (Scenario 5) (Hartmann et al. 2009). The ability of ENPs to disrupt cell membranes (prerequisite

for Scenario 4) has been studied (Leroueil et al. 2007) and found to be related to the size and charge of the nanoparticles. That ENPs can act as carriers for other contaminants into aquatic organisms has been demonstrated by Zhang et al. (2007). The presence of TiO2 nanoparticles (Degussa P25) was found to increase the accumulation of cadmium in carp (Cyprinus carpio) compared with the accumulation in the presence of natural sediment particles. It was found that TiO2 nanoparticles have a stronger sorption capacity for cadmium than the natural soil particles and that cadmium accumulated in the carp together with TiO2 (Scenario 5). This increase may partly reflect the presence of cadmium sorbed onto TiO2 and the increase in cadmium concentration could be partly reversible through excretion and desorption of TiO2 from the intestine, skin, and scales, respectively. However, an increase in muscle bioconcentration of cadmium indicates actual uptake (Zhang et al. 2007).

Nanoparticles decreasing bioavailability and/or toxicity of environmental contaminants The presence of black carbon (an anthropogenic carbonrich nanoparticle originating from incomplete combustion) has been shown to reduce the toxicity of diuron (a widely used herbicide) to green algae (Knauer et al. 2007). This reduction in toxicity is explained by sorption, leading to


reduced bioavailability of the contaminant (Scenario 6). Also, the addition of 300 nm TiO2 particles was found to reduce toxicity of cadmium to algae through sorption (Hartmann et al. 2009). Finally, the release of metal ions from ENPs (e.g., from Ag and ZnO) can lead to a competition with other metal ions. The presence of dissolvable metal or metal oxide nanoparticles is, therefore, likely to reduce the bioavailability and toxicity of other metals (Scenario 7). Changes in overall toxicity will, however, depend on the toxicities of the 2 individual metal ions.

Future perspectives As we have shown in this paper, ENP interactions with other environmental contaminants need to be addressed experimentally. While low exposures usually equal low effect levels, this may not be the case for ENPs, because there is a risk that they can potentiate the effects of other contaminants occurring concomitantly. However, it is a huge challenge to address this issue experimentally, because we not only have to deal with controlling exposure of ENPs—something that has proven very difficult in the traditionally used ecotoxicity test methods—but also need to develop appropriate test designs for mixture effects. It is our experience that a proper test design for these types of studies should, as a minimum, consider: ENP aggregation behavior in the test medium; Physical interactions between 1) particles and contaminants (sorption, desorption) and 2) particles and organisms (attachment, uptake, excretion); Potential release of free metal ions; and Potential for cell membrane damage and/or increased cell permeability. We strongly encourage more researchers to move into this so far almost overlooked area of nanoecotoxicology.

REFERENCES Baun A, Sørensen S, Rasmussen R, Hartmann NB, Koch C. 2008. Toxicity and bioaccumulation of xenobiotic organic compounds in the presence of aqueous suspensions of aggregates of nano-C60. Aquat Toxicol 86:379–387. Hartmann NB, Kammer FVD, Hofmann T, Baalousha M, Ottofuelling S, Baun A. 2009. Algal testing of titanium dioxide nanoparticles: Testing considerations, inhibitory effects and modification of cadmium bioavailability. Toxicology. AvaiIable from: http://www.sciencedirect.com/science/journal/0300483X. DOI: 10.1016/j.tox.2009.08.008. Hu X, Liu J, Mayer P, Jiang G. 2008. Impacts of some environmentally relevant parameters on the sorption of polycyclic aromatic hydrocarbons to aqueous suspensions of fullerene. Environ Toxicol Chem 27:1868– 1874. Knauer K, Sobek A, Bucheli TD. 2007. Reduced toxicity of diuron to the freshwater green alga Pseudokirchneriella subcapitata in the presence of black carbon. Aquat Toxicol 83:143–148. Leroueil PR, Hong S, Mecke A, Baker JR, Orr BG, Banaszak Holl MM. 2007. Nanoparticle interaction with biological membranes: does nanotechnology present a Janus face? Accounts Chem Res 40:335–342. Zaleska A. 2008. Doped-Ti O2: A review. Recent Pat Eng 2:157–164. Zhang X, Sun H, Zhang Z, Niu Q, Chen Y, Crittenden JC. 2007. Enhanced bioaccumulation of cadmium in carp in the presence of titanium dioxide nanoparticles. Chemosphere 67:160–166.

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IS THERE A DISTINCT TROPICAL ECOTOXICOLOGY? E. M. da Silva*y and A. M. V. M. Soaresz yUniversidade Federal da Bahia, Salvador, BA, Brazil zCESAM, Universidade de Aveiro, Aveiro, Portugal *[email protected] DOI: 10.1002/ieam.45

Introduction To discuss the existence of a tropical ecotoxicology, we must recognize first that ecology plays an important role. The basic principles of toxicology, developed in the 70s remain: experimental testing, analysis of concentration and/or dose– effect relationships, and estimation of effect concentrations, such as the exposure concentration at which an x% effect is observed within a certain period (ECx). However, they have been enhanced by the addition of ecological considerations (i.e., ecotoxicology). Robinson (1978) first asked the question of whether or not there is a distinct tropical ecotoxicology. A no answer would indicate that ecological phenomena and mechanisms change as a continuum from temperate to tropical areas. A yes answer would indicate an abrupt change from temperate to tropical areas. However, verifying a yes or no answer is not a simple task. Below, we consider this question in terms of ecology, stressors, and toxicology respectively and then address environmental management issues.

Ecology Although inter- and intraspecific relationships between species and ecological mechanisms do not show major differences along latitudinal gradients, there is an increase in biodiversity from the poles to the tropics, with an increase in structural complexity. Whether increased species diversity may affect ecosystem function or lead to greater functional redundancy is still unclear (Johnson et al. 1996); however, it is even more unclear how this is affected by the presence of environmental stressors. Biological complexity is greater in the tropics than in temperate regions, but it remains an open question how this species diversity can be influenced further by contaminant and noncontaminant stressors in the tropics. To study megadiverse communities, one needs an ecology associated with an appropriate theoretical approach; however, it is still very complex to measure spatial and temporal variation in such communities, let alone biological and evolutionary traits, and establish resilience patterns.

Stressors Sources of chemical contamination and other stressors in tropical environments are similar to those in temperate ecosystems. However, their manifestation differs between tropic and temperate areas. For example, the final effect of eutrophication is the same for tropical and temperate waters: massive changes in biological diversity and ecosystem services (Smith et al. 1999). However, differences are encountered in the trajectory towards these massive changes, as tropical waters tend to have constant high temperatures with high light input. Biological processes play a determinant role regarding eutrophication, bacterial activity on the recycling of organic matter, zooplankton grazing and excretion, macrophyte interaction, and fish traits, to cite some. In temperate waters, where temperature and light vary seasonally, physical

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processes dominate over biological processes (Kilham and Kilham 1990).

Ecotoxicology Most standard ecotoxicology test species are from temperate regions and, as a rule, they have been accepted as an adequate tool to assess pollution in the tropics. However, as noted by Baird et al. (1995), tropical species may differ markedly from these standard test organisms in their response to numerous types of contaminants, particularly given that tropical environmental conditions do not match standard test conditions for temperate species. In tropical ecosystems, biological variables play an important role in determining whether contamination becomes pollution. For example, due to the higher temperatures in the tropics than in temperate regions, decomposition rates, organic matter recycling, and nutrient remobilization all proceed at accelerated rates year-round, affecting the bioavailability and thus the toxicity of chemical contaminants. For example, the production of MeHg in surface sediments is approximately 30 times lower than in macrophyte roots, and its bioavailability is probably limited, as well as the sediment water flux of MeHg (Guimara˜es et al. 2000). Clearly, this again demonstrates a higher level of biological complexity in the tropics.

Environmental management Environmental management of tropical ecosystems, in comparison to temperate ones, has a myriad of complex issues resultant from natural complexities: habitat and species diversity; species interactions; population dynamics; and community relationships. To start with, environmental risk assessments (ERA) are not a mandatory protocol in tropical countries; instead, environmental impact assessments are typically conducted. The number of scientists and environmental managers is still proportionally lower than in countries from the Northern Hemisphere. Moreover, politicians legislate for the present, typically without concern for the future. Stakeholders involved in the management of tropical ecosystems are also diverse in terms of cultures, language, scale, nature, and impact of organization, ability to access a growing knowledge base, financial resources, and technical skills. Environmental managers largely continue the practice of transferring technology developed in the Northern Hemisphere to tropical ecosystems, without taking into account the fact that the relative sensitivities of tropical and temperate species are noticeably different to different stressors. Additionally, ecological differences of the tropics are often overlooked: higher temperatures, higher organic matter turnover with faster oxidation-reduction activities, and also a possible multi-stressor scenario with anthropogenic modifying factors.

Concluding remarks Robinson (1978), basing his reasoning on conditions in coral reefs and tropical forests, concluded that there is a tropical ecology; however, his focus was on the higher level of complexity in the tropics. We are certain that this does not constitute an argument for the existence of a tropical ecology. One of the tools for achieving an ecological understanding is generalization; this is critical in ecological research, because without pattern, we cannot determine the significance of a prospective explanation (Pickett et al. 2007), and there is no


specific generalization for a tropical ecotoxicology that does not hold for ecology. Moreover, inter- and intraspecific relationships between species and ecological mechanisms do not show major differences along latitudinal gradients as hypothesized by Robinson (1978). Ecotoxicology is a subdiscipline of ecology, integrated with toxicology, and its development in the tropics is dependent upon the use of appropriate tools that recognize the biological complexities and develop local management models to prevent the loss of valuable ecological services. In fact, most of the ecotoxicology that is being carried out in the tropics is still based on the toxicology testing-based approach that can serve to derive maximum acceptable chemical concentrations, but has not been effective as far as ecological understanding is concerned. Although we cannot agree with the arguments of Robinson (1978), we cannot deny that biological processes play a more important role in the tropics, and the task of the scientist is to convince environmental managers and policymakers that adequate methodologies should be generated to recognize this, adapting the ERA protocols to the specific conditions of different sites. The uniqueness of tropical ecotoxicology represents an enormous challenge, because the vast majority of the world’s most threatened biodiversity hotspots are found in the tropics. Therefore, appropriate tools need to be determined or developed to prevent the loss of valuable ecological services.

REFERENCES Baird DJ, Finlayson CM, Camillero C. 1995. Ecological impact of contaminants on wetlands: Towards a relevant method of risk assessment. In: Finlayson CM, editor. Wetland Research in the West Dry Tropics of Australia: workshop, Jabiru, NT, 22–24 March 1995. Barton (ACT): Supervising Scientist. pp 242–246. Guimara˜es JRD, Meilib M, Hylanderc LD, Silva EC, Roulete M, Mauroa JBN, Lemosa RA. 2000. Mercury net methylation in five tropical flood plain regions of Brazil: high in the root zone of floating macrophyte mats but low in surface sediments and flooded soils. Sci Tot Environ 261:99–107. Johnson KH, Vogt KA, Clark HJ, Schmitz OJ, Vogt DJ. 1996. Biodiversity and the productivity and stability of ecosystems. TREE 11:372–377. Kilham P, Kilham SS. 1990. Endless summer: internal loading processes dominate nutrient cycling in tropical lakes. Freshw Biol 23:379–389. Pickett STA, Kulasa J, Jones CG. 2007. Ecological Understanding. The Nature of Theory and Theory of Nature. 2nd ed. London: Academic. 248 p. Robinson MH. 1978. Is tropical ecology real? Trop Ecol 19:30–50. Smith VH, Tilman GD, Nekola JC. 1999. Eutrophication: impacts of excess nutrient inputs on freshwater, marine, and terrestrial ecosystems. Environ Pollut 100:179–196.

EPIGENETICS: AN EMERGING FIELD IN ENVIRONMENTAL TOXICOLOGY Juliette Legler* Institute for Environmental Studies, VU University Amsterdam, Amsterdam, The Netherlands *[email protected] DOI: 10.1002/ieam.40

Epigenetics is an emerging field in biology that has widespread implications for environmental toxicology, in terms of experimental design and analysis, as well as risk


assessment of chemicals. Epigenetics refers to heritable changes in gene expression and gene function without accompanying alterations in the DNA sequence. In parallel to the term ‘‘genome’’ that defines the complete set of genetic information contained in the DNA of an organism, ‘‘epigenome’’ refers to the complete set of characteristics of epigenetic pathways in an organism. The epigenome provides the instructions to genes for what to do, and where and when to do it. By regulating gene expression patterns, epigenetic mechanisms are essential in normal development and differentiation, and crucial for health. Environmental factors, including contaminants, may alter epigenetic control of gene expression, with important consequences for development and susceptibility to disease. In this article, I briefly introduce the field, provide an overview of the evidence of disturbances of epigenetic mechanisms by chemicals, and discuss the implications for environmental toxicology. Up to now, 4 types of epigenetic pathways have been identified: DNA methylation, histone modification, nucleosome remodeling, and noncoding RNA-mediated pathways. The most widely study epigenetic mechanism is DNA methylation, in which a methyl group is covalently added by DNA methyltransferases to the fifth carbon of the cytosine ring to form 5-methyl cytosine. Along the linear DNA chain, CpG sites occur where a cytosine is followed by and linked via a phosphate to guanine. DNA methylation occurs predominately on the so-called CpG islands, regions of DNA that have a high density of CpG sites. DNA methylation is actively involved in regulating cell differentiation and function. A methylated gene is usually silenced; DNA methylation is thus referred to as the switch that turns genes on and off. Importantly, the methylated cytosine modifications can be copied to newly synthesized DNA, and so passed on to daughter cells. Alterations in epigenetic pathways have been implicated in aging and common human diseases, such as cancer, cardiovascular diseases, type 2 diabetes, and obesity (see below). In cancer, epigenetic changes have been observed in many steps of tumor development and progression. Too little DNA methylation may initiate chromosome instability and activate oncogenes, while too much DNA methylation may initiate the silencing of tumor suppressor genes. Developmental exposure to environmental toxicants can also induce epigenetic changes. As reviewed in Reamon-Buettner et al. (2008), animal studies have demonstrated altered DNA methylation patterns and histone modifications due to exposure to nongenotoxic chemicals and carcinogenic metals, such as phenobarbital, inorganic arsenic, and nickel, as well as cigarette smoke. Epigenetic mechanisms also play a role in endocrine disruption, in particular in the transgenerational effects of some hormonally active chemicals such as diethylstilbestrol (DES) (reviewed in Crews and McLachlan 2006). Mice treated during development with DES show altered DNA methylation patterns in a specific uterine gene, LF. It is possible that the transgenerational epigenetic effects seen with DES may also occur with other environmentally relevant EDCs. A recent study by Anway et al. (2005) showed that rats treated with the antiandrogenic fungicide vinclozolin during pregnancy produce male offspring that have increased infertility, and that this effect on fertility is passed through the adult male germ line for 4 generations. These transgenerational effects were accompanied by altered

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patterns of DNA methylation in the germ cells. However, this study is controversial, because other laboratories have not been been able to reproduce these findings. Research groups worldwide are studying the implications of early life exposure to environmental chemicals on longterm human health via altered epigenetics, to test the socalled ‘‘developmental origins of health and disease’’ hypothesis. We have recently started a European Commissionfunded project called OBELIX (OBesogenic Endocrine disrupting chemicals: LInking prenatal eXposure to the development of obesity later in life; www.theobelixproject.org), in which we will examine the hypothesis that prenatal exposure to EDCs in food plays a role in the development of obesity later in life. What about the role of epigenetics in environmental toxicology? Novel publications are starting to appear, in which techniques from mammalian epigenetics are being applied to ecotoxicological model species such as Daphnia (Vandegehuchte et al. 2009) and fish (Aniagu et al. 2008) to examine changes in global methylation by chemicals. These are the first steps necessary to determine the methylation status of invertebrates and fish. The next steps will be to determine if chemicals alter the epigenome, and if these changes lead to long-term, heritable effects. Given the evidence from mammalian studies, I would be surprised if they did not. The advent of epigenetics in all areas of biological sciences means that we as environmental toxicologists should consider altering experimental protocols to include developmental exposure, epigenetic analysis and transgenerational studies. Importantly, epigenetic inheritance implies that epigenetic alterations due to chemical exposure in one generation may have effects on subsequent nonexposed generations. This is a concept with far-reaching implications for (ecological) risk assessment.

REFERENCES Aniagu SO, Williams TD, Allen Y, Katsiadaki I, Chipman JK. 2008. Global genomic methylation levels in the liver and gonads of the three-spine stickleback (Gasterosteus aculeatus) after exposure to hexabromocyclododecane and 17-betaoestradiol. Environ Int 34:310–317. Anway MD, Cupp AS, Uzumcu M, Skinner MK. 2005. Epigenetic transgenerational actions of endocrine disruptors and male fertility. Science 308:1466– 1469. Crews D, McLachlan J. 2006. Epigenetics, evolution, endocrine disruption, health, and disease. Endocrinology 147:S4–S10. Reamon-Buettner SM, Mutschler V, Borlak J. 2008. The next innovation cycle in toxicogenomics: Environmental epigenetics. Mutat Res 659:158–165. Vandegehuchte MB, Kyndt T, Vanholme B, Haegeman A, Gheysen G, Janssen CR. 2009. Occurrence of DNA methylation in Daphnia magna and influence of multigeneration Cd exposure. Environ Int 35:700–706.

CRADLE TO CRADLE: OLD WINE OR NEW SPIRITS? Jose´ Potting*y and Carolien Kroezey,z yWageningen University & Research, Wageningen, The Netherlands zOpen University of The Netherlands, Heerlen, The Netherlands *[email protected] DOI: 10.1002/ieam.42

The founding fathers of Cradle to Cradle, McDonough and Braungart (2002), consider present environmental policy to be playing too much on guilt and reduction of consumption.

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Everyone can produce and consume as much as they want, according to the ideas of Cradle to Cradle. Problems caused by overpopulation and depletion of resources would disappear if people would learn, just as termites, to convert all used materials in food cycles. In other words, we should design all products such that they can be fully reused and recycled. Waste becomes food in this way. Waste ¼ Food. That is the appealing metaphor summarizing the philosophy of McDonough and Braungart (2002). The Cradle to Cradle Certification Program (MBDC 2007) claims Cradle to Cradle as a new and revolutionary approach taking its basis in the regenerative productivity of nature that enables creating industry that continuously improves and sustains life and growth. In this Learned Discourse, we reflect upon the originality of Cradle to Cradle by taking a look at the history of the already existing and strongly overlapping philosophy of closing cycles. Closing cycles The first National Environmental Policy Plan in the Netherlands (NEPP 1989) marked a turning point in Dutch environmental policy and served as a source of inspiration for similar plans in several other countries as well as for the Fifth Environmental Action Program from the European Commission in 1993 (Lieffering and van der Zouwen 2004). The NEPP sets out an encompassing strategy for the development of a more sustainable economy by controlling substance flows through the entire cycle of production and consumption. Closing substances cycles, portrayed in Figure 1, is one of the leading principles in NEPP (1989). Figure 1 visualizes the substance streams between and within the ecological system and economic system (from resource extraction, production, and consumption, to disposal and waste management). The policy plan considers opening or changing of substance cycles, more intensive use of energy, and neglect of quality aspects in production processes and products as the common causes for environmental problems. Separating the economic and ecological substance cycles by preventing leaks from the economic substances cycle, as represented by the valves in

Figure 1, is put forward as a main strategy to preserve the environment’s carrying capacity and its natural resources. This is basically what McDonough and Braungart (2002) advocate, and stimulated us to make a closer comparison of the Cradle to Cradle philosophy with the philosophy of closing (substance) cycles. In doing so, we draw from a long history of being actively involved in the discussions about closing cycles, and in method development and case studies in the field of life cycle assessment since its infancy.

DISCUSSION There is clearly a strong overlap between the Cradle to Cradle philosophy and the philosophy of closing cycles. There is an important difference, however, between the analytical supports of the 2 philosophies. Cradle to Cradle provides a design strategy within clearly set boundaries. These boundaries involve human and ecological health, material reutilization, use of renewable energy, water conservation and discharge, and social criteria. There is a strong emphasis in Cradle to Cradle on human and ecological health, and toxic substances in particular (McDonough et al. 2003; Braungart et al. 2007). Life cycle assessment (LCA), the analytical tool developed in support of the philosophy of closing cycles, is more encompassing than Cradle to Cradle in the environmental issues covered (ISO [International Standards Organization] 14044 dated 2006). LCA does not use clearly set boundaries and targets, but rather analyzes product systems to facilitate system optimizations by evaluating benefits of process or product alternatives. Braungart et al. (2007) explicitly take distance from such optimization strategy as ‘‘less bad is no good.’’ Several criticisms exist about Cradle to Cradle, in particular its seemingly belief in the nonhazardousness of natural substances and materials (Reijnders 2008). Rather than hairsplitting about these criticisms, however, we focus here on what we consider as the major contribution of Braungart and McDonough (2002) and their philosophy: the way in which they empower stakeholders to implement Cradle to

Figure 1. Substance streams with ‘valves’ (NEPP 1989).


Cradle in their own practice. This is of crucial importance as policy makers, industry, and consumers are the carriers of sustainable change. The philosophy of Cradle to Cradle now is just as appealing as the philosophy of closing cycles and LCA were in their early days. Both the philosophy of closing cycles and LCA have experienced a diminishing of societal and political attention over the years. Other societal and political problems became important at the expense of the environment. At the same time, the philosophy of closing cycles got thoroughly rooted in our environmental policies, whereas LCA moved up to a highly appreciated member of the environmental toolbox. Both the philosophy and analytical tool grew from innovative eye-catchers into less visible business-as-usual, which in itself, of course, is a very positive development. The difficulties in practice of closing cycles, on the other hand, have also manifested themselves over the course of time. Many products involve complex manufacturing networks that tend to value aspects like company profits and employment more than the environment. Identifying improvement measures also turned out to be not unambiguous. Finally, we may question ourselves whether LCA as an analytical tool has become too sophisticated and comprehensive, and therewith too long winded and complex, specifically against the background of the ongoing debate about uncertainties in results. Cradle to Cradle could learn from that history and avoid some pitfalls of closing cycles and of LCA. The enormous reservoir of knowledge and experience acquired with LCA in particular can help Cradle to Cradle to evaluate whether waste really can become food in an infinitive eco-effective cycle. Another, more crucial lesson, in our opinion, is the importance of remaining appealing for stakeholders. Cradle to Cradle presently functions as an extremely powerful framing for communicating and mobilizing societal and political action. It should not forfeit its appealing character in the

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course of its further refinement. LCA, on the other hand, could contemplate on possibilities of streamlining its present complexity into a simpler though evenly robust methodological framework. This might help to (re-)build enthusiasm for the tool.

Conclusion Cradle to Cradle is an extremely powerful framing for communicating and mobilizing societal and political action. We consider it more relevant to recognize and use this fact, than to split hairs about similarities and differences with the earlier philosophy of closing cycles and LCA as its supporting analytical framework. It would be a terrible loss to forfeit (again) the energy and enthusiasm for change once embodied in closing cycles and LCA and now reinvigorated in the Cradle to Cradle approach.

REFERENCES Braungart M, McDonough W, Bollinger RA. 2007. Cradle-to-cradle design: Creating healthy emissions and a strategy for eco-effective product and system design. J Cleaner Prod 16: 1337–1348. Lieffering D, van der Zouwen M. 2004. The Netherlands: the advantages of being ‘mr average’?. In: Jordan A, Liefferink D. Environmental Policy in Europe. New York (NY): Routledge, Taylor and Francis Group. p. 136–153. McDonough W, Braungart M. 2002. Cradle to Cradle. Remaking the Way We Make Things. New York (NY): North Point. 197 p. McDonough W, Braungart M, Anastas PT, Zimmerman JB. 2003. Applying the principles of green engineering to Cradle-to-Cradle design. Environ Sci Technol 37: 434–441. [MBDC] McDonough Braungart Design Chemistry. 2007. Cradle to Cradle Certification Program. Charlottesville (NC): 25 p. [NEPP]. 1989. National Environmental Policy Plan. To Choose or to Lose. The Hague (NL): SDU-Uitgevery. 257 p. Reijnders L. 2008. Are emissions or wastes consisting of biological nutrients good or healthy? J Cleaner Prod 16: 1138–1141.