part 1 - SETAC Europe 22nd Annual Meeting / 6th SETAC World

Transcription

A comparison of deterministic and stochastic matrix
population models to evaluate ecological risk of chemicals
1
Niklas Hanson1 and John Stark2
University of Gothenburg, Dept. of Plant and Environmental Sciences, Sweden
2
Washington State University, Puyallup Research and Extension Center, USA
E-mail contact: [email protected]
1. Introduction
The two most commonly used metrics to estimate ecological risk of chemicals are acute LC50 and
chronic NOEC. However, environmental protection goals are mainly defined on the population level
[1], such as population size, sustainable harvest and risk for extinction. To be useful for decision
makers, the observations on individuals, therefore, have to be extrapolated to the population level.
The most frequently used method for extrapolation is fixed safety factors [2], which are set to avoid
unacceptable risk. However, it has been shown that this involves a large degree of uncertainty, and
can lead to overprotective and underprotective assessments [3, 4].
To provide more realistic measures of population level risk caused by toxic chemicals, it has been
suggested that population models should be used more frequently [5-7]. However, before population
models can be used routinely in ecological risk assessment (ERA), it has to be evaluated which types
of population models that are most suitable to estimate risk in different management scenarios.
In the present study, deterministic and stochastic matrix population models were used to estimate the
population level impact of toxic chemicals for two species of fish (eelpout, Zoarces viviparous, and
perch, Perca fluviatilis). The toxic effect was simulated based on previously published dose-response
studies. The results of the models were also compared to traditional ERA based on acute LC50 and
chronic NOEC, according to current approaches for ERA in the EU. The objective of the study was to
determine if simple deterministic models provide an improvement compared to traditional measures of
ecological risk (given the same data), or whether more complicated models that include environmental
stochasticity are preferable.
2. Materials and methods
The data that were used to parameterize the models covered 10 years of catch data for eelpout and
17 years for perch. The development of the models is described in two earlier papers [8, 9].
Previously published studies that examined toxic effects on survival and reproduction for fish were
used to develop dose-response relationships for a metal mixture, 2,3,7,8-tetrachlorodibenzo-p-dioxin
(TCDD), and tributyltin (TBT). Dose-response curves were created by fitting the logistic function to the
data, where survival and fertility were presented as proportions of control. The study was not intended
to estimate the risk these chemicals pose to the two modelled species. Instead, the aim was to
compare different methods (two types of models and traditional LC50/NOEC) using the same data.
For this purpose, data from real fish studies are more informative than hypothetical dose-response
curves.
The predicted no effect concentration was retrieved from LC50 (PNEC acute ) and NOEC (PNEC chronic ) in
accordance with the EU technical guidance document [2]. PNEC is determined by dividing LC50 and
NOEC with safety factors. Depending on the amount of available data, the factors vary. Therefore a
higher and a lower PNEC was determined. From the models, a higher PNEC model was determined
from the deterministic model as the concentration where the population is stable. Assuming that the
risk for error is not biased, this represents a 50% probability for population decline. The lower
PNEC model was determined from the stochastic model as the concentration where there is a 5%
probability for population decline.
3. Results and discussion
Figure 1 shows the results for eelpout (the results were similar for perch). For the metal mixture and
TCDD, the models provided PNECs that were clearly higher than the traditional risk estimates.
However, for TBT, one of the traditional risk estimates (LC50 and the safety factor 100) was higher
than one PNEC model . Based on acute data, a TCDD concentration that causes a probability for
population decline between 5% and 50% would, thus, be acceptable. The traditional safety factors
based on LC50 rendered clearly overprotective results in two cases, and underprotective results in
one case. Using the deterministic model, and a safety factor, would reduce this error. Using stochastic
models, and a percentile, would eliminate it. However, in most situations the data is not available to
develop such models.
Figure 1: Deterministic (λ d ) and stochastic (λ s5 ) population growth rates (solid and dotted curves, respectively). The
solid lines represent PNEC acute (dotted), PNEC chronic (dash-dotted) and PNEC model (solid).
4. Conclusions
In the present study, it was clearly seen that simple deterministic models reduced uncertainty in ERA
considerably compared to traditional, organismal-level, endpoints. However, population models that
include environmental stochasticity reduce uncertainty further, and may be useful for higher tiers of
ERA, if the data and the expertise is available.
5. References
1. Hommen, U.; Baveco, J.; Galic, N.; van den Brink, P. J., Potential application of ecological models
in the European environmental risk assessment of chemicals I: Review of protection goals in EU
directives and regulations. Integrated Environmental Assessment and Management 2010, 6, (3),
325-337.
2. European Commission Technical Guidance Document in support of Commission Directive 93/67
EEC on risk assessment for new notified substances, Commission Regulation (EC) no. 1488/94 on
risk assessment for existing substances and Directive 98/8/EC of the European Parliament and of
the Council concerning the placing of biocidal products on the market; European Chemical Bureau:
Ispra, Italy, 2003.
3. Forbes, V. E.; Calow, P.; Sibly, R. M., Are current species extrapolation models a good basis for
ecological risk assessment? Environ. Toxicol. Chem. 2001, 20, (2), 442-447.
4. Stark, J. D., How closely do acute lethal concentration estimates predict effects of toxicants on
populations? Integrated Environmental Assessment and Management 2005, 1, (2), 109-113.
5. Ferson, S.; Ginzburg, L. R.; Goldstein, R. A., Inferring ecological risk from toxicity bioassays. Water
Air and Soil Pollution 1996, 90, (1-2), 71-82.
6. Stark, J. D.; Banks, J. E., Population-level effects of pesticides and other toxicants on arthropods.
Annu. Rev. Entomol. 2003, 48, 505-519.
7. Forbes, V.; Calow, P.; Grimm, V.; Hayashi, T. I.; Jager, T.; Katholm, A.; Palmqvist, A.; Pastorok, R.
A.; Salvito, D.; Sibly, R.; Spromberg, J.; Stark, J. D.; Stillman, R. A., Adding value to ecological risk
assessment with population modelling. Hum. Ecol. Risk Assess. 2011, 17, 287–299.
8. Hanson, N., Population level effects of reduced fecundity in the fish species perch (Perca fluviatilis)
and the implications for environmental monitoring. Ecol. Model. 2009, 220, (17), 2051-2059.
9. Hanson, N.; Åberg, P.; Sundelöf, A., Population-level effects of male-biased broods in eelpout
(Zoarces viviparus). Environ. Toxicol. Chem. 2005, 24, (5), 1235-1241.
An ecosystem model for risk assessment of aquatic
environments impacted by endocrine disrupters
Ludiwine Clouzot1, Mike Paterson2, Alain Dupuis2, Paul Blanchfield2, Mike Rennie2,
Karen Kidd3 and Peter A. Vanrolleghem1
1
modelEAU, Université Laval, 1065, avenue de la Médecine, Québec G1V 0A6, QC, Canada
2
Fisheries & Oceans Canada, 501 University Crescent, Winnipeg, R3T 2N6, MN, Canada
3
Canadian Rivers Institute, University of New Brunswick, Saint John, E2L 4L5, NB, Canada
1. Introduction
Endocrine disruption has been measured in many aquatic environments across the world but the
consequences on the whole ecosystem are still unclear. Experimental approaches for characterizing the
ecological impact of such disturbances are costly and time-consuming. Therefore, ecological models are
currently being developed to support risk assessors in their decisions. Ecosystem models are required for
assessing effects at high levels of organization, but the need for extensive calibration for a specific
ecosystem limits their application in ecological risk assessment (ERA). This study aims at providing risk
managers with an ecosystem model able to predict critical changes in aquatic environments impacted by
endocrine disrupters. The objective is to find a compromise between the data available and the complexity
required for the model.
The ecosystem model is developed with field data obtained from a multi-year whole-ecosystem study
performed at the Experimental Lake Area (ELA, Ontario, Canada): (i) two years of reference data (ii) three
years of exposure to environmentally-relevant concentrations of the synthetic hormone 17α-ethinylestradiol
(EE2) and (iii) five years of recovery data [1]. EE2 was chosen because it is one of the most widespread and
potent endocrine disrupters. Indeed, the fathead minnow population collapsed after the second year of EE2
additions and endocrine disruption was observed in the other fish species as well. The developed ecosystem
model considers direct effects of EE2 on fish species but also the consequences on the whole ecosystem
through ecological interactions i.e. feeding and competition relationships. The model is developed with the
reference data and calibrated with the data collected during EE2 addition.
2. Experimental and modelling approaches
2.1. Ecosystem modelling of endocrine disruption
An ecosystem model was previously developed to predict the impact of metals on lentic ecosystems [2]. The
model was based on simplified equations of the AQUATOX model (USEPA, 2002). The study presented in
this abstract consists of adding equations for endocrine disruption. The model was built with the populations
naturally present in the experimental lake, i.e fish, benthic invertebrates, zoo & phytoplankton (Figure 1).
Fathead minnow
Epi
Redbelly dace
Lake chub
White sucker
Pearl dace
Finescale dace
Slimy sculpin
Hypo
Lake trout
Figure 1: Ecosystem of the experimental lake
In AQUATOX, reproductive endpoints exist for fish with different age- or size-classes. However, modelling
endocrine disruption requires additional classes to differentiate males and females. For the moment, intersex
fishes are not considered in the model because the important endpoint is reproduction and intersex fish can
still reproduce. Instead, a variable representing the reproductive ability was associated with each fish class.
Three fish classes are used in this ecosystem model; juveniles, females, and males. Further model
development will allow for predicting the percentage of intersex fish, males, and females.
2.2. Dynamic modelling of the biomass and the nutrients
The experimental lake stratified during summer and thus different profiles appeared (Figure 2). The
epilimnion (the top layer) was warm and well-oxygenated whereas the hypolimnion (the deep layer) was cold
with low oxygen. The two layers were separated by a rapid temperature change with an oxygen peak
(metalimnion). Twice a year (spring & fall), these layers were mixed, and then oxygen entered the deep zone
and the nutrients accumulated in the hypolimnion during summer were recycled to the epilimnion.
T (°C)
O2 (µmol/L)
25
12
DOC (µmol/L)
600
8
400
4
200
N (µg/L)
800
20
Hypo
15
600
400
10
5
Epi
Epi
0
0
0
200
Met
5
10
Depth (m)
15
Met
Hypo
0
0
0
5
10
Depth (m)
15
Figure 2: Temperature, oxygen, and nutrient profiles during the lake stratification (07-25-2002)
The ecosystem model was built to consider the annual dynamics presented in Figure 3. The stratification
was represented by the epilimnion (including the metalimnion) and the hypolimnion. Organic matter (OM)
and nutrients (N and P) are modelled separately in the two layers with sedimentation of the particulate OM.
EE2 was added to these pools and the temperature and oxygen effects on decomposition are considered.
Transport equations between the layers were developed for the lake mixing during spring and fall (in bold).
Figure 3: Model structure for the lake dynamics
Phytoplankton in the epilimnion was influenced by photosynthetically active radiation (PAR), photoperiod,
turbidity (secchi measurement), temperature, and oxygen (Figure 3). EE2 was added as input to consider
bioaccumulation. Mechanistic equations connect phytoplankton to the detritus/nutrient pool through biomass
growth (Bio), death, respiration (Resp), excretion (Excr) and assimilation (Assim). Phytoplankton sink directly
to the hypolimnion. A similar framework was used for zooplankton with additional transport equations.
Indeed, zooplankton was in the hypolimnion during the day and migrated upwards to the epilimnion at night.
3. Conclusions
This ecosystem model was developed with a focus on EE2-endpoints reliable for risk assessment. The
experimental lake study gave acess to a high quality data set and additional data from similar studies will be
used to validate the model with different contaminations (e.g. metals, high nutrients load, acidification, etc.).
4. References
[1] Kidd KA, Blanchfield PJ, Mills KH, Palace VP, Evans RE, Lazorchak JM and Flick RW. 2007. Collapse of
a fish population following exposure to a synthetic estrogen. Proceedings of the National Academy of
Sciences 104(21):8897-8901.
[2] De Laender F, De Schamphelaere KAC, Vanrolleghem PA and Janssen CR. 2008. Validation of an
ecosystem modelling approach as a tool for ecological effect assessment. Chemosphere 71:529-545.
PRiME – An approach for the comparative risk assessment
of pesticides
Pierre Mineau1, Susan E. Kegley2, Michael Guzy3, Chuck Benbrook4, Paul C.
Jepson5, Karen Benbrook6, Tom A. Green7, Wade D. Pronschinske7, Amrita Batra7,
Leigh Presley7, Jonathan Kaplan8.
1
Science and Technology Branch, Environment Canada, Ottawa K1A 0H3
2
Pesticide Research Institute, 2768 Shasta Road, Berkeley, CA 94708
3
Biological and Ecological Engineering, Oregon State University, OR, 97331
4
The Organic Center, Boulder Colorado, CO 80301
5
Integrated Plant Protection Centre, Oregon State University, Corvallis, OR 97331
6
BCS Ecologic, 90063 Troy Rd., Enterprise, OR 97828
7
IPM Institute of N.A., 4510 Regent Street, Madison, WI 53705
8
Natural Resources Defense Council, 111 Sutter St. San Francisco, CA 94104
[email protected]
__________________________________________________________________________________
Most approaches for the comparative assessment of pesticides are hazard-based. They typically
score combinations of fate and toxicity endpoints from standard laboratory tests. Application rates
may or may not factor in this comparison. The best approaches derive exposure estimates through
models, as regulatory authorities typically do. However, even those approaches that attempt to
combine toxicity and exposure into an actual measure of risk, fail in a number of key areas: 1)
Adequately addressing inter-species differences in toxicological susceptibility, a key source of
uncertainty; 2) Including local conditions such as rainfall, proximity to water bodies and soil type to
produce a context-specific risk score; 3) Adjusting risk scores for application methodology; 4)
Calibrating the estimated risk score against documented field impacts. The Pesticide Risk Mitigation
Engine (PRiME) was designed to address all of these key areas. It was developed with the support of
a Conservation Innovation Grant administered by the U.S. Department of Agriculture’s National
Resource Conservation Service as well as with time and resource contributions from numerous
partners in the agricultural and food industry sectors. PRiME uses acute and other toxicity endpoints
from both regulatory and non-regulatory tests on a wide range of species. Species-sensitivity
distributions are used to derive taxon-specific endpoints. Fate and runoff models are run to assess
likely environmental exposures and risk scores are calculated for a number of environmental and
human receptors across a selection of pesticides. Most of those scores are estimated probabilities of
harm or undesirable effect. Risk is presented on a low /moderate/high scale for final output to the
user. Risk score indices designed to date are described below. Others (e.g. pollinators) are in
development.
Avian Acute
• Undesirable effect: Visible mortality defined as mortality that would be detected in the course of an
intensive and competent search of the treated area.
• Index score: The probability that a given application will give rise to visible bird mortality. This
probability has been calibrated against actual field studies carried out with a variety of pesticides.
Avian Chronic
• Undesirable effect: Birds breeding in and around fields are prevented from breeding because of
persistent residues in food sources.
• Index score: The proportion of the typical breeding season (90 days) with residue levels such that
reproduction is compromised in a reasonably sensitive 15g songbird. In other words, the worst
possible score of 1 would be an application that potentially inhibits avian reproduction for the entire
length of the “normal” breeding season. There is no field information with which this probability can
be verified; however, the determination of risk is based on laboratory studies of reproductive effects
as well as accepted principles of exposure recently devised by a number of European expert groups.
Small Mammals
• Undesirable effect: Population-level effects on mortality, fertility or recruitment.
• Index score: The probability that residues will persist long enough at a toxic level to cause changes
in the population trajectory of small mammals directly exposed to the spray. This probability has been
calibrated against actual field studies carried out with a variety of pesticides.
Earthworms
• Undesirable effect: Loss of 35% or more of earthworm biomass.
• Index score: The probability of biomass loss calibrated against actual field studies carried out with a
variety of pesticides.
Fish
• Undesirable effect: Fish in nearby receiving waters are prevented from breeding due to levels of
pesticide residues exceeding the Maximum Acceptable Toxicant Concentration (MATC).
• Index score: The proportion of the typical fish breeding season (30 days) where residue levels
exceed the MATC. There is no field information with which effects can be verified. However, the
determination of risk is based on laboratory studies of reproductive effects as well as modeled
estimates of exposure.
Aquatic Invertebrates
• Undesirable effect: More than approximately 10-20% of invertebrate taxa are significantly impacted
by treatment, meaning that their population is often decreased by 10 fold or more.
• Index score: The probability that an application will give rise to this level of impact calibrated against
actual field and mesocosm studies carried out with a variety of pesticides.
Algae
• Undesirable effect: More than approximately 20-30% of algal species are significantly impacted by
treatment, meaning that their growth is impeded, usually by 10 fold or more.
• Index score: The probability that an application will give rise to this level of impact calibrated against
actual field studies carried out with a variety of pesticides.
Inhalation Risk
• Undesirable effect: A vulnerable individual (i.e. 1-year-old child) spending a significant amount of
time (e.g. living) within 100 feet of a treated field is exposed to volatilized pesticide at a concentration
exceeding the short-term Reference Exposure Level (REL) based on US EPA’s or California
Department of Pesticide Regulation’s Level of Concern.
• Index score: This score is given as a hazard quotient defined as the predicted dose divided by the
REL.
Dermal Risk
• Undesirable effect: Workers reentering a treated field after the re-entry interval are exposed to a
dermal dose of pesticide that exceeds US EPA’s short-term dermal Reference Dose (RfD)
• Index score: This score is given as a hazard quotient defined as the predicted dose (given specific
conditions of crop, clothing, and work time) divided by the RfD.
Dietary risk
This index differs from the others in being pesticide-specific but not application-specific. Commodity
survey data are used to establish probable residue levels in common foodstuffs.
• Undesirable effect: Consumption of pesticide residues in the food supply that approaches or
exceeds the EPA-determined chronic Reference Dose (cRfD) or chronic Population Adjusted dose
(cPAD).
• Index score: The likelihood that the cRfC for a 16 kg child will be exceeded given consumption at the
th
95 percentile of consumption surveys.
An approach for environmental risk assessment of
pharmaceuticals by using fuzzy AHP
Emel Topuz1, Egemen Aydin1, Ilhan Talinli1
1
Istanbul Technical University, Faculty of Civil Engineering, Environmnetal Engineering Department, 34469
Maslak, Istanbul, TURKEY
1. Introduction
Pharmaceuticals in the environment is one of the mostly studied topics for three decades. They are
introduced to environment via discharge or reuse of wastewater since most of the pharmaceuticals are
relatively more resistant towards conventional treatment [1]. There are significant amount of studies on
occurrence, fate, transport and effects of pharmaceuticals [2]. However, such studies are effect directed and
do not indicate conclusive results on risks caused by pharmaceuticals. Moreover, everyday new
pharmaceuticals are released to markets and it is difficult to predict environmental effects of them, quick
enough to take proper measures. Therefore, a robust environmental risk assessment (ERA) approach is
needed. Although there are some approaches such as PEC/PNEC used for this purpose, they contain
uncertainities and subjectivenesses.
Risk is the combination of likelihood (RL) and probability (RP) of an event and the strenght of the reults (RS)
of this event. In order to assess the risks of pharmaceuticals, it is needed to evaluate criterias related with
these concepts. Fuzzy analytic hierarchy process (FAHP) provides to organize these criterias systematically
and conclude priority numbers for the risk management activities [3]. Risk magnitude (RM) can be inferred by
assesssing expert opinions instead of strict formulatios. Fuzzy inference (FI) systems are used for complex
matrices that cannot be regulated or formulated. It is aimed to propose an approach for ERA of
pharmaceuicals based on FAHP and FI system as an objective and sensitive method in order to overcome
uncertainities and compansate data gap in literature.
2. Proposed ERA Approach
A separate hiererchy (Figure 1) was developed for RL, RS and RP in order to get the scores of AHP for
inference of RM. Characteristics of the pharmaceuticals and environment under evaluation were consşdered
as the subfactors of the RL in order to assess fate/transport of the pharmaceutical in the environment and
possible exposure pathways for the ecosystem. Effect assessment is conducted by using RS hierarchy
including subfactors of ecotoxicological effects of pharmaceuticals and environmental characteristics that
contribute the magnitude of these effects. RP hierarchy contains the sources of the pharmaceuticals and
factors related with treatibility indicating the possibility of release to environment. AHP scores of these factors
were combined by using fuzzy inference (intersection and union operations) based on expert opinions and
RM for the case evaluated is achieved. Calculation steps of the methodology is provided in Topuz et al.,
2011 [4].
3. Conclusions
Risk assessment of pharmaceuticals in the environment by using AHP provide to consider all of the factors
contribute to risk both related with pharmaceuticals and environment in terms of risk components that are
RL, RS and RP. FI method enables to use expert opinions instead of ampric formulations which is beneficial
for the complex assessments that are affected by numerous factors and cannot be completely formulated
like ERA. Proposed approach reduces uncertanities and subjectiveness and provides more rigid RM that can
be used as a guide for the risk management attributes.
4. References
[1] Ternes, T. A., Joss, A., editors. 2006. Human pharmaceuticals, hormones and fragrances : the challenge
of micropollutants in urban water management. Seattle, WA: IWA Pub.
[2] Santos, L. H.M.L.M., Araújo A.N., Fachini, A , Pena, A., Delerue-Matos C., Montenegro, M.C.B.S.M.
2010. Ecotoxicological aspects related to the presence of pharmaceuticals in the aquatic environment.
Journal of Hazardous Materials. 175. 45-95
[3] Saaty TL. 2001. Decision making for leaders: the analytic hierarchy process for decisions in a complex
world. Pittsburgh: RWS Publications.
[4] Topuz, E., Talinli, I., Aydin, A., 2011. Integration of environmental and human health risk assessment for
industries using hazardous materials: A quantitative multi criteria approach for environmental decision
makers. Environment International. 37. 393-403
Figure 1. Risk assessment hierarchy for pharmaceuticals in the environment
A new methodology for PBT prioritization of chemical
inventories
Denitsa Georgieva1, Sabcho Dimitrov1, Nadezhda Dimitrova1 and Ovanes Mekenyan1
1
Laboratory of Mathematical Chemistry, University "Prof. As. Zlatarov", 8010 Bourgas, Bulgaria
1. Introduction
A new PBT prioritization methodology has been developed in view of regulatory requirements for quantitative
assessment of persistence, bioaccumulation potential and toxicity of chemicals. The approach uses
CATALOGIC/OASIS [1-3] and EPI Suite QSAR models for calculating P, B and T endpoints. A classification
scheme is introduced based on endpoint-specific cut-off criteria and applicability domains of the used
models. PBT classification results are further used for grouping chemicals focused on: 1) identification of
high-priority chemicals based on PBT criteria, 2) identification of data gaps and possible further research.
The methodology is implemented in a software system for computerized PBT classification of large chemical
inventories.
2. Methodology
The developed approach introduces a two-stage PBT classification scheme – primary classification of parent
chemicals and secondary classification based on stable degradants, as shown on Fig. 1.
Figure 1: Classification scheme – general workflow
Primary classification of parent chemicals provides a PBT profile based on:
•
•
•
Endpoint-specific cut-off criteria identifying levels of concern
Uncertainty defined as a bright-line around regulatory cut-off criteria
Confidence of obtained classification results based on applicability domain of the used models
2.1. Cut-off criteria
Primary classification of parent chemicals applies cut-off criteria for each of the estimated endpoints. The
approach addresses the uncertainty of obtained results by introducing bright lines around the central
threshold values. Fig. 2 exemplifies the adopted persistence classification. CATALOGIC model predictions
for biodegradation as percentage of theoretical BOD are used. Four persistence classes are suggested,
based on a regulatory threshold and the 20% BOD variation of experimental results (regulatory threshold
±10%). The adopted persistence classes are: P, Marginally P, Marginally notP, notP.
Figure 2: Persistence thresholds and uncertainty bright line
The uncertainty bright line approach is also adopted for B classification. Toxicity classification is based on
thresholds for acute aquatic toxicity. The used thresholds and associated uncertainty will be presented in
detail.
2.2. Confidence of obtained results
The model applicability domain of each used QSAR model is incorporated in the classification scheme in
order to evaluate confidence of obtained estimates. The approach used the applicability domain of
CATALOGIC/OASIS models [4] in order to identify low confidence results based on parametric ranges,
structural similarity and mechanistic understanding of the modelled endpoints.
2.3. Secondary classification
Secondary classification is based on stable degradation products identified by CATALOGIC biodegradation
models. It provides PBT assessment of stable degradants, which can be used to complement the PBT
profile of parent chemicals. Following a more conservative approach, identification of a hazardous stable
degradant can be considered grounds for inclusion of the parent chemical in a higher priority class.
3. Discussion
The PBT prioritization methodology can be used for grouping chemicals into four broad classes – PBT; PB,
BT or PT; P, B or T; notPnotBnotT, based on meeting fate/hazard classification criteria. The approach also
focuses on the rarely addressed uncertainty of model-based prioritization procedures. It allows identification
of “gray area” results, such as Marginally P and Marginally notP classification, and low confidence of results,
which can be used to guide further research and identify data gaps.
4. Conclusions
The adopted PBT classification approach offers a transparent procedure for fate/hazard assessment of
chemicals. Grouping of chemicals can be completed following an agenda of identification of chemicals of
concern and uncertainty analysis of obtained results. The methodology can be used for computerized PBT
classification of large chemical inventories.
5. References
[1] Dimitrov S, Pavlov T, Dimitrova N, Georgieva D, Nedelcheva D, Kesova A, Vasilev R, Mekenyan O.
2011. Simulation of chemical metabolism for fate and hazard assessment. II. CATALOGIC simulation of
abiotic and microbial degradation. SAR and QSAR in Environ Res 22: 719-755.
[2] Dimitrov S, Dimitrova N, Georgieva D, Vasilev K, Hatfield T, Straka J, Mekenyan O. Simulation of
chemical metabolism for fate and hazard assessment. III. New Developments of the bioconcentration factor
base-line model. SAR QSAR Environ Res DOI: 10.1080/1062936X.2011.623321.
[3] Dimitrov S, Mekenyan O, Schultz T. 2000. Interspecies modeling of narcotics toxicity to aquatic animals.
Bull Environ Contam Toxicol 65: 399-406.
[4] Dimitrov S, Dimitrova G, Pavlov T, Dimitrova N, Patlewicz G, Niemela J, Mekenyan O. 2005. A stepwise
approach for defining applicability domain of SAR and QSAR models. J Chem Inf Model 45: 839-849.
Comparative risk assessment of arsenic trioxide and its
substitutes for occupational exposure in Murano (Venice, Italy)
artistic glass production
Valentina Faggian1, Lisa Pizzol2, Elisa Giubilato1, Petra Scanferla2, Andrea Critto1,2,
Antonio Marcomini1 and Nicola Favaro3
1
Department of Environmental Sciences, Informatics and Statistics, University Ca’ Foscari Venice,
Dorsoduro 3246 - 30123 Venice;
2
CVR – Consorzio Venezia Ricerche, c/o PST VEGA, Via della Libertà 12 – 30175 Venice;
3
SSV – Stazione Sperimentale del Vetro, via Briati 10, 30141 Murano, Venice;
1. Introduction
In the district of Murano artistic glass (Venice, Italy), arsenic trioxide (As 2 O 3 ) has being used since centuries
as an additive for refining and bleaching melting glass and it is estimated that about 8 ton/year of this
compound are currently used in Murano glassworks. As 2 O 3 is applied as a fine white powder, which can be
easily inhaled by workers during several stages of the hand-based production. Arsenic trioxide is included in
the Candidate List of Substances of Very High Concern (SVHC) and soon it could be subject to authorization
under the REACH regulation [1], because of its properties of carcinogenicity, mutagenicity and toxicity for
reproduction (CMR substance). As 2 O 3 elimination from the glass production is thus advisable since it would
allow to significantly reduce occupational health risks. These reasons have lead many glass companies to
carry out research programs to replace As 2 O 3 with other oxides [2]. The SSV-Stazione Sperimentale del
Vetro, a test laboratory specialized in applied research for glass industrial development, recently began a
collaboration with Murano glass artisans to test alternative mixtures. Cerium oxide (CeO 2 ) and blast furnace
slag turned out to be the most promising substitutes of As 2 O 3 . A comparative study has been performed with
the aim of evaluating the occupational health risks associated with exposure to arsenic trioxide in
comparison with the identified substitute substances. Specifically, the objectives of this study are: i) to
assess arsenic trioxide health hazard in comparison with cerium oxide and blast furnace slag; ii) to develop
for all production phases suitable exposure scenarios for workers, representative of a typical glassworks; iii)
to quantify inhalation and dermal contact exposure to arsenic trioxide and its substitutes through the
application of different occupational exposure models and to compare models’ performance for inorganic
substances in the selected exposure scenarios; iv) to estimate and compare health risks for glass workers
associated to inhalation and dermal contact for the substances of interest.
2. Methods
The comparative risk assessment (CRA) has been developed according to the following conceptual
framework:
a) identification and selection of the most relevant glass production phases;
b) development of the conceptual model for each identified production phase;
c) comparative risk assessment (hazard assessment, exposure assessment,
characterisation).
risk
In step (a), the main phases of the glass production process have been determined and the most dangerous
ones, in terms of potential human exposure, have been identified and further analysed. In step (b), for each
selected glass production phase, two occupational exposure scenarios have been developed: one including
the application of Preventive and Protective Measures (PPM scenario) and one without PPMs (NoPPM
scenario) which represents the “worst-case scenario”. Accordingly, data and information needed for the
characterization of such scenarios have been collected. In step (c), for each selected glass production
phase, a comparative risk assessment has been carried out for As 2 O 3 and CeO 2 , for both the PPM scenario
and the NoPPM scenario. In the Hazard Assessment phase, available toxicological data for As 2 O 3 , CeO 2
and blast furnace slag have been collected and critically analyzed and dose-response parameters have
been selected. In the Exposure Assessment phase, occupational exposure models proposed by European
guidelines have been reviewed and the most suitable for the case study of interest have been selected, i.e.
the models ECETOC worker tool [3], MEASE [4] and ART [5]. The selected models have then been applied
to estimate inhalation and dermal contact exposure to arsenic trioxide and its substitutes. Finally, in the Risk
Characterisation phase, the Risk Characterisation Ratio (RCR), corresponding to the ratio between exposure
concentration and the related dose-response parameter, has been calculated for all scenarios and for all
considered exposure routes (i.e., inhalation and dermal contact). An overall RCR as the sum of individual
RCRs has also been estimated.
As2O3 is highly toxic and surely carcinogenic [6]. Differently there are no sufficient data for demonstrating
the carcinogenic effects of CeO 2 , while there is some evidence that it is toxic through inhalation [7]. The
toxicological dose-response parameters of interest (Derived No Effect Level, DNEL) have been defined only
for toxic effects and for chronic inhalation and dermal contact exposure. Lastly, the studies currently
available on blast furnace slag highlight its non-toxic and carcinogenic nature, so this substance has not be
considered as dangerous and has not been included in the exposure assessment.
The phases identified as dangerous for potential exposure to the chemicals of interest are six:
transport/weighing of dust, mixing of the glass mixture, loading of mixture in the furnace, fusion of glass
mixture, processing and finally finishing of glass. The most significant routes of exposure for glassworkers
turned out to be inhalation of dusts, fumes and/or aerosols and dermal contact with dust, particulate and
aerosols. Exposure estimates obtained for the two substances, considering the same operational conditions,
are very similar and, for some phases, even identical, except for processes at high temperatures where
CeO 2 evaporates less than As 2 O 3 . Exposure estimates obtained with the models ECETOC TRA and
MEASE are quite different. Anyway when they are compared with real measured and site-specific data, they
result both conservative and, particularly those derived with MEASE, in line with the real trend. The ART
model has been used only to better investigate those phases of glass production for which risk was not
acceptable, even with use of CeO 2 .
The risk characterization has shown that the use of As 2 O 3 causes unacceptable risk for inhalation exposure
for all stages of glass production in NoPPM scenario, and for 5 phases out of 6 for the PPM scenario.
Furthermore also the risk for dermal contact is not acceptable. On the contrary the use of CeO 2 poses no
significant risk for all stages of production and for all scenarios of exposure, except for inhalation for the
weighing phase in NoPPM scenario.
4. Conclusions
The comparative risk assessment procedure proves to be effective in demonstrating that the use of cerium
oxide and blast furnace slag would reduce the health risks associated to occupational exposure. The
difference in the estimated risks is related to the different toxicological profiles of the substances of interest
rather than to the levels of exposure. CeO 2 and blast furnace slag are therefore confirmed as interesting
candidates to replace As 2 O 3 in artistic glass production.
5. References
[1] Directive number 1907 of 2006, Official Journal of 30 December 2006, L 396, page 1.
[2] Montagnani R, Campagna M, Gasparello S, Hreiglich A, Apostoli P. 2006. L’esposizione ad arsenico
nella produzione artigianale della bacchetta di vetro. Risultati del monitoraggio biologico e indicazioni
preventive. G Ital Med Lav Erg 28(2):158-162.
[3] ECETOC. 2004. Target Risk Assessment. Brussels, Belgium, European Centre for Ecotoxicology and
Toxicology of Chemicals. Technical Report n. 93.
[4] EBRC. MEASE - Occupational Exposure Assessment Tool for REACH. http://www.ebrc.de/industrialchemicals-reach/projects-and-references/mease.php. Accessed 07/03/2012.
[5] Fransman W, Cherrie J, van Tongeren M, Schneider T, Tischer M, Schinkel J, Marquart H, Warren N,
Kromhout H, Tielemans E. 2009. Development of a mechanistic model for the Advanced REACH Tool
(ART). TNO Quality of Life (The Netherlands) Report n. V8667.
[6] ATSDR. 2007. Toxicological Profile for Arsenic. http://www.atsdr.cdc.gov/ToxProfiles/tp.asp?id=22&tid=3
Accessed 07/03/2012.
[7] US EPA. 2009. Toxicological review of Cerium Oxide and Cerium Compounds.
http://www.epa.gov/iris/toxreviews/1018tr.pdf. Accessed 07/03/2012.
Acknowledgement - This work is part of a national project focus on the arsenic substitution in artistic glass
production developed by Stazione Sperimentale del Vetro, a high qualified glass research institute with the
aim of promoting the technological progress of glass industry. The authors are particularly grateful to the
Environmental Italian Ministry for its significant financial support.
A Fish Tissue Archive for monitoring chemical pollution in UK
rivers. How it operates and its application to EU priority
substances
Monika D. Jürgens1, Andrew C. Johnson1, Alan Lawlor2, Dave Hughes3,
Askin Birgul3, Athanasios Katsogiannis3, and Kevin Jones3
¹Centre for Ecology and Hydrology (CEH) Wallingford, OX10 8BB, UK;
² Centre for Ecology and Hydrology (CEH) Lancaster, LA1 4AP, UK
³Lancaster University, LA1 4YQ, UK
1. Introduction
Where a concern exists over river contamination by persistent and bioaccumulative chemicals a valuable
approach to monitor pollution is through biomonitoring. Higher organisms, such as fish, can integrate
pollutants over a number of years whilst they live in the contaminated area. Not only can fish make good
‘passive samplers’ they are also relevant target organisms whose health is matter of importance to the
ecosystem. A practical benefit from using ‘nature’s passive samplers’ is they can greatly increase the
chance of positive detections compared to direct water measurements. In addition, for long term monitoring
of pollutants, it is extremely helpful to have historic as well as current samples so direct comparisons can be
made using the same methods.
For this study we focused on chemicals for which a biota standard is set in the Priority Substances Directive
of the EU (Directive 2008/105/EC, entered into force 13.1.2009). This lists freshwater environmental quality
standards (EQS) for over 30 substances and includes an (optional) biota standard for 3 of these
(article 3, 2a) requiring member states to “apply, for mercury and its compounds, an EQS of 20 μg/kg, and/or
for hexachlorobenzene, an EQS of 10 μg/kg, and/or for hexachlorobutadiene, an EQS of 55 μg/kg, these
EQS being for prey tissue (wet weight), choosing the most appropriate indicator from among fish, molluscs,
crustaceans and other biota;”
Mercury was used in many industrial and agricultural applications in the past and is still released as a trace
component of fossil fuels and some electrical components. Hexachlorobenzene (HCB) is a fungicide that is
no longer used in the EU and Hexachlorobutadiene (HCBD) was used in the past as a solvent in polymer
production and as a fungicide and seed dresssing. Both HCB and HCBD are also generated in small
quantities as by-products or trace contaminants of other chlorinated organics [1, 2].
The Environment Agency (EA) in England and Wales carries out an annual fish survey at many river sites,
whereby all the fish in a stretch are caught and numbers, lengths, and species are recorded before releasing
them back into the river. In 2007 scientists from CEH agreed with the EA to build a fish tissue archive by
collecting 10 fish (normally roach, typically 3-4 years old) per year from a number of these survey sites along
the river Thames in the South of England and the rivers Nene, Glen and Welland in Eastern England.
The Thames is the second-longest river in the UK (346 km) with predominantly rural upper reaches and
densely populated urban areas lower down. About 1/5 of the UK population live in the Thames catchment.
The samples were taken along several stretches from the point where it becomes navigable at Cricklade
2
3
(19 km from the source, catchment area ca. 320 km , mean flow ca. 2.4 m /s) to Kingston (3 km above the
2
3
tidal limit and 233 km from the source, catchment area 9948 km , mean flow 78.2 m /s) [3]. The Nene is the
th
10 longest river in the UK (161 km, catchment area and mean flow near the tidal limit in Peterborough are
2
3
1634 km and 9.3 m /s). The Welland is 105 km long and the Glen is the largest tributary of the Welland,
3
3
they have a mean flow of 2.1 m /s and 1.2 m /s at the nearest gauging stations upstream of the sampling
sites [3].
The fish are frozen on site in the gas phase of liquid nitrogen and stored long term at -80°C. That way over
time a valuable resource for future retrospective monitoring of a large number of pollutants can be built up.
To provide information about the current status of English rivers a subset of these fish was analysed for
persistant organic pollutants (POPs) using solvent extraction followed by extensive cleanup and GC/MS
analysis, and for metals using a microwave digestion with nitric acid followed either by ICPMS or a
GALAHAD mercury analyser. The results were compared to EU environmental quality standards.
52
46
24
21< 25
33
36
25
< 26
42
< 24
28
10
30
22
22
35
30
22
23
27
32
29
25
23
23
29
36
22
24
13
< 13
15
12 < 22
12 < 26
12 < 23
15 < 22
13 < 20
6 < 14
13< 17
17
15< 18
12 < 24
41
41
2
< 26
15 < 25
17
25
4
< 25
15
17 < 25
24
29
18 < 26
51
Hg <LOQ, ICPMS
Hg, LOQ ignored, ICPMS
Hg GALAHAD
41
26
26
16 < 24
25
29
23
< 25
24
< 26
24
< 24
22< 25
23
< 25
19 < 24
6
< 25
4
< 23
1
< 26
2
< 25
12 < 23
24
26
EQS
70
60
50
40
30
20
10
0
trout 1
trout 1 ground
trout 2
trout 2 ground
trout 3
trout 3 ground
trout 4
trout 4 ground
trout 1 skin
trout 1 skin ground
trout 2 skin
trout 2 skin ground
TH08-0002
TH08-0002
TH08-0004
TH08-0012
TH08-0014
TH08-0020
TH08-0021
TH08-0023
TH08-0068
TH08-0069
TH08-0070
TH08-0071
TH09-0050
TH09-0052
TH09-0053
TH09-0056
TH09-0058
TH09-0097
TH09-0098
TH09-0112
TH09-0113
TH09-0114
TH09-0115
TH09-0116
TH09-0117
TH09-0118
TH09-0119
NE08-0011
NE08-0012
NE08-0013
NE08-0014
NE08-0015
NE08-0017
NE08-0018
NE08-0019
NE08-0020
NE08-0001
NE08-0002
NE08-0003
NE08-0004
NE08-0005
NE08-0006
NE08-0007
NE08-0008
NE08-0009
NE08-0010
NE08-0021
NE08-0023
NE08-0024
NE08-0025
NE08-0026
NE08-0027
NE08-0028
NE08-0029
NE08-0030
GL09-0008
GL09-0009
GL09-0015
GL09-0016
GL09-0017
Hg [µg/Kg]
fresh weight
EU food limit = 500 or 1000 µg/kg for fish
68
trout
trout skin fr.
fillets fillets w. trout
w/o skin skin
fillets
Thames 08,
Cav er shamSonning
Thames Thames Thames 09, Molesey- Nene 08, Cogenoe
08, Bray- 09, BrayKingston
Boveney Boveney
Nene 08, Thrapston
Nene 08, Oundle Glen 09, P.
West
Figure 1: Mercury concentrations in individual roach samples from the UK fish tissue archive compared to the EQS of
20 µg/Kg. Samples from the same river are shown with the furthest upstream on the left. Also included are some store
bought trout fillets used for method development. LOQ= limit of quantification.
3.1. Mercury
Mercury exceeds the EQS of 20 µg/kg wet weight in about half of the samples analysed so far (Figure 1), but
all are about an order of magnitude below the limit for human consumption which is either 500 or 1000 µg/kg
depending on the type of fish (Commission Regulation (EC) No 1881/2006).
3.2. Hexachlorobenzene
Hexachlorobenzene was detectable in all fish but was always below the EQS of 10 µg/kg. The median
concentration in eel was 1.8 µg/kg and the maximum 6.4 µg/kg compared to 0.9 and 4.5 µg/kg in the roach
analysed so far.
3.3. Hexachlorobutadiene
Hexachlorobutadiene was not detected in the majority of samples (only roach analysed).
4. Conclusions
•
The samples stored in the Fish Tissue Archive are well suited for monitoring of priority substances,
especially those for which an EQS has already been set.
•
50% or more of roach in the Thames, Nene and Glen rivers for 2008 and 2009 exceed the 20 µg/L
EQS for Mercury. No fish contain levels of HCB, or HCBD which exceed the EU EQS levels in those
rivers sampled so far.
•
As the Fish Archive grows it will become possible to determine temporal and spatial trends of these
and other substances
5. References
[1] Eurochlor 2002. Euro Chlor Risk Assessment for the Marine Environment OSPARCOM Region - North
Sea. Hexachlorobutadiene.
[2] Eurochlor 2002. Euro Chlor Risk Assessment for the Marine Environment OSPARCOM Region - North
Sea. Hexachlorobenzene.
[3] Marsh TJ and Hannaford J. 2008. UK Hydrometric Register. Hydrological data UK series, Centre for
Ecology and Hydrology, UK, 210 pp.
Acknowledgement - The authors thank the Fisheries teams of the Environment Agency in England and
Wales for support in collecting the fish.
Environmental monitoring data: support for an effectiveness
assessment and a success control under REACH
Rita Groß1, Dirk Bunke1, Reinhard Joas2, SonjaBauer2, Yvonne Floredo2, Martin Führ3,
SilkeKleihauer3, Julian Schenten3, Lars Tietjen4, Lena Vierke4, Michael Neumann4
1
Öko-Institut e.V., P.O. Box 17 71, 79017 Freiburg, Germany
2
BiPRO GmbH, Grauertstr. 12, 81545 München, Germany
3
sofia/Hochschule Darmstadt, Haardtring 100, 64295 Darmstadt, Germany
4
Federal Environment Agency, Wörlitzer Platz 1, 06844 Dessau-Roßlau, Germany
Keywords: Environmental monitoring, REACH, effectiveness assessment, success control
1. Introduction
Environmental specimen banks support competent authorities in the assessment of chemicals. [1] Together
with national and EU-wide monitoring systems they can give information for the identification of new
substances of regulatory concern. Beyond the assessment of individual chemicals, monitoring data can also
be used to develop indicators to evaluate the implementation of different chemical regulations by providing
information on the impact of chemicals to human health and different compartiments of the environment.
By June 2012, the European Commission has to present a first evaluation of the effectiveness and success
of the REACH regulation. Human biomonitoring and environmental monitoring programmes may provide
valuable data to underpin this evaluation.In Germany as well as at EU-level, there is a variety of environmental monitoring activities which have been established for various reasons. [2] They are – in different
degrees – suitable to indicate the interim results of the implementation of REACH. A research project –
funded by the German Environment Protection Agency (UBA) – analyzed the available monitoring programs
(including environmental specimen banks ) to identify those who can contribute to evaluate the effectiveness
and success of different REACH task, starting from specific tasks related to specific parts of REACH (e.g.
registration) up to tasks related to the effectiveness evaluation of REACH as a whole.
In order to determine appropriate indicators and methods to evaluate the impact of the European chemicals
legislation REACH, a detailed and complete review on ongoing environmental monitoring activitieshas been
conducted. This overview included both the international and the European dimension as well as the
activities undertaken at Federal State level, those implemented in industrial companies (self-monitoring of
companies within or outside the business), within science and in other organizations. Furthermore, human
biomonitoring programmes on a German, European and international level as well were examined with
regard to aspects that can be transferred to environmental monitoring.
Considering all environmental compartments and the indirect exposure of humans, methods and indicators
of the exising programmes were checked and evaluated to find out whether they can be used for regulatory
activities under REACH. Furthermore, a guidance document for the use of environmental monitoring data
with a view to the evaluation of chemicals has been developed and presented. This guidance document
includes both the company’s self-monitoring as well as government monitoring and important regulatory focal
points under REACH. Exemplary case sheets on selected chemicals were documented for the monitoring / identification of substances of concern, thereby using the Environmental Specimen Bank (ESB).
3.1. Review and evaluation of existing monitoring programmes
A total of 313 programmes including databases, tool, networks, etc. were identified. Thereof, 13 monitoring
programmes were identified which are conducted by companies or industrial associations and 47 relate to
human biomonitoring activities. All identified programmes, except HBM programmes, were evaluated
according to a set of predefined evaluation criteria. The evaluation resulted in a number of 35 programmes
and databases as well as one additional company programme with high relevance for regulatory tasks of
authorities. Included are programmes/databases covering the Arctic area (1), the Baltic Sea region (2) as
well as the North-/East Atlantic (3). Seven pan-European (7) and several national programmes / databases
were identified as relevant and added to the final list. Among the programmes / databases selected at
national level, countries like Germany (18), Sweden (1), Switzerland (1), UK (1) and USA (1) are covered.
3.2. Methods and indicators
In order to structure the broad range of tasks for which environmental monitoring data can be used, tasks
were allocated to one of three groups:
I
REACH Regulation as a whole (related to the total impact of all chemicals on human health
and the environment);
II Specific REACH mechanisms (mostly related to specific substances);
III Success control (mostly related to specific substances).
Different types of indicators are required to fulfill the different tasks. The DPSIR approach, developed by the
EEA, has been used to clarify which type of indicators are needed for an effectiveness assessment of the
REACH regulation. [3], [4]. Up to now, neither the REACH regulation nor the Guidance Documents provide
guidance for the performance of environmental or human biomonitoring studies or for the use of data from
monitoring studies for the different REACH tasks. Consequently, there are no consistent or standardised
requirements (e.g. related to methods and indicators) regarding monitoring programs or data from such
programmes which could be used within the different REACH processes.
The evaluation of the selected relevant environmental monitoring programmes clearly shows that the majority of the programmes report concentration values of single substances in the different media. However, at
least the programmes monitoring organic substances such as PCBs, PAHs, HCHs and PCDD/F do not only
report concentrations of single substances but also concentrations of groups of substances/congeners (e.g.
∑HCH/kg) and/or toxicity equivalents (ng TEQ). The latter represent a first step toward aggregated values or
indicators. Few programmes/databases do report aggregated values for inorganic substances such as a
‘multi-metal index’. Based on the experiences from the assessment of existing monitoring data a practical
guide was developed how to use these data for the assessment of chemicals and the tasks of authorities
under REACH.
4. Conclusions
A large number of well-established environmental monitoring programmes including specimen banks is
already available. Currently, data from environmental monitoring are used, however, only to a very limited
extent to evaluate the effectiveness and success of different REACH task or to derive REACH indicators. A
guidance document developed within the framework of the project is intended to give the required instruction
for the use of environmental monitoring data under REACH. Existing programmes and environmental
specimen banks could be the starting point for the development of REACH state and impact indicators,
which are not available yet. Future challenges in regulatory risk assessment go beyond time trends of
individual substances. Effectiveness assessments of regulations such as REACH require indicators for the
total environmental burden due to the wide dispersive uses of a large number of substances – and its effect
on biodiversity. Environmental specimen banks can support such assessments and the elaboration of
adequate indicators.
5. References
[1] Federal Environment Agency. 2011. Environmental Specimen Bank. http://www.umweltprobenbank.de.
Accessed 29.11.2011
[2] European Environment Agency. 2007. Towards a European chemicals information system. EEA
Technical Report No 6/2007
[3] European Environment Agency. 2007. The DPSIR framework used by the EEA.
http://ia2dec.ew.eea.europa.eu/knowledge_base/Frameworks/doc101182
[4] Jenseit, W.; Bunke, D.; Rheinberger, U.; Öko-Institut e.V. in cooperation with FoBIG; DHI, INERIS. 2007.
REACH BASELINE Study. Commissioned by: EU Commission, Eurostat, Luxemburg
Acknowledgement – This project was commissioned by the Federal Environment Agency under the environmental research plan, ref. no. 3710 63 404, and financed by federal funds.
Temporal trends in dioxins and dl-PCBs from Baltic
herring (Clupea harengus)
Aroha Miller1, Jenny Hedman1, Peter Haglund2, Karin Wiberg3, Anders Bignert1
1
Department of Contaminant Research, Swedish Museum of Natural History, Frescativägen 40, 114 18
Stockholm, Sweden
2
Department of Chemistry, Umeå University, SE-90187, Umeå
3
Department of Aquatic Sciences and Assessment, Swedish University of Agricultural Sciences, 750 07,
Uppsala, Sweden
1. Introduction
The unintentional release of dioxins (dibenzo–p–dioxin (PCDD), dibenzofuran (PCDF), and dioxin like PCBs
(dl-PCB)) into the environment was at its peak prior to the 1980s [1]. Since then, extensive measures have
been taken to reduce dioxin emissions within the EU e.g., the Helsinki Convention (1974, 1992), the
Stockholm Convention on Persistent Organic Pollutants (POPs), the plan for integrated pollution prevention
and control (IPPC) [3]. However, dioxin concentrations continue to be higher than expected in Baltic fatty
fish. A large proportion of fish caught in the Baltic Sea exceed the limit for marketing fish within the EU [1] (4
pg WHO 05 -TEQ/g ww (∑PCDDs+PCDFs); 8 pg WHO 05 -TEQ g/ww (∑PCDDs+PCDFs+dlPCBs)) [2] and
thus, ongoing environmental monitoring occurs in many Baltic countries.
Within the Swedish National Monitoring Programme, Baltic herring, Clupea harengus, a fatty fish consumed
by humans, are analysed yearly for dioxin concentrations. As dioxins are lipophilic, fatty fish tend to
bioaccumulate greater concentrations of these contaminants compared to lean fish species. Here, yearly
monitoring data of herring from three sites on the east and one site on the west coast of Sweden are
presented. The aim of this work is to monitor and detect temporal trends and changes in dioxin
concentrations over time in a fish species of importance for human consumption and top marine predators.
Results are used to inform policy and regulation makers so dioxin guidelines can be adjusted accordingly.
2.1 Sites and sample matrice
Herring were sampled from four sites, including Harufjärden, located in the northern Bothnian Bay, sampled
since 1990; Ängskärsklubb, in the southern Bothnian Sea, sampled since 1979; Utlängan in the southern
Baltic Proper, sampled since 1988; and Fladen, on the Swedish west coast, sampled since 1990. Muscle
tissue from herring was sampled and sent to the Department of Chemistry, Umeå University, for analysis.
Herring were age determined using their scales. Analysed specimens were female, generally aged between
2 – 5 years. Total body weight and length, and reproductive phase were determined for each individual.
Extra individuals from each year are stored in the Swedish Museum of Natural History’s Environmental
Specimen Bank.
2.1 Chemical analyses
Analysis of dioxins and dl-PCBs were carried out at the Department of Chemistry, Umeå University. The
extraction method is described in [4], clean-up method in [5], and instrumental analysis in [6]. Between 1995
– 2000 a simplified method was used at three of the four sites, whereby only four congeners were analysed.
The other congener concentrations were estimated based on their ratios to the four analysed congeners.
This method is not considered comparable and as such these years are not included in the analyses.
2.1 Data treatment and statistical analyses
Data quality control was conducted. Geometric means were calculated for congener concentrations and
biological variables for each year. Congener profiles for each site were graphed. Biological variables were
graphed with TEQ values to examine relationships. The dominant congeners, as well as TEQ (PCDD/F and
dlPCBs) values, and year, were analysed using the non-parametric Mann-Kendall trend test to examine
changes over time at each site.
3.1. Congener profiles and biological variables
Congener patterns are similar between all four sites. At Ängskärsklubb, the total PCDD/F and dl-PCB values
are considerably higher than at the other sites. Herring from this site have often been older than those
sampled from the other sites, and the area is known for its local pulp/paper mill industry [7]. As high levels of
dioxins and dl-PCBs are still found in surface sediments near paper/pulp mill industrial areas [in 1], this
would probably explain the higher concentrations.
Six biological variables (fat %, fishing date, reproductive phase, weight, total length, and age) were plotted
alongside TEQ values for PCDD, PCDF and dl-PCB respectively. None of these variables show any
consistent relationship to the TEQ values at any site.
3.2. Temporal changes in dioxin concentrations
At Ängskärsklubb and Fladen, the dominant congeners and the TEQ (PCDD/F and dl-PCB) values showed
significant decreasing trends over time (Mann-Kendall, p<0.050). Ängskärsklubb has the longest temporal
data set of these four sites, and is located in a pulp/paper mill area, thus the decreases are likely a result of
processing changes to this industry [1] decreasing local dioxin emissions, even though overall
concentrations remain elevated compared to the other sites. Fladen is located on the Swedish west coast
(Kattegatt). Decreases here are likely a result of restrictions to dioxin emissions established within the EU [3].
Harufjärden and Utlängan both showed significant decreases in TCDD and TEQ (PCDD), with Utlängan also
showing a significant decrease in PeCDD and TEQ (dl-PCBs). However, all other dominant congeners and
TEQ values at these two sites were not significant (Table 1).
(TEQ 05- values)
PCDD
PCDF
dl-PCB
PCDD/F
Ängskärsklubb
-0.67, p<0.01
-0.53, p<0.01
-0.65, p<0.01
-0.61, p<0.01
Fladen
-0.43, p<0.03
-0.49, p<0.02
-0.49, p<0.03
-0.56, p<0.01
Harufjärden
-0.49, p<0.01
0.12, NS
0.03, NS
-0.12, NS
Utlängan
-0.39, p<0.04
-0.22, NS
-0.58, p<0.01
-0.33, NS
Table 1: Mann-Kendall trend and p-values (to 2 dp) for TEQ 05 PCDD, PCDF, dl-PCB at the four sites. Significant at p<0.050. NS
is non-significant.
4. Conclusions
Dominant PCDD congeners in herring are showing significant decreases in concentration over time at all
sites examined here, in line with earlier reports [in 1]. PCDFs are showing significant declines only at
Ängskärsklubb and Fladen, but not Harufjärden or Utlängan. Dl-PCBs show significant declines at
Ängskärsklubb, Fladen and Utlängan, but not Harufjärden. Reasons for this are still under investigation, but
monitoring remains the best way to inform policy and regulation makers.
5. References
[1] Wiberg K, McLachlan M, Jonsson P et al. 2009. Sources, transport, reservoirs and fate of dioxins, PCBs
and HCB in the Baltic sea environment. Swedish EPA. Report 5912. 145 p.
[2] EFSA 2010. Scientific Report of EFSA. Results of the monitoring of dioxins levels in food and feed. EFSA
Journal 8:1385-1421.
[3] HELCOM 2012. Development of a set of core indicators: Interim report of the HELCOM CORESET
project. PART B: descriptions of the indicators. Submitted.
[4] Wiberg K, Oehme M, Haglund P et al. 1998. Enantioselective analysis of organochlorine pesticides in
herring and seal from the Swedish marine environment. Mar. Poll. Bull. 36:345-353.
[5] Danielsson C, Wiberg K, Korytar P et al. 2005. Trace analysis of polychlorinated dibenzo-p-dioxins,
dibenzofurans and WHO polychlorinated biphenyls in food using comprehensive two-dimensional gas
chromatography with electron-capture detection. J. Chromatography A. 1086:61-70.
[6] Liljelind P, Soederstroem G, Hedman B et al. 2003. Method for Multiresidue Determination of
Halogenated Aromatics and PAHs in Combustion-Related Samples. Environ. Sci & Tech. 37:3680-3686.
[7] HELCOM 2010. Hazardous substances in the Baltic Sea – An integrated thematic assessment of
hazardous substances in the Baltic Sea. Balt. Sea. Envir. Proc. No. 120B.
Acknowledgement - The authors thank the Swedish Environmental Protection Agency for funding the
Swedish National Monitoring Programme, and the involvement of the BalticPOPS project and collaborators.
Title: Temporal trends in human exposure to fluorosurfactants and
related chemicals in two cities from Germany
Leo W.Y. Yeung, Shona Robinson, and Scott A. Mabury
Department of Chemistry, University of Toronto,
80 St George Street, Toronto, M5S 3H6, ON, Canada
E-mail: contact: [email protected]
______________________________________________________________________________
1.
Introduction
Fluorinated chemicals have been used in various industrial and daily applications for over 5 decades. Their unique oilrepellence and high surface activity make them excellent surface-protectors and surfactants. Perfluorooctanesulfonte
(PFOS) and perfluorooctanoate (PFOA) are two well-known fluorinated chemicals which, due to their persistency and
bioaccumulation potential, have restricted use in several countries. Recently, a class of commercial fluorinated
surfactants, the polyfluoroalkyl phosphoric acid diesters (diPAPs), has received growing attention. Unlike PFOS and
PFOA, diPAPs are direct consumer products, used in food contact paper, personal care products, cosmetics,
wetting/levelling agents. Human exposure to diPAPs may occur through use of diPAPs-containing products. In fact,
diPAPs have recently been observed in blood samples from the U.S [1,2]. In the present study, n = 420 university
student blood samples collected from two German cities, Münster and Halle, over a 30-year period (1981 – 2009) were
analyzed for
di-PAPs, a structurally-related surfactant from another manufacturer (i.e., Nethylperfluorooctanesulfonamidoethanol-based polyfluoroalkyl phosphoric acid diester (SAm-PAP)), and a series of
perfluoroalkyl acids (PFAAs: e.g., PFOS, PFOA, etc). The aims of the present study are to investigate if there are any
temporal trends of PFAAs in the blood samples from students living in these two German cities during the period of
1981 – 2009; to evaluate any spatial differences in those target analytes between the two cities; and to measure those
newly identified fluorinated compounds (i.e., DiPAPs) in those samples, and investigate any trends for those
compounds.
O
N
F F F F F F F F
F
F
S
F
F F F F F FO
O
N
OF F F F F F
F
S
F
F
F F F F F F F F
O
P
OH
O
O
x = 4, 6, 8, 10
y = x or x+2
If y=x, x:2 diPAP
Fig 1. Structures of DiPAP (left) and SAmPAP (right)
2.
Material and Methods
Human blood sera samples were archived from the German Environmental Specimen Bank (ESB)[3]. Sera samples
were stored at -80oC up to year 2005, and samples from 2005 onwards were stored at -150oC at the facilities. The sera
samples were shipped with dry ice and were stored at -20oC in the University of Toronto. A total of 420 sera samples
from the period of 1981-2009, from two cities in Germany: Münster (10 samples (5 male and 5 female) per year from
1981-2009 (no sample for 1994), for a total of 270 samples) and Halle (10 samples (5 male and 5 female) per year from
1995-2009, for a total of 150), were analyzed for a suite of fluorinated analytes. One – three mL of sera samples were
extracted by a modified ion-pair extraction method, as developed by Hansen et al. [4]. In brief, 0.5M
tetrabutylammonium sulfate (TBAS) was adjusted to pH 10 using 30% aqueous ammonia, instead of using sodium
buffer solution. Methyl tert butyl ether (MTBE) was used to extract the ion-pair. Sample extracts were evaporated under
a gentle stream of nitrogen and reconstituted with methanol for instrumental analysis. The samples were analyzed using
an Acquity UPLC (Waters) and Waters Xevo-TQ S MS-MS system in ESI negative mode. Chromatographic separation
was performed using a UPLC BEH C18 column (2.1 x 50 mm, 1.7 µm, 100Å, Waters). Analytes were quantified either
by using mass-labeled internal standards or by external calibration. Analytical standards were donated by Wellington
Laboratories (Guelph, ON).
Three extraction blanks using MilliQ water and three matrix blank (i.e. calf sera) were carried out for every batch of
extraction to evaluate any contamination introduced during the extraction process. Two samples of standard reference
material (SRM1957) non-fortified human blood sera from NIST were also extracted for every batch of extraction to
evaluate inter-day analytical variation. Most of the extraction and matrix blanks were below corresponding limits of
quantifications (LOQs ranged 2 – 50 pg/mL). The values of the SRM across several batches agreed to those certified
values, and the relative standard deviations across batches were less than 15%.
3.
Results and Discussion
Among the 420 samples analyzed here, the perfluorosulfonates (PFSAs: i.e., C6 (0.053 - 3.83 ng/g) and C8 (0.318 79.0 ng/g)), the perfluorooctanesulfonamidoacetates (N-MeFOSAA (<0.0031 - 8.11 ng/g), N-EtFOSAA (0.0058 - 9.00
ng/g), FOSAA (<0.0011 - 8.25 ng/g)), and the perfluorocarboxylates (PFCAs: i.e., C8 (0.176 - 31.7 ng/g), C9 (0.020 2.70 ng/g), C10 (0.020 - 0.880 ng/g), C11 (0.003 - 0.555 ng/g) were detected in over 80% of the samples. The longer
chain PFCAs (C12 (<0.0017 - 0.056 ng/g) and C14 (<0.0017 - 0.049 ng/g) PFCAs), however, were only detected in
20% of the samples. Temporal trends can be observed for some of the analytes: PFOS concentrations peaked in 1986
(around 30 ng/g) and reached a plateau before they began to decrease starting in the year 2000 until 2009 (around 4 ng/g)
in Münster samples (Fig 2). A similar decline in PFOS concentrations (1995-2009) was also observed for samples from
Halle. The temporal trend observed here for human PFOS contamination mirrors industrial production patterns, with the
post-2000 decline in PFOS sera concentrations occurring concurrently with the phase-out of PFOS and related
chemicals starting in 2000.
Halle n = 150
Münster n = 270
Concentration ng/g
Halle samples begin
Fig 2. PFOS concentrations in human blood sera from two cities (i.e. Halle and Munster) in Germany from 1981-2009.
A total of 320 samples were analyzed for DiPAP and SAmPAP, although quantifications had been done only for 6:2/6:2
and 8:2/8:2 DiPAPs due to the limited availability of analytical standards. Further confirmation using matrix matched
calibration curve and standard addition will be carried out for 4:2/4:2, 10:2/10:2, and SAmPAPs. The 6:2/6:2 (<0.00048
- 0.762 ng/g) and 8:2/8:2 (<0.0004 - 0.285 ng/g) diPAPs were detected in 46% and 32% of the samples, respectively.
No distinct was observed for 8:2/8:2 DiPAP, however, an increasing trend could be observed for 6:2/6:2 DiPAP after
year 2000.
4.
Reference
1
Lee H, Mabury S. A. 2011. A pilot survey of legacy and current commercial fluorinated chemicals in human sera from United States
donors in 2009. Environ Sci Technol 45: 8067-8074.
2
D’eon J C, Mabury SA. 2009. Observation of a commercial fluorinated material, the polyfluoroalkyl phosphoric acid diesters, in human
sera, wastewater treatment plant sludge, and paper fibers. Environ Sci Technol 43: 4589-4594.
3
Wiesmüller GA, Eckard R, Dobler L, Günel A, Oganowski M, Schrӧter-Kermani C, Schlüter C, Gies A, Kemper FH. 2007. The
environmental Specimen Bank for human tissues as part of the German Environmental Specimen Bank. Int J Hyg Environ. Health 210:
299-305.
4
Hansen KJ, Clemen LA, Ellefson ME, Johnson HO. 2001. Compound-specific, quantitative characterization of organic fluorochemicals in
biological matrixes. Environ Sci Technol 35: 766−770
Acknowledgements -The authors would like to thank Jan Koschorreck of the Federal Environmental Agency, Berlin,
Germany for the help in getting the samples from the German Environmental Specimen Bank and Wellington
Laboratories for donating the analytical standards for the project.
New challenges for environmental specimen bank applications
- banking for marine mammal health research
Paul R. Becker1, Rebecca S. Pugh1, Amanda J. Moors1, John R. Kucklick1, and Teri K.
Rowles2
1
National Institute of Standards and Technology, Analytical Chemistry Division, Hollings Marine Laboratory,
Charleston, South Carolina, USA
2
National Oceanic and Atmospheric Administration, National Marine Fisheries Service, Office of Protected
Species, Silver Spring, Maryland, USA
1. Introduction
In 1979, the National Institute of Standards and Technology (NIST) in collaboration with the U.S.
Environmental Protection Agency began a program of banking human liver specimens for contaminant trend
monitoring. The protocols, procedures, and specialized equipment developed for this program have provided
the basis for all subsequent banking components that NIST has instituted for other research and monitoring
programs over the last 33 years. NIST banking has emphasized long-term storage under cryogenic
conditions to minimize sample degradation over time, use of special materials and procedures for specimen
collection and handling, to minimize the possibility of introducing an artifact into the sample and thus biasing
the analytical results, and well documented and published standard protocols and procedures covering all
steps in sampling and banking including collection procedures, sample handling and processing, specimen
transport, bank specimen log-in and tracking procedures, and continuous recording of specimen storage
conditions over time.
Presently, the banking program of NIST is centered at its Marine Environmental Specimen Bank (Marine
ESB) that is located at the Hollings Marine Laboratory, Charleston, South Carolina, USA. As is suggested by
the formal name of the bank, the great majority of environmental specimens banked by NIST are from the
marine environment and the greatest number of banked specimens with the widest geographic coverage is
marine mammal tissues. The banking of marine mammal tissues began in 1992 as part of the National
Oceanic and Atmospheric Administration’s (NOAA’s) Marine Mammal Health and Stranding Response
Program (MMHSRP). This program began as a response to a specific mortality event and the lack of banked
specimens needed to address questions associated with the event.
2. The National Marine Mammal Tissue Bank
During 1987-1988, a massive die-off of dolphins occurred on the Atlantic coast of the US. Although a red-tide
algal toxin was implicated as a causative factor, high levels of PCBs and chlorinated pesticides were also
found in the dead animals and were proposed by some investigators as contributing to this event [1]. As
there were no baseline data on levels of contaminants in these populations of dolphins and banked tissue
specimens were not available for analysis, the question of the role of contaminants in this die-off was not
resolved. Based on this event, the MMHSRP began with one component being the National Marine Mammal
Tissue Bank (NMMTB) that is maintained by NIST as part of its Marine ESB. Marine mammal tissues are
collected and banked from fresh-stranded marine mammals (single animal strandings and mass strandings),
animals that are taken accidentally in commercial fishing operations and, in Alaska, from Alaska Native
subsistence harvests. From the beginning the NMMTB was designed to establish a resource of marine
mammal tissue samples that could be used for chemical analysis to determine exposure of the animals to
bioaccumulative contaminants and retrospective research on exposure history of populations to emerging
contaminants.
3. Bank Expansion for Health Assessments
Although originally intended for contaminant measurements, banked tissue samples have proven useful for
other kinds of marine mammal investigations. Vitamin D measurements on banked tissues from ringed seals,
bearded seals, and beluga whales were used to evaluate nutrient deficiencies in captive polar bears (ringed
and bearded seals and beluga whales are natural prey of polar bears in the Arctic) [2]. Samples from the
bank have been used to determine the degree of genetic separation of beluga whale populations in Alaska
[3]. Genetic information has been useful for interpreting differences in contaminant concentrations in these
populations [4, 5, 6]. Information on position of individual animals in the food web is also important in
interpreting both geographic and temporal patterns of contaminant exposure in these animals; thus, C and N
isotope analysis of these banked tissues have proven to be particularly useful.
In 2003, NIST began collaborating with organizations and researchers on the U.S. east coast in studies to
assess the health of bottlenose dolphins, Tursiops truncates [7]. These studies are being conducted on eight
populations of this species and involve the periodic capture and release of live animals during which health
measurements and samples are collected for determining incidence of disease, overall health conditions,
and exposure to biotoxins. As part of these studies, blood (plasma, serum, and whole blood), milk, and
skin/blubber biopsy samples are also routinely collected using NIST protocols and banked in the Marine ESB
primarily for future contaminant research. Supplemental dart biopsy samples of skin plus blubber are also
collected and incorporated into these studies. This health assessment and banking approach is being
expanded to other marine mammal species and to other regions of the US. It is also being used in a study of
response of dolphins to the Gulf of Mexico Deepwater Horizon Oil Spill as part of a natural resources
damage assessment program. The Marine ESB has become an important resource as a chain-of-custody
biorepository for marine mammal samples collected as part of the spill response. Recently, banking for
“disaster response” has been an ongoing item of discussion among some members of the specimen banking
community.
4. Future Direction
Based on the experience gained through the bottlenose dolphin health assessment project and the ongoing
response to the recent Gulf of Mexico oil spill incident, NIST is working with NOAA and its collaborating
partners to establish the marine mammal specimen bank as a major resource of samples for integrative
animal health research. This expansion will be emphasizing the banking of specimens for wildlife disease
studies, determining exposure to biotoxins, and developing health biomarkers, in addition to determining
contaminant exposures. This expansion is requiring additional banking expertise, identification of additional
kinds of matrices to bank, and the storage conditions required for long-term banking. These and other
aspects of the expansion will be discussed.
5. References
[1] Lillestolen TI, Foster N, Wise SA. 1993. Development of the National Marine Mammal Tissue Bank. Sci.
Total Environ. 139/140:97-107.
[2]Kenny DE, O’Hara TM, Chen TC, Lu Z, Tian X, Holick MF. 2004. Vitamin D content in Alaskan Arctic
zooplankton, fishes, and marine mammals. Zoo Biol. 23:33-43.
[3] O’Corry-Crowe GM, Suydam RS, Rosenberg A, Frost KJ, Dizon AE. 1997. Phylogeography, population
structure and dispersal patterns of the beluga whale, Delphinapterus leucas, in the western Nearctic by
mitochondrial DNA. Mol. Ecol. 6:955-970.
[4] Krahn MM, Burrows DG, Stein JE, Becker PR, Schantz MM, Muir DCG, O’Hara TM, Rowles T. 1999.
White whales (Delphinapterus leucas) from three Alaskan stock: Concentrations and patterns of persistent
organochlorine contaminants in blubber. J. Cetacean Res. Manage. 1(3):239-249.
[5] Becker PR, Krahn MM, Mackey EA, Demiralp R, Schantz MM, Epstein M, Donais MK, Porter B, Muir
DCG, Wise SA. 2000. Concentrations of polychlorinated biphenyls (PCBs), chlorinated pesticides, and heavy
metals and other elements in tissues of beluga whales (Delphinapterus leucas) from Cook Inlet, Alaska. Mar.
Fish. Rev. 62(3):81-98.
[6] Reiner JL, O’Connell SG, Moors AJ, Kucklick JR, Becker PR, Keller, JM. 2011. Spatial and temporal
trends of perfluorinated compounds in beluga whales (Delphinapterus leucas) from Alaska. Environ. Sci.
Tech. 45:8129–8136.
[7] Kucklick J, Pugh R, Becker P, Keller J, Day R, Yordy J, Moors A, Christopher S, Bryan C, Schwacke L,
Goetz C, Wells R, Balmer B, Hohn A, Rowles T. 2010. Specimen Banking for Marine Animal Health
Assessment. In, Interdisciplinary Studies on Environmental Chemistry Vol 4: Environmental Specimen Bank:
Exploring Possibility of Setting-up ESBs in Developing Countries. Eds., Tomohiko Isobe, Kei Nomiyama,
Annamalai Subramanian and Shinsuke Tanabe. TerraPub, Tokyo, Japan. pp 15-23.
Midpoint and endpoint indicators for global scale terrestrial
acidification: a dilemma for decision-making
Louise Deschênes1, Pierre-Olivier Roy1, Ligia B. Azevedo2, Mark J. Huijbregts2, Rosalie
van Zelm2 and Manuele Margni3
1
CIRAIG, Chemical Engineering Department, École Polytechnique de Montréal, P.O. Box 6079, Montréal,
Québec, H3C 3A7, Canada
2
Radboud University Nijmegen, Institute for Water and Wetland Research, Department of Environmental
Science, P.O. Box 9010, 6500 GL, Nijmegen, The Netherlands
3
CIRAIG, Mathematical and Industrial Engineering Department, École Polytechnique de Montréal, P.O. Box
6079, Montréal, Québec, H3C 3A7, Canada
1. Introduction
Life cycle assessment (LCA) uses characterization factors (CFs) to evaluate potential impacts. A CF is a
mathematical representation of the cause effect chain of an impact category. They are evaluated somewhere
along the cause effect chain and represent a midpoint indicator whereas endpoint indicators are obtained
when modelling the entire chain up to the consequences on areas of protection. While an endpoint CF is
considered to be more environmentally relevant, it is nevertheless considered more uncertain as it is
modelled further up the cause effect chain. For some impact categories, such as climate change, the link
between midpoint and endpoint indicators is based on a proportionality factor, which adds relevance, but
does not add further discrimination between emitted pollutants [1]. However, for other impact categories,
such as acidification or (eco) toxicity, all three modeling steps: fate, sensitivity/receptor and effect, add
discrimination between chemicals in addition to relevance. Whether to make a decision based on a midpoint
or endpoint indicator is an old debate. This paper aims at contributing to this discussion by showing for the
first time, the lack of correlation between midpoint and endpoint indicators for acidification, and in light of
these results, discuss the relevance of using midpoint and/or endpoint indicators and the implication in
decision making.
Terrestrial acidification is the process in which atmospheric deposition, related to emissions of SO 2 /SO 4 ,
NO x and NH 3 , cause changes in soil acidity that can eventually harm terrestrial flora. Midpoint CFs [(mol × L
-1
1
) × ha × (kg S or N emitted ) × yr] results from the multiplication and subsequent summation of an atmospheric
-1
fate factor (FF) and a soil sensitivity factor (SF) (Eq. 1). Endpoint CFs [ha × (kg S or N emitted ) × yr] were
obtained by adding a biome specific vegetation effect factor (EF) to the midpoint CF. CFs were obtained at a
o
o
worldwide 2 x2.5 (latitude x longitude).
int
int
CFiendpo
= CFi ,midpo
× EFi , j =
p
,p
∑ FF
i, j , p
× SFi , j , p × EFi , j
(1)
j
-1
-1
FF [(kg S or N deposited ) × yr × (kg S or N emitted ) × yr] describes the atmospheric impact pathway from the
-1
emission location i of pollutant p to the corresponding deposition in the receiving soil j. SF [(mol × L ) × (kg S
-1
or N deposited ) × yr × ha] translates the change of pH in the soil according to a change in emission. A marginal
-1 -1
change of +10% of the emission was used. Changes in the vegetation EF [(mol × L ) ] is a function of biome
specific coefficients. The effect factor takes into account an “optimal pH” value; where effects begin to occur.
For pH values higher than the “optimal pH”, EF is equal to 0. The midpoint CF quantifies the concentration of
H+ ions (evaluated through pH) and assumes that any change in pH will have consequences on the
receiving environment while the endpoint CF evaluates the decrease in species richness of the biome
vegetation related to this change in pH.
Figure 1 shows results of the midpoint and the endpoint assessment for an emission of NO x . As can be
observed, the highest midpoint impacts are located in the eastern parts of North America and central Asia.
However, the highest endpoint impacts occur in central Asia, central Africa and southern Europe. This shows
the importance of introducing a further modelling step to highlight biomes having vegetation with lower
resistance to pH change, such as found in the (Sub)tropical grasslands, savannas and shrublands biome
(mostly located in central Asia). Direct comparison of the NO x midpoint and endpoint CFs, based on
2
geographical emission location, shows a medium-low correlation (R = 0.60) between the midpoint and
-1
-6
endpoint indicators. As an example, a midpoint equal to 1×10 [ha × (kg S or N emitted ) × yr] has a variety of
-4
-3
-1
concordant endpoint values which can be as low as 6.6×10 or as high as 5.8×10 [(mol × L ) × ha × (kg S
-1
or N emitted ) × yr]. In other words, the conclusions of an LCA study could be inverted (i.e. product A
preferable to product B or vice versa) whether a midpoint or endpoint approach is selected. This raises the
question about which indicator should be promoted to support decision making.
A
B
C
Midpoint
Endpoint
Figure 1: Midpoint (A) and endpoint (B) CF results for NO x . (C) Boxplot analysis of the midpoint and endpoint 2x2.5 CFs: The
box is delimited by the lower quartile and upper quartile of the CF distribution values, the whiskers show the extent of the rest
of the data and outliers (red crosses) are data with values beyond the ends of the whiskers.
The boxplot analysis, from Figure 1 C, infers that endpoint assessment could increase the spatial
discrimination between emission locations since, contrary to midpoint CFs, outliers extend up to an order of
magnitude over the upper whisker. Of course, endpoint assessment, by definition, is more environmentally
relevant as it provides information that directly matters to society [2] and therefore is more accurate to what
we want to protect. However, it does bear more uncertainty. In this specific case, the endpoint CFs assumes
that EF=0 if pH ≥ pH optimum. However, it would be hard to believe that a receiving environment within a
biome with an optimum of 6.9, who varies, for example, from pH 10 to 7 does not suffer any impacts. This
abrupt change in pH would have been severely evaluated with the midpoint approach, which quantifies the
concentration of H+ ions. However, this indicator would also consider areas with naturally low pHs, as
significant, although it would only be marginally affected.
4. Conclusions
This work raises questions about which indicator is more relevant and/or should be used since a further
modelling step from midpoint to endpoint is no longer synonym of proportionality between the elementary
flows and thus LCA conclusions can be inverted. The trade-off between precision and accuracy is still far
from being resolved, but at least science comes closer in offering the decision maker the opportunity to
perform a sensitivity analysis to clarify whether or not they are in a situation where the conclusion of a study
might be inverted by the choosing an indicator over another.
5. References
[1] Goedkoop, M., Heijungs, R., Huijbregts, M., De Schryver, A., Struijs, J., van Zelm, R. 2009. ReCiPe 2008:
A life cycle impact assessment method which comprises harmonised category indicators at the midpoint
and the endpoint level: Report 1: Characterization. 1-132
[2] van Zelm, R., Huijbregts, M., van Jaarsveld, H. A., Jan Reinds, G., de Zwart, D., Struijs, J., van de Meent,
D., 2007. Time horizon dependant characterization factors for acidification in Life-Cycle Assessment
based on forest plant species occurence in Europe. Environmental science & technology, 41(3), 922-927.
Acknowledgements – We acknowledge the financial support of the industrial partners of the International
Chair in Life Cycle Assessment and the European Commision (through the LC-Impact project).
Global life cycle impact assessment on marine eutrophication
Henrik Fred Larsen1, Gao Yang2
1
2
QSA, DTU Management Engineering, Technical University of Denmark (DTU), Lyngby, Denmark
School of Mechanical & Automobile Engineering, Hefei University of Technology, Hefei, Anhui, China
Email contact: [email protected]
1. Introduction
As part of the ongoing EU FP7 project LC-Impact (www.lc-impact.eu) new life cycle impact assessment
(LCIA) methods are going to be developed and tested on industry cases. Among the life cycle assessment
(LCA) impact categories in focus are aquatic eutrophication. As related to especially the marine environment
very few and restricted attempts have yet been done on trying to include eutrophication in LCA. The aim of
LC-Impact is to develop both a global and a spatial (and temporal) differentiated model, as both central fate
processes, sensitivities of receiving environments (e.g. differences in limiting nutrient and variations in this
over the year) and the resulting damage can show important spatial variations. Both midpoint and endpoint
(damage) modelling are to be included and the aim is to base the damage modelling on dose-response
curves expressing the correlation between the (increase in) nutrient concentration and the potentially
affected fraction of species in the marine ecosystem. This paper presents the first draft on the midpoint
model for global marine eutrophication due to nitrogen emissions.
2. Methodology
The fate model used here is highly based on recent model work by Bouwman and co-workers [1; 2; 3 and
more], especially the results from the Integrated Model to Assess Global Environment, IMAGE 2.4 [1], and
research by Wollheim and co-workers regarding the fate of N in river basin systems [4; 5]. The IMAGE model
use information at 0.5° by 0.5° spatial resolution (grid cells) at annual time steps but have been aggregated
to country level [6]. Regarding the distribution (deposition) of ammonia and NOx emissions to air results from
the recently developed atmospheric model by Roel et al. [7] is used. By adaption to the framework of life
cycle impact assessment a LCIA fate model on anthropogenic nitrogen emissions reaching coastal marine
water is developed. The fate model is done in both a global version and a spatially differentiated version
based on a country level. Only the global model is presented here.
Manure N
Fertilizer N
Agricultual
N-fixation
Atmosph.
N-depos.
101 Tg TN/year
82 Tg TN/year
30 Tg TN/year
35 Tg TN/year
Harvest/
grazing
Total terrestrial N input
Ammonia
volatilisation
93 Tg TN/year
248 Tg TN/year
33 Tg TN/year
Denitrification
in top soil
74 Tg TN/year
Leaching
49 Tg TN/year
Denitrification
in ground
water
31 Tg TN/year
Total aquatic (river) input
17 Tg TN/year
Figure 1: Global agricultural fate model for N-input (year 2000) based on [1; 6; 8]. Only input with corresponding fate factors on
denitrification and leaching included.
The result of modelling the fate of agricultural total-N input for the year 2000 is shown in Figure 1. The part
shown comprises the fate of total-N (modelled as nitrate) until it reaches surface freshwater on its way to the
sea. If the models for distribution of air emissions [7] and fate of tot-N in surface fresh water before river
mouth [4] are included in the calculations the balance shown in Table1 is obtained.
Table 2: Full balance for global total N-emissions to marine water (in percentages, %)
Tot-N emission source
Loss
to
stratosphere
Removed
by
crops/ grazing
in agriculture
Denitrified in
topsoil
and
ground water
Lost
in
surface
fresh water
Ending up
in marine
water
Fertilizer in agriculture
0,0666
38,7
48,3
4,35
8,64
Manure in agriculture
0,109
35,8
47,9
4,42
11,8
N-fixation in agriculture
0
43,1
48,9
4,24
3,81
Air emissions (NH 3 )
0,495
10,0
44,7
5,03
39,8
Air emissions (NO x )
9,53
8,42
37,5
4,23
40,3
Water emissions (SW*)
0
0
0
52,7
47,3
* Point source emissions to surface water (direct sewage water (SW) emissions or emissions from waste water treatment systems)
4. Conclusions
The figures in the last column of Table 1 are used to calculate the potential increase in the total-N
concentration in an average global marine ecoregion for each emission source after emission of 1 ton per
year. By reference to the Redfield ratio these results are used as global mid-point characterization factors.
5. References
[1] Bouwman AF, Beusen AHW, Billen G. 2009. Human alteration of the global nitrogen and phosphorus soil
balances for the period 1970–2050. Global Biogeochem. Cycles, 23, GB0A04, doi:10.1029/2009GB003576.
[2] Bouwman AF, Van Drecht G, Knoop JM, Beusen AHW, Meinardi CR. 2005. Exploring changes in river
nitrogen
export
to
the
world’s
oceans.
Global
Biogeochem.
Cycles,
19,
GB1002,
doi:10.1029/2004GB002314.
[3] Van Drecht G, Bouwman AF, Knoop JM, Beusen AHW, Meinardi CR. 2003. Global modeling of the fate of
nitrogen from point and nonpoint sources in soils, groundwater, and surface water. Global Biogeochem.
Cycles, 17(4), 1115, doi:10.1029/2003GB002060.
[4] Wollheim W M, Vörösmarty CJ, Bouwman AF, Green P, Harrison J, Linder E, Peterson BJ, Seitzinger SP,
Syvitski JPM. 2008. Global N removal by freshwater aquatic systems using a spatially distributed, withinbasin approach, Global Biogeochem. Cycles, 22, GB2026, doi:10.1029/2007GB002963.
[5] Wollheim W M, Vörösmarty CJ, Peterson BJ, Seitzinger SP, Hopkinson CS. 2006. Relationship between
river size and nutrient removal, Geophys. Res. Lett., 33, L06410, doi:10.1029/2006GL025845.
[6] Bouwman AF. 2011. Personal communication with AF Bouwmann. Excel file: “nutdata2000_out”, version
1.2.3, 7 May 2007.September 22, 2011.
[7] Roy P-O, Huijbregts M, Deschênes L, Margni M. 2011. Spatially-differentiated atmospheric sourcereceptor relationships for nitrogen oxides, sulfur oxides, and ammonia emissions at the global scale for life
cycle impact assessment, submitted to Atmospheric Environment September 22, 2011.
[8] Bouwman AF. 2011. Personal communication with AF Bouwmann. DAT-files with leaching factors,
denitrification factors and more in 0.5° by 0.5° spatial resolution. Calculated for the 1995 situation by IMAGE
2.2 as described in Bouwman et al. 2005.
Acknowledgement - This study is part of the EU LC-Impact project: Development and application of
environmental Life Cycle Impact assessment Methods for imProved sustAinability Characterisation of
Technologies (Grant agreement No.: 243827 – LC-IMPACT), which is financially supported by grants
obtained from the EU Commission within the SEVENTH FRAMEWORK PROGRAMME ENVIRONMENT
ENV.2009.3.3.2.1: Improved Life Cycle Impact Assessment methods (LCIA) for better sustainability
assessment of technologies.
Accounting for greenhouse-gas emissions in LCA from the
degradation of chemicals in the environment
Ivan Muñoz, Giles Rigarlsford, Llorenç Milà i Canals, Henry King
1
Safety and Environmental Assurance Centre, Unilever, Sharnbrook MK44 1LQ, UK
1. Introduction
In Life Cycle Assessment (LCA) and carbon footprinting, GHG emissions need to be inventoried across all
life cycle stages. In product disposal, models have been developed for waste treatment processes such as
landfilling and wastewater treatment. Nevertheless, there are cases in which waste materials enter the
environment. In these cases, degradation may take place in the receiving environmental compartments (air,
soil, water, sediments), and chemicals containing carbon and nitrogen will contribute to the formation of GHG
emissions, through the formation of carbon dioxide (CO 2 ), methane (CH 4 ), and nitrous oxide (N 2 O) (Fig. 1).
In practice, these GHG emissions are seldom accounted for in LCA studies. In this work, we address this
methodological gap by providing a method for LCA to account for GHG emissions produced by the
degradation of chemicals in the environment.
2.1. Method description
The method is described as an impact assessment method, and consists of a set of equations to quantify
emissions of CO 2 , CH 4 and N 2 O for chemicals entering the air, soil or water compartments of a generic
modelled environment. These emissions are converted by means of Global Warming Potentials (GWP) to an
overall GWP from degradation (GWP deg ), measured in CO 2 -eq per kg of chemical released. Chemical
partitioning in the environment is determined using a level III fugacity model [1], and the following
assumptions are made: 1) air and soil are considered as aerobic compartments, 2) water and sediments are
partly aerobic and partly anaerobic with the extent of anaerobic conditions being modelled using available
literature data, and 3) CO 2 emissions are linked to aerobic conditions, CH 4 to anaerobic conditions, and N 2 O
to both.
A time horizon of 100 years is taken for the accounting of degradation emissions. There is no specific
allowance for delayed emission effects, hence degradation and its derived emissions are assumed to occur
at time 0. If a chemical is stable enough to remain intact after 100 years in the environment, then no GHG
emissions are attributed to its release.
The method proposes several consistent carbon accounting rules. Biogenic CO 2 can be considered neutral
(Global Warming Potential, GWP = 0) or not, but the former requires that CO 2 sequestration (fraction of
biogenic carbon in chemical that is not degraded after 100 years) be attributed a GWP of -1. Finally, the
GWP for CH 4 is proposed to be adjusted to account for its oxidation to CO 2 in the atmosphere, in order to
keep consistency with the IPCC guidelines for national GHG inventories [2].
2.2. Application to a group of chemicals
The proposed method was applied to a set of nine organic chemicals, namely ethanol (CAS 64-17-5),
propane (CAS 74-98-6), a generic enzyme (no CAS available), the surfactant linear alkylbenzene sulfonate
(LAS, CAS 25155-30-0), the fluorescent whitening agent 1 (FWA-1, CAS 16090-02-01), melamine (108-781), and the pesticides diuron (CAS 330-54-1) and pendimethalin (CAS 40487-42-1).
The calculated GWP deg values were compared to the GWP associated to producing these chemicals (cradle
to gate boundaries), using mainly data from the ecoinvent database [3]. The GWP deg of these chemicals
when released to water were compared to the expected GWP when these chemicals are subject to
treatment in a wastewater treatment plant (WWTP), including the energy use in the latter.
The magnitude of GWP deg for the set of nine chemicals ranges from 0.1 to 4.6 kg CO 2 -eq/kg chemical
released (Fig. 2). The lowest value corresponds to biogenic ethanol released to air, and this is caused
because the main breakdown product from degradation is biogenic CO 2 which is considered neutral. On the
other hand the highest emissions correspond to degradation of propane released to water (not shown in Fig.
2), due to the potential CH 4 emissions from the fraction of water compartment under anaerobic conditions.
However water is an unlikely release compartment for propane, given its high volatility. The GWP deg values
are sensitive to the initial release compartment, but values for the air and soil compartments are very similar,
due to the fact that both are assumed to be strictly aerobic.
The GWP deg of chemicals released to water is compared to the emissions expected for these chemicals
when water is treated in a WWTP prior to its discharge to the environment. Results suggest that the overall
emissions are similar: when a chemical degrades in a WWTP the modelled GHG emissions are expected to
be lower than if this process takes places in the environment, as in the latter a higher probability of CH 4
formation exists. On the other hand, the WWTP process involves GHG emissions from energy use, which
make the two scenarios appear similar overall.
To understand the relative relevance of GHG emissions from the degradation of chemicals in the
environment, they were compared to those associated with producing the chemicals (cradle to gate
boundaries). For each chemical, a likely release compartment (air, soil, water) and direct discharge into the
environment were assumed. The results of this comparison show that GWP deg is in many cases in the same
order of magnitude as cradle to gate emissions, and for three chemicals (ethanol, propane, LAS) they are
even higher.
Currently these emissions are not taken into account in LCA studies, and yet missing them can be compared
to omit combustion emissions in a cradle to grave LCA of fuels. Exclusion of such emissions in cases where
the chemical is emitted to air after use leads to a clear underestimation of the life cycle GHG emissions
associated with the chemical or the product using that chemical. The same applies when chemicals enter the
environment via water without previous treatment in a WWTP, a common situation in developing countries.
Figure 1: Current and proposed scopes of GHG accounting.
Released to water
CO2
CH4
Diuron
Pendimeth
alin
Melamine
Enzyme
FWA-1
Propane
Ethanol
(fossil)
Released to air
LAS
N2O
4.0
3.5
3.0
2.5
2.0
1.5
1.0
0.5
0.0
Ethanol
(biogenic)
GWPdeg (kg CO2-eq kg-1)
The main limitations of this method are: 1) the potential lack of high-quality chemical specific data to model
the environmental partitioning of chemicals, and 2) it has been designed to describe a generic environment
and thus it might not be appropriate to describe local conditions.
Released to soil
Figure 2: GWP deg per kg for assessed chemicals.
4. Conclusions
The presented method allows to account for GHG emissions that up to date have been omitted in LCA
studies and carbon footprint standards, and provides consistent carbon accounting rules, including a
proposal to increase the GWP of CH 4 to account for its oxidation to CO 2 in the atmosphere. The case study
on nine organic chemicals shows that these emissions can be important in the life cycle of such chemicals,
and therefore also in the life cycle of products and services using them. Such an omission can be crucial
when comparing fossil-based with bio-based chemicals, as the source of carbon greatly determines the GHG
impact associated with end-of-life emissions.
5. References
[1] Mackay, D. (1991) Multimedia Environmental Models: The Fugacity Approach, pp 67-183. Lewis
Publishers/CRC Press: Boca Raton, FL USA.
[2] Gillenwater M, Saarinen K, Ajavon ALN, Smith K (2006) Precursors and indirect emissions. In: Eggleston
HS, Buendia L, Miwa K, Ngara T, Tanabe K. (eds.) IPCC Guidelines for National Greenhouse Gas
Inventories. IGES, Japan. Vol 1, Chapter 7.
[3] Swiss Centre for Life Cycle Inventories. 2011. http://www.ecoinvent.com/ (accessed 11/11/2011).
Water Footprint and Life Cycle Assessment frameworks:
synergies and hurdles
Manuele Margni1, Anne-Marie Boulay1, Sébastien Humbert2 and Cécile Bulle1
1
CIRAIG, Department of Chemical Engineering, P.O. Box 6079, École Polytechnique de Montréal (Qc), Canada H3C 3A7
2
Quantis, Lausanne, Switzerland
1. Introduction
Water footprinting has come a long way in the past 10 years, starting with simple water volume inventories,
to scarcity assessment, to damage oriented life cycle impact assessment (LCIA) modelling of several impact
1
pathways up to endoints . With this increase of methodological developments and applications, the scientific
and industrial communities are now debating about the ultimate meaning(s) of the water footprint indicator.
This paper aims to evaluate the synergies and hurdles between the so far developed water footprint
2
framework(s) and approaches, with the traditional LCIA midpoint-endpoint framework . We therefore propose
to demystify: a) Standalone versus full life cycle assessment (LCA) water footprint; b) Single indicator versus
multiple indicators; c) Midpoint versus endpoint assessment; d) Impacts of water use versus impacts on the
water resource and e) Availability of water resource versus water pollution. This discussion is illustrated with
a simplified example of aluminium production, where available methods are compared and synergies and
contradictions are presented.
1
The extensive review by Kounina et al , identified 8 inventory methods, 4 general so-call midpoint method, 3
specific midpoint, one for each of the areas of protection (human health, ecosystems and resources), and 11
specific endpoint methods (4 for human health, 5 for ecosystems and 2 for resources) addressing water use
3
that can potentially be included within LCA following the framework presented by Bayart et al . However,
these solely assess the potential impacts related to the availability of the water resource (impacts from water
use). Potential impacts from pollution of the resource water (impacts on the water resource) are considered
by the traditional LCIA models of, for example, aquatic ecotoxicity, eutrophisation, thermal pollution and
acidification. Despite the fact these are two distinct concepts, we can demonstrate that both lead to common
endpoints. To do so the different water footprint methods are grouped into three aggregation level: I) a Water
Availability Footprint (WAF), a standalone method which address availability issues from water use, including
lowered availability from pollution (but no other impact related to water pollution), II) a Water Footprint
Assessment (WFA) as standalone, which includes all impacts of an activity on the water resource or III) as
part of an LCA methodology. The WFA models the potential impacts along all possible impact pathways up
to the respective endpoints as shown by Figure 1.
Stand alone
Stand alone or as
part of complete LCA
Goal and Scope
Inventory I
Inventory II
Water Availability Footprint
(WAF)
Water Footprint Assessment (WFA)
Preliminary stress
assessment of water
consumption and
degradation
Modeling of impacts on the water
resource using all available models,
(eutrophisation, ecotox, availability
for human health, etc...)
ex: Veolia, Boulay et al,
Ridout and Pfister, WFN
sustainability
Water
Footprint
Indicator
Water Footprint Profile
(midpoint or endpoint)
Human
Health
WF
Ecosystems
WF
Weighting
Water
Footprint
Assessment
Result
Figure 1: Water Footprint definition framework
Resources
WF
The water footprint results for the three aggregation levels are calculated for an illustrative example, which
accounts for the five main process stages of primary aluminium production (bauxite mining, alumina
production, anode production, electrolysis and ingot casting). Available data on water withdrawal and
releases, and contaminant emissions into air, water and soil are used to assess impacts from water use and
4,5
impacts on the water resource. Level I: To calculate the WAF, Boulay et al is used, thereby assessing
water availability issues from water consumption and degradation. Level II: The WFA consists of a profile of
three damage categories: human health, ecosystem quality and resources. The WFA-human health is
5
calculated applying Boulay et al for the assessment of human health impacts from water use and it is further
complemented with human health impacts related to water pollution from ionising radiation and human
toxicity (but solely impacts through “aquatic routes of exposure” such as ingestion of water and fish).
Similarly, WFA-ecosystem are calculated combining impacts from water use on ecosystem with impacts on
the resource water from aquatic pollution, such as aquatic eutrophication, aquatic ecotoxicity, aquatic
ionising radiation, aquatic thermal pollution and aquatic acidification, as well as other impacts influencing
water recharge and filtration related to land use . Level III: Level II WFA profile is complemented with other
types of impact to obtain a comprehensive LCIA profile.
3. Results
Results are provided for the three levels: WAF, WFA and LCIA along the five different stages of aluminium
production. The results of this illustrative example show that the WFA profile is fully compatible to the more
comprehensive LCA profile, the impact on water resource being a fraction of each LCIA damages on human
health, ecosystem quality and resources. The WFA profile, can be further disaggregated in such a way that
impacts from water use are distinguished from the ones generated by other water quality-related impact
pathways, putting in perspective the relative contribution of each impact pathway to the overall impacts on
the water resource within each area of protection.
4. Discussion and Conclusions
This paper presents an attempt to integrate several water footprint methods and concepts within the LCIA
framework articulating the concepts and respective methods into two distinct water footprint types: the WAF
and WFA, both serving distinct purposes. A simplified indicator related to scarcity (WAF) seems to be the
most straightforward way to answer the need clearly expressed by the industry asking for a simple, easy to
communicate, single number However, this is made at the cost of a less relevant and comprehensive scope
when addressing the impact on the water resources. The second option given by the WFA, allows including
impacts from pollution on the resource water. It offers more environmentally relevant water footprint
indicators, by modelling the most important environmental mechanism, including chemical fate, and
pathways up to endpoints consistent with the one traditionally used in LCIA. The increased relevance,
however, is obtained at the expenses of an increased uncertainty and sophistication of the water footprint
results through a multi indicator profile. Last but not least, taking advantage of the current scientific
developments of LCIA methodologies, the WFA can be fully integrated in a LCIA profile with the advantage
of covering a comprehensive scope and therefore reduce the risk of burden shifting when only focusing on a
WFA profile or even more a single WAF.
5. References
(1)
Kounina, A.; Margni, M.; Humbert, S. Review of methods addressing water in a life cycle perspective
International Journal of Life Cycle Assessment 2011, in progress.
(2)
Jolliet, O., R. Müller-Wenk, et al. (2004). "The LCIA Midpoint-damage Framework of the
UNEP/SETAC Life Cycle Initiative." International Journal of Life Cycle Assessment 9(6): 394-404
(3)Bayart, J.-B.; Margni, M.; Bulle, C.; Deschênes, L.; Pfister, S.; Koehler, A.; Vince, F. Framework for
assessment of off-stream freshwater use within LCA International Journal of Life Cycle Assessment 2010,
15, 439.
(4)
Boulay, A.-M.; Bouchard, C.; Bulle, C.; Deschênes, L.; Margni, M. Categorizing water for LCA
inventory The International Journal of Life Cycle Assessment 2011, 16, 639-651.
(5)
Boulay, A.-M.; Bulle, C.; Bayart, J.-B.; Deschenes, L.; Margni, M. Regional Characterization of
Freshwater Use in LCA: Modeling Direct Impacts on Human Health Environmental Science & Technology
2011, 45, 8948-8957.
Chemical footprint from point sources in Sweden
Louise Sörme1, Yevgenia Arushanyan2, Anders Wadeskog1, Göran Finnveden2, Hanna
Brolinson1
1
2
Statistics Sweden, Box 24300, 104 51 Stockholm
KTH Royal Institute of Technology, 100 44 Stockholm
1. Introduction
There is a large demand in Sweden and elsewhere to quantify a chemical footprint. This study is part of a
larger study financed by the Swedish EPA, performed by Statistics Sweden together with KTH, Royal
Institute of Technology. As a start Swedish EPA decided to develop indicators for the greenhouse gas
emissions, other emissions to air and chemicals. This part on chemicals is more of a prestudy.
This study has identified a method to quantify the chemical footprint from point sources with data from the EPRTR (European Pollutant Release and Transfer Register) togheter with USEtox. USEtox is often used in
Life Cycle Impact Assessment.
There is an EU regulation since a few years saying that industrial facilities (e.g. industries) emitting over
certain thesholds have to report to a register, E-PRTR, annually. The emission data used in this study comes
from E-PRTR. Countries report emission data according to an EU regulation [1] and a guidance document
[2]. The industrial facilities covers 65 economic activities within the following 9 industrial sectors: energy,
production and processing of metals, mineral industry, chemical industry, waste and waste water
management, paper and wood production and processing, intensive livestock production and aquaculture,
animal and vegetable products from the food and beverage sector, and other activities.
Data is provided in the register for 91 pollutants falling under the following 7 groups: greenhouse gases,
other gases, heavy metals, pesticides, chlorinated organic substances, other organic substances and
inorganic substances. Data was collected from the EEA website [3], where all E-PRTR data is available for
all EU27 countries. The emissions were listed by amount to water and air respectively. In most cases CAS
numbers are specified in the regulation.
A calculation on the impact is made using the USEtox method [4] as implemented in SimaPro. The USEtox
model is an environmental model for characterization of human and ecotoxic impacts in Life Cycle Impact
Assessment and for comparative assessment and ranking of chemicals according to their inherent hazard
characteristics. The method is for example described in [5]. In this study the impacts human toxicity (cancer
and non-cancer) and ecotoxicity are included. The calculations are performed in SimaPro. The results are
given in CTU (comparative toxic units). Only emissions from those substances that are relevant for human
toxicity and ecotoxicity are included from the emissions registred in E-PRTR.
The results are for the year 2008. Sweden has delivered emissions for 53 substances, some to air or water
and some to both. The total emission to air is very much larger than the total emission to water.
The expected results are:
•
•
•
•
An aggregated measure of the toxicity from the sources included in E-PRTR
Contribution to toxicity by different industry sectors
Information on each substance’s contribution to toxicity
Identification of the most important substances
For these four the results are divided in contribution to human toxicity and to ecotoxicity.
A next step could be to also include emissions in other countries from the E-PRTR register to be able to
estimate the emissions in other countries caused by Swedish consumption, to include a consumption
perspective. This is a focus of the Swedish EPA, to try to get data on emissions in other countries caused by
Swedish consumption. Emissions from countries outside Europe would still be missing with this data source
E-PRTR. For greenhouse gas emissions calculations of emissions in other countries caused by Swedish
consumption will be made. E-PRTR includes large industrial facilities, it would therefore be valuable to
investigate how much of the emissions are excluded using this data source.
4. Conclusions
This study makes it possible to discuss how useful this method is to quantify the chemical footprint by using
TM
data from E-PRTR and the USEtox method. Diffuse emissions are often important, at least in urban areas
and they are still missing with this approach which makes the footprint underestimated.
5. References
[1] Regulation (EC) No 166/2006 Of the European Parliament and of the council of 18th January 2006
concerning the establishment of a European Pollutant Release and Transfer Register and amending
Council Directives 91/689/EEC and 96/61/EC
[2] European Commission 2006. Guidancedocument for the implementation of the European PRTR.
http://prtr.ec.europa.eu/docs/EN_E-PRTR_fin.pdf
[3] http://rod.eionet.europa.eu/obligations/538/deliveries
[4] http://www.usetox.org/
[5]Rosenbaum RK, Bachmann TM, Gold LS, Huijbregts MAJ, Jolliet O, Juraske R, Koehler A, Larsen HF,
MacLeod M, Margni M, McKone TE, Payet J, Schuhmacher M, Van de Meent D, Hauschild MZ. 2008.
USEtox—the UNEP-SETAC toxicity model: recommended characterisation factors for human toxicity and
freshwater ecotoxicity in life cycle impact assessment. International Journal of Life Cycle Assessment 13 (7):
532–546.
Acknowledgement – Linnaeus University, Sweden is acknowledged for contribution to Louise Sörmes time to
work on this conference paper
Probabilistic Environmental Hazard Assessment of
Implementing Green Chemistry Property Design Guidelines to
Reduce Acute and Chronic Aquatic Toxicity
Kristin A. Connors1, Adelina M. Voutchkova2, Paul Anastas2, Julie B. Zimmerman2,
Bryan W. Brooks1
1
2
Baylor University, Waco, TX, USA
Yale University, New Haven, CT, USA
E-mail contact: [email protected]; [email protected]
1. Introduction
Probabilistic environmental hazard assessment (PEHA) is employed to determine the probability of
encountering compounds or toxicological thresholds below specific values. Chemical Toxicity Distributions
(CTDs), a type of PEHA modelling approach, are ideal for predicting toxicological thresholds when
information is lacking for new chemicals. CTDs are ideally suited for large and complex datasets. This
1
approach has recently been utilized to predict antibiotic concentrations of concern in model aquatic plants ,
2
3
describe the effects thresholds of fish and invertebrates to parabens , pesticides and surfactants , to
4
establish thresholds of ecotoxicological concern for various organic chemicals , to identify pharmaceuticals
5
with potentially large acute-to-chronic ratios , and to compare the sensitivity of common in vitro and in vivo
6
models of estrogen agonist activity .
One of the Twelve Principles of Green Chemistry emphasizes the need to synthesize safer chemicals.
These chemicals should be capable of performing their desired function but would be designed to elicit
minimal toxicity. In support of this effort, two recent studies have explored the relationships between
7
8
chemical properties and acute or chronic toxicity as measured through standardized OECD and EPA
protocols. Mechanistically-rationalized guidelines were derived resulting in two practical design guidlines.
Specifically, Voutchkova et al suggested acute and chronic toxicity could be minimized if compounds had an
7,8
octanol-water partition coefficient (log P ow ) below 2 and HOMO-LUMO gap (dE) greater than 9 eV .
In this study, we examined the potential utility and effectiveness of these design guidelines to reduce aquatic
toxicity of common industrial chemicals. PEHAs were performed: 1. to predict the likelihood of encountering
industrial chemicals exceeding established US EPA thresholds of acute and chronic toxicity to standardized
algae, cladoceran and fish models; 2. to predict the likelihood of exceeding these thresholds if chemical
safety guidelines were followed; and 3. to examine acute and chronically toxic chemicals, chemical classes
and modes of action of chemicals that may not be “designed out” by chemical property guidelines.
Large and diverse experimental toxicity data were gathered for industrial chemicals from the Japanese
Ministry Database. Acute toxicity (LC50) data for a 96 hr fathead minnow assay (570 compounds), a 96 hr
Japanese medaka assay (276 compounds), a 48 hr Daphnia magna assay (343 compounds) and a 72 hr
green algae (300 compounds) were collected. Chronic toxicity data included a 504 hr Daphnia magna assay
(317 compounds), and a 336 hr and 504 hr medaka assays (66 and 27 compounds, respectively).
CTDs were created and four types of PEHAs were performed: 1. representative of the current chemical
universe; 2. modelling chemicals conforming to only one design guideline (logP ow or dE); or 3. modelling
chemicals that met both design guidelines. The percentage of chemicals predicted to be encountered at
acute and chronic toxicity thresholds of high, moderate, low or no level of concern (0-1, 1-100, 100-500 and
>500 mg/L, respectively), as outlined by US EPA guidelines, were determined.
PEHAs employing the CTD approach to examine model datasets were used to predict the percentage of
industrial chemicals that are expected to elicit acute toxicity within the US EPA’s established categories of
concern for acute toxicity. For example, in the absence of chemical design guidelines, our model predicts
that 14.5% of chemicals would be classified by the EPA as being of “High Level of Concern” for acute toxicity
(LC 50 of 0-1 mg/L) to the fathead minnow. However, if log P ow and dE guidelines are employed during
chemical development, only 3.3% of industrial chemicals are predicted to be of High concern to the fathead
minnow (Figure 1).
99.8
Percent Rank
99
98
95
90
80
70
50
30
20
10
5
2
1
0.5
0.2
0.1
10-5
10-4
10-3
10-2
10-1
100
101
102
103
104
105
106
LC50 mg/L
Figure 1. Chemical toxicity distribution of acute fathead minnow data.
(●)All chemicals (○)Chemicals that obey both design guidelines.
7,8
If the two green chemistry design guidelines put forth by Voutchkova et al are employed, the present study
predicts that reduced acute toxicity to the fathead minnow could be achieved for over 10% for industrial
chemicals classified to possess High acute toxicity.
4. References
[1] Brain RA, Sanderson H, Sibley PK, Solomon KR. 2006. Probabilistic ecological hazard assessment:
Evaluating pharmaceutical effects on aquatic higher plants as an example. Ecotox Environ Safety 65:128135.
[2] Dobbins LL, Usenko S, Brain RA, Brooks BW. 2009. Probabilistic ecological hazard assessment of
parabens using Daphnia magna and Pimephales promelas. Environ Toxicol Chem. 28(12):2744-2753.
[3] Williams ES, Berninger JP, Brooks BW. 2011. Application of chemical toxicity distributions to
ecotoxicology data requirements under REACH. Environ Toxicol Chem 30:1943-1954.
[4] de Wolf W, Siebel-Sauer A, Lecloux A, Kock V, Holt M, Feutel T, Comber M, Boeije G. 2005. Mode of
action and aquatic exposure thresholds of no concern. Enviro Tox Chem 24(2):479-485.
[5] Berninger JP, Brooks BW. 2010. Leveraging mammalian pharmaceutical toxicology and pharacology data
to predict chronic fish responses to pharmaceuticals. Toxicol Lett 193:69-78.
[6] Dobbins LL, Brain RA, Brooks BW. 2008. Comparison of the sensitivities of common in vitro and in vivo
assays of estrogenic activity: Application of chemical toxicity distributions. Environ Toxicol Chem 27: 26082616.
[7] Voutchkova AM, Kostal J, Steinfeld JB, Emerson JW, Brooks BW, Anastas P, Zimmerman JB. 2011.
Towards rational molecular design: derivation of property guidelines for reduced acute aquatic toxicity.
Green Chem 13: 2373-2379.
[8] Voutchkova AM, Kostal J, Connors KA, Brooks BW, Anastas P, Zimmerman JB. Towards rational
molecular design for reduced chronic aquatic toxicity. In review.
Exergetic footprint as indicator to assess the environmental
sustainability of processes.
Alba Saravia1, Marta Herva1, Carlos García-Diéguez1 and Enrique Roca1
1
Sustainable Processes and Products Engineering Group, Department of Chemical Engineering, University
of Santiago de Compostela, Campus Vida, 15705 Santiago de Compostela, Spain
1. Introduction
Exergy analysis has two key attributes for being used as an environmental indicator: first, given that the
environment is used as a reference state, exergy is a measure of any thermodynamic deviation with respect
to its normal state; second, it allows comparisons between inflows and outflows, regardless they are mass or
energy streams, using the same physical units (exergy) for their analysis [1].
Besides, the exergetic footprint (ExF) bears in mind the exergy that still remains in products and highlights
the use of exergy as a convenient unit of measure and comparison. In this paper, the (ExF) was used as a
resource/waste accounting indicator for the environmental assessment of a production process.. The
proposed methodology was applied to a wood-based particleboard production process from a factory located
in Galicia (NW Spain).
2.1. Particleboard production process
The study was proposed as a gate to gate analysis, considering the production process of wood-based
particleboard (from the entering of raw material to the process to the particle board obtaining of the finished
product). The incoming streams were categorised as raw materials, secondary materials and energy
3
sources, while the output streams were the generated waste and the produced particleboard. 1 m of
produced particleboard was selected as functional unit.
Furthermore, with the aim of assessing the influence of the use of recycled wood as raw material on the
ExF,, some scenarios of the production process were built considering and increasing ratio of recycled wood
used until the complete substitution of the roundwood. Consequently, as the recycled wood ratio increased,
the energy consumption associated with debarking, chipping and drying processes necessary when
roundwood is employed, decreased.
2.2. Exergetic footprint methodology
The ExF is determined considering the exergy content of the following categories: material and energy
resources, generated waste and the resulting product.
In order to calculate the ExF, the consumed exergy (CEx) in the process was calculated. The CEx is the
exergy required to carry out the process from energy/material inputs plus the exergy required to neutralize
the potential environmental harm of wastes as it is calculated by Equation 1:
CEx = ∑i m i ⋅ Exch ,i + ∑ j n j ⋅ β
ex − e ( k , p , n , r ,t ) j
+ ∑w Exch , w ⋅ m w
1
where m i is the mass of material resource i [kg]; Ex ch,i is the chemical exergy per kg of material i [MJ∙kg ]; n j
is the quantity of energy by energy source j [MJ]; β ex-e(k,p,n,r,t)j is the exergy to energy ratio (quality factor) by
-1
energy source j [MJ∙MJ ] (k, p, , n, r and t are kinetic, potential, nuclear, solar and thermal exergies,
respectively) and, finally, Ex ch,w and m w are the chemical exergy and mass of wastes, respectively.
-1
In this accounting methodology, the chemical exergy of wastes represents an approach to the potential
environmental impact of wastes which is based solely on its exergy content [2]. Furthermore, the proposed
ExF takes into account the exergy remained in products, as it is shown in Equation 2:
ExF = CEx − ∑ p Ex ⋅ m p
where Ex ch,p and m p are the chemical exergy per mass of product and mass of product.
2
3.1. Inventory of the process
The raw material was wood from pine (75%), black poplar (10%), mixed wood (5%) and recycled wood
-3
waste (10%). The energy demand for particleboard manufacturing was 2823.82 MJ∙m particleboard, where
heat supply represented 84.5%, electricity 13.5% and fossil fuel 2%. The thermal energy, used mainly during
the drying process, was provided by wooden dust (48.2%) and natural gas (51.8%). Wooden dust provided a
40.8% of the total energy use. With regard to other studied scenarios, the total energy requirements ranged
-3
from 3017.38 to 1081.78 MJ∙m of particleboard when the percentage of recycled wood employed was
increased from 0% to 100%.
3.2. Exergy footprint assessment
The exergy values for the different categories considered are shown in Table 1.
Category
Input
Raw material
Secondary materials
Energy sources
Output
Generated wastes
Product
Exergy [MJ per functional unit]
12911.62
1616.05
3033.02
<0.01
12780.82
Table 1: Consumed exergy byselected categories in the particleboard production process
The total exergy consumption for the particleboard manufacturing process was determined as the sum of
each input category exergy plus the exergy content of wastes following equation [1]. It represented the
requirements demanded from the process production to the natural environment, as well as the requirements
for restoring the natural environment conditions damaged by the generated wastes. For the case study it
reached a value of 17.56 GJ per functional unit. However, as mentioned above, to determine the ExF the
exergy that still remains in the produced particleboard is considered and included in the assessment. Hence,
the proposed ExF of the process was estimated in 3.5 GJ per functional unit. Similarly to the result obtained
in terms of Ecological Footprint in a previous work [3], the material resources was identifyed as the main
contribution to the overall CEx and ExF, representing more than 70% in all estimations.
As regard other escenaries, the CEx of process was ranged from 17.76 to 15.73 GJ ∙ m particleboard, and
-3
the ExF from 3.69 to 1.66 GJ ∙ m for 0% to 100% of recycled wood wastes, showing the positive impact on
environment of using recycled material as particleboard component-3
4. Conclusions
The process ExF was considered a convenient tool for the assessment of the environmental sustainability of
a production process. It is understandable and allows for a thermodynamic approach over the environmental
appraisal of production processess. In this case, the material resources was the main contributor to the
overall CEx and ExF. Thus, a clear improvement in the sustainability of the process could be attained by
increasing the feasibility of using different waste materials in the design of the particleboard.
5. References
[1] Velásquez HI, Ruiz AA, Oliveira S. 2009. Ethanol Production from Banana Fruit and its Lignocellulosic
Residues: Exergy and Renewability Analysis. Int J of Thermodynamics 12:155-162.
[2] Rosen MA. 2009. Indicators for the environmental impact of waste emissions: Comparison of exergy and
other indicators. Trans. Can Soc Mech Eng 33:145-160..
[3] Saravia AM, Herva M, García-Diéguez C, Roca E. 2011 Sustainability appraisal of particleboard
production with recycled wood waste by ecological footprint. Proc 1st Int Conference on Wastes:
Solutions, Treatments and Opportunities 2011
Thermodynamic resource indicators and footprint in LCA: A
case study of titania in China
Wenjie Liao1, Reinout Heijungs1and Gjalt Huppes1
1
CML, Leiden University, POBox 9518, 2300RA Leiden, The Netherlands
1. Introduction
While the LCA community has standardized methods for assessing emission impacts, some comparable
methods for the accounting or impact assessment of resource use exist but are not as mature or
standardized. In the ILCD handbook (EC JRC 2010a, 2010b), resource depletion is the only impact category
for which no single recommended method has been identified.This study contributes to the existing research
by offering a comprehensive comparison of similarities and differences of different resource indicators, in
particular those based on thermodynamics, developing possible concepts of thermodynamic footprint, and
testing them in a case study on the titania (titanium dioxide pigment) produced in China.
2.1. Thermodynamic resource indicators and footprint
The system boundary for resource indicators is defined using a thermodynamic hierarchy at four levels and
the case data for titania also follow that hierarchy. Seven resource indicators are applied. Four are
thermodynamics based: cumulative energy demand (CED), solar energy demand (SED), cumulative exergy
demand (CExD), and cumulative exergy extraction from the natural environment (CEENE), and three have
different backgrounds: abiotic resource depletion potential (ADP), environmental priority strategies (EPS),
and eco-indicator 99 (EI99). The normalization of various resource indicators to the total anthropogenic
resource demand is implemented and the concept of entropy footprint is applied.
2.2. Case study: the titania produced in China
The chloride route and the sulphate route, as two current mature routes for the commercial production of
titania, are analyzed in this study as two alternatives for the titania produced in China. This study is a cradleto-gate analysis of the titania system. The functional unit in this study is defined as 1 kg of titania at plant.
Inventory data for the foreground system has been collected through on-site interviews and visits.
Background inventory data are from the database ecoinvent v2.2. Characterizations factors are based on
CML-IA database covering all major methods. Computations are with the CMLCA software.
3.1. Resource scores
Table 1 gives the scores of seven resource indicators addressed in this study for 1kg of titania produced via
both the chloride route and the sulphate route. Within the four thermodynamic resource indicators, CED,
CExD, and CEENE have similar scores with each other while their scores are five orders of magnitudes
lower than the score of SED.
Route
CED
SED
CExD
CEENE
ADP
EPS
EI99
MJ
MJ se -eq
MJ ex -eq
MJ ex -eq
kg Sb -eq
ELU
MJ-eq
Chloride route
106
7.91E+07
129
123
0.0536
1.50
8.66
Sulphate route
117
6.63E+07
151
143
0.0735
2.32
10.6
Table 1: Scores of various resource indicators of 1kg of titania in China
3.2. Resource contributions
The relative contribution of each resource group to the scores of different resource indicators is represented
in Figure 1. Atmospheric resources do not contribute to the SED or CEEND score. Land resources account
for a negligible percent to the SED scores, and have a small contribution to the CEENE scores. Nonrenewable resources have a dominant contribution to the scores of all seven resource indicators in both
routes while renewable energy sources have a small contribution (less than 2%). Fossil resources have a
relatively high contribution to the scores of CED, CExD, CEENE, ADP, and EI99 (more than 74%) in all types
of non-renewable resources. Metal ores and fossil fuels have comparable contributions to the score of EPS.
The score of SED is dominated by the demand of metal ores and minerals.
Figure 1: Contribution of resource groups to the different resource indicators (left: chloride route; right: sulphate route)
3.3. Normalization
The cumulative non-renewable energy demand (CED NRR ) and exergy demand (CExD NRR ) can be defined as:
CED NRR = CED fossil + CED nuclear and CExD NRR = CExD fossil + CExD nuclear + CExD metals + CExD minerals ,
respectively. Considering the CED NRR and CExD NRR of 1kg of titania (110 MJ and 127 MJ ex -eq,
respectively), the global production of titania (4.5 E+09 kg/yr) would correspond to the CED NRR and the
CExD NRR of global titania of 4.92 E+11 MJ/yr and 5.72 E+11 MJ ex -eq/yr, respectively. The normalization to
the total anthropogenic non-renewable energy demand and exergy demand (4.11 E+14 MJ/yr and 4.15 E+14
MJ ex -eq/yr, respectively) shows that the global production of titania would account for about 0.12% of the
total anthropogenic non-renewable energy demand and about 0.14% of the total anthropogenic nonrenewable exergy demand. Similar normalizations in other resource indicators can also be implemented if
corresponding values of the total anthropogenic demand are available.
3.4. Entropy footprint
It is assumed that the exergy extracted from non-renewable natural resources is completely consumed in
various production and consumption processes. The CExD NRR of global titania corresponds to a heat
emission to the ecosphere of 5.72 E+11 MJ/yr, which generates entropy at a rate of 6.32 E+08 W/K at 288K.
If one relates the entropy production to the global net entropy generation density (1.21 W/K∙m2), the result
indicates the entropy footprint in terms of an area of Earth’s surface that is occupied by the technological
system under consideration. For global titania, its entropy footprint would be 5.2 E+07 m2.
4. Conclusions
First we demonstrate the feasibility of thermodynamic resource indicators and the concept of entropy
footprint. We recommend CEENE as the most appropriate one within the four thermodynamic resource
indicators for accounting and characterizing resource use because it accounts for the largest number of
resource groups while highlighting fossil and nuclear energy resources. As for the three non-thermodynamic
resource indicators, they take the different resource issue as their key problem, and, as pointed by the ILCD
handbook (EC JRC 2010a, 2010b), have higher environmental relevance in terms of expressing the
resource scarcity and depletion than the thermodynamic ones. Regarding the case study on the titania
produced in China, all the resource indicators except SED show that the sulphate route demands more
resource use than the chloride route. As the conceptual basis for the several indicators differs fundamentally,
different cases might well show more diverging outcomes.
5. References
[1] European Commission Joint Research Centre (EC JRC) (2010a): International Reference Life Cycle Data
System (ILCD) Handbook - Framework and Requirements for Life Cycle Impact Assessment Models and
Indicators. First edition March 2010. EUR 24586 EN. Luxembourg. Publications Office of the European
Union
[2] European Commission Joint Research Centre (EC JRC) (2010b): International Reference Life Cycle Data
System (ILCD) Handbook - Analysing of existing Environmental Impact Assessment methodologies for use
in Life Cycle Assessment. First edition March 2010. EUR 24586 EN. Luxembourg. Publications Office of the
European Union
European Guidelines for Measuring the Environmental
Footprint of Products and Organisations
Rana Pant1, Kirana Chomkhamsri1, Nathan Pelletier1, Simone Manfredi1, Michele
Galatola2, and Imola Bedo2
1
European Commission, Joint Research Centre, Ispra, Italy
European Commission, DG Environment, Brussels, Belgium
2
1. Introduction
In close co-operation, DG Environment and the Joint Research Centre of the European Commission are
developing two harmonised European methodologies for measuring the Environmental Footprint of products
(covering goods and services) and organisations [1].
The objective is to support decision-making processes in business and policy with a comprehensive multicriteria set of indicators based on the robust and quality assured measurement of environmental
performance. This is to overcome some identified shortcomings of single indicator measurements and other
existing methods. The overarching purpose is to reduce the environmental impacts of products and
organisations in a more resource-efficient and sustainable Europe and address the right of European citizens
to receive reliable environmental information.
The presented work relates to one of the building blocks of the Flagship initiative of the Europe 2020
Strategy – “A Resource-Efficient Europe.” The European Commission's “Roadmap to a Resource Efficient
Europe” proposes ways to increase resource productivity and to decouple economic growth from both
resource use and environmental impacts, taking a life-cycle perspective. It states specifically as one of the
objectives to: “Establish a common methodological approach to enable Member States and the private
sector to assess, display and benchmark the environmental performance of products, services and
companies based on a comprehensive assessment of environmental impacts over the life-cycle
('environmental footprint')” [2]. The abstract is based on drafts before ongoing stakeholder consultations and
before results from onging pilot studies could be incorporated.
The Product Environmental Footprint (PEF) and Organisation Environmental Footprint (OEF) guidelines
describe on how to calculate an Environmental Footprint, as well as how to create product category or sector
specific requirements for use in Product Footprint Category Rules (PFCRs) or Organisation Footprint
Sectorial Rules (OFSRs).
Each requirement specified in the PEF and OEF guidelines has been chosen taking into consideration the
recommendations of similar environmental accounting methods and guidance documents. Specifically, the
methodology guides considered were:
•
ISO 14044: Environmental management -- Life cycle assessment -- Requirements and guidelines
•
ISO 14067: Carbon footprint of products
•
ILCD: International Reference Life Cycle Data System
•
Ecological Footprint
•
Product and supply chain standards, Greenhouse Gas Protocol (WRI/ WBCSD)
•
Méthodologie d'affichage environnemental (BPX 30-323)
•
Bilan Carbon
•
DEFRA - Carbon Disclosure Project (CDP)
•
CDP water
•
GRI: Global Reporting Initiative
•
Specification for the assessment of the life cycle GHG of goods and services (PAS 2050)
Although these documents align on some of the methodological guidance they provide, discrepancies and/or
lack of clarity remains on a number of important decision points, which significantly reduces the consistency
and comparability of the results, for example the coverage of impact categories and the models used for the
calculation of impacts. The intention of the PEF and OEF guidelines is (wherever feasible) to identify a single
requirement for each decision point to support more consistent, robust and comparable studies.
3.1.
The guidelines set requirements for each relevant decision point in an Environmental Footprint study,
covering among others goal and scope definition, definition of relevant impact categories, identification of
system boundaries and environmentally significant versus insignificant processes, completing the resource
use and emissions profiles (including a decision hierarchy for dealing with multi-functionality problems), the
use of generic data sources as well as defining data quality and review requirements.
For the selection of relevant impact categories, a “default list” of 14 midpoint Life Cycle Impact Assessment
(LCIA) categories is provided together with the models and characterisation factors to be used. The models
and factors in the default list are taken from the International Reference Life Cycle Data System (ILCD)
Handbook recommendations for LCIA [3] and needs to be considered as the starting point for PEF and OEF
studies. If it can be demonstrated that a certain impact category is not relevant for the product or
organisation under study, it can be neglected. If it can be demonstrated that other impacts are relevant or
that for a specific impact category a better model is available, this information can be provided under
“additional environmental information” similar to the way it is possible in Environmental Product Declarations
(EPDs) according to ISO 14025. However, the impacts have to be calculated as “baseline” according to the
model from the default list for comparability and consistency reasons. This is a compromise between
scientifc rigor and allowing for ongoing scientific development. Such an approach strives for comparability of
results and a ceratin stability in the models and factors to be applied.
3.2.
Product Environmental Footprint Category Rules (PFCRs) and Organisation Environmental Footprint
Sectorial Rules (OFSRs) are a necessary extension of and complement to the more general guidelines for
PEFs and OEFs. PFCRs and OFSRs will play an important role in increasing the reproducibility, consistency,
and relevance of PEF and OEF studies. PFCRs and OFSRs will facilitate focusing on the most important
parameters, thus also aiming at reducing time, efforts, and costs for completing an EF study. To ensure
consistency, the PEF and OEF guidleines also specify requirements for the development of related PFCRs
and OFSRs that should be developed as a complement to the more general guidelines.
4. Conclusions
In the creation of the guidelines, a balance had to be found between different objectives and principles: to
ensure a more robust and reproducible decision support of environmental footprints comparability was often
given priority over flexibility, making it necessary to go beyond the level of guidance provided in existing
national or international accountancy documents. A balance also had to be found between latest scientific
developments and practicability / required efforts. The stakeholder consultation process on the guidelines
finished in December 2011 and pilot studies were ongoing until spring 2012. While draft versions are already
publically available on the DG ENV website [1], the intention is to have the two methodologies formally
adopted by the European Commission by the end of 2012.
5. References
[1] Product guidelines available at: http://ec.europa.eu/environment/eussd/product_footprint.htm and
Organisation guidelines at http://ec.europa.eu/environment/eussd/corporate_footprint.htm .
[2] European Commission 2011. A Resource-Efficient Europe. The European Commission's “Roadmap to a
Resource Efficient Europe”
[3] European Commission-Joint Research Centre 2011. International Reference Life Cycle Data System (ILCD)
Handbook - Recommendations for Life Cycle Impact Assessment in the European context. First edition November
2011. EUR 24571 EN. Luxemburg. Publications Office of the European Union
Speciation in the Transformation/Dissolution Examination of
Antimony and Tungsten Metal and Compounds
J.M. Skeaff1, R. Beaudoin1 and D.J. Hardy2
1
Mine Closure and Ecosystem Risk Management Program, Mining and Mineral Sciences Laboratories,
CANMET, Natural Resources Canada, 555 Booth St., Ottawa, Ont., Canada K1A 0G1
2
Present address: 2313 Rembrandt Rd., Ottawa, Ont., Canada K2B 7P6
1. Introduction
Speciation is held to be a key factor in controlling the human health and environmental effects of metals in
solution [1]. Moreover, a potential hazard classification in the EU CLP (Classification, Labelling and
Packaging) Regulation [2] can be changed to a less severe level or even annulled if a metal is >70%
transformed to a non-available species within 28 days. In this study, we present the results of the T/D
(Transformation/Dissolution) examination at pH 6 and 8.5 of tungsten metal and several of its compounds, and
of antimony metal and several of its compounds in terms of the concentrations of total dissolved metal and of the
concentrations of the corresponding metal-bearing species. We also show how the T/D data have been used to
derive UN GHS (United Nations Globally Harmonized System of Hazard Classification) or EU CLP hazard
classification outcomes for submission to regulatory authorities.
Using the UN T/DP (T/D Protocol) [3], we have examined the T/D behaviour of Na 2 WO 4 .2H 2 O, WO 3 , WC,
W metal; 5(NH 4 ) 2 (H 2 W 12 O 42 ).4H 2 O (ammonium paratungstate, APT); 6(NH 4 )(H 2 W 12 O 40 ).6H 2 O (ammonium
metatungstate, AMT) and WO x (x = 2.9). In addition to total dissolved tungsten, we also measured the
2concentrations of the tungstate ion, WO 4 . We have also determined the T/D behaviour of Sb metal,
NaSb(OH) 6 , Sb 2 O 3 , Sb 2 S 3 , NaSbO 3 , Sb 2 (C 2 H 4 O 2 ) 3 , SbCl 3 , Sb(CH 3 COO) 3 and Sb 2 O 5 with respect to total
dissolved antimony, and for the 1 mg/L loadings, the speciation of dissolved antimony in terms of the
concentrations of Sb(III) and Sb(V) solutions. All substances were in the powder form.
2.1. Analytical determinations
The Analytical Services Group of CANMET-MMSL developed and validated the speciation methodologies.
2For WO 4 speciation and total dissolved W, the T/D solution samples were essentially split into two streams:
one stream introduced into a Dionex ICS-3000 HPLC (high performance liquid chromatograph) column onto
2which the WO 4 anions were loaded and then eluted with a weak solution of KOH, and a second stream for
total dissolved W that passed directly into the HPLC eluent. The coupled solution advanced directly to a
Perkin Elmer Elan 6100 simultaneous inductively-coupled plasma mass spectrometer (ICP-MS) interfaced
with a concentric nebulizer, and system retention times of 1.7 and 4.3 minutes for tungsten and tungstate,
respectively, resolved into two separate peaks, the areas under which were proportional to their
5+
concentrations. For Sb(III) and Sb(V) speciation, the method was based on the existence of Sb as
3+
Sb(OH) 6 over the pH range 2.7-10.4, and the reaction of dissolved Sb with Na 2 EDTA (sodium
ethylenediaminetetraacetic acid) to form Sb(EDTA) . Once separated on the HPLC column by approximate
peak retention times for Sb(V) and Sb(III) of 4.6 and 6.6 minutes, respectively, we converted the areas under
the peaks to concentrations as above. We measured the concentrations of total dissolved Sb using either a
Varian Vista RL ICP-AES for concentrations in the mg/L range, or a Thermo-Fisher X-series II ICP-MS or the
Perkin-Elmer Elan 6100 ICP-MS ICP-MS for concentrations in the µg/L range. The limits of quantification
depended on the extent of T/D sample dilution, but were usually in the range 1.0-2.0 µg/L.
3.1. Tungsten and compounds
2-
For the W compounds examined, the analytical data revealed that all dissolved W existed as WO 4 . One
hundred mg/L loadings of Na 2 WO 4 .2H 2 O, APT and AMT were readily and completely soluble, yielding
2measured W concentrations of about 67 mg/L, and WO 4 concentrations in the range 75 to 93 mg/L which
2correspond to all W dissolved as WO 4 . A typical plot of the average concentrations of total dissolved W
2and of WO 4 as a function of time for 100 mg/L loadings of yellow WO 3 at pH 8.5 is in Figure 1. We see that
the rates of dissolution of both increase rapidly over the first 72 hr after which this rate declines, yielding
2measured 168-hr W and WO 4 values of 62.4 and 83.5 mg/L, respectively. The calculated concentrations
2WO 4 and W were in good agreement with the measured. Compared to an acute ERV (Ecotoxicity
Reference Value) of 31 mg/L, their 168-hr concentrations would classify Na 2 WO 4 .2H 2 O, APT, AMT, WO3
and WO 2.9 as GHS Acute 3, but would not classify them under the EU CLP scheme. W metal and WC
would not classify under either scheme.
3.2. Antimony and compounds
With NaSb(OH) 6 and Sb 2 O 3 , the T/D data revealed, within detection limits, that Sb dissolved entirely as
Sb(V) and Sb(III), which are its respective valences in these compounds. With Sb metal, Sb dissolution was
primarily as Sb(III), again as expected since metallic Sb was always present in the 1 L jars to equilibrate with
antimony in solution. Nonetheless, the speciation data suggested some oxidation of Sb(III) to Sb(V). For
Sb 2 S 3 , the speciation data suggested a significant degree of Sb(III) oxidation to Sb(V) over the course of the
28-day tests. For NaSbO 3 , Sb dissolved as Sb(V), since antimony is pentavalent in this compound. For
2Sb 2 (C 2 H 4 O 2 ) 3 , the C 2 H 4 O 2 ligand appeared to stabilize Sb(III) in solution, with only a moderate amount of
oxidation at both pHs. Similar comments apply to the CH 3 COO ligand in Sb(CH 3 COO) 3 , particularly at pH
6. On the other hand, Sb(III) released from SbCl 3 , Figure 2, was readily oxidized to Sb(V) at both pHs. With
pentavalent Sb in Sb 2 O 5 , the speciation data indicated an initial small amount of Sb(III) that oxidized to
Sb(V) over the 28 days. A comparison of the T/D data with the 6.9 mg/L acute ERV for dissolved antimony
of revealed that none of Sb metal and its compounds would classify under the EU CLP.
4. Conclusions
In examining the T/D characteristics of W metal and several of its compounds, in terms of the concentrations
22of total dissolved W and of the WO 4 anion, we found that essentially all W dissolved as the WO 4 anion.
Na 2 WO 4 .2H 2 O, APT, AMT, WO 3 and WO 2.9 would all classify as GHS Acute 3, but would not classify under
the EU CLP scheme. W metal and WC would not classify under either scheme.
Due to complexing, the trivalent organic Sb compounds Sb 2 (C 2 H 4 O 2 ) 3 and Sb(CH 3 COO) 3 exhibited little or
no oxidation of Sb(III) to Sb(V) in their T/D behaviour. However, oxidation of Sb(III) to Sb(V) was evident for
the trivalent inorganic Sb compounds Sb 2 S 3 and SbCl 3 , although not in the trivalent Sb 2 O 3 , being relatively
insoluble, nor in Sb metal. For the pentavalent Sb compounds NaSb(OH) 6 , NaSbO 3 and Sb 2 O 5 , Sb(V) was
stable in solution. None of Sb metal and its compounds would classify under the EU CLP.
5. References
[1] Krachler M and Emons H. 2001. Urinary antimony speciation by HPLC-ICP-MS. J Anal At Spectrom 16:20-25.
[2] EU (European Union). 2006. Regulation (EC) No 1907/2006 of the European Parliament and of the
Council of 18 December 2006 concerning the Registration, Evaluation, Authorisation and Restriction of
Chemicals (REACH). http://ec.europa.eu/environment/chemicals/reach.htm. Accessed 14/Feb/2007.
[3] UN (United Nations). 2007. The Globally Harmonized System of Classification and Labelling of
Chemicals. ST/SG/AC.10/30/Rev. 2. http://www.unece.org/trans/danger/publi/ghs/ghs_rev02/02files_e.html.
Accessed 28/Feb/2008.
Acknowledgements - We are grateful to the Tungsten Consortium and the International Antimony
Association (i2a) for funding this research. The data in this study cannot be freely used to comply with
regulatory requirements such as REACH without the formal agreement of the Tungsten Consortium or i2a.
Error! Not a valid
100000
100
100 mg/L
measured W
Sb total
W(aq) or WO42-(aq), µ g/L
calculated W
Sb(III)
2measured WO4
80000
80
calculated WO4
Sb(III) + Sb(V)
Sb total for 10% dissolution of 1 mg/L
60000
40000
60
Sb,
µg/L
40
20000
20
0
0
0
24
48
72
96
120
144
168
time, hr
link.
Sb(V)
2-
Figure 1: Net T/D reaction kinetic data for 100 mg/L loadings of
yellow WO3 powder at pH 8.5.
0
168
336
504
672
time, hr
Figure 2: Net T/D reaction kinetic data for 1 mg/L loadings of
SbCl3 powder at pH 6.
Transformation/Dissolution of Nickel Metal and Sparingly Soluble
Nickel Compounds
E. Rogevich-Garman1, C. Schlekat1, P. Huntsman-Mapila2 and J. Skeaff2
1
Nickel Producers Environmental Research Association (NiPERA), 2525 Meridian Parkway, Durham, North
Carolina, 27713, USA
2
Mine Closure and Ecosystem Risk Management Program, Mining and Mineral Sciences Laboratories,
CANMET, Natural Resources Canada, 555 Booth St., Ottawa, Ont., Canada K1A 0G1
1. Introduction
The environmental fate and toxicology assessment of nickel metal and nickel compounds is based on the
observation that adverse effects to aquatic, soil- and sediment-dwelling organisms are a consequence of
exposure to the bioavailable nickel-ion, as opposed to the parent substances. The result of this assumption
is that the ecotoxicology is the same for all nickel substances in which ecotoxicity is caused by the
2+
bioavailable nickel ion (Ni ) and should all be used in the derivation of ecotoxicological endpoints (aquatic
toxicology, sediment toxicology, terrestrial toxicology) and environmental fate endpoints (bioaccumulation,
adsorption/desorption), based on read-across from the Ni-ion. The justification for using this read-across
approach for nickel metal and nickel compounds is that at present it is not justifiable to differentiate among
2+
soluble nickel compounds (Ni salts) when using the general data set, simply because other factors are
more important for determining the NOEC than the actual nickel compound used.
For the purposes of identifying the hazards of nickel metal and nickel compounds in the aquatic environment,
2+
it is necessary to examine the rate and magnitude of metal ion (Ni ) release when the substance is
introduced in a solution. The transformation/ dissolution protocol (T/D P) (Annex 10, UN GHS, 2009) was
designed to examine the extent to which the nickel metal or sparingly soluble nickel compound will react with
1
the media and transform to the bioavailable, toxic ion . For classification under the Globally Harmonized
System of Classifying and Labeling Chemicals (GHS) and the Regulation (EC) No 1272/2008 on
Classification, Labeling and Packaging of substances and mixtures (CLP), T/D P testing is used to determine
the release of bioavailable metal ions to the aquatic environment at mass loadings of 1 mg/L, 10 mg/L, 100
mg/L (acute) and 0.01 mg/L, 0.1 mg/L and 1 mg/L (chronic), in order to estimate ecotoxicity and determine
the environmental hazard. Transformation/ dissolution rates are dependent on several factors, including the
size and specific surface area of the particles, test duration, pH or test media, temperature, and mass and
specific loading rates. To establish hazard classifications for nickel metal or a sparingly soluble nickel
compound, data from the T/D P at each mass loading are compared to ecotoxicity reference values (ERVs)
to determine if the nickel ion dissolution is greater or less than the corresponding ecotoxicity threshold value
(Skeaff et. al., 2011).
T/D P data can also be used to derive hazard classifications for metal mixtures or special preparations, such
as metal alloys. The aquatic toxicity classification of a metal mixture or special preparation can be derived
by comparing the Transformation/Dissolution data of the metal mixture or alloy with the acute and chronic
ecotoxicity reference values for the individual components of the mixture or alloy. Resulting classifications
are based on comparison of the dissolution of the individual metal ions with their respective ERVs, as well as
the summation of the net toxic effect using the Toxic Unit (TU) approach. The TU approach assumes that
toxicity following exposure to multiple metals is additive.
Using the UN T/DP (T/D Protocol), we have examined the T/D behavior of several nickel substances;
including nickel metal powder, nickel metal massive, nickel monoxide, nickel mattes, and ferronickel alloys,
for the purpose of classification under the UN GHS and EU CLP systems. As nickel toxicity is pH
dependent, low and high pH acute toxicity reference values are needed to compare to T/D data at pH 6 and
8, respectively. In addition to total dissolved nickel concentrations, we also measured the concentrations of
other dissolved metals (e.g., copper, cobalt, iron, arsenic), where appropriate.
1
Although the intent of the T/D P is to quantify the rate of release of bioavailable, toxic metal species, in practice the dissolved fraction
is measured. The dissolved fraction can include free metal ions as well as dissolved metal complexes.
Based on the results of the T/D testing, the following classifications were derived:
Substance
GHS Classification
Based on T/D Data
CLP Classification
Based on T/D Data
Harmonized Classification
st
in 1 ATP to CLP
Nickel Metal Powder
R52/53
Aquatic Chronic 3
Aquatic Chronic 3
Nickel Metal Massive
No Classification
No Classification
No Classification
Nickel Monoxide (Green)
No Classification
No Classification
Aquatic Chronic 4
Nickel Monoxide (Black)
R51/53
Aquatic Chronic 2
Aquatic Chronic 4
Nickel Metallic Matte
R52/53
Aquatic Chronic 3
Aquatic Acute 1/Chronic 1
Ferronickel Alloy
R52/53
Aquatic Chronic 3
N/A
3.1. Nickel Metal
The results of the T/D P indicate that nickel metal powder presently should be classified as R52/53 (Aquatic
Chronic 3) using the data presented above in the EU and GHS classification systems for metals and metal
compounds and that there should be no chronic classification for nickel metal massive.
3.2. Nickel Monoxide
T/D tests were conducted with two samples of nickel monoxide (green and black). Green nickel monoxide is
formed at a high calcining temperature, while black nickel monoxide is a less refractory form of nickel
monoxide. Results of the testing on the green nickel monoxide found it to be essentially unreactive while the
test results show that the black nickel monoxide was considerably more reactive than the green. T/D P tests
indicated that the appropriate classification for black nickel monoxide is R51/53 (Aquatic Chronic 2) and that
there should be no classification for green nickel monoxide.
3.3. Sparingly Soluble Nickel Compounds
The results of the T/D P indicate that the appropriate classification for both a metallic nickel matte and a
ferronickel alloy are R52/53 (Aquatic Chronic 3).
4. Conclusions
T/D testing can be used to derive appropriate hazard classifications under the Globally Harmonized System
of Classifying and Labeling Chemicals (GHS) and the Regulation (EC) No 1272/2008 on Classification,
Labeling and Packaging of substances and mixtures (CLP). Several nickel substances underwent T/D
testing in the context of REACH and GHS to evaluate the extent of transformation and dissolution in order to
determine the appropriate hazard classifications. Results were compared to the existing harmonized
st
classifications reported in the 1 ATP to the CLP regulation. T/D testing indicated that in some instances the
harmonized classifications were appropriate, while in others, the harmonized classifications were either overor under-protective compared to the appropriate hazard classification derived from the T/D data.
5. References
EU (European Union). 2006. Regulation (EC) No 1907/2006 of the European Parliament and of the Council
of 18 December 2006 concerning the Registration, Evaluation, Authorisation and Restriction of
Chemicals (REACH). http://ec.europa.eu/environment/chemicals/reach.htm.
Skeaff, J., Adams, W., Rodriguez, P., Brouwers, T., Waeterschoot, H. Advances in Metal Classification
under the United Nations Globally Harmonized System of Classification and Labeling. IEAM, Vol.7,
No. 4, pp. 559-576.
UN (United Nations). 2007. The Globally Harmonized System of Classification and Labelling of Chemicals.
ST/SG/AC.10/30/Rev. 2. http://www.unece.org/trans/danger/publi/ghs/ghs_rev02/02files_e.html.
Transformation/dissolution of copper concentrates: Effect of
mineral composition on metal solubility
Katrien Delbeke1, Paola Urrestarazu2, Margaret Opazo2, Jose Arbildua2 and
Patricio H Rodriguez2 , Jussi Liipo3
1
European Copper Institute, Avenue de Tervueren 168 Box 10, B-1150 Brussels, Belgium
Chilean Mining and Metallurgy Research Center, Av. Parque Antonio Rabat 6500, Vitacura, Santiago, Chile
3
Outotec Research Center, P O Box 69 - Kuparitie 10, FIN-28101 PORI, Finland
2
1. Introduction
The environmental classification of metals and sparingly soluble inorganic metal compounds is based on
comparing the soluble fraction, measured after Transformation/Dissolution (T/D), with the ecotoxicity reference
values of the soluble metal ions. The Transformation/Dissolution protocol (TDp) is the outcome of an
international effort under the OECD to develop a standard operating procedure (SOP) to assess the rate and
extent of metal-ion releases from metals and sparingly soluble inorganic metal compounds [1, 2] .
Copper concentrates are complex metal bearing substances, whose composition varies depending on the
geological origin of the extracted ore body. Elemental analysis of copper concentrates demonstrate that
copper concentrations range between 10 - 40% with other metals such as iron, arsenic, cadmium, cobalt,
lead, molybdenum, nickel and zinc present at various concentrations. Considering the heterogeneity of the
copper concentrates, a profound understanding of their mineralogy and potential transformation/dissolution
is necessary to assign appropriate classification entries.
The aims of this study are:
- Determine the mineralogical and elemental composition and the releases of metals during 7 days TD for 5
copper concentrates, obtained from different ore bodies.
- Determine the releases of metals during 7 days T/D of pure copper minerals present, in copper
concentrates
- Evaluate the possible correlations between the copper releases from copper concentrate samples and the
release of copper from the minerals constituents.
The elemental composition of the five copper concentrates, obtained from different copper ore bodies, were
assessed through ICP-OES (inductively coupled argon plasma optic emission spectrometer) after total
dissolution. The mineralogy of the samples was determined by X-ray diffraction and detailed microscopical
investigations. The mineralogical studies demonstrate the presence of one or more of the following copper
minerals: Bornite (Cu 5 FeS 4 ), Chalcocite (Cu 2 S), Chalcopyrite (CuFeS 2 ), Covellite (CuS), Digenite (Cu 9 S 5 ),
Enargite (Cu 3 AsS 4 ) and Tennantite ((Cu,Fe) 12 As 4 S 13 ). Information on the mineral composition (Table 1)
and elemental composition (Table 2) illustrates the variable composition of the selected copper concentrates.
Mineral
Mineral composition of copper concentrates (%)
CuC1
CuC2
CuC3
CuC4
CuC5
Chalcopyrite (CP)
46.81
78.33
8.69
2.12
54.80
Covellite (CV)
0.72
0
2.10
Bornite (BN)
2.53
8.21
Enargite (EN)
0.66
0.56
Tennantite (Tn)
0.2
Chalcocyte-dininite (CC-DG)
7.37
28.43
Table 1. Copper containing mineral composition of copper concentrates (%)
Considering the composition of copper concentrates, the following experimental design was used: TD tests
were carried out at pH 6, a sample loading of 100 mg/L and test duration of 7 days. The releases of six
metals were measured: Cu, As, Zn, Pb, Co, and Cd.
The T/D assay was performed following the SOP [Skeaff et al., 2006] prepared for the ad hoc OECD
Validation Management Group (VMG). Dissolved metal concentrations were measured in a Class 1000 clean
room using an ICP-MS Perkin Elmer model 9000 instrument.
Table 2 summarizes the releases of metals from the copper concentrates and demonstrates low release
rates for copper and arsenic (0.2-4%), intermediate release rates for zinc (0.9-9%) and highest release rates
for lead (7- 35%). The releases of Cd and Co are often below detection limits. The detailed kinetics from the
T/D tests show that the metal release rates decrease as a function of time, therefore metal concentrations
are expected to reach pseudo-equilibrium concentrations.
Table 3 summarizes the copper releases after 7days T/D of the selected minerals. The data demonstrate
mineral-specific solubility with lowest release rates for chalcopyrite (0.3%) and highest release rates for
chalcocite (5.3 %).
Figure 1 shows the relation between the measured dissolved copper concentration (µg/L) after 7 days TD of
the 5 selected copper concentrates (100 mg/L, pH 6) and the dissolved copper concentration predicted to be
released from the mineral composition of the samples (Table 1) and the minerals-specific copper releases
(Table 3). Four of the five predicted values are close to the measured, the fifth predicted value exceeds the
measured by a factor of about two.
CuC 1
Constituents
CuC 2
CuC 3
CuC 4
CuC 5
Element, %
7d, %
Element, %
7d, %
Element, %
7d, %
Element, %
7d, %
Element, %
7d, %
Cu
16.70
0.2
27.20
0.3
14.00
3.98
23.90
3.29
20.00
BDL
As
0.20
BDL
0.16
BDL
0.36
0.53
0.13
1.15
0.22
0.82
Cd
0.06
9.8
0.01
5.9
ND
Zn
6.90
9.1
3.27
4.4
0.87
2.87
0.54
2.41
9.00
0.92
Pb
5.80
35.4
0.33
19.8
3.80
21.26
1.60
23.03
12.50
7.55
Co
0.00
BDL
0.01
4.0
0.12
4.93
0.02
7.55
<0.004
BDL
ND
ND
Table 2. Metal releases from 5 copper concentrates (CuC1 to CuC5) after 7 days TD tests at 100 mg/L loading and at pH 6.
BDL= below detection limit
Copper releases from minerals after 7 days transformation/dissolution tests
Chalcopyrite
Digenite
Tennantite
Enargite
Bornite
Covellite
Chalcocite
% Cu released
0.3
0.6
2.3
2.3
2.6
3.4
5.3
Table 3. copper releases from 7 copper minerals after 7 days TD tests at 100 mg/L loading and at pH 6. BDL= below detection
limit; ND: not determined.
Figure 1. Copper releases observed from a 7 day transformation/dissolution test at pH 6 and a loading of 100 mg/L versus the
calculated copper releases using the the copper releases of the minerals and the mineral composition of the concentrate.
From the data, it can be concluded that it will not be possible to derive one environmental classification entry
applicable to all copper concentrates. Comparison of release data among concentrates and their mineralogy
clearly indicate that the metal release rates, depend on the elemental and mineral composition of the copper
concentrate. The results further indicate that, as demonstrated for copper, that the metal releases from
copper concentrates can be predicted from the mineral composition and the mineral- specific metal releases.
The releases of lead remain difficult to interpret and are possibly related to the association of lead with the
small particle size fraction of the concentrate.
4. References
[1] [EU] European Union CLP guidance. 2008. Guidance on labelling and packaging in accordance with
regulation (EC) No 1272/2008.
[2] Skeaff JM, Ruymen V, Hardy DJ, Brouwers T, Vreys C, Rodriguez PH, Farina M. 2006. The standard
operating procedure for the transformation/dissolution of metals and sparingly soluble metal compounds—
revised, June 2006. Natural Resources Canada. CANMET-MMSL Division Report MMSL 06-085 (TR). 555
Booth St., Ottawa, Canada, K1A 0G1.
Metal Classification using a Unit World Model: Assessing
Removal Rates from the Water Column and Remobilization
from Sediment with TICKET-UWM
Kevin J. Rader1,2, Richard F. Carbonaro1,2, and Kevin J. Farley2,
1
2
Mutch Associates, LLC, Ramsey, NJ, 07446 USA
Department of Civil and Environmental Engineering, Manhattan College, Riverdale, NY, 10471 USA
1. Introduction
European Union (EU) regulations pertaining to Classification, Labelling, and Packaging (CLP) of chemical
substances follow the United Nations Globally Harmonized System (UN GHS). Annex 9 of the GHS
document on classification and labelling recognizes speciation, partitioning, and precipitation as critical
elements in metal hazard classification. It also provides an initial discussion on how transformation of metals
to potentially less toxic forms (via partitioning and precipitation) and subsequent removal from the
environment (i.e., the water column) can be analogous to degradation of organic chemicals in the context of
classification [1]. Further practical guidance on environmental transformation of metals was provided in the
EU CLP guidance document [2]. This guidance document includes a provision for demonstrating removal
from the water column to assess the “persistence” or lack of degradation of metal ions, responsible for the
toxicity of metals and metal compounds. In analogy to organic chemicals, “rapid degradation” for metals
requires greater than 70% removal within 28 days. However, unlike organic chemicals, where removal from
the water column occurs via degradation, metal removal occurs through changes in speciation (partitioning
and precipitation) followed by sedimentation which transfers metal to the sediment. Therefore, in line with
the GHS guidance, “rapid degradation” for metals requires one to demonstrate not only rapid loss from the
water column, but also limited remobilisation potential from sediment.
A unit world model for metals in lakes, TICKET-UWM, has been developed that considers key processes
affecting metal transport, fate, and toxicity including complexation by aqueous inorganic and organic ligands
(e.g., DOC), adsorption to particulate organic carbon (POC), binding to biological receptors (biotic ligands),
and transport of dissolved metals and solids between the water column and sediment. The TICKET-UWM
was used to assess the rate at which metals (Cu, Pb, Zn, Ni, Co, and Cd) are removed from the water
column in a generalized lake system through partitioning and settling. The model was also used to assess
metal speciation changes in the sediment and the potential for metal remobilization from sediment.
Model simulations were made using a generalized lake system based upon the EUSES lake [3, 4, 5]. Water
chemistries for pH 6, 7, and 8 used in water column simulations were based on available directives/guidance
documents [1, 2, 3]. Metal addition was modelled as a single instantaneous addition of soluble metal salt at
loading concentrations ranging from the L(E)C 50 /acute ecotoxicity reference values for each metal to the cutoff value of 1 mg/l. Two methods were used to quantify the binding of metals on particles in the water
column removal calculations: (1) the linear partitioning method and (2) the speciation model method. In the
former, a single fixed distribution coefficient (K D ) was applied based on empirical values. In the latter, the
WHAM V speciation model [6] was used to calculate the K D value at each time step in the simulation through
consideration of metal binding to DOC and POC.
Sediment chemistry for the remobilization potential simulations used bulk and porewater sediment chemistry
from a number of field studies [Personal communication with Marnix Vangheluwe, 7]. The speciation model
method was used to calculate metal speciation in the sediment. Simulations were made for an anoxic
sediment with acid volatile sulfide (AVS) and POC present and for an oxic sediment with ferric/manganese
oxides and POC present as binding phases. The simulation time was 365 days. Metrics used to assess
remobilization potential include: long-term maintenance of water column concentrations below 70% removal
concentration, relative magnitudes of water column and sediment K D values, prevailing direction of the
sediment/water column diffusive flux, and speciation changes.
3.1. Water column partitioning and sedimentation results
Model simulations with the linear partitioning method indicate that for Cu, Zn, Pb, Ni, Co, and Cd, greater
than 70% removal from the water column occurred for all loading concentrations and pH values. Results for
the speciation model method (Table 1) show that some metals (Cu, Pb, and Zn) showed rapid removal for all
pH/loading scenarios tested, while others (Ni, Co and Cd) were removed rapidly under a subset of the
scenarios.
Metal
Cu
Pb
Zn
Ni
Co
Cd
Loading
(µg/L)
29, 1000
9, 17.8, 52, 73.6, 1000
19, 82, 136, 413, 1000
2.4, 68, 120
1000
4.9, 90.1
1000
0.21, 18
1000
Water Chemistry
pH 6
pH 7
pH 8
Y
Y
Y
Y
Y
Y
Y
Y
Y
Y
Y
N
Y
N
N
Y
Y
N
Y
N
N
Y
Y
N
Y
N
Y
A comparison of the model-predicted K D to
empirical
values
indicates
that
some
discrepancies exist, particularly for Ni, Co, and
Cd where calculated values are significantly
lower than measured values. This motivates
future research into refining the partitioning
model used by TICKET-UWM
3.2. Sediment speciation and
remobilisation potential results
For both anoxic and oxic sediments and all
loadings considered, pseudo-steady state
Y Greater than 70% removal achieved in 28 days
water column concentrations (sustained by
N Less than 70% removal achieved in 28 days
feedback from the sediment) were still
Table 1: Results from Speciation Model Method
markedly below the concentration representing
70% removal of the initial concentration. In anoxic sediment, the interaction of metals with AVS produced: 1)
a larger K D in the sediment (compared to the water column), 2) integrated diffusive fluxes over the 365-day
simulation time directed into the sediment, and 3) a major change in metal speciation (i.e., from binding to
POC to precipitation as metal sulfides). The sequestration indicated by these metrics supports limited
remobilization potential for metals where sufficient AVS is present.
4. Conclusions
−
Metal removal from the water column (analogous to degradation of organics): Under the linear
partitioning method, metals considered showed greater than 70% removal in 28 days. Under the
speciation model method, some metals showed rapid removal under all pH/loading scenarios tested,
while other metal showed rapid removal under a subset of the scenarios.
−
Sediment results suggest that remobilization potential is limited, particularly when sufficient AVS is
present.
−
TICKET-UWM represents a useful tool for evaluating metal “degradation” in the aquatic environment.
5. References
[1] United Nations, 2009. Globally Harmonized System of Classification and Labelling of Chemicals (GHS),
rd
3 Edition. United Nations. ST/SG/AC.10/30/Rev.3
[2] ECHA. 2011. Guidance on the Application of the CLP Criteria. Helsinki, Finland: European Chemicals
Agency. ECHA-11-G-06-EN.
[3] European Chemicals Bureau, 2003. Technical Guidance Document on Risk Assessment – Part II.
European Commission Joint Research Centre EUR 20418 EN/2.
[4] EC. 2004. European Union System for the Evaluation of Substances 2.0 (EUSES 2.0). Bilthoven, The
Netherlands: National Institute of Public Health and the Environment (RIVM). Report 601900005.
[5] ECHA. 2010. Guidance on Information Requirements and Chemical Safety Assessment: Chapter R.16:
Environmental Exposure Estimation.
[6] Tipping E and Hurley MA. 1992. A unifying model of cation binding by humic substances. Geochim
Cosmochim Acta 56:3627–3641.
[7] Besser J, Brumbaugh B, Kemble N, and Ingersoll C. 2010. Preliminary Data Summary for USGS Toxicity
Studies with Nickel in Water and in Nickel-Spiked Sediments. Durham, NC, USA: US Geological Survey.
Acknowledgement - European Copper Inst., International Zinc Assn., International Lead Zinc Research Org.,
Nickel Producers Environmental Research Assn., The Cobalt Development Inst., European Aluminium
Assn., Eurometaux, Rio Tinto, Bill Stubblefield (Oregon State University).
A BLM probabilistic approach to integrate variability in the
derivation of Water Quality Criteria at regional and local
scales
Ciffroy P.1, Charlatchka R.1, Ferreira D.2
1
EDF, R&D, 6 quai Watier, 78400 Chatou, France
Veritas, 2,Boulevard Vauban - 78067 SAINT QUENTIN EN YVELINES
2
1. Introduction
The biotic ligand model (BLM) is one of the more promising models allowing to derive Water Quality
Criteria (WQC) based on true bioavailable fractions of metals. Generalizing the concept of the free ion
activity model, BLMs investigate indeed how metal speciation and the competing cations (majors
cations and/or pH) alter metal binding with biological sites.
However, several physico-chemical parameters must be assigned for running BLMs (i.e. pH, major
cations, DOC, temperature, alkalinity, sulphur compounds) and the operational use of BLMs for
deriving site-specific WQC can suffer several flaws because of:
(i) the lack of comprehensive data. National monitoring programs of water bodies were indeed not
originally designed for BLMs application and, as a consequence, some datasets can be incomplete;
(ii) natural variability. Rivers and lakes can indeed show great spatial and temporal variations (e.g.
daily or seasonal variations of pH, seasonal autochthonous and/or allochthonous sources of DOC,
effects of tributaries). So far however, only single values were generally considered for describing
variables of concern, and the high natural variability with time and along a river watershed was
ignored.
The main objective of this study was then to investigate the potential of probabilistic approaches able
to capture the uncertainty of variables of concern and the incompleteness of datasets for BLMs
application on large scales.
For testing the methods described below, the BLM model developed by the UK Environment Agency
was selected because it was designed to use a limited number of variables, while remaining reliable.
The methodology was tested on the Loire river watershed for copper, where different spatial scenarios
were considered.
The BLM tool developed by the UK Environment Agency requires three physico-chemical variables,
i.e. pH, Ca concentration and Dissolved Organic Concentration (DOC), in addition to copper
-1
concentration in filtered water. The generic PNEC proposed in this tool was fixed at 1 µg.l for copper.
As a case study, the Loire river watershed was selected. Data needed for running the selected BLM
model were extracted from the database freely put available on the web by the ‘Loire-Bretagne Water
Management Agency’. The analysis of this database showed that data were available at 48 different
monitoring stations for the period 1999-2010.
Considering all data available for pH, DOC and Ca concentration (i.e., 5018, 3223 and 2493 data
®
respectively in filtered water), normal PDFs were fitted using the CrytalBall software.
For copper, two types of data were available, i.e. total concentration in raw water and dissolved
concentration. Two flaws were identified in data treatment: (i) to convert total copper concentrations
into dissolved concentrations, a Distribution coefficient K d was required. Instead of considering a
single K d value taken from the literature, a Bayesian approach was developed to build a Probability
Density Function (PDF) actually relevant for the Loire system variability, merging prior estimation of
K d s from a geochemical model (ECOSAT) and actual monitoring data; (ii) a significant fraction of
dissolved concentrations (about 50%) were below the limit of detection (LOD – data called ‘nondetects’). In order to impute realistic values to these non-detects, ‘distributional’ methods were used
(Cohen, 1991; Baccarelli et al, 2005): assuming that dataset of detects represents a truncated normal
distribution, estimates of mean and standard deviation of the ‘true’ distribution were calculated from
the mean and variance of ‘detects’ data, corrected by a tabulated bias (Kuttatharmmakul et al,
2001).Distributional methods were thus used to reconstruct a complete dataset affecting potential data
to non-detects (Distribution-based imputation).
In addition, probabilistic risk assessment was also conducted at local scale (i.e. considering data
collected at a given monitoring station only). In such a case, the number of available data can be
reduced dramatically and frequentist statistics can be more subject to criticism. To overcome the
problem of data scarcity, a Bayesian approach was proposed. For this purpose, it was assumed that
data collected at stations spatially close to the targeted station provide a prior PDF representing what
is expected for the ‘copper concentration’ variable at the targeted station. Bayes theorem was used to
update this prior distribution with information contained in data specifically collected at the investigated
station.
Once the PDFs were obtained for each of the variables, 10000 combinations were randomly
generated by a Monte Carlo procedure and the BLM tool was run for each of them. Risk index were
calculated for each combination as the ratio between the copper concentration in water and the
‘combination-specific’ PNEC.
Results and discussion
It was observed that: (i) pH values in the Loire river can show high basic levels, exceeding the
validation limit of the BLM model used here (i.e. 8.5). From this observation, it can then be expected
that a certain proportion of simulations will be outside of the validity domain and will thus lead to
-1
minimum generic PNECs (here 1 µg.l ); Ca concentrations show an increasing trend when going
downstream of the Loire river, suggesting that the protective effect of cationic competition depends on
the investigated spatial scenario; (iii) DOC does not show significant differences according to the
investigated monitoring station.
As far as Cu concentrations are concerned, PDFs significantly depend on the method chosen for
treating data extracted from the database. For example, at the watershed scale, the mean Cu
concentration was 2.75 µg.l-1 if non-detects are ignored. This value overestimates the actual mean Cu
concentration as about half of the data were discarded from the data analysis. When a distributional
-1
approach is considered instead, the mean Cu concentration was calculated as 1.77 µg.l . If no further
probabilistic BLM analysis is considered (i.e. if only the mean PEC is compared to the generic PNEC),
these values would significantly modify the conclusion of the risk analysis: if non-detects are ignored
indeed, the risk index (i.e. the PEC/PNEC ratio) is above one (situation at risk), while it is below one if
a distributional approach is considered (situation at no risk).
Similarly, in case of scarcity of data at local scale (i.e. at a given monitoring station), frequentist and
Bayesian approaches led to results significantly different.
These observations showed that, beyond the application of the BLM approach, some guidelines are
necessary to better treat input data.
References
[1] Baccarelli A., Pfeiffer R., Consonni D., Pesatori A.C., Bonzini M., Patterson Jr D.G.., Bertazzi P.A.,
Landi M.T., 2005. Handling of dioxin measurement data in the presence of non-detectable values:
Overview of available methods and their application in the Seveso chloracne study
Chemosphere, Vol. 60, Issue 7, 898-906
[2] Cohen, A.C., 1991. Truncated and Censored Samples: Theory and Applications, Marcel Dekker,
New York
[3] Kuttatharmmakul S., Massart D.L., Coomansb D., Smeyers-Verbeke J., 2001. Comparison of
methods for the estimation of statistical parameters of censored data. Analytica Chimica Acta 441
(2001) 215–229
Accounting for both local aquatic community composition and
bioavailability in setting local quality standards for metals
Adam Peters1, Peter Simpson1
1
wca environment ltd, Faringdon, Oxfordshire, UK.
1. Introduction
Recent years have seen considerable developments in the ability to make water quality standards for trace
metals more ecologically relvant by taking account of the effect of local water chemistry conditions on
bioavailability. This prevents situations where a different level of risk is considered to acceptable at different
sites due to changes in bioavailability which are not accounted for in the standard. The present study
describes preliminary efforts to address an additional issue in the development of water quality standards
which are specific to particular locations, by taking account of the composition of the local ecological
community (the ultimate protection objective). This has occasionally been addressed through the use of field
measurements to derive species sensitivity distributions (SSD) in sediments [1, 2]. An alternative approach,
which combines a quality assessed ecotoxicity dataset with field measurements of the abundance of benthic
macroinvertebrates to derive an SSD based on the community which is either expected to be present, in the
absence of anthropogenic pressures, or the community which is present at the sites. Site specific standards
are derived for zinc in an area impacted by historic mining activities. Site-specific targets for zinc, based on
the macroinvertebrate ecology predicted or observed at a site, can be derived and can result in improved
compliance compared to the use of both conventional and bioavailability-based EQS. In addition to zinc, the
approach is likely to be applicable to other metals and possibly other types of chemical stressors (e.g.
pesticides). However, the methodology for deriving site-specific targets requires additional development and
validation before they can be robustly applied during surface water classification.
There are 25 species in the SSD for Zn that was used for EQS derivation, of which three are algal species,
nine are fish or amphibian species, and 13 are invertebrate species. Of the invertebrate species represented
there are four sponges, three cladocerans (i.e. daphnids), two rotifers, one amphipod, one snail, one mussel,
and one midge larvae. Two potential approaches were identified for establishing a SSD for zinc whose
composition is based on the community predicted to be present at a site by RIVPACS. The first approach
considers applying the F BL NOEC values (fractional occupancy of Zn at the biotic ligand (target receptor)
associated with no observed effects) derived from ecotoxicity tests to the most taxonomically similar
RIVPACS scoring families. The second approach simply estimates a distribution of taxa sensitivities based
on the observed distribution of sensitivities from ecotoxicity testing and an empirical ranking of taxa
sensitivity derived from macroinvertebrate monitoring data at zinc impacted sites.
The first approach is limited by the range of invertebrates tested in the laboratory compared to the range of
RIVPACS scoring families, and the dissimilarity of the riverine RIVPACS families to some of the more
typically lacustrine species used for laboratory ecotoxicity testing, such as daphnids (water fleas). Identifying
which of the available tested laboratory species best represents a relatively dissimilar family, which could be
from an entirely different insect order is a difficult judgement, and the resulting assignment of F BL NOEC
values to RIVPACS scoring families is likely to be rather arbitrary. In addition, there is a limited range of
F BL NOEC values from which to select an appropriate one.
The second approach requires an appropriate basis for assigning relative sensitivities of different RIVPACS
scoring families, and information about the distribution of sensitivities of species included in the SSD. This
approach could either use all species, or just invertebrate species. It is typical in Europe to consider that the
sensitivity of a tested species can be considered as being representative of the sensitivity of an untested
species, even if the tested species is not considered to be directly relevant to the assessment in question.
Thus non-European species are taken as being representative of untested European species. Adopting this
approach suggests that taking the distribution of all species may provide a better reflection of a more diverse
invertebrate ecosystem in the field.
The ranking of the sensitivity of invertebrate taxa was performed based on the abundance of taxa relative to
their expected abundance. This used a selection of high Zn exposure sites, from which sites with elevated
levels of other potential contaminants had been removed as they could have confounded the ranking due to
Zn. This resulted in a selection of 29 sites where Zn was considered likely to be the main contaminant. The
maximum levels of other potential pressures are shown in Table G.1. Dissolved Zn concentrations at these
-1
-1
sites ranged from 3.6 to 284 µg l , with mean and median concentrations of 79 and 61 µg l , respectively.
Following this approach it was possible to define a ranking on the basis of O/E for abundance at the 29 Zn
impacted sites for 64 of the RIVPACS scoring families. In several cases very similar values of O/E for
abundance were calculated for several different families, and in order to take account of this fact that some
taxa have very similar sensitivities the taxa were sorted into categories, each covering 0.02 O/E units. This
resulted in 32 groups with different sensitivity, with each group containing between one and five families. The
F BL NOEC values were assigned based on the position in the distribution, and were fitted to the same lognormal distribution that was derived from the SSD based on ecotoxicity test data.
Site specific SSDs were developed depending upon the expected community composition predicted by
RIVPACS for each site. The SSDs based on the expected community composition were compiled following
two different approaches. In the first approach those families with a predicted probability of capture of greater
than 0.5 were used as the basis for the SSD, and in the second approach those families with a predicted log
abundance of greater than 0.5 were used as the basis for the SSD. The second approach results in an SSD
with a greater number of taxa. In both cases the same, bioavailability normalised, NOEC values were applied
to each taxon. Site specific SSDs were also derived for some sites based on the community which was
observed to be present in the biological samples. In these cases taxa with an observed log abundance score
of 1 or greater were included in the SSD (except for non-BMWP scoring taxa and any taxa which were not
included in the ranking system).
The approach used here to derive site specific SSDs only includes invertebrate taxa, and does not include
any vertebrates (fish or amphibians) or plants (algae or macrophytes). For the purposes of the specific cases
considered in this study this is not a significant problem due to the limited range of water chemistry
conditions encountered across these sites. There is considerable overlap between the sensitivities of
invertebrates and fish in the EQS ecotoxicity database, and algae are particularly sensitive under high pH
conditions, which are not encountered at these sites. Whilst the simplification of the site specific SSDs to
include only invertebrate taxa may be acceptable for the sites included within this study, such simplifications
of the approach currently limit the utility of site specific targets for more general application, due to the likely
variability in water quality conditions which means that plants (and possibly also vertebrates) need to be
included to ensure sufficient sensitivity.
The validity of the sensitivity ranking for RIVPACS scoring invertebrate taxa to Zn is key to the approach
taken towards deriving site specific quality targets in this study. Any discrepancy between the sensitivity rank
of any individual taxon and its true sensitivity to Zn may lead to inaccuracies in the derived SSD. Table G.1
shows the maximum concentrations of several potential contaminants that were measured at the sites used
for the ranking. Approaches such as these are limited by the extent of the available information on the
presence and levels of other contaminants at the sites used. Very few of the other potential contaminants
were measured at all sites, and there are also numerous other potential contaminants that could affect the
abundance of invertebrates that were not included in the analysis. Unknown contaminants could potentially
result in reduced O/E values for taxa which are interpreted as Zn sensitivity in this assessment.
Significant relationships were observed between the ranking of taxa used for the site specific SSDs and the
observed O/E values for abundance of those taxa at two sites (p = 0.030 in both cases), although in one
case only when the taxon Heptageniidae were removed from the analysis.
4. Conclusions
This indicates that taxa which were expected to be insensitive to zinc tended to have higher O/E values than
taxa which were expected to be sensitive to zinc, and that this information can be used for the derivation of
site specific quality standards for zinc which take account of local ecological community composition.
5. References
[1] Leung K, Bjorgesaeter A, Gray J, Li W, Lui G, Wang Y, Lam P. 2005. Deriving sediment quality guidelines
from field-based species sensitivity distributions. Environ Sci Technol 39:5148-5156.
[2] Kwok K, Bjorgesaeter A, Leung K, Lui G, Gray J, Shin P, Lam P. 2008. Deriving site specific sediment
quality guideleines for Hong Kong marinbe environments using field based species sensitivity
distributions. Environ Toxicol Chem 27:226-234.
Bioavailability and beneficial use as primary demands for a
management guidance of contaminated dredged sediments
Wolfgang Ahlf, Barbara Wirska and Alexander Scheffler
Technical University of Hamburg-Harburg, Institute of Environmental Technology and Energy Economics,
Eissendorfer Strasse 40, 21073 Hamburg, Germany
1. Introduction
Dredging operations are mainly carried out in the coastal or marine environment; therefore ports and
navigation authorities worldwide face the continuous effort of dredging in order to maintain the required water
depth. In Europe, the volume of dredged material is roughly estimated at 200 million cubic meters per year.
The problem of dredging works becomes more complicated when dredged material is considered to be
contaminated. Sediments act as a sink for many contaminants, and consequently, bottom sediments in many
locations nationwide have became polluted because of municipal and industrial discharges and non-point
sources. Options for remediation contaminated sediments include no action (involves natural processes to
gradually improve conditions), non removal (treatment or isolation in place) and removal (environmental
dredging followed by treatment or disposal of contaminated sediments). However, for any dredging operation
contaminated sediment needs to be an integral part of the total environmental management of the system,
with respect to management issues about costs, ecological and social impact.
The purpose of this presentation is to present the draft guideline on sustainable management of
contaminated sediments for dredging projects all around the Baltic Sea. This applies to the whole sea area,
regardless of its jurisdiction, as well as to the project undertaken within the water bodies placed at the land of
whom the dredged material is planned to be deposited at the Baltic Sea.
This guideline comprises knowledge and practice including sustainability assessment and decision support
in relation to the handling alternatives for dredged sediments and its disposal. Sustainability means that the
environmental, economic and social aspects must be integrated and interventions should not result in
unwanted impacts. Environmental, economic and social criteria are considered to assess dredged material
treatment process under this concept (Fig.1).
This proposal is a result of the joint EU-project SMOCS under consideration of actual national and
international regulations, integrating other projects as well as interests of stake-holders. For example, the
participants of a workshop preferred a guidance document giving new ideas and an overview of options for
dredged material management. Exact definitions for the condition of dredged material (when to call dredged
material “clean”, “contaminated” or “hazardous”) and action levels were required.
Emphasis was given to fulfill political demands as to avoid waste production. Therefore it was primarily
recommended to assess the feasibility of beneficial use of the sediments. Bioavailability concept for the
derivation of sediment quality standards has a challenge to foster this approach, because not the total
amount of contaminants is of importance rather than the Such risk assessment based on in situ or field
studies meets the request of the European Water Framework Directive to develop a toxicity-based
bioavailability model to estimate the risk of sediment-associated contaminants.
Life Cycle Impact Assessment (LCIA) is an environmental assessment tool that aims to quantify
environmental impacts associated to the complete inventory of material and energy flows of a process over
its entire life-cycle. Integrated approaches have a challenge also to connect different types of information
relevant to those three factors to support decision-making procedures (e.g. integrating environmental factors
such as chemical concentrations, effect based in bioassays, energy consumption).
Fig. 1: Framework of a Guidance for sustainable management of contaminated sediments
The purpose of an LCIA probably in combination with a Risk Assessment is to ensure that dredging activities
are performed in an environmentally acceptable manner, use sound engineering techniques, which they are
economically warranted and take sufficient consideration of long term effects. The scope of an LCIA should
include a survey and review of prior studies to describe the physical, chemical, and biological conditions of
the study area; an evaluation of regional dredging and disposal needs; and a characterization of physical
and chemical properties of typical regional dredged materials. The LCIA will also formulate alternatives for
dredged material management, perform a preliminary evaluation of alternatives management methods, and
make recommendations for further evaluation.
On this basis a long term monitoring strategy has to be designed for the final management option.
3. Conclusions
The development of a guidance for sustainable management of contaminated sediments is currently an ongoing process, whereby scientists and problem-owner have the opportunity to be part of the process.
Natural attenuation in sediments
M. van Bemmel1, M. Wagelmans1
1
Bioclear BV, PO Box 2262, 9704 CG, Groningen
1. Introduction
Due to polluted groundwater in urban areas, sediments in urban water streams become polluted as well.
This can cause pollution of surface waters. Natural attenuation of chlorinated hydrocarbons in sediment can
be used to sustainably manage or remediate groundwater pollution. Due to the change in remediation
procedures in the Netherlands from site specific remediation to area specific remediation, an approach
based on natural attenuation will be increasingly considered. Provided that the risks of this approach can be
controlled, natural attenuation processes in sediment can be a valuable contribution to area specific
management and remediation plans.
However, there are still some knowledge gaps about these natural processes. Also research methods to
quantify these processes are lacking. And the possible ecological effects in sediment or surface waters are
still unknown.
At many sites in NL polluted groundwater is being managed by
seepage through sediments (especially sludge) to surface waters.
This management option is more and more accepted as
management option provided that biodegradation of pollutants
occurs (or can be expected) in the sediment (sludge). Due to high
organic matter contents in sludge, mostly methanogenic
conditions are present, which is favorable for biodegradation of
chlorinated ethenes. As a result of this, chlorinated ethenes can
be degraded before they reach the surface water. Besides
anaerobic degradation, aerobic degradation of DCE and VC could
theoretically take place in the upper sediment layers, the sludge.
This means that sludge from ditches, creeks, canals and rivers
can play an active role in removing chlorinated ethenes from the
environment.
2. Research questions and methods
2.1. Research questions
From the knowledge gaps, research questions have been defined:
1. Hoe are the most important processes related (sorption, transport, biodegradation of chlorinated
ethenes) and can these processes be modeled?
2. Does biodegradation occur in the entire sediment zone or is a vertical zoning present?
3. How fast are biodegradation processes in sediment/sludge? What is the most important factor?
4. How important is aerobic degradation at the surface from sludge to surface water?
5. Can biodegradation processes be stimulated?
6. Which techniques can be used to assess and monitor natural attenuation processes in sludge?
7. Which management and monitoring measurements are needed in order to sustainably use these
processes?
8. Will ecological risks be present in sediment, sludge and surface water? How can risks be prevented
or dealt with without disturbing the biodegradation processes?
2.2. Methods
Core samples (up to 1 m length) were taken from six polluted sites in three cities in the Netherlands,
including 1 reference sample per site. Samples were either frozen in liquid nitrogen or the individual
sediment layers were sliced and sampled in the field. Analyses have been performed according to table 1.
Chemical analyses in sediment
Microbiology in sediment (qPCR)
Dry weight
Eubacteria
Ethene mono-oxygenase
Organic matter
Archea
Polaromonas
Chlorinated ethenes or benzenes
(pore water)
Dehalococcoides sp.
Nitrificating bacteria
Nitrate
VC reductase
Denitrificating bacteria
Sulphate
Sulphite reductase
Iron oxidizing bacteria
Methane
MCRA
Iron reducing bacteria
Chlorinated ethenes or chlornated
benzenes
Sulphur oxidating bacteria
Sulphur reducing bacteria
Table 1: Analyses
3. Preliminary results
20
15
10
5
1,0E+06
0
0,00-0,19
0,19-0,38
1,0E+04
1,0E+03
1,0E+02
1,0E+01
1,0E+00
0,00-0,19
organic matter
Figure 1 a: Dehalococcoides sp and organic matter content,
and methane concentration.
0,19-0,38
0,38-0,72
Depth (m)
0,38-0,72
Depth (m)
Dehalococcoides sp.
45
40
35
30
25
20
15
10
5
0
1,0E+05
Methane (µg/l)
25
cells/ g sediment ( ww)
2,0E+05
1,8E+05
1,6E+05
1,4E+05
1,2E+05
1,0E+05
8,0E+04
6,0E+04
4,0E+04
2,0E+04
0,0E+00
organic matter content (%)
cells/ g sediment ( ww)
At this moment, all samples have been taken. However, results of both chemical and molecular analyses are
only available of one site yet. For the other sites, results of the molecular analyses are expected to be
available during december 2011.
Dehalococcoides sp.
etnE
methane
etnC
Vinylchloride reductase (vcrA)
b: Dehalococcoides sp, Ethene mono-oxygenase, VC reductase
The first prelimary results (see graphs in figure 1a and 1b) show that in the toplayer of the sediment, which
contains the highest organic matter content, anaerobic conditions are present. In this layer the microbial
activity is higher than in deeper sediment layers. The dechlorinating capacity in this layer is higher than in
deeper layers. However, in the toplayer, not only anaerobic organisms are present, but also aerobics. This
implies that microaerophilic conditions are present and chlorinated ethenes can be both anaerobically and
aerobically be degraded. In the deeper layers the dechlorinating capacity is lower. This has implications for
dredging managent in this specific canal. Dredging of the canal would, at least temporary, decrease the
dechlorinating capacity. This could cause an increase in chlorinated ethenes in surface water and result in
ecological risks during a certain period of time.
4. References
[1] Evaluating the fate of chlorinated ethenes in streambed sediments by combining stable isotope,
geochemical and microbial methods, Abe, Aravena, Zopfi, Parker, Hunkeler, Journal of contaminant
hydrology 107 10-21, 2009.
Acknowledgement - The authors thank the authorities of Haarlem, Amersfoort and Den Haag.
Assessing the Impacts of Climate Change on the Fate and
Transport of HCB and Cd in the Elbe River Basin (Germany)
Kari Moshenberg1, Susanne Heise
1
Hamburg University of Applied Sciences, Lohbrügger Kirchstraße 65, D-21033 Hamburg
1. Introduction
The Elbe River is the third largest river in Central Europe, starting in the Czech Republic and running through
the cities of Dresden and Hamburg before empting into the North Sea. Due to extensive historical
contamination and redistribution of contaminated sediments throughout the basin, the Elbe River transports
significant loads of contaminants downstream, particularly during flood events (Heise et al. 2008). These
contaminated sediments are ultimately deposited on floodplains or in the coastal environment.
This study focus on two representative contaminants, Hexachlorobenzene (HCB) and Cadmium (Cd), which
have both previously been identified as contaminants of concern in the Elbe River Basin. Sediment-sorbed
concentrations of both HCB and Cd significantly exceed maximum allowable concentrations established by
the European Commission in almost all sections of the Elbe River. Though sources and sinks of both
contaminants differ, high water events, such as the flooding that occurred in 2002, have been shown to be
the dominant process controlling contaminant transport downstream the Elbe River (Heise et al. 2008).
Conversely, low discharge events result in the accumulation of sediments in still water zones that are
available for resuspension during future floods.
In the KLIWAS program (“Impacts of climate change on waterways and navigation”) funded by the German
Federal Ministry of Transport, Building and Urban Development, the impact of projected climatic changes on
discharges of German rivers is being investigated. On the basis of 19 projections between 1971 and 2100,
and applying the A1B-scenario, a decrease of precipitation during summer months and an increase during
winter months is likely. For the Rhine river, average winter discharge are expected to increase up to 25%
(2021-2050) (Enno Nilsen, personal communication). For the Elbe, discharge projections are still being
simulated. However, a precipitation increase during winter months of up to 50% in 2071 to 2100 compared to
1961-1990 in some parts of the catchment has been projected (Wechsung & Becker 2005). Due to rising
temperatures, this precipitation will increasingly be rain instead of snow, increasing the probability of high
discharge events in winter.
The impact of these changes in hydrologic regime on the long-term fate and transport of HCB and Cd were
assessed using simulated discharge at 25 and 50 years in the future.
In this study, the one and two-dimensional flexible mesh (FM) modeling systems, MIKE 11 and 21FM,
respectively, were applied to quantify the dynamics of cohesive sediments along a 230 km segment of the
Middle Elbe River between Zehren and Magdeburg. The modeling system uses a hydrodynamic module as
basis for a number of add-on modules. The complexity of sediment transport within the numerous groyne
fields that line the banks of the Elbe necessitated the use of a two dimensional model. Boundary conditions
were established using measurements of river flow and suspended sediment as well as bi-weekly
measurements of sediment-sorbed HCB and Cd concentrations.
The 1D model was calibrated for a ten year period (1995-2005) using water level, discharge, and suspended
sediment data from gauging stations. Long-term decreases in HCB and Cd were used to calibrate the
contaminant modules. A 2D model was calibrated for a smaller, overlapping section of the model, between
Aken and Barby based on specific discharge events. This 2D model was used to evaluate the effect of
groyne fields and floodplains on HCB and Cd sorbed-sediment.
3.1. Model Validation
After the successful calibration of the 1D and 2D models, their performance was analyzed over a 2 year time
period (2006-2007) using a variety of indices. Boundary conditions and suspended sediment concentrations
were used to predict HCB and Cd concentrations. The numerical model was generally in agreement with
the measured values.
3.2. Climate Change Scenarios
To evaluate the sensitivity of contaminant transport to climate change, two sets of scenarios were
developed. The objective of the first set of scenarios was to evaluate the short-term effect of single floods on
contaminant transport. This was achieved by modeling recent floods (2006, 2011) separately, and then in
succession, and quantifying the differing impacts on HCB and Cd concentrations downstream, on
floodplains, and within groyne fields.
The second set of scenarios involved using theoretical discharge datasets at 25 and 50 years in the future
that encompass lower, moderate, and upper range of potential future discharge variability. The theoretical
discharge datasets have decreased summer discharge and an increased frequency of peak flows during the
winter and spring. The goal of this scenario is to evaluate how the combined effect of long periods of
undisturbed sedimentation and an increased frequency of high water events will contribute to downstream
HCB and Cd concentrations, as well as the volume of contaminated material transported to floodplains and
groyne fields.
4. Conclusions
The results from this modeling effort provide insight into climate-mediated changes in the transport of
sediment-sorbed contaminants, areas of historically contaminated sediment, and sources of secondary
contamination, as well as aid in identifying potential source management measures.
5. References
Heise, S., F. Krüger, M. Baborowski, B. Stachel, R. Götz, and U. Förstner. 2008. Assessment of risks from
particle bound substances in the Elbe river basin (in German). Commissioned by Hamburg Port
Authority and River Basin Community (FGG) Elbe. Hamburg, Mai 2008, p 349.
Wechsung, F. and A. Becker, Eds. (2005). Auswirkungen des globalen Wandels auf Wasser, Umwelt und
Gesellschaft im Elbegebiet. Berlin, Weißensee Verlag.
Acknowledgement - The authors wish to thank Jos van Gils (Deltares) and Monika Donner (DHI) for their
assitance with model development. This work has been funded by the German Federal Hydrology Ministry
(BfG)
Dioxin-like activity of sediment samples from the Elbe
River and soil samples from the Elbe associated flood area
Kathrin Eichbaum1, Thomas-Benjamin Seiler1, Steffen H. Keiter1, Kerstin Winkens1,
Markus Brinkmann1, Gunther Umlauf2, Burkhard Stachel3, Georg Reifferscheid4,
Sebastian Buchinger4 & Henner Hollert1
1
Institute for Environmental Research, RTWH Aachen University, Worringerweg 1, 52074 Aachen,
Germany
2
Rural, Water and Ecosystem Resources Unit, Institute for Environment and Sustainability, European
Commission - Joint Research Centre, T.P. 290, Via E. Fermi 2749, I-21027 Ispra (VA), Italy
3
Ministry of Urban Development and Environment Hamburg, Dep. of Water Management, Billstrasse
84, 20539 Hamburg, Germany
4
German Federal Institute of Hydrology (BfG), Am Mainzer Tor 1, 56068 Koblenz, Germany
[email protected]
While the water quality of many European streams improved in recent years, sediments as sinks and
potential sources of secondary pollutions are underestimated by the European Water Framework
Directive [1]. Extreme flood events like the Elbe flood of August 2002 pose a risk to remobilize
sediment associated persistent organic pollutants (POPs) like dioxins and other dioxin-like compounds
that can adversely affect wild life and humans[2,3,4]. The aim of the present study was to compare the
dioxin-like activity of Elbe sediments from the year 2008 and soil samples of the Elbe associated flood
area, sampled after the Elbe flood of August 2002.
The Sediments from the year 2008 originated from Czech and German river sections as well as from
the North Sea. They were considered to reflect the present pollution of the Elbe River without a
mentionable flood influence. In contrast, the soil samples of the Elbe associated flood area – taken
after the Elbe flood of August 2002 – rather represented the remobilization potential of POPs during
extreme flood events. The soil samples originated from three transects. Transect Glinde was located
downstream the confluences of the Elbe tributaries Mulde and Saale, whilst the transects Mulde und
Wörlitz were located close to and upstream the Mulde confluence, respectively. Hence, they also
might indicate the influence of the tributaries on the contamination of the Elbe River [5].
All samples were freeze-dried and extracted with n-hexane:aceton (1:1, v/v) by means of pressurized
liquid extraction. To assess their dioxin-like activity, two different in-vitro bioassays were used, the
7-ethoxyresorufin-O-deethylase (EROD) assay with liver cells from the rainbow trout (Waterloo 1;
RTL-W1) and the H4IIE-luc assay with the eponymous transfected rat hepatoma cell line. Both
bioassays are based on the activation of the Aryl hydrocarbon receptor (AhR) [5]. This receptor is
activated by many POPs, which are structurally related to the most potent AhR-inducer 2,3,7,8tetrachloro-dibenzo-p-dioxin. This study is focused on polychlorinated dibenzo-p-dioxins and
polychlorinated dibenzofurans (PCDD/F), dioxin-like polychlorinated biphenyls (DL-PCB) as well as
polycyclic aromatic hydrocarbons (PAH), which are all known to activate the AhR [6,7].
All samples showed elevated dioxin-like effects in both bioassays, except for two samples taken from
the North Sea, which only showed dioxin-like activity by means of the EROD assay. In general, both
bioassays gave a high correlation (r p = 0.77), but EROD data were constantly higher. The calculated
Bio-TEQs ranged between 1307 pg/g dw and 10462 pg/g dw and identified the Czech sampling site
Lysa nad Labem as the highest contaminated site, which was probably caused by several industrial
plants in this area – such as the chemical plant Spolana [8]. The lowest effectiveness was found for
the samples taken at Klavary (Czech Republic) and Elbhafen Brunsbüttel (Germany) in the EROD and
H4IIE-luc assay, respectively. The floodplain soils (Bio-TEQs from 1058 pg/g dw to 13244 pg/g dw)
were contaminated at levels, which were in the same order of magnitude compared to the sediments
of the Elbe River. A flood influence, however, could only be supposed due to missing pre-flood data
and soil sampling depths (10-30 cm) that gave no information about freshly deposited matter during
the Elbe flood of August 2002. Moreover, an influence of the Elbe contamination through its tributaries
could not be found, because all transects exhibited comparable Bio-TEQs. A normalization of the Bio-
TEQ values to their respective total organic carbon contents revealed two North Sea samples to be
the highest contaminated samples of the entire sampling campaign. This contamination is probably
caused by former dumping of sewage sludge from Hamburg in this area, but could also reflect a
contamination of the North Sea by the whole Elbe River [5,9].
A further comparison of Bio-TEQs with their respective Chem-TEQs [2,3,4] for the sum of PCDD/F and
DL-PCB was performed. The Chem-TEQs accounted between 0.1 % and 11.9 % for the observed
Bio-TEQs, indicating that the majority of the observed in-vitro effects was caused by non-priority and
non-persistent pollutants. Moreover, distinctly lower percentages were found for sediments than for
soils, which gives evidence that PCDD/F and DL-PCB are more important for the contamination of
floodplains.
Three extracts of the transects Glinde and Mulde were chosen for a multilayer fractionation to
eliminate the moderately persistent pollutants. In doing so, 73 % (Glinde) and about 20 % (Mulde) of
the calculated Bio-TEQs could be explained by the respective Chem-TEQs, more precisely by the sum
of PCDD/F and DL-PCB. An additional active coal-celite fractionation divided the extracts in PCB and
PCDD/F-containing fractions. The PCB-containing fractions showed no effects, whilst between 8 %
and 57 % of the Bio-TEQs accounted for the Chem-TEQs of the PCDD/F-containing fraction. The
fractionation gave clear evidence that the majority of dioxin-like activity in Elbe sediment and soil
samples of the associated flood areas was caused by non-priority and non-persistent pollutants.
Several samples of the entire campaign will further be investigated via the Micro-EROD assay with
H4IIE cells to support the results from the former bioassays. Further investigations also will have the
aim to close the gap between in-vitro and in-vivo investigations by focusing on the influence of
bioavailability and ingestion pathways in fish.
The present study shows the importance of ecotoxicological sediment assessment, especially in times
of climatic change and increasing flood events. It also clearly shows evidence that chemical
investigations can underestimate the toxic potential of sediments due to missing data of non-classical
AhR-inducers. As a consequence, bioanalytical endpoints should also be considered when assessing
the toxicity of complex environmental samples.
Keywords: Bio-TEQs, Chem-TEQs, Micro-EROD, multilayer and active coal-celite fractionation
[1]
[2]
[3]
[4]
[5]
[6]
[7]
[8]
[9]
Hollert H, Heise S, Keiter S, Heininger P, Förstner U (2007): Wasserrahmenrichtlinie —
Fortschritte und Defizite. Umweltwissenschaften und Schadstoff-Forschung 19, 58-70
Stachel B, Jantzen E, Knoth W, Kruger F, Lepom P, Oetken M, Reincke H, Sawal G, Schwartz
R, Uhlig S (2005): The Elbe flood in August 2002 - Organic contaminants in sediment samples
taken after the flood event. Journal of Environmental Science and Health Part aToxic/Hazardous Substances & Environmental Engineering 40, 265-287
Stachel B, Christoph EH, Götz R, Herrmann T, Krüger F, Kühn T, Lay J, Löffler J, Päpke O,
Reincke H, Schröter-Kermani C, Schwartz R, Steeg E, Stehr D, Uhlig S, Umlauf G (2006):
Contamination of the alluvial plain, feeding-stuffs and foodstuffs with polychlorinated dibenzop-dioxins, polychlorinated dibenzofurans (PCDD/Fs), dioxin-like polychlorinated biphenyls (DLPCBs) and mercury from the River Elbe in the light of the flood event in August 2002. Science
of The Total Environment 364, 96-112
Umlauf G, Bidoglio G, Christoph EH, Kampheus J, Krüger F, Landmann D, Schulz AJ,
Schwartz R, Severin K, Stachel B, Stehr D (2005): The Situation of PCDD/Fs and Dioxin-like
PCBs after the Flooding of River Elbe and Mulde in 2002. Acta hydrochimica et hydrobiologica
33, 543-554
Umlauf G, Mariani G, Skejo H, Mueller A, Amalfitano L, Stachel B, Goetz R (2010): Dioxins
and dioxin-like PCBs in solid material from the river Elbe, its tributaries and from the North
Sea. Organohalogen Compounds Vol. 72, 95-99
Hilscherova K, Machala M, Kannan K, Blankenship A, Giesy J (2000): Cell bioassays for
detection of aryl hydrocarbon (AhR) and estrogen receptor (ER) mediated activity in
environmental samples. Environmental Science and Pollution Research 7, 159-171
Safe S (1990): Polychlorinated Biphenyls (PCBs), Dibenzo-p-Dioxins (PCDDs), Dibenzofurans
(PCDFs), and Related Compounds: Environmental and Mechanistic Considerations Which
Support the Development of Toxic Equivalency Factors (TEFs). Critical Reviews in Toxicology
21, 51-88
Heinisch E, Kettrup A, Bergheim W, Wenzel S (2007): persistent chlorinated hydrocarbons
(PCHCS), source-oriented monitoring in aquatic media 6. strikingly high contaminated sites,
Freising
Stachel B, Mariani G, Umlauf G, Götz R (2011): Manuskript, Dioxine und PCBs in Feststoffen
aus der Elbe, ihren Nebenflüssen und der Nordsee (Längsprofilaufnahme 2008)
Can flood events affect rainbow trout? The biomarkercascade after exposure to PAHs in sediment suspensions
Markus Brinkmann1, Sebastian Hudjetz1, Jochen Kuckelkorn1, Michael Patrick Hennig1,
Catrina Cofalla2, Sebastian Roger2, Ulrike Kammann3, Markus Hecker4, John P. Giesy4,
Holger Schüttrumpf2, Amdreas Schäffer1, Henner Hollert1
1
Institute for Environmental Research, RWTH Aachen University, Aachen, Germany
Institute of Hydraulic Engineering and Water Resources Management, RWTH Aachen University, Aachen,
Germany
3
Johann Heinrich von Thünen-Institute (vTI), Hamburg, Germany
4
Toxicology Centre, University of Saskatchewan, Saskatoon, Canada
2
1. Introduction
In context of the ongoing scientific discussion about the potential ecotoxicological impacts of flood events, it
is of vital importance to understand the detailed mechanisms of contaminant uptake from suspended
particles and related effects in aquatic biota. An initial attempt to provide a comprehensive experimental
methodology for toxicity testing of sediment suspensions under realistic flood-like conditions was made by
the interdisciplinary Pathfinder project Floodsearch, which was funded by the German Excellence Initiative
[1]. Rainbow trout (Oncorhynchus mykiss) were exposed to artificial sediment that was spiked with polycyclic
aromatic hydrocarbons (PAH) under simulated 5 day flood conditions in an annular flume. A set of
biomarkers was investigated after exposure and the hypothesis that resuspension of sediments can lead to
accumulation of contaminants that are bound to particles and affect aquatic biota was verified [2].
A major limitation of the annular flume, however, was the fact that no animals can be collected during the
simulated flood-event. Thus, additional experiments were conducted to be able to follow temporal trends in
responses of biomarkers. As part of the follow-up project Floodsearch II, rainbow trout were exposed to
PAH-spiked and unspiked suspensions of natural sediment from the River Rhine in batch experiments.
Juvenile rainbow trout were exposed to suspensions of a natural sediment from the River Rhine
(Ehrenbreitstein Harbour), either un-spiked or spiked with a mixture of the PAHs pyrene (4.1 mg/kg d.w.),
phenanthrene (5.0 mg/kg d.w.), chrysene (3.3 mg/kg d.w.), and benzo[a]pyrene (8.3 mg/kg d.w.).
Experiments were conducted in 750 L glass fibre-reinforced plastic containers in which submersible pumps
were used to constantly suspend the sediments at a nominal concentration of 10 g/L with additional aeration
at a rate of 25 L/min. While temperature was at average 24 °C in the first experiment, tanks were cooled
using submersible coolers controlled by plug-in thermostats in the second experiment and kept at an
average of 12 °C for a second experiment. In each of the two experiments, physicochemical water
parameters were monitored. Sediment suspensions and fish (n=10) were sampled after 0, 1, 2, 4, 6, 8, and
12 days, respectively. Sediment suspensions were used to determine concentrations of suspended
particulate matter (SPM) and PAH concentrations in SPM. Several biomarkers, i.e. 7-ethoxyresorufin
O-deethylase (EROD) activity and lipid peroxidation (LPO) in liver, as well as PAH metabolite concentrations
in bile and frequency of presence of micronuclei in peripheral erythrocytes, were assessed in exposed fish.
Instrumental chemical analyses revealed that pyrene and phenanthrene concentrations in suspended solids
decreased over time, while no significant degradation was observed for chrysene and benzo[a]pyrene.
Concentrations of biliary PAH metabolites in fish increased slightly in the treatment without addition of PAHs
-1
-1
at 24 °C, while average levels increased to 166 µg ml for 1-hydroxypyrene (control value 4.6 µg ml ) and
-1
-1
17 µg ml for 1-hydroxyphenanthrene (control value 0.1 µg ml ) in the spiked treatment within two days,
followed by a decrease. In the experiment conducted at 12 °C accumulation of PAHs was slower (Figure 1).
With a latency of two days, the peak of metabolism was followed by a peak of LPO and micronucleus
frequency, indicating oxidative stress and DNA damage caused by the PAHs (Figure 2).
EROD was not significantly induced by the treatments. Significant differences were observed between the
1-Hydroxypyrene /
µg ml-1 bile
250
200
+
150
+
+
+ +
100
Unspiked
Spiked
+
50
10
5
*
*
*
4
6
8
*
0
0
2
10
12
Time / d
Figure 1 Absolute biliary metabolite concentrations during the
experiments (12°C) in the un-spiked (■) and the spiked treatments
(●). Symbols give the mean of n=10 animals, error bars the 95 %
confidence intervals. Asterisks denote significant differences
between control and un-spiked treatments (Kruskal-Wallis oneway ANOVA on ranks with Dunn’s post-hoc test, p ≤ 0.05), plus
symbols between control and spiked treatments.
Micronucleus frequency /
‰
bioavailability of freshly spiked and field-aged PAH contamination.
4
Unspiked
Spiked
3
2
*
1
0
0
2
4
6
8
10
12
Time / d
Figure 2 Micronucleus frequency in peripheral erythrocytes of
exposed animals during the experiments (12°C) in the un-spiked
(■) and the spiked treatments (●). Symbols give the mean of n=10
animals, error bars the SEM. The asterisk denotes a significant
difference between the un-spiked and the spiked treatment (pair
wise parametric t-test).
4. Conclusions
The biomarkers assessed in the present study responded in a cascade-like pattern. We thus recommend
that sediment resuspension experiments in the annular flume should be accompanied by selected batch
exposures to improve the understanding of the biomarker induction kinetics. Furthermore, field-aged
contaminated sediments should be preferred over the use of spiked sediments for mechanistic studies, since
the bioavailable fractions differ significantly. Temperature was also shown to be an environmental stressor
that could potentially lead to increased effects caused by particle-bound contaminants during flood events.
Availability of field-aged PAH contamination of the sediment from Ehrenbreitstein was comparably high,
which supported the hypothesis that even moderately contaminated sediments can pose a risk to aquatic
biota during resuspension and flood events.
5. References
[1] Wölz J, Cofalla C, Hudjetz S, Roger S, Brinkmann M, Schmidt B, Schäffer A, Kammann U, Lennartz G,
Hecker M, Schüttrumpf H, Hollert H. 2009. In search for the ecological and toxicological relevance of
sediment re-mobilisation and transport during flood events. Journal of Soils and Sediments 9:1-5.
[2] Brinkmann M, Hudjetz S, Cofalla C, Roger S, Kammann U, Zhang X, Wiseman S, Giesy J, Hecker M,
Schüttrumpf H, Wölz J, Hollert H (2010) A combined hydraulic and toxicological approach to assess resuspended sediments during simulated flood events. Part I–multiple biomarkers in rainbow trout. Journal
of Soils and Sediments 10:1347–1361
Acknowledgement - This study has been generously supported by a Boostfunds project of the Exploratory
Research Space (ERS) at RWTH Aachen University, as part of the German Excellence Initiative. Analyses of
hepatic gene expression at the University of Saskatchewan were perfomed in context of a travel grant given
by the German Academic Exchange Service (DAAD).
Reach Regulation :
Communication Behind the Information Needs
Philippe Garrigues1, Françoise Lafaye2 and Nicolas Léca3
1
2
ISM, Université Bordeaux 1, CNRS UMR 5255, 351 Cours de la libération, 33 400 Talence, France
ENTPE, Université de Lyon, CNRS UMR 5600, rue Maurice Audin, 69 518 Vaulx en Velin cedex, France
3
CRDEI, Université Montesquieu Bordeaux 4, avenue Léon Duguit, 33 608 Pessac, France
1. Introduction
The implementation of REACH aims a high level of protection of human health and the environment. Industry
has the burden to prove the safety use of products and to ensure control of any risk. If the information to
general public is also one of the major goals of the regulation, the first steps of implementation show different
ways to experience information diffusion [1], [2].
These works were conducted in the framework of a research project « How to use REACH. How the
stakeholders approach REACH through technical classes". Discussion meetings have been organised with
the parties who have designed the regulation and who are assuming the control of its implementation, but
also people who are implementing REACH in professional organisations and private companies: REACH
regulation designers, chemical producers, chemical industry trade-unions, state-member representative.
Meetings have been planned with the parties who have designed the regulation (members of European
parliament) and who are assuming the control of its implementation (ECHA, ANSES and industrials), but also
people who are applying REACH in professional organisations and private entities (professional sectors,
commercial companies, CEFIC and CONCAWE) and to have interviews with them about encountered
difficulties, found or expected solutions.
During interviews, it appears that problems are not shared on a common basis by all the actors. Some of
them are focused on articles, others on UVCB, and still others
The objectives of that presentation is to show that questions related to the informations on chemicals are
present in the REACH regulation , but communication that supposes interaction with multiple stakeholders
with different aims and strategies is at the early stage : which stakeholders, which topics, how these
exchanges have influence on decision making ?
Communication is the way to transfer informations to somebody. This is the means to diffuse informations to
various ans specific people. Finally this is also the way for somebody or for an entity to inform and to
promote its actions towards the général public.
Information on/into REACH, at this early stage (registration), is not a simple question and REACH proceeds
by the "Learning by doing » approach. Information on REACH and products is circulating to some extend
into topical focused forums (SIEF) and into chemical consortia, but it appears that different stakeholders
don't share the same points of view (what kind of questions for which recipients?) [3], [4]. Communication to
public which is also an important objective of the regulation is far beyond what maybe expected at that stage.
The next steps (evaluation, authorization) surely need to plan communication towards the public, outside
from the closed and confined space of acting stakeholders [5], [6], [7].
In an other way, stakeholders will implement actions as they concieve phenomenons. So, the actors’
analyses of the implementation of REACH is different. If we ask for difficulties of implementation, the
answers are different and are related to their own interests and positions.
4. Conclusions
Communication between ECHA and stakeholders seems to be effective through the various Committeees
and the Forum . However downstream communication into the information chain, particularly towards the
general public, needs to be developped. The participation of public into debate about technical and scientific
issues of REACH regulation is also questionnable.
5. References
[1] Béal S. et al., « Les informations exigées par la législation REACH: Analyse du partage des coûts»,
Revue d'économie politique, 2010/6 Vol. 120, pp. 991-1014.
[2] Jouzel J-N. et Lascoumes P., « Le règlement REACH : une politique européenne de l'incertain. Un détour
de régulation pour la gestion des risques chimiques », Politique européenne, 2011/1 n° 33, pp. 185-214.
[3] Romi R., "A propos du projet REACH : quel rythme pour quelle révolution sanitaire", RDSS 03/2006, n°2,
pp. 238-247.
[4] Espuny Pascual C., « Controverses autour de Reach » Illustration de communication, débats et
controverses sur « Reach » dans le processus de responsabilisation des entreprises sur le double plan
social et environnemental, Revue internationale de Psychosociologie, 2010/38 Vol. XVI, pp. 85-98.
[5] Jacquet C., Saint Girons A-L., « REACH » : un monument réglementaire et son impact sur l'activité
juridique des entreprises, JDE, 10/2008, n°152, pp.233-239.
[6] Gisquet E., Goldberg S., Canet C., Brixi O. "REACH: un programme européen de gestion renouvelée des
risques sanitaires", Santé Publique, HS2008, Vol.20, pp 191-200.
[7] Hershtl J., Harrouet F., REACH, conséquences du défaut d'enregistrement : « Quand le Mieux reste
l'Ennemi du Bien », BDEI, 09/2010, n°29, pp.38-42.
Acknowledgement - The authors thank the CNRS for funding that project on REACH (PIR ISCC/CNRS2011)
Research findings and decision making: the case of renewable
energy
Valentina Castellani1, Andrea Piazzalunga1 and Serenella Sala1
1
University of Milano Bicocca, Dpt. of Environmental Science, Piazza della Scienza 1, 20126 Milano, Italy
1. Introduction
Scientific research can have a role in the promotion of more sustainable patterns of consumption and
production because it can provide information aimed to raise awareness about the impacts of different
behaviours and to support more sustainable choices from different kind of actors. The challenge posed to
science in this context is to provide information that is effectively supporting for decision making processes at
different scales and that can be easily understood by all the stakeholders involved in the process (policy
makers, entrepreneurs and citizens). In recent years the attention of citizens on the issues of sustainability,
environmental impacts and sustainable behaviour has grown considerably [1]. In parallel the demand for
scientifically sound and transparent information upon which to base consumption choices and behaviours is
growing among citizens.
Nonetheless, it is difficult to think that there may be a direct contact between those who do research and who
makes the decisions: firstly, because in most cases there are no opportunities for direct contact (e.g. to
citizens) and, secondly, because is necessary to translate the information resulting from scientific research in
a language that is understandable. The lack of on communication between science, policy and citizens
communication can lead to not evidence-based decision making, lack of trust and unsustainable behaviour
due to low level of information and awareness. One possible way of success in environmental
communication could be represented by the presence of those intermediaries who have relationships with
key stakeholders and are able to translate information for them so that they become understandable and
translatable into action.
2. Research findings and decision making in the wood energy sector
An interesting example in this context is the use of renewable energy sources, particularly from woody
biomass. Energy resources play a fundamental role in the world's future. There are many alternative
renewable energy sources which can be used instead of fossil and conventional fuels. The decision as to
what types of energy source should be utilized must, in each case, be made on the basis of economic,
social, environmental and safety considerations. In this perspective, the extensive use of biomass, both in
the residential sector and for electricity production in power plants, is considered as an alternative option.
The attention of the scientific community and policy towards the use of wood as a heating source is steadily
increasing. However, if not managed in a sustainable way, the use of wood can cause direct impacts from
extraction and distribution (such as excessive harvesting leading to clearance, soil erosion, loss of forest
area and loss of biodiversity in plantations and secondary forest) and disturbance to material cycles (such as
the reduction of the carbon storage function). Moreover, while the contribution of this source of energy can
be considered zero impact on the emission of climate-altering gases, although its use, particularly in low-tech
plants, appears to contribute significantly to air quality both in urban than in rural areas. Therefore it is
necessary that the choice to promote wood use as a renewable energy source is made taking into account
sustainability in a broader sense, considering, for instance, the availability of resources at local scale and the
environmental impacts throughout the whole supply chain, in a life cycle perspective [2].
As far as Italy is concerned, preliminary results of a study carried out by ENEA and APAT show that the
expansion of domestic burning of wood logs, briquettes, chips and pellet will contribute for a significant quote
to the renewable share of the energy budget [3]. However, it is well known that biomass combustion in
domestic appliances is a significant emission source for air pollutants. Together with NOx and CO, wood
smoke consistently contains particulate matter (PM), contributing for a relevant share to the total PM2.5 and
PM10 emitted [4]. This ambiguity between the positive and the negative aspects of wood burning, if not
adequately integrated by information about the specific conditions that influence pollution levels, can lead to
opposite political decisions about the use of wood in local energy plans. In Italy, for instance, Piedmont
Region has a strong policy of incentives for wood energy appliances with no restriction criteria (e.g. about
the distance of supply or the type of appliance), while Lombardy Region applies a strict regulation that
prohibits wood burning in the whole Po valley, without differences between old and new boilers or the
presence of filters, to prevent PM emissions.
Thus the most effective approach to tackle this problem should be to identify which are the critical aspects
and which are the operative solutions that can help to maximize the benefits from using wood as a
renewable energy source, i.e. defining warnings and guidelines that should be followed by decision makers
at several levels in order to ensure sustainability of these systems. Table 1 takes the wood-energy supply
chain as an example to explain what should be the actions to be undertaken in order to make the best use of
scientific research findings as a support to decision making in all the stages of the supply chain.
Elements of
the supply
chain
Critical issues (as result
from scientific research)
Fossil fuel consumption
(chainsaws and other
machinery)
Wood cutting
Burning
systems
Guidelines
Choose the most efficient
harvesting system (e.g.
debranching before cable
logging helps to reduce fuel
consumption)
Impacts on the forest
ecosystems (e.g. oil
emission, clearance, soil
erosion, etc.)
Follow sustainable forestry
criteria (e.g. selective cutting
instead of clearcutting)
Deforestation
Respect the carrying
capacity of the forest
ecosystem
Inefficient combustion
leads to higher emission
Develop new and more
efficient technologies
PM air emissions
Install appropriate filters
Oversizing can lead to
request more biomass
than what is available in
the area
Prefer distributed generation
instead of big plants
Inefficient combustion
leads to higher emission
Prefer boilers instead of
open fireplaces
PM air emissions
Substitute old appliances
with new ones (safer and
more efficient)
Wood burning
Wood logs/chip/pellets
transport implies fossil
fuel consumption
Target
audience
Forest
managers and
wood cutters
Possible mediator
Forestry consortium
(including foresters and
forestry firms)
Forest
managers
Boilers
producers
Users
Producers association
Citizens/consumers
association
Buy wood from a short
supply chain
Table 1: Critical issues, guidelines for sustainable use of wood as energy source and actors to be involved
3. Conclusions
The communication of the results of scientific research is an important element in decision making. However,
often does not happen because there is no contact between those who produce information and who should
use and why those who do research do not always bother to translate their findings into practical information
to support decisions (e.g.” wood burning can be a source of pollutants if not made properly; the guidelines to
be follow in order to make a sustainable use of wood resource are...”). Therefore, for each field of
environmental research is necessary to: identify the critical aspects (through scientific research); identify the
relevant actors in the sector / industry and possible existing intermediaries (e.g, trade associations,
consumer associations, etc.); define guidelines for decision makers and communicate them to the actors
previously identified.
4. References
[1] Hargreaves T, 2011. Practice-ing behaviour change: applying social practice theory to pro-environmental
behaviour change. J. of Consumer Culture, 11 (1), pp. 79–99.
[2] Sala S, Castellani V. 2011. Energy production from biomass: technology sustainability assessment to
support decision making at local scale. Int. J. Sustainable Development and Planning, in press.
[3] Pignatelli T, D’Elia I, Vialetto G, Bencardino M, Contaldi M. 2008. The use of bio-mass: synergies and
trade-offs between Climate Change and Air Pollution, in Italy. 17th Annual International Emission
Inventory Conference, Portland, OR, US Environmental Protection Agency.
[4] Piazzalunga, A., Belis, C., Bernardoni, V., Cazzuli, O., Fermo, P., Valli, G., Vecchi, R., 2011: Estimates
of wood burning contribution to PM10 in Lombardy (Po Valley, Italy) using different approaches, Atmos.
Environ., 45, 6642-6649.
Challenges of Integrating Science and People within a Network
of Excellence
T.G. Hinton1, L. Bouhouch2, O. Brenna,2 D. Echenique,2 Z. Jones,2 F. Siracusa,
1
Institute of Radiation Protection and Nuclear Safety, Cadarach, France
2
IAE School of Change Management, Aix-en-Provence, France
1. Introduction
1
A memorandum of understanding was recently signed by the heads of nine organizations indicating their
intention to bring together, with an aspiration of sustainability, part of their respective R&D programmes into
an integrated trans-national programme to support radioecology. Radioecology is a multidisciplinary science
concerned with the fate and effects of environmental radioactive contaminants on humans and biota.
Motivation for the integration was recognition that the science was suffering from several problems:
•
•
•
•
•
fragmentation, with poor coordination among the national strategies and programmes,
a steadily decreasing funding base,
closure of important infrastructures,
retirement of key personnel,
poor recruitment of young scientists.
A Network of Excellence in radioecology was formed (called STAR: Strategy for Allied Radioecology) and
structured to substantially reduce the above threats and prevent the further decline of radioecology within
Europe.
Management personal within STAR were confident of the science associated with the NoE, but were less
sure of the mechanisms required for sustained integration. Integration on the scale envisioned by STAR
requires new ways of thinking by scientists, supported by an effective management and governance
structure. The integration within the NoE resembled, in the business world, a partial merger of divisions from
within different organizations. To assist the NoE, advice was sought from the IAE business school in Aix-enProvence, France. The IAE curriculum specializes in change management. The IAE faculty also recognized
the challenges faced by STAR and decided to make it a special project for five of its MBA students. This
presentation highlights some of their findings on the challenges of integrating science and people.
2. Methods
The MBA team recognized the need for STAR to embrace some of the fundamental principles of change
management. Studies of past attempts at mergers and acquisitions in the business world indicated that
failure is often due to overlooking and/or taking for granted the fundamentals of change management. Such
fundamentals include a shared vision, an understanding of the cultural differences involved in the project,
and that all participants have similar expectations about what they as individuals and their respective
organizations might obtain from the integration. Five concepts were highlighted by the MBA team as key
change management principles that the NoE needed to develop in order to have a successful sustained
integration. Their advice is applicable to other organizations that are attempting to integrate science and
people.
1
Institute for Radioprotection and Nuclear Safety (IRSN) France; Radiation and Nuclear Safety Authority (STUK) Finland; Belgium
Nuclear Research Center (SCK-CEN) Belgium; National Institute for Energy, Environment and Technology (CIEMAT) Spain; Stockholm
University (SU) Sweden; Federal Office of Radiation Protection (BfS) Germany; Norwegian Radiation Protection Authority (NRPA)
Norway; Norwegian University of Life Sciences (UMB), Norway.
Natural Environmental Research Council, (NERC), UK
3. Results
3.1.
Define a vision:
A clear vision can help everyone understand why you're asking them to do something. When people see for
themselves what you're trying to achieve, then the directives they're given tend to make more sense.
3.2.
Quick wins:
Nothing motivates more than success. It is important to give the participants in the NoE a taste of victory
early in the change management process. Within a short time frame have results that the staff can see (this
could be a month or a year, depending on the type of change). Without this, critics and negative thinkers
might hurt the change management progress. When positive changes have been observed, let the team
know about them, and appreciate those who contributed the most: Acknowledge the work and the people
involved that make things happen. Facebook, Blogs and newsletters are a great way to do this. Celebrate
your early victories, no matter how small they might be.
3.3. Make glue:
Glue is the social link and understanding between people in the organization. The MBA team strongly
believed that the NoE scientific ambitions would be difficult to achieve if glue elements were not properly
addressed. To develop stronger bonds among participants:
•
publish a newsletter for NoE members, highlighting important work, people and ideas. This will give
people recognition and possibly stimulate future ideas and work.
•
build a sense of community among STAR participants, have t-shirts made with the NoE logo or
bumper stickers on your cars (“Radioecologists do it in ionizing ways”)
•
Reinforce the need for social bonding. Use team and trust building sessions in plausible ways..
•
Loosen up social stiffness. Find a way to facilitate communication in meetings. Using small work
groups, Post-Its, brainstorming sessions without criticism, etc.
3.4. Create a sense of urgency:
To make change and integration happen, the entire organization must be looking for it. Therefore, a sense of
urgency is crucial to change. The sense of urgency will make people realize that change is mandatory and
this motivation will get things moving. For a NoE to reach its full potential, ways must be found to integrate
on a person-to-person level. Without the sense of urgency, this won‘t happen.
STAR is a NoE concerned with environmental radiation. The nuclear accident at Fukushima had both good
and bad effects on radioecology and STAR. Before Fukushima, the science of radioecology was in a state of
decline. Funding was decreasing, and recruiting new and talented students was becoming more difficult.
There was a sense of urgency that this needed to change, and therefore STAR was created. Fukushima
changed all of this. It immediately raised the profile for radioecology and reminded the world why this science
is important. Therefore, for radioecology as a whole, it had tremendously positive effects. Impact of
Fukushima: For STAR, however, the effects are quite the opposite. Instead of feeling the pressure to build a
sustainable organization to get funding in the future, now funding is practically guaranteed. Therefore, people
have lost the sense of urgency, or the incentive to really put in the extra effort to integrate on anything more
than a scientific level. Also, if STAR were to fail, the individual members would still have jobs at their
respective organizations. This hurts their motivation to really exert the energy and time to commit 100% to
STAR‘s efforts.
3.5.
Build a coalition for change:
Change management requires champions. We need people who believe that the NoE can only reach its true
potential if we integrate on a person-to-person level. Ideally, these would be people that have some influence
within the NoE (e.g. work package leaders). However, it does no good to use these people if they don‘t
believe in the integration. It is crucial to form a “guiding coalition” to spread these ideas throughout their
organizations. If it is not the WP leaders, then at least one member from each of the member organizations
should be found to champion the value of integration. The process of building a coalition involves:
• identifying the true leaders in the NoE
•
asking for an emotional commitment from these key people.
•
working on team building among the change coalition members.
•
checking your team for weak areas, and ensure that you have a good mix of people from different
departments and different levels within the NoE.
4. Conclusions
Such people attributes of integration are foreign to most NoE coordinators that come from a scientific
background. Likewise, they are not principals that most of the scientific staff will be aware of, or even willing
to embrace. Change is difficult. To make integration and change effective people must recognize the
benefits and be willing to think and act in new, innovative ways. The lessons learned from the business world
on change management could be of great value to those trying to integrate science and people.
Acknowledgement - The authors thank the EC FP7 Euratom programme for funding the STAR NoE......
Keep your boots muddy
Rüdiger Berghahn1 and Anne Rinn2
1
German Federal Environment Agency, Schichauweg 58, 12307 Berlin, Germany
2
Atelier: Langhansstraße 7A, 13068 Berlin, Germany
1. Introduction
The German Federal Environment Agency runs a site for aquatic simulation in the very south of Berlin, which
also includes a set of outdoor and indoor artificial pond and stream mesocoms (FSA) for ecotoxicological
research (http://www.umweltbundesamt.de/wasser-und-gewaesserschutz/fsa/ ). Up to the present, 30
studies were carried out in that facility during the last 10 years. The results have been popularised via the
official homepage, guided tours, conference posters, talks and sessions, scientific journals, magazines,
newpapers, and TV features. In search for further means to reach a wider audience apart from
ecotoxocological professionals and people interested in natural sciences, formerly gained positive
experience in the cooperation between science and art (Mohr 2007) was picked up and the idea of the artist
Anne Rinn (www.anne-rinn.de) to stage an exhibition and to create a film entitled ‘Keep Your Boots Muddy’
was supported by the FSA team. Both exhibition and film pivot on the triangle nature-simulation-art with
simulation trying to create artificial nature. The presentation will be the 10 minutes version of the German film
with English subtitles.
The film views the FSA with different eyes, not with the eyes of a natural scientist or engineer but the eyes of
the artist Anne Rinn. Pictures of the device and the divers work activities were set to offstage voices of the
FSA team members explaining their tasks and their philosophy. The viewer has the impression that they are
talking to themselves. This inner dialog allows us to gain insight into their ritualized and repetitive work
processes and to understand how intuition might also be a leading force in science.
The German 2 minutes version of the film was used as a teaser for the exhibition on youtube from November
to December 2010. It was viewed more than 500 times and the exhibition was a success even though the
location in Berlin was not optimal with regard to accessability with local transport and opening hours.
Moreover, the film was presented at the German 28.Kasseler DokFest 2011 (www.filmladen.de). It has also
routinely been shown as introduction to public visitors of the FSA, for example on open days.
Figure 1: Nature-Simulation-Art.
4. Conclusions
The combination of science and art may help to open up ecotoxicological issues to the wider public. If this is
done by means of contemporary media in a popular way, one can avoid the risk to just move from one ivory
tower to the next and may create a win-win situation for both science and art.
5. References
Mohr, S. (2007): Dialog zwischen Wissenschaft und Kunst. BIOspektrum 4, 462.
Acknowledgement - The authors thank Gabriele Coan for the translation of the subtitles.
Benefits of ‘EFSA Risk Assessment for Birds and Mammals’
guidance document for ecological refinements
Rachel Sharp, Domenica Auteri, Javier Herranz Montes, Stephanie Bopp, Csaba Szentes,
Franz Streissl
European Food Safety Authority (EFSA), Pesticides Unit, Largo N. Palli 5/A, I - 43121 Parma, Italy
1. Introduction
In order to gain EU Level approval of an active substance under Regulation (EC) 1107/2009, and previously
EU Directive 91/414/EEC, it is necessary to demonstrate that there are no unacceptable risks to bird and
mammals. Since the early 2000’s to 2009 the main guidance document used was the ‘SANCO/4145
Guidance Document on Risk Assessment for Birds and Mammals Under Council Directive 91/414/EEC’,
1
European Commission (2002) (hereafter referred to as SANCO/4145). However, in 2009 the European
Food Safety Authority (EFSA) issued a new guidance document (‘Risk Assessment for Birds and Mammals’,
2
EFSA (2009) ) which aimed to update and improve bird and mammal risk assessment in the EU. Both
guidance documents follow a tiered risk assessment whereby a first tier risk assessment seeks to address
low risk situations. However, should a first tier assessment fail to be sufficient to demonstrate a low risk then
it is necessary to refine the risk. There are various approaches outlined in both guidance documents
including the use of ecological data to gain a better understanding of whether key focal species are at risk.
This paper will outline the key differences between the two guidance documents and highlight the benefit of
risk assessment methodology in EFSA (2009) in relation to ecological refinements. An analysis of a
hypothetical case will demonstrate how the first tier risk assessment in EFSA (2009) provides valuable
information for the refined assessment. An example will be given to show how improvements in the use of
ecological data and uncertainty analysis will be used in the overall characterisation of the risk.
2. Outline of benefits for first tier risk assessment in EFSA 2009
2.1. Key differences
The main principles of first tier risk assessment methodology in both SANCO/4145 and EFSA 2009 are
similar. Toxicity Exposure Ratios (TER) are calculated and compared to trigger values. Previously
SANCO/4145 used general ‘indicator species’ and, mostly, did not differentiate between crop growth stages.
However, the EFSA (2009) risk assessment introduces multiple ‘generic focal species’, mixed diet, spray
intercept and uses the crop growth stage for a more targeted, specific first tier assessment. The outcome of
the first tier risk assessment in EFSA (2009) indicates the feeding guilds of concern which require higher tier
risk assessment.
2.2. Example first tier risk assessment under SANCO/4145 and EFSA (2009)
Consider the example where the representative use is a single application to maize at 0.5 kg a.s./ha in the
UK. The application can be made any time between growth stages BBCH 30-45. The reproductive NOEL
for birds is 20 mg a.s./kg bw/day.
The first tier risk assessment according to SANCO/4145 results in TER values of 2.5 and 1.3 for a medium
herbivorous bird and an insectivorous bird, respectively. A higher tier risk assessment is therefore triggered
for both indicator species. To define the key focal species for the higher tier assessment the risk assessor
must now consider what birds utilise the crop at the growth stage and time of year. It is anticipated that
these will include multiple focal species including herbivorous, omnivorous and insectivorous birds.
The first tier risk assessment according to EFSA (2009) requires TER values for seven generic focal species
to be calculated but, in this case, only one of the TER value is less than the trigger value. A refined
assessment is therefore also triggered; however, the first tier assessment is significantly more informative
and indicates that the risk assessment should focus on a medium herbivorous bird such as a pigeon
potentially consuming 100% leaves. In this instance, there is no need to further consider an omnivorous bird
and an insectivorous bird due to the mixed diet and crop intercept used in the risk assessment.
Furthermore, the assessment indicates that a simple change to the proposed timing of application, to growth
stages BBCH >40, will negate the need for any higher tier risk assessment.
3. Outline of key improvements for refined risk assessment in EFSA 2009
3.1. Ecological parameters
Refinement of ecological parameters was a commonly used approach to higher tier risk assessment under
SANCO/4145. Information was used to identify key focal species, defined as species of birds or mammals
which are most abundant and prevalent in the crop. Ecological information was then used to refine
ecological parameters PT and PD. PT is defined as the proportion of an animal’s daily diet obtained in the
habitat treated with the pesticide. PD is defined as the composition of diet obtained from the treated area.
Limited guidance was provided in SANCO/4145 and resulted in difficulties for risk assessors designing and
interpreting ecological studies or utilising existing information in the literature. EFSA (2009) provides much
more detailed guidance on the conduct and interpretation of higher tier ecological studies and also the
application of such studies to risk assessment (focal species, PT and PD). It is hoped that this will aid the
development of more robust risk assessments and consistency in evaluation.
3.2. Risk characterisation and uncertainty analysis
A further improvement to risk assessment methodology outlined in EFSA (2009) is that it is necessary to
conduct a risk characterisation and uncertainty analysis for every refined risk assessment. It is, of course,
common practice for experts to consider the quality, scientific value and appropriateness of data or evidence
for use in a risk assessment but the guidance given in EFSA (2009) will formalise this step. Communication
of risk is a critical aspect of any refined risk assessment where comparison to an assessment factor is no
longer suitable for ensuring that protection goals are met. Identifying uncertainty and characterising the risk
in relation to protection goals will allow for better dialogue between risk assessors and risk managers. In
addition it will allow risk assessors to make best use of data. For example, should a risk assessor dismiss
evidence from studies conducted in different crops for not being entirely appropriate, or is it better to utilise
the information, acknowledging the uncertainty, in an overall characterisation of the risk?
3.3. Example of risk characterisation and uncertainty analysis
Consider an example of a refined risk assessment for a woodpigeon in maize. Good quality focal species
and PT studies are available and were conducted in maize fields in the UK at the correct time of year and
th
growth stage. The mean and 90 percentile PT values are 0.7 and 0.8. A good quality residue decline study
is also available and indicates that the DT 50 is 2 days; however, the study was conducted in cereal foliage.
Using these data the resulting refined long-term TER value is 12.8.
The overall risk characterisation should take in to account all of the available information. The focal species
and PT studies are good quality and are representative; therefore, it is possible to be reasonably confident in
the data. However, as the PT values are quite high they have little impact on the TER value. The refinement
of only PT is not sufficient to conclude a low risk. The residue decline data indicate significantly faster
degradation than assumed in first tier risk assessment. However, it must be borne in mind that the data were
derived in cereal shoots and are therefore not entirely representative of the foliage a pigeon would consume
in maize fields at BBCH 30 - 40. Overall, whilst the appropriateness of the residue decline study creates
some uncertainty with the quantified TER value, the refined risk assessment for woodpigeon is considered to
demonstrate a low long-term risk.
4. Conclusions
The new EFSA guidance document will help applicants and risk assessors to produce reliable, robust and
informative risk assessments. It provides more clarity at the first tier, for refined assessments and for the
overall communication of the risk.
5. References
[1] European Commission, 2002. Guidance Document on Risk Assessment for Birds and Mammals Under
Council Directive 91/414/EEC. SANCO/4145/2000.
[2] EFSA (European Food Safety Authority), 2009. Guidance Document on Risk Assessment for Birds and
Mammals on request of EFSA. EFSA Journal 2009; 7(12):1438.
Risk assessment of birds and mammals exposed to plant
protection products in the Nordic and Baltic countries
Nina Åkerblom1, Alf AAgaard2, Leona Mattsoff3, Dace Bumane4, Zita Varanaviciene5, Marit
Randall6
1
Swedish Chemicals Agency, P.O. Box 2, SE-172 13 Sundbyberg, Sweden
2
Danish EPA, DK-1401 Strandgade 29, Copenhagen K, Denmark
3
Finnish Safety and Chemicals Agency, PL 66, 00521 Helsinki, Finland
4
State Plant Protection Service, Lielvardes street 36/38, Riga, LV-1006, Latvia
5
State Plant Service, Ozo 4a, LT - 08200, Vilnius, Lithuania
6
Norwegian Food Safety Authority, P.O. box 3, N-1431 Ås, Norway
1. Introduction
Regulation EC 1107/2009 concerning the placing of plant protection products (PPP) on the market in the EU
entered into force 14 June 2011. A central aspect of the new regulation is the view of the principle of mutual
recognition; authorizations granted by one Member State should be accepted by other Member States where
agricultural, plant health, and environmental (including climatic) conditions are comparable. Therefore, the
Community is divided into zones (Northern, Central, and Southern zone) with such comparable conditions in
order to facilitate mutual recognition.
To facilitate the mutual recognition and the work-sharing process within the Northern zone, the Nordic and
Baltic countries have regular meetings discussing harmonization of the risk assessments. Additionally,
several pilot product applications have been evaluated according to the new principals with a zonal
rapporteur Member State prior to the entry into force of EU regulation EC 1107/2009. Agreements among
the Member States during this process are compiled in the guidance document for the Northern zone [1]
which is available at all Member States authorities’ home pages.
2. Harmonized risk assessment of birds and mammals
To harmonize the initial risk assessment of birds and mammals, the Northern zone has agreed upon using
the EFSA Guidance document (2009) [2]. If the initial assessment by using the screening step and the tier I
approach in the EFSA Guidance document indicates a potential risk, a more refined risk assessment can be
provided based on more realistic assumptions regarding exposure and effects. Several of the refinement
options are related to which species that should be used in the risk assessment (focal species) and the
ecology of these species, e.g. proportion of an animal’s daily diet obtained in habitat treated with pesticide
(PT) and composition of diet obtained from treated area (PD). Other refinement options consider the
properties of the substance, e.g. degradation time (DT50) in/on treated crops and weeds (residue decline).
The Northern zone Member States are continuously working in order to further harmonize also the refined
risk assessment and has so far agreed upon the following:
2.1. Residue decline in plant material (DT50)
To replace the default DT50 value of residue decline used in the initial risk assessment in the EFSA
guidance document, results must be available from at least 4 study sites. This is considered to be consistent
with the data requirements for degradation in soil and residue trials (where 4 studies are required). If results
from 4-10 sites are presented, the longest DT50-value should be used. If more than 10 values are available,
generally the mean value can be used. In cases where a lower number of data are available that indicate a
very rapid residue decline, also other relevant information may be used as supporting data, such as
information on hydrolysis rate or volatilization. For herbicides, also information on wilting rate may be useful
for the estimation of possible exposure of herbivorous animals. Due to the uncertainties regarding the
relevance of such data under field conditions, a significant margin of safety must however be demonstrated
in the risk assessment.
Residue decline studies may only be used for refinement of DT50 if results are evaluated regarding their
validity for Nordic/Baltic conditions according to the EFSAs guidance document. For example, the following
parameters have to be declared and considered acceptable; experimental design, growth stages (to consider
dilution effects), and climatologically factors (e.g. rain and/or irrigation - the reliability of studies including
substantial rain/irrigation within degradation half time will be questioned due to the wash-off effect).
2.2. Proportion of diet obtained in treated fields (PT)
In the EFSA guidance document the default proportion of diet obtained in treated fields (PT) of birds and
mammals is 100%. This default value could be replaced by referring to e.g. radio-tracking studies including,
th
as a minimum, 10 individual animals. The 90 percentile PT value from these studies should be used in the
refined risk assessment. Studies to refine the PT may only be used in the risk assessment if they are
evaluated regarding their validity according to the EFSA guidance document.
2.3. Focal species and their feeding strategy (PD and PT)
The need for a common strategy in the Northern zone regarding the choice of focal species and their PD and
PT at higher tier risk assessments was discussed at the Nordic workshop, Copenhagen, 7-9 June, 2011. It
was agreed that a harmonization of focal species, PD and PT for higher tier risk assessment for birds and
mammals will enhance uniform and agreed assessments and ease evaluation of registration reports. The
latter is very important with the strict time lines in the new regulation (EC 1107/2009). At the meeting it was
agreed that the Danish guidance document [3], including focal species and relevant scenarios for different
crops could be a starting point for the development of a Northern zone guidance and a generic tool for higher
tier risk assessments.
The plan for the nearest future is that each Member State in the zone evaluate if the focal species used in
the Danish document could be considered to cover relevant species in the countries. Available information of
the species’ food selection (PD) and foraging time in treated fields (PT) will also be discussed. Agreed focal
species and endpoints will be used in harmonized scenarios for the refined risk assessments. The scenarios
will focus on main crops in the Northern zone and focal species releted to these crops. It will not be possible
to derive scenarios for all minor crops, as they may be more country specific and there will be a general lack
of knowledge on the relevant focal species and related parameters (e.g. dietary composition and foraging
activity).
Our goal is that a guidance for the choice of focal species and their PD and PT for refined risk assessment in
the Northern zone will be available in May 2012 when the SETAC conference is held. At a later stage a
generic tool for higher tier risk assessments will be developed.
2.4. Ecological modelling and body burden modelling
The Member States in the Northern zone has agreed that the use of ecological modelling in higher tier risk
assessments are generally not considered appropriate until commonly agreed models are available at
European level and Guidance Documents with criteria for assessing model output are available.
Similarly, refinements based on body burden modelling (ADME studies) are generally not considered
appropriate for the Northern zone until validated models and guidance for use are available.
3. Conclusions
The Nordic and Baltic countries aim at harmonizing risk assessments for birds and mammals by discussing
and agreeing on refinement options. A Northern zone gidance for refined risk assessment is under
development and agreed defualt refinement options (focal species, PD, and PT) will be accepted by Member
States, in order to faciliate a zonal risk assessment for birds and mammals.
4. References
[1] Guidance document on work-sharing in the Northern zone in the registration of plant protection products,
July 2011.
[2] Risk Assessment for Birds and Mammals, EFSA Journal 2009; 7(12):1438.
[3] Pesticide risk assessment for birds and mammals, Selection of relevant species and scenarios for higher
tier risk assessment in accordance with the EFSA draft guidance document under Directive 91/414,
Miljøministeriet, Miljøstyrelsen, 2010.
Identification of bird focal species in France for refined risk
assessments for Plant Protection Products
Véronique Poulsen1 and Camila Andrade2
1
ANSES – French Agency for Food, Environmental and Occupational Health and Safety, 253 avenue du G
Leclerc, 94701 Maisons Alfort, France
2
National Museum of Natural History, 55 rue Buffon 75005 Paris
al
1. Introduction
The refined risk assessment for birds and mammals might be based on different parameters such as
measured residues in food items, and/or on behaviour and diet of focal species. The EFSA Opinion (2009)
provides generic focal species for the major crops, but when further refinement is necessary, focal species
defined in the opinion as “real species that actually occur in the crop when the pesticide is being used”
should be identified.
As long as no agreed focal species are identified, when submitting a refined risk assessment to national
authorities, the notifiers can provide their own studies, and/or argumentations based on literature review.
Differences in the focal species choice between notifers make evaluations more difficult and non
homogeneous. In order to produce more harmonised refined risk assessment for birds, a working group has
been built in France, including the National Museum of Natural History and industry, in order to identify
agreed focal species identified by a robust methodology.
The National Museum of Natural History in France has got a huge data base containing registrations of bird
presence on the territory, based on field observations made by a large network of naturalists: the French
Breeding Bird Survey program (FBBS). A methodology has been worked out to identify the focal species on
several groups of crops by crossing the information issued from the FBBS database, the repartition of crops
in France, protected studies provided by industry, and a literature review.
The FBBS database contains around 15,000 observation points in France. An inventory has been made
once a year since 2001. For permanent crops such as vineyards, the localisation of these points can be
correlated to the crop repartition recorded in Corine Land Cover database.
For non permanent crops, a correlation between the FBBS database and a national mapping of crops based
on geographical pixels (named PAC), provided by the French ministry of agriculture, has been made. The
FBBS observation point is considered relevant for a crop when the percentage of presence of the target crop
in a pixel is higher than 60%.
Additional information was issued from dedicated protected studies, owned by industry, conducted in field
and based on transect observations.
A working group has been comparing and evaluating the FBBS and industry bird data in France and
developed criteria for the selection of focal bird species in France. The criteria generally follow the
recommendations by EFSA 2009 with the addition on a novel specific criterion derived from the use of the
FBBS data (IndVal, see below). The selection criteria, which are used in a stepwise manner, are:
-
Merging data on species and their frequency issued from industry and the FBBS,
-
Only species with a frequency higher than 10% in both data sets were considered,
-
Use of the following parameters as second trigger: diet, body weight, feeding behaviour, IndVal
value,
-
All diet guilds observed in the crop should be represented: the smallest and the most specialised
species (based on IndVal value) are selected.
The criteria of frequency of occurrence (FO), diet, foraging strata, and weight are described in EFSA 2009.
Specific for France is the IndVal calculated as follows. For each species that has been observed in relevant
plots, two parameters have been considered:
•
Frequency = Number of plots where the species has been observed / total number of plots
•
Abundance = Number of individuals of the considered species in the target crop / total number of
individuals of the species in all crops
The IndVal is Frequency x Abundance X 100
The study has been conducted for vines, orchards and a large number of crops such as cereals, vegetables,
maize, oilseed rape, sunflower, etc... For each crop or group of crops, a ranking of frequencies values of all
recorded species was made. Additional information on behaviour and feeding habits of each species, issued
from the literature, was taken into account to select the representative species (see Table.1). The objective
is to select focal species which should cover all diets and feeding behaviour. For example, two insectivorous
species may have been simultaneously considered as relevant focal species if one feeds on foliage insects
when the other one feeds on soil dwelling insects as exposure scenarios are different.
When based on the EFSA 2009 criteria there were two or more species, the one with the highest IndVal
value was considered as the relevant focal species for the corresponding diet guild.
Diet
Feeding behaviour
Frequency
Granivorous
Soil
0.39
Insectivorous
Foliage
0.17
Insectivorous
Soil or low foliage
0.51
Granivorous
Soil or lowest bough
0.31
Omnivorous
Soil
0.24
Table 1: List of the selected focal species
Additionally, for each focal species, an individual descriptive form is provided including detailed feeding
behaviour, a detailed diet when sufficient data are available, the presence periods in France for migratory
species, the reproduction period and any other available information of population level.
4. Conclusions
The focal species for the major crops or groups of crops, covering diets and feeding behaviour identified as
relevant, are selected. The outcome of this work is presented as a compiled document, and contains
individual descriptive forms for all focal species. The document also includes information on agronomic
practice and plant protection product applications regimes detailed by crop type.
This work will allow a more harmonised refined risk assessment for birds based on agreed and robust focal
species. We hope that this approach could offer a common, scientific and defensible document that can be
employed by stake-holders.
Additional information may still be provided for further refinement when needed.
Acknowledgement - The authors thank the experts involved in the working group and the DGAL of the
French Ministry of Agriculture for partial funding of this project. We warmly thank all volunteers who
participate to the Breeding Bird Survey in France.
Bird focal species in crops according the EFSA - How to find the right
candidates
Jan-Dieter Ludwigs1, Christian Wolf2, Christian Dietzen1 and Peter Edwards3
1
RIFCON GmbH, Zinkenbergweg 8, 69493 Hirschberg, Germany
tier3 solutions GmbH, Am Wallgraben 1, 42799 Leichlingen, Germany
3
Syngenta, Jealotts Hill International Research Centre, Bracknell, Berkshire, UK.
2
1. Introduction
According to EU Council Directive 91/414/EEC and Regulation 1107/2009, the effects of crop protection
products on wild mammals and birds have to be assessed. The concept of the focal species are based on
EPPO (1992) and developed further by SANCO (2002). In the most current Guidance of EFSA (‘Risk
Assessment for Birds and Mammals’; 2009) the option of defining focal species for the bird risk assessment
is presented in detail. Applying this approach gives information about what species should be considered,
i.e. what species to focus on in particular crops at the time the pesticides are applied. Within the past 10
years field research was conducted to investigate focal species in major European crops by some pesticide
manufacturers. This presentation will explain the method used in the field, and how the focal speciescandidates are determined from the data by examples from the comprehensive field program.
A range of fields were selected in one or several countries to represent typical fields of the crop in question.
The study fields were visited at pre-defined periods representing different BBCH growth stages for the
respective crop plants of interest (i.e. of which data was needed) and usually three surveys were conducted
within the study period representative of the majority of crop protection chemical usage. Exact locations of
each field surveyed were recorded together with field size, crop growth stage, date, field surroundings and
weather conditions.
All bird species were recorded in each study field by walking slowly along a defined longitudinal transect line
whose length was equal to the length of the respective fields. Individual birds registered either visually or
acoustically were assigned to the so-called ‘in-crop transect band’
Only birds present (foraging, roosting, calling, singing etc.) in the ‘in-crop transect band’ were included for
data analysis. Birds not directly associated with the study field were omitted from data analysis.
The total data base consists of 72 field studies of plant protection product producing companies who register
their products within Europe (BASF, BCS, Cheminova, DOW, GOWAN, Makhteshim, Irvita, Isagro,
Monsanto, Sharda and Syngenta). The studies have been conducted in eight different countries spread
across member states in the central and southern registration zones of Europe. The total period during which
the surveys of each study took place ranges from a minimum of two weeks to up to four months. Here we will
present data of two crops (cereals and pome fruit), and the values calculated on the basis of the raw data
are as follows:
Frequency of occurrence per field (FO Field ) – denotes the number of fields where a species was recorded
as percentage of the total number of fields regardless of the number of individuals observed. This approach
serves as a measure for the spatial frequency of occurrence, or the proportion of fields a species is present
on. A FO Field of 100% for one species indicates that this species was observed in all fields during at least one
survey.
FO field (% ) =
Number of fields in which species was recorded
× 100
Total number of fields
These frequency-of-occurrence-values are used as the main parameter to indicate what potential ‘focal
species’ may occur in the crop using an FO threshold of 20%. An FO < 20% is considered suitable to filter
out species that are unlikely to represent focal species. By definition of the risk assessment prerequisites
(EFSA 2009), selected ‘focal species’ should cover all other species which will be potentially present in the
field. Therefore it is necessary to consider issues such as diet guild, feeding strata, body weight and food
intake rate to ensure that species are representative to provide a robust risk assessment.
The selection of focal species obeys the following general procedure:
Select species with FO Field >20% and rank in order of the highest FO field
Species are then categorized into dietary guilds, i.e. insectivorous, granivorous, herbivorous or
omnivorous species
Species body weight is taken into account because this may reflect food ingestion rates.
If there are no clear candidates focal species are determined in a weight-of-evidence-approach
based on the information given under points 1) to 3) above and in some circumstances more
than one focal species may be identified.
1)
2)
3)
4)
Table 1: Overview on main results of focal bird species transect count studies in cereal and pome fruit
Crop Category
Crop
Cereals
Cereal
Orchards
Pome fruit
a)
Zone
No. of
Studies
C
S
C
S
2
5
4
6
No. of
Species
total
37
49
33
37
No. of Focal
Species
a)
Candidates
6
24
24
31
Focal bird species candidates based on frequency of occurrence per field (FO field ) ≥ 20%
Furthermore, to the FO field data, important criteria to determine the correct focal species candidates in terms
of environmental risk assessments for birds in Europe is diet guild of the species found and its body weight.
Focal species selection methodology is discussed using cereals and pome fruit orchards in Europe as case
studies.
4. Conclusions
Exposure assessments for individual species have to be protective of the bird community (i.e. also other
species) which may forage in the same crop and geographic region. The focal species defined in this way is
expected to achieve this. These same focal species may be used to further refine exposure through field
evaluation of the factors that may affect their exposure.
5. References
[1] EPPO (1992) Decision-making scheme for the environmental risk assessment of plant protection
products. Bulletin OEPP/EPPO Bulletin 24: 37-87
[2] SANCO (2002) Guidance Document on Risk Assessment for Birds and Mammals Under Council
Directive 91/414/EEC; Sanco 4145/2000, rev. 6, 25.09.2002
[3] EFSA 2009. European Food Safety Authority; Guidance Document on Risk Assessment for Birds &
Mammals on requestfrom EFSA. EFSA Journal 2009; 7(12):1438
Acknowledgement - The authors thank all the data owners within European Crop Protection Association for
access to these data and the field workers for collecting the data in the field.
European Distributions of Farmland Birds used as Focal
Species for Pesticide Risk Assessment
Peter Edwards1, Jan Dieter-Ludwigs2, Christian Wolf3
1
Syngenta, Jealotts Hill International Research Centre, Bracknell, Berkshire, UK
2
RIFCON GmbH, Zinkenbergweg 8, 69493 Hirschberg, Germany
3
tier3 solutions GmbH, Am Wallgraben 1, 42799 Leichlingen, Germany
1. Introduction
Risk Assessment guidance for birds and mammals in Europe has been described by EFSA [1]. The process
is tiered. To refine risk assessment it is neccessary to define focal species in different crops at relevent
growth stages appropriate for the use of the chemical applied. Focal species have been studied in 16 EFSA
crop categories, mostly by Industry, in several EU member states. Here we present the data from field
studies conducted by Industry. For most major crops the focal species have been determined in several
MS’s and this provides evidence of their distributions. Studies have not been measured in all crops in all
Member States (MS) since there is room for extrapolation between zones and crops. The EU have
established voluntary zonal work sharing for the registration and reregistration of plant protection products
which have been described by SANCO/6896/2009 [2] stating it is reasonable to place MS in geographical
zones (Zonal Approach) where there is greatest similarity in climate, crops and general conditions that may
influence exposure to pesticides and hence the risk. The purpose of the presentation is to demonstrate how
an evaluation of the distribution of focal species from these data can be done and to compare the distribution
of focal species from these studies with the distributions for the same species provided by the European
Birds Atlas database. To illustrate this approach, one crop example (cereals) will be used. The approach
developed may allow extrapolation in line with the Zonal Approach, such that MS risk managers can have
confidence in using refined exposure data for a single set of focal species applied to a crop within a zone or
even across zones.
Focal species have been determined based on EFSA [1] guidance and presented for cereals as an example
using transect methods and criteria based on the prevalence/frequency of occurence in fields. Focal species
studies for cereals comprised 7 studies from 5 Member States, Germany, Poland, France (2), Italy and Spain
(2). For the purpose of this demonstration of regional distributions only selected species with the highest
frequencies of occurence (FO) have been used.
The European Bird Census Committee (EBCC) have been responsible for publishing the European Birds
Altas distributions [3] from 40 European countries. More recently the EBCC have established a Pan
European Common Bird Monitoring Scheme where point counts are made along a fixed transect in the same
location each year and at the same time through the landscape every year in each participating country.
These data are used to evaluate annual trends in the population and have been combined regionally to
provide supranational indices. The rationale for these regional indices are that countries in the same region
share similar climates and socio-economic developments.
North, Central and Southern zones described by SANCO/6896/2009 [2] are listed as North (Denmark,
Estonia, Latvia, Lithuania, Finland and Sweden), Central (Belgium, Czech Republic, Germany, Ireland,
Luxenbourg, Hungary, Netherlands, Austria, Poland, Romania, Slovenia, Slovakia, United Kingdom) and
South (Bulgaria, Greece, Spain, France, Italy, Cyprus, Malta, Portugal). The Zonal composition and
supranational regions are similar for the north and south while the Central zone countries are separated into
West and East in the supranational regions.
The yellow wagtail is ubiquitous across Europe and may be considered a candidate for a focal species of
insectivore throughout all zones (north, central and south). The fantailed warbler has a very restricted
distribution and is only present in late growth stages in cereals and the yellow wagtail may be considered to
be protective of the fan tailed warbler dependent on higher tier exposure considerations like the diet.
The skylark is also ubiquitous throughout Europe, while the crested lark and corn bunting have more
restricted distributions. The skylark may be considered the candidate for focal species in the north and
central zones and the corn bunting in the southern zone. The corn bunting may be preferred to the crested
lark in the southern zone because the distribution is a slightly better fit and the frequency of occurence
markedly higher at the western side of this zone.
1
2
Species
Diet guild
Germany
Poland
France
north
France
south
Italy
Spain
Spain
Fantailed warbler
insectivore
0
0
0
0
25
52
0
Yellow wagtail
insectivore
20
67
86
7
25
71
0
Skylark
omnivore
96
96
81
73
95
14
0
Crested lark
omnivore
0
0
0
0
100
33
40
Corn bunting
omnivore
0
38
38
20
95
86
80
Table 1: Frequency of occurrence (%) for 2 species of insectivore and 3 species of omnivore measured in separate studies in 5
MS’s (Spain 1&2 – different studies, location, time).
Figure 2: European Atlas breeding distribution of yellow wagtail (left) and fan tailed warbler (right)
Figure 2: European Atlas breeding distribution of skylark (left), crested lark (centre) and corn bunting (right)
4. Conclusions
The frequency of occurence of these candidate focal species are consistent with the European Bird Atlas
distributions. While there are no data currently available for birds in cereals in the Northern zone it seems
reasonable to extrapolate focal species based upon their distribution (e.g. yellow wagtail and skylark). This
same approach may be applied to other crops and candidate focal species in those crops.
5. References
[1] EFSA 2009. Risk Assessment for Birds and Mammals. Journal 2009, 7 (12) 1438
[2] SANCO/6896 2009. Guidance document on the process for inter and inter-zonal work sharing to facilitate
the registration of plant protection products following inclusion of an active substance in Annex I of the
council directive 91/414/EEC. http://www.ec.europa.eu
[3] European Bird Census Committee http://www.ebcc.info/atlas.html
Acknowledgement - The authors thank all the data owners within European Crop Protection Association for
access to these data.
Small mammal communities in agricultural landscape in
central Europe: review of long-term field data
Blanckenhagen, F. v.1, Heroldová, M.2, Janova, E. 2, Bryja, J. 2, Konecny A. 2, Zejda, J. 2
and Städtler, T.1
1
RIFCON GmbH, Im Neuenheimer Feld 517, 69120 Heidelberg, Germany
2
Institute of Vertebrate Biology, AS CR, 60365 Brno, Czeck Republik
1. Introduction
Little is known about general small mammal communities in agricultural landscape in central Europe. Most
published data represent only a few months’ data in a specific habitat type focussing on a small region. This
presentation will review data from several studies performed in the last decades in agricultural landscape
across different regions in Germany and the Czech Republic. Data on the distribution of small mammal
species in agricultural landscape including cropped fields, meadows and adjacent field structures like
hedgerows and woodland are presented. The results of three data sets in comparison will focus on general
conclusions and may help to interpret the spatial and temporal composition and distribution of small mammal
communities in agricultural landscape in central Europe.
For the data evaluation three long-term data sets were available. The first data set is from southern Moravia
in the Czech Republic. Samples of small mammal communities were taken from different habitats in the
agricultural landscape by regular snap trapping in a three month interval from 1983 to 1989, which resulted
in 51,480 trap nights. The second data set is also from southern Moravia in the Czech Republic and contains
data about small mammal communities in 170 fields of different agricultural crops, sampled two times per
year at the beginning of the vegetative season of each crop and before harvest from 2004-2006 by snap
trapping, which resulted in 8,500 trap nights. The third data set is from Germany and consists of 111
independent trapping grids distributed over 7 federal states in the years 2001-2010. Trapping was mainly
conducted with multi-capture live traps. The trapping during all studies was performed in agricultural fields
and adjacent off-crop habitats. In the data set from Germany more than a quarter million trap checks were
conducted, >150,000 of these in agricultural fields, >60,000 in grassland and >40,000 in hedges and
woodland.
In order to get comparable values across all studies an easy available and calculated parameter, the
trapping efficiency (calculated as captures/100 trap checks) is used and the temporal scale of the data will
be presented according to meteorological seasons. For the presentation these data were assigned to the
following habitat types: (1) agricultural field, (2) woodland, (3) hedges and (4) perennial crop (including
meadows, pastures, the ground plant cover in orchards and vineyards, alfalfa and fallow land).
The trapping data revealed that 18 different species were captured in Germany and 14 in the Czech
Republic in agricultural landscape. In Germany 13 species belonged to the order Rodentia, four to the order
Insectivora and one to the order Carnivora. In the Czech Republic nine species belonged to the order
Rodentia and five to the order Insectivora.
Considering the mean trapping efficiency, the four most abundant species in Germany in downward order
were the bank vole (Myodes glareolus), the yellow-necked mouse (Apodemus flavicollis), the common vole
(Microtus arvalis) and the wood mouse (Apodemus sylvaticus). In the Czech Republic the five most
abundant species were the wood mouse, the pygmy field mouse (Apodemus uralensis), the common vole,
the yellow-necked mouse and the bank vole. In southern Moravia the pygmy field mouse, close related to the
wood mouse, was trapped regularly in agricultural landscape. However, in the south east of the Czech
Republic the pygmy field mouse reaches the western limits of the species' range.
The woodland habitat in Germany was inhabited by 14 small mammal species and in the Czech Republic by
ten small mammal species. The bank vole and yellow-necked mouse were the dominant species for this
habitat in both countries especially in summer. Besides these two species the wood mouse was trapped
regularly.
Hedges with diverse microhabitats were inhabited by nine small mammal species in Germany and seven
small mammal species in the Czech Republic. In both countries the bank vole, the yellow-necked mouse and
the wood mouse were dominating this habitat type. In summer the bank vole and the yellow-necked mouse
were dominating, whereas the wood mouse was the dominant species in spring for both countries and also
in autumn for the Czech Republic.
Perennial crops, in which little agricultural practice took place, like alfalfa, permanent grassland (e.g.
meadows) and fallow land, were dominated in both countries by the common vole in summer and autumn.
In agricultural fields with intensive agricultural practice (e.g. cereal fields and leafy crops) the trapping
efficiency was reduced, especially pronounced in Germany, compared to the other habitats in the agricultural
landscape. In Germany the common vole and the wood mouse showed the highest trapping efficiency. The
common vole was especially dominating in summer. In the Czech Republic in downward order the pygmy
field mouse, the wood mouse and the common vole were dominating the agricultural fields.
4. Conclusions
Considering the three long-term data sets from Germany and the Czech Republic, the small mammal
species in central Europe most exposed to the use of pesticides in agricultural fields and perennial crops are
the wood mouse and the common vole.
5. Research needs
1
Considering the other zones given by EFSA (2009) , the southern Zone and the northern Zone, no sound
data is yet published according the small mammal communities in agricultural landscape and the species
possibly exposure to pesticides.
6. References
[1] EFSA 2009. European Food Safety Authority; Guidance Document on Risk Assessment for Birds &
Mammals on requestfrom EFSA. EFSA Journal 2009; 7(12):1438.
Acknowledgement – Many thanks to all field biologists involved in the data collection.
Mercury emissions in large Hg-polluted floodplain areas in
Germany are an underestimated problem: challenges for
reliable risk assessments and implications for authorities
Jörg Rinklebe1
1
University of Wuppertal, Soil- & Groundwater-Management, Pauluskirchstr. 7, 42285 Wuppertal, Germany
SETAC 6th World Congress / SETAC Europe 22nd Annual Meeting, Berlin – Global Mercury Session
F11 - Global Mercury: Bridging science and policy
Oral presentation
Abstract
Environmental pollution by mercury is a world-wide problem. Particularly floodplain ecosystems are
frequently affected (e.g., Devai et al. 2005, Du Laing et al. 2009). One example is the Elbe River in Germany
and its catchment areas; large amounts of heavy metals including Hg from a range of anthropogenic and
geogenic sources have been accumulated in the soils of these floodplains (Overesch et al. 2007, During et
al. 2009, Rinklebe et al. 2009, 2010). They serve as sink for Hg originating from the surface water of
adjacent river. Today, the vastly elevated Hg contents of the floodplain soils at the Elbe River often exceed
even the action values of the German Soil Conservation Law (BBodSchG 1998, BBodSchV 1999). This is
especially important as Hg polluted areas at the Elbe River achieve several hundred square kilometres.
Thus, authorities are coerced by law to conduct an appropriate risk assessment and to implement practical
actions to eliminate or reduce environmental problems. A reliable risk assessment particularly with view to
organisms (vegetation as green fodder and hay production, grazing and wild animals) to avoid the transfer of
Hg into the human food chain, requires an authentic determination of Hg fluxes and their dynamics since
gaseous emissions from soil to atmosphere are an important pathway of Hg. However, reliable estimates of
Hg fluxes from the highly polluted floodplain soils at the Elbe River and its tributaries, and its influencing
factors are scarce. For this purpose, we have developed a new method to determine mercury emissions from
soils at various sites (Böhme et al. 2005, During et al. 2009, Rinklebe et al. 2009).
Our objectives were i) to quantify seasonal variations of total gaseous mercury (TGM) fluxes for floodplain
soils of the Elbe River and its tributary Saale in Germany, ii) to provide insights into physico-chemical
processes regulating these TGM fluxes, and iii) to quantify the impacts of the controlling factors soil
temperature and soil water content on Hg volatilization from a typical contaminated floodplain soil within soil
microcosm experiments under various controlled temperature and moisture conditions.
Soil temperature and soil water content related to flood dynamics are considered as important factors
affecting seasonal dynamics of TGM fluxes. We quantified seasonal variations of TGM fluxes and conducted
a laboratory microcosm experiment to assess the effect of temperature and moisture on TGM fluxes in
-2 -1
heavily polluted floodplain soils. Observed TGM emissions ranged from 10 to 850 ng m h and extremely
exceeded the emissions of non-polluted sites (Rinklebe et al., 2009; 2010). TGM emissions increased
2
2
exponentially with raised air and soil temperatures in both field (R : 0.49-0.70) and laboratory (R : 0.99)
experiments. Wet soil material showed higher TGM fluxes, whereas the role of soil water content was
affected by sampling time during the microcosm experiments. However, mercury concentrations and stocks
in top soil, C org content, pH, and soil texture, however, did not show significant correlations with TGM.
Rather, we have detected seasonal variations of TGM fluxes due to large fluctuations of soil temperature,
soil water content and flood dynamics. Hg emission rates increased exponentially with raised air and soil
temperatures in both field and laboratory experiments. The effect of soil water content was interfered by the
effect of soil temperature during field measurements. Wet soil material showed higher TGM fluxes compared
to dry soil, whereas the role of soil water content was also affected by sampling time during microcosm
experiments. Our study provides insight into TGM emissions from highly Hg-polluted floodplain soils in
Germany and that those emissions are an underestimated problem. Current needs for reliable risk
assessments, the induced implications for authorities, and future challenges will be discussed. Moreover,
this presentation will contribute to a better understanding of seasonal dynamics of Hg fluxes and its
controlling factors. The presented data as well as the discussed problems should be of large interest for a
wide international audience, such as environmental scientists and managers, applied ecologists,
environmental and technical engineers, and authorities.
Keywords: Hg soil-air exchange, TGM seasonality, risk assessment, wetland soils, Elbe River
References
BBodSchG - Bundes-Bodenschutzgesetz mit Erläuterungen. 1998. In: Bundes-Bodenschutzgesetz/ BundesBodenschutz- und Altlastenverordnung. Handkommentar. 2. Aufl., Bodenschutz und Altlasten. 47-273
(Holzwarth, F., Radtke, H., Hilger, B., Bachmann, G. 2000), Erich Schmidt Verlag.
BBodSchV - Bundes-Bodenschutz- und Altlastenverordnung mit Erläuterungen. 1999. In: BundesBodenschutzgesetz/ Bundes-Bodenschutz- und Altlastenverordnung. Handkommentar. 2. Aufl.,
Bodenschutz und Altlasten., 275-448 (Holzwarth, F., Radtke, H., Hilger, B., Bachmann, G. 2000), Erich
Schmidt Verlag.
Böhme F, Rinklebe J, Stärk HJ, Wennrich R, Mothes S, Neue HU. 2005. A simple field method to determine
mercury volatilisation from soils. ESPR – Environ Sci & Poll Res.12(3): 133-135.
Devai I, Patrick WH Jr, Neue HU, Delaune RD, Kongchum M, Rinklebe J. 2005. Methyl mercury and heavy
metal content in soils of Rivers Saale and Elbe (Germany). Analytical Letters 38(6): 1037-1048.
Du Laing G, Rinklebe J, Vandecasteele B, Meers E, Tack FMG. 2009. Trace metal behaviour in estuarine
and riverine floodplain soils and sediments: A review. Science of the Total Environment 407: 3972-3985.
During A, Rinklebe J, Böhme F, Wennrich R, Stärk HJ, Mothes S, Du Laing G, Schulz E, Neue HU. 2009.
Mercury Volatilization from Three Floodplain Soils at the Central Elbe River (Germany). Soil and Sediment
Contamination: An International Journal 18(4): 429-444.
Overesch M, Rinklebe J, Broll G, Neue HU. 2007. Heavy metals and arsenic in soils and corresponding
vegetation at Central Elbe river floodplains (Germany). Environmental Pollution 145: 800-812.
Rinklebe J, During A, Overesch M, Wennrich R, Stärk HJ, Mothes S, Neue HU. 2009. Optimization of a
simple field method to determine mercury volatilization from soils - Examples of 13 sites in floodplain
ecosystems at the Elbe River (Germany). Ecological Engineering 35: 319-328.
Rinklebe J, During A, Overesch M, Du Laing G, Wennrich R, Stärk HJ, Mothes S. 2010. Dynamics of
mercury fluxes and their controlling factors in large Hg-polluted floodplain areas. Environmental Pollution
158: 308-318.
Mercury in the Mediterranean: status and mass balance
Milena Horvat1, Nicola Pirrone2, Francesca Sprovieri2, Sergio Cinnirela2, Jože Kotnik1,
Nives Ogrinc1, Dušan Žagar3
1
Jozef Stefan Institute, Slovenia,
Institute of Atmospheric Pollution Research, Italy,
3
Univesity of Ljubljana, Slovenia
2
1. Introduction
An interesting feature of mercury biogeochemistry in the Mediterranean is that several fish species from the
Mediterranean show higher concentrations of Hg than same fish species in the Atlantic ocean, although the
concentrations of total mercury in the open waters of both oceans are similar. Elevated Hg levels have been
noted in environmental matrices from the Mediterranean regions adjacent to known mercury anomalies, yet,
the data do not clearly indicate that the effects of these anomalies have been transmitted to open waters or
to lower trophic level species living in these waters. Recent studies indicated that the main source of MeHg
in organisms in the coastal areas is related to methylation in sediments, while net mercury methylation in the
open ocean occurs in the water column and is linked to organic matter regeneration promoted by the
presence of small-sized nano- and picophytoplankton, that dominate under oligotrophic conditions. Relatively
large portion of mercury in waters is present as dissolved gaseous mercury (DGM), originating from
photochemical, biologically mediated mechanisms and/or diffusion from deeper layer either due to biological
and/or to tectonic activity which is typical of the Mediterranean region. Several attempts have been made in
order to evaluate Hg mass balance in the Mediterranean Sea. Cossa et al. [1] published the first
approximate balance for the Western Mediterranea and Rajar et al. [2] established the balance for THg,
where the trends were also evaluated. However, a mass balance for MeHg has not yet been calculated.
During an extensive research from 2002–2009, measurements of different mercury forms were carried out in
the Mediterranean Sea by the Italian research vessel Urania as a part of the Med Oceanor and MERCYMS
projects funded by the EU Framework programme and form the basis for the mass balance presented in this paper.
Several attempts have been made in order to evaluate Hg mass balance in the Mediterranean Sea. Cossa et
al. [1] published the first approximate balance for the Western Mediterranea and Rajar et al. [2] established
the balance for THg, where the trends were also evaluated. However, a mass balance for MeHg has not yet
been calculated.
In this work, the THg values [2] were only slightly corrected and divided among the surface layer,
thermocline and the deep sea, while MeHg vales were evaluated on the basis of numerous measurements
and publications form recent years. Therefore, only the sources and evaluation of the MeHg values are
described in detail. However, it must be mentioned, that the recent research on bromine Hg chemistry in the
air suggests a possibility of higher halogene-mediated deposition in marine boundary layer (MBL). The
possible elevated deposition from the MBL (and immediate re-evasion) could be a reason for the
discrepancy between the modelled [3] and the measured values [4] of mercury evasion from the
Mediterranean sea-surface.
The MeHg inventory in water was evaluated on the recent measurements by [5]. Lower concentration (0.1
pmol/l) were measured in euphotic zone and in lower layers the average concentrations were twice as high
(0.2 pmol/l).
Exchange with the atmosphere represents the most important source/sink of THg for the Mediterranean Sea,
while MeHg exchange is only of minor significance. Numerous authors suggest that MeHg deposition should
not exceed 1 % of the GEM deposition. Due to low precipitation and absence of gaseous MeHg species in
the surface layer [6,7] only 0.5 kMoles/yr of MeHg is deposited, while the evasion was found to be negligible.
River inputs were estimated from numerous measurements in the Isonzo/Soča River. MeHg concentrations
in the river outflows range from 5 – 50 pg/l were observed; taking into account the median and the discharge
of the Mediterranean rivers, the quantity was estimated to 0.5 kMoles/yr.
The surface Atlantic water inflowing through the Gibraltar contributes about 2.5 kMoles/yr MeHg into the
euphotic zone, while the Levantine Imtermediate Water (LIW) carries about twice as much MeHg out of the
thermocline.
The most important physical source/sink of MeHg in the Mediterranean Sea is exchange with the bottom
sediment, particularly the coastal sea. The accumulation and diffusive fluxes (separately for the coastal area
and the deep sea) were evaluated [8]. The calculated value of accumulation in the coastal area (500
kMoles/yr) is very likely somewhat too high, as the measurements and calculations of fluxes were mostly
performed in sites with higher particulate THg (Gulf of Trieste, Thau Lagoon) and high sedimentation ratio.
Such high values are difficult to extrapolate to the entire Mediterranean coastal zone (15 % of the area) and
are significantly higher than the quantity published [2]. Coastal sediment, on the other hand, represents a
significant source of MeHg (more than 50 kMoles/yr). In the deep sea, bottom sediment is source of both
THg and MeHg, contributing 15 kMoles/yr and 7.5 kMoles/yr, respectively. Taking into account the average
euphotic zone concentrations the THg and MeHg flux to the deep sea is approximately 90 and 6 kMoles/yr,
respectively.
Approximately 1.5 million tons fin-fish are caught annually, while there is no data on molluscs and decapods.
If we assume an average concentration of 0.20 mg/kg, the total output of Hg with fish is 1.2 kmol/year, which
is rather negligible amount [2].
Two important zones of MeHg productivity are
reported in the Mediterranean Sea: one at the
bottom of the euphotic layer and the other at the
oxygen minimum in the thermocline [5]. The
intensity of MeHg production and degradation was
evaluated [9] and shows some seasonal
variability, which is higher in the coastal zone. The
proposed methylation and demethylation rates
-1
-1
vary between 0.3-6.3 % day and 6.5 -25 % day ,
respectively. Based on these values the estimated
production in the euphotic zone is between 500
and 1000 kMoles/yr, and degradation was
estimated to 750-100 kmoles/yr.
Figure 1. Preliminary mas balance for total and MeHg in the
Mediterranean
4. Conclusions
Mass balance is presented in Figure 1. It has been shown that the total mercury exchanges at the straits are
not unbalanced, while mercury entering the western Mediterranean is mainly in inorganic Hg forms and is
exported to the Atlantic partially as methylated species.
5. References
[1] D. Cossa, Martin, J. M., Takayanagi, K., and Sanjuan, J., 1997. The distribution and cycling of mercury species in the western
Mediterranean, Deep-Sea Res. II, 44, 721–740.
[2] R. Rajar, R., M. Cetina, M. Horvat and D. Zagar 2007. Mass balance of mercury in the Mediterranean Sea. Mar. Chem. 107(1): 89102.
[3] Zagar, D., G. Petkovsek, et al. 2007. "Modelling of mercury transport and transformations in the water compartment of the
Mediterranean Sea." Mar. Chem. 107(1): 64-88.
[4] F. Sprovieri, Hedgecock, I. M., and Pirrone, N. An investigation of the origins of reactive gaseous mercury in the Mediterranean
marine boundary layer, Atmos. Chem. Phys., 10, 3985–3997.
[5] D. Cossa, D., Averty, B., Pirrone, N., 2009. The origin of methylmercury in opean Mediterranean waters. Limnol. Oceanogr. 54, 837844.
[6] J. Kotnik, Horvat, M., Tessier, E., Ogrinc, N., Monperrus, M., Amouroux, D., Fajon, V., Gibicar, D., Zizek, S., Horvat, N., Sprovieri,
F., and Pirrone, N. 2007. Mercury speciation in surface and deep waters of the Mediterranean Sea, Mar. Chem., 107, 13–30, 2007.
[7] M. Horvat, M., Kotnik, J., Fajon, V., Logar, M., Zvonaric, T., and Pirrone, N. Speciation of Mercury in Surface and Deep-Sea waters
in the Mediterranean Sea, Atmos. Environ., 37(S1), 93–108, 2003.
[8] N. Ogrinc N., Monperrus M., Kotnik J., Vidimova K., Horvat, M., Fajon, V., Amouroux, D., Kocman, D., Tessier, E., Žižek, S., 2007.
Distribution of mercury and methylmercury in deep-sea surficial sediments of the Mediterranean Sea. Mar. Chem. 107, 31-48.
[9] M. Monperrus, Tessier, E., Amouroux, D., Leynaert, A., Huonnic, P., Donard, O.F.X., 2007b. Mercury methylation, demethylation and
reduction rates in coastal and marine surface waters of the Mediterranean Sea. Mar. Chem. 107, 49–63.
Mercury exposure in relation to selenium and glutathione Stransferase gene deletion variants in pregnant women from
Mediterranean
Janja Snoj Tratnik1, Simona Jurković Mlakar6, Ana Miklavčič1, Darja Mazej1, Mladen
Krsnik2, Joško Osredkar2, Fabio Barbone3, Marika Mariuz3, Francesca Valent3, Katia
Sofianou4, Zdravko Spirić5, Janja Marc6, Milena Horvat1
1
2
“Jožef Stefan” Institute, Department of Environmental Sciences, Ljubljana, Slovenia
University Medical Centre Ljubljana, Institute of Clinical Chemistry and Biochemistry, Ljubljana, Slovenia
3
Unit of Hygiene and Epidemiology, University of Udine, Udine, Italy
4
Institute of Child Health, “Aghia Sophia” Children's Hospital, Athens, Greece
5
Institute for Applied Ecology, Oikon ltd., Zagreb, Croatia
6
University of Ljubljana, Faculty of Pharmacy, Department of Clinical Biochemistry, Slovenia
1. Introduction
It is well known that an antagonistic effect exists between selenium (Se) and mercury (Hg), and that Se can
play a protective role against Hg toxicity in organisms. Children are at particular risk for Se deficiency due to
the several factors as: food quality and intake, regional geological and environmental factors.
In order to find an evidence for such connection, total Hg and Se were determined in umbilical cord blood,
maternal blood and breast milk in pregnant women from Slovenia, Croatia, Italy and Greece. Methyl mercury
(MeHg) was determined in 30% of the samples. In addition, to observe gene-environment interactions,
GSTT1 (Glutathione S-transferase theta 1) and GSTM1 (Glutathione S-transferase Mu 1) gene deletion
th
variants were studied in a subset of women. This study was implemented within the EU 6 framework
programme Public health impact of long-term low-level mixed element exposure in susceptible population
strata (PHIME).
Women living in Slovenia, Croatia, Italy and Greece were recruited randomly in 2008 through their
gynaecologists in the last month of their pregnancy or in the maternity hospital just before the delivery. The
participating women filled out a food frequency questionnaire to obtain information on frequency of
consumption of different food items. Cord blood and breast milk samples were collected from 1654 and 1051
women, respectively. Venous blood was collected from 1081 mothers.
Total Hg in blood was determined by thermal combustion, amalgamation and atomic absorption
spectrometry using Direct Mercury Analyser (DMA-80, Milestone Srl., Italy), while total Hg in urine and breast
milk was determined by chemical digestion and atomic absorption spectrometry using cold-vapour atomic
absorption spectrometer (CVAAS). MeHg was determined in a subset of samples: 468 umbilical cord blood
samples, 364 maternal blood samples, and 268 breast milk samples. MeHg in blood and breast milk was
determined measured by acid dissolution, solvent extraction, aqueous phase ethylation, isothermal gas
chromatography (GC) and cold vapour atomic florescence detection (CVAFS). Selenium was determined by
inductively coupled plasma mass spectrometry (ICP-MS, Agilent, Japan). A complete description of the
methods is given by Miklavicic et al [1].
Genomic DNA was isolated from peripheral blood leukocytes using the High Pure PCR Template preparation
kit (Roche Diagnostics GmbH, Mannheim, Germany). Three pairs of oligonucleotide primers (GSTT1:
Forward (F): 5-ATGTGACCCTGCAGTTGC-3, Reverse (R): 5-AGATGTGAGGACCAGTAAGG-3; GSTM1:
F: 5-GCTTCACGTGTTATGGAGGTT-3, R: 5-CGGGAGATGAAGTCCTTCAGA-3; GPX1 as positive control:
F: 5-AGCCCAACTTCATGCTCTTC-3, R: 5-AGATGTGAGGACCAGTAAGG-3) were designed based on the
sequences of GSTT1, GSTM1 and GPX1 genes available in GenBank (accession no. NM_000637.2).
Fragments for the all three genes were amplified simultaneously by triplex-PCR method with the following
cycling conditions for all primer pairs together in the same reaction: an initial 10 min at 95ºC followed by 35
cycles of 1min at 95ºC, 1min at 57ºC, 1min at 72ºC, and finally 8 min at 72ºC. The PCR mixture (20μL)
contained genomic DNA (100 ng), 1x PCR buffer, 0,4 mM of each of the four deoxyribonucleotides, 1.5 mM
MgCl 2 , 0.20 μmol/L of each primer and 3 U of AmpliTaq Gold polymerase (Applied Biosystems, Roche
Molecular Systems Inc., Brandenburg, New Jersey, USA). Aliquots of PCR products were electrophoresed
on 2% agarose gel to check their quality and quantity. GPX1 was used as positive control in each reaction.
The absence of an amplification product combined with the presence of a positive control band indicated the
null (variant) type for both polymorphisms. The method is described in detail by Jurkovic Mlakar et al [2].
The samples with null polymorphisms were repeated in PCR reaction with only one related pair of primers of
each gene, to confirm its deletion polymorphism.
Kolmogorov-Smirnov normality test was conducted before association analysis and data transformation was
performed where appropriate. One-way ANOVA and ANCOVA, with adjustment for Se, in a logarithmic scale
where appropriate, and followed by post hoc test, was used to assess the statistical differences between Hg
and MeHg concentrations on individuals with different deletion genotypes of GSTM1 and GSTT1
polymorphisms. Kruskal-Wallis non-parametric test was used to evaluate the association of not normally
distributed parameters with different genotypes. P values less than 0.05 were considered statistically
significant. All statistical analyses were performed using the statistical software package SPSS 17.0 for
Windows (SPSS Inc., Chicago, IL, USA).
3.1. Mercury exposure and selenium
Total Hg and Se levels in cord blood of Slovenian women (median 1.52 ng/g and 76 ng/g, respectively) were
significantly lower compared to the levels in women from Croatia (median 2.94 ng/g and 96 ng/g,
respectively), Italy (median 3.94 ng/g and 113 ng/g, respectively) and Greece (median 5.81 ng/g and 104
ng/g, respectively). Total Hg levels in women from Greece were significantly higher than the levels in women
from other countries, while the Se levels were the highest in women from Italy.
Positive and significant linear correlation between total Hg and Se was found in cord blood (r=0.312),
maternal blood (r=0.212) and breast milk (r=0.644). Correlation between methyl Hg and Se in cord and
maternal blood was similar as for total Hg (r=0.229 and 0.269, respectively). Overall, the strongest
correlation was observed between inorganic Hg and Se (r=0.801). No correlation between MeHg and Se was
observed in milk (r=0.023, p=0.710). These results indicate strong association of Se with inorganic Hg.
Hg and Se were found to be associated positively and significantly also in blood of Slovenian women aged
50-59 (r=0.529) and children aged 6-11 from mercury mine area (r=0.251), but not in children from other
areas in Slovenia, confirming the association of Se with inorganic Hg, which is the predominant species
people are exposed to in the contaminated site.
Se in maternal and cord blood, but not in milk, was significantly correlated with the intake of many food items
in pregnancy. The strongest direct associations regarded cheese (r S =0.266), artichokes (r S =0.214) and
fennels (r S =0.249). In addition, both Hg and Se were significantly associated with fish consumption, possibly
explaining correlations between these two elements found in selected biomarkers.
3.2. GSTT1 and GSTM1 polymorphism
Interestingly, pregnant women with homozygous deletion of GSTT1 gene showed significantly higher MeHg
concentrations in cord blood compared to women with the presence of GSTT1 gene (p=0.028), whereas the
significance was not observed in the maternal blood. Moreover, no significant results were obtained for total
Hg and Se concentrations in the same tissues. However, strong associations were observed for MeHg
concentrations again in cord blood and additionally in maternal blood, when adjusted to Se levels. No
significant differences of MeHg, total Hg and Se concentrations between GSTM1 gene deletion variants
subgroups were obtained.
4. Conclusions
Association between Hg and Se in breast milk indicates that mainly inorganic Hg is conjugated with Se in the
selected population of pregnant women. Lower exposure to MeHg was observed to be related to glutathione
S-transferase activity.
5. References
[1] Miklavcic A, Cuderman P, Mazej D, Snoj Tratnik J, Krsnik B, Planinsek P, Osredkar P, Horvat M. 2011.
Biomarkers of low-levelmercury exposure through fish consumption in pregnant and lactating Slovenian
women. Environmental Research 111:1201-1207.
[2] Jurkovic Mlakar S, Osredkar J, Prezelj J,Marc J. 2011. Opposite effects of GSTM1 - and GSTT1 - gene
deletion variants on bone mineral density. Dis Markers: 31(5):279-87.
Mercury and methyl mercury in the trophic chain of the Lagoon
of Venice, Italy
Janusz Dominik 1,2, Davide Tagliapietra1, Andrea Garcia Bravo2, Marco Sigovini1, Jorge
Spangenberg3, David Amouroux4 and Roberto Zonta1
1
Istituto di Scienze Marine - Consiglio Nazionale delle Ricerche, Arsenale - Tesa 104, Castello 2737/F 30122
Venezia, Italy,
2
F.-A. Forel, Université de Genève, CP 416, 1290 Versoix, Switzerland,
3
Institute of Mineralogy and Geochemistry, University of Lausanne, CH 1015 Lausanne, Switzerland,
4
IPREM-LCABIE, CNRS UMR, 5254Hélioparc, 2 av P Angot, Pau, France
1. Introduction
Mercury contamination originating from both global and local sources is of concern because this metal
accumulates in biota. In many aquatic systems a part of inorganic mercury (Hg) is converted to
monomethylmercury (MMHg) which is biomagnifying along the food web. The mercury level in aquatic
organisms is not a simple function of mercury input, but depends of many environmental variables and
processes, such as methylation rate, trophy, food chain structure and many others, which complicate risk
assessment. The Lagoon of Venice is one of the numerous coastal zones contaminated by mercury emitted
from chemical industry (mainly alkali-chlor)[1]. Because of a long residence time of mercury, perpetually
recycled between water and sediments, the transfer of mercury to biota in lagoons persists a long time after
the reduction or elimination of Hg point sources. In such a dynamic system with a complex food chain it is
particularly difficult to predict the fate of mercury contamination and the associated risk. Previous studies
have demonstrated an elevated methylation potential [2], tide-driven MMHg transfer from sediments to
water column [3] and accumulation in some organisms [4]. Here we report for the first time the initial results
on Hg and MMHg bioaccumulation in the food web in the Lagoon of Venice.
2. Sampling and methods
Biota and sediment samples were collected between 22 June and 14 July 2011 in the northern part of the
2
Lagoon of Venice (Paluda di… GPS coordinates) within the area of xx m , except of fish obtained from a
local fisherman. Biota samples included pelagic and benthic organisms from all trophic levels from primary
producers to fish. All samples were freeze-dried and pooled by species. Carbon and nitrogen isotope were
determined at the Stable Isotopes Laboratory of the of Lausanne, Switzerland by flash combustion on a
Carlo Erba 1108 elemental analyzer (EA) connected to a Thermo Fisher Scientific Delta V (Bremen,
13
15
Germany) isotope ratio mass spectrometer (IRMS). Results were expressed as δ C and δ N values as the
per mil deviations of the isotope ratio relative to standards (C: Vienna Pee Dee Belemnite; N: atmospheric
nitrogen). Total Hg was analyzed using thermal combustion method (AMA 254, Leco) at the Institute F.-A.
Forel, University of Geneva. The Hg species extraction was achieved with a focused microwave method and
analyzed by species-specific isotope dilution and capillary gas chromatography (Focus GC, ThermoFinnigan)
hyphenated to inductively coupled plasma mass spectrometer (X7 II,ThermoElectron) at IPREM-LCABIE,
CNRS, Pau, France). The sum of inorganic Hg and MMHg measured with ICP-MS agreed well with the total
2
Hg concentration measured with AMA (R =0.95, slope=1.00)
3.1. Trophic levels
13
15
.δ C and δ N values in biota varied from– 24 to-11 ‰ and from 5.4 to 12.3‰, respectively. Excluding few
2
outliers, there was a significant relationship between the two variables (R =0.51) with the regression slope
15
close to 0.5. The lowest δ N values (5.4‰) were found in seegrass (Zostera marina ) and bivalve mollusc
(Mytilus galloprovincialis), followed by fine seston, mostly phytoplankton (6.0‰). The highest values were
15
found in some fish and shrimps, although the range of N enrichment in fish varied considerably. Based on
15
δ N results the sampled organisms covered 3 trophic levels.
3.2. Relation between trophic level and mercury content
Total mercury concentrations in biota varied by nearly three orders of magnitude from 0.030 µg g-1 d.w. (dry
weight) in seagrass (Zostera marina) to 2.3 µg g-1 d.w. in tissue of shrimp (Palaemon elegans). There was
a tendency of increasing Hg concentration with increasing trophic level, but the relation was not significant if
all species were considered. This was partly due to a relatively high Hg concentration in bivalve molluscs
15
and a low concentration in some benthic algae relative to their δ N signature. If Mytilus galloprovincialis is
taken as a base for determination of the trophic levels, and excluding two fish species (migratory?) the
2
relation between total Hg and the higher trophic levels improved (R =0.36, slope =0.87).
-1
MMHg concentrations varied between 5 (seegrass, phytoplankton) and about 2000 ng g d.w (shrimps, fish)
and the proportion of MMHg in total Hg was increasing with the trophic position of organisms. For the full set
15
of data (fig.1), the relation between MMHg and δ N values was best expressed by an exponential function
2
(R =0.59), as it was also found in the Gulf of St.Lawrance [5].
15
Fig 1. MMHg concentration vesus δ N in all organism groups and in sediments from the Lagoon of Venice.
Splitting the data set into benthic and pelagic food web may largely improve the interpretation of MMHg
accumulation relative to trophic level. For example, benthic filter feeder can accumulate as much Hg (and
also MMHg) as the organisms of the higher trophic levels.
4. Conclusions
On average, the accumulation of total Hg and MMHg in the organisms in a moderately polluted area of the
Lagoon of Venice increased by one order of magnitude for each of the three trophic levels. However, the
deviations from this rule can be considerable for benthic filter feeders (bivalve mollusks) and some fish
apparently migrating from less contaminated areas.
5. References
[1] Zonta, R., Botter, M., Cassin, D., Pini, R., Scattolin, M., Zaggia, L., 2007. Sediment chemical
contamination of a shallow water area close to the industrial zone of Porto Marghera (Venice Lagoon,
Italy). Marine Pollution Bulletin 55, 529-542
[2] Han, S., Obraztsova, A., Pretto, P., Choe, K.Y., Gieskes, J., Deheyn, D.D., Tebo, B.M., 2007.
Biogeochemical factors affecting mercury methylation in sediments of the Venice Lagoon, Italy.
Environmental Toxicology and Chemistry 26, 655-663
[3] Guédron, S., Huguet, L., Vignati D.A.L., Liu B., Gimbert, F., Ferrari B.J.D., Zonta R. and Dominik J.2012.
Tidal cycling of mercury and methylmercury between sediments and water column in the Venice Lagoon
(Italy), Marine Chemistry, in press.
[4] Bloom, N.S., Moretto, L.M., Scopece, P., Ugo, P. 2004. Seasonal Cycling of Mercury and Monomethyl
Mercury in the Venice Lagoon (Italy). Marine Chemistry 91, 85-99.
[5] Lavoie A.L., Hebert C.E., Rail J.-F.,Braune, B.M.,Yumvihoze E.,Hill L.G.and LeanD.R.S. 2010. Trophic
structure and mercury distribution in a Gulf of St. Laurence (Canada) food web using stable isotope
analysis. Science of Total Environment.
Dietary selenium at environmental concentrations reduces
methyl mercury retention in some aquatic organisms at the
lower trophic levels
Poul Bjerregaard, Narges A. Biuki, Alan Christensen, Tanja St.John
Institute of Biology, University of Southern Denmark, Campusvej 55, DK-5230 Odense, Denmark
1. Introduction
It is a well established fact that methyl mercury biomagnifies along aquatic food chains and that certain top
predators in such food chains - including humans with a high intake of fish or aquatic mammals – risk
neurological symptoms because of the high methyl mercury content in the food items. Generally, methyl
mercury is taken up into organisms quite efficiently, whereas - once assimilated – methyl mercury is
eliminated very slowly. Thereby, the long retention times (of course together with the efficient uptake) are the
main contributing factors in the biomagnification process.
It is also a well known fact that selenium interacts with methyl mercury and mercury in multiple, complex and
not fully understood ways. Methyl mercury contents in fish tend to be low in selenium rich (either natural,
manipulated or polluted) ecosystems and it has recently been shown [1] that dietary selenium increases the
elimination of methyl mercury from freshwater fish in laboratory experiments.
Lower retention times in organisms at the various trophic levels will inevitably result in reduced
biomagnification of methyl mercury along aquatic food chains but the role of dietary selenium on methyl
mercury biokenitics in aquatic invertebrates is poorly known.
The aim of the present experiments was therefore to obtain a better understanding of the role selenium in
the food plays for the retention of methyl mercury in aquatic invertebrates.
Principle of the experiments: Initially, the various types of organisms were exposed to one or several pulses
203
of radiocatively labelled methyl mercury (CH 3 - Hg-Cl) in their food. After a while (t 0 in the experiments), the
radioactivity of the animals was determined in a gamma counter (Wallac) and set to 100%.Thereafter feeding
the organisms with either control food or food amended with selenium was commenced. Subsequently, the
radioactivity of the animals was determined regularly (either repeated monitoring on the same live animals or
– for copepods – determination of the radioactivity of ten animals from each group).
Food for filter feeding organisms (copepods Arcartia tonsa and mussels Mytilus edulis) was marine algae
(the red alga Rhodomonas salina or the diatom Thalassiosira weissflogii) and food for the predatory or
omnivorous decapod crustaceans (the brown shrimp Crangon crangon and the shore crab Carcinus
maenas) was homogenised soft parts of blue mussels solidified by addition of gelatine.
The algae used as food for the filter feeding organisms were exposed to CH 3 copepods were exposed to labelled algal cells.
203
Hg-Cl whereafter he
Algal cells were exposed to selenium (as selenite) concetrations of 500 µg/l for and subsequently used as
food for the selenite exposed copepods. Selenium concentrations in the algae were determined..
203
Stock solutions of CH 3 - Hg-Cl and selenium were added directly to the mussel homogenate used as food
for the decapod crustaceans. To obtain a diet especially low in selenium, chicken breast muscle was
homoginised and used as basis for the food in one experiment. Selenium concentrations in all of the types of
food were determined by AAS-hydride generation (Perkin-Elmer) according to standard procedures.
st
Elimination of methyl mercury from the brown shrimps C. crangon could be described by 1 order kinetics
with half lives for methylmercury in brown shrimps fed the homogenised mussel soft parts without any
addition of selenium (0.34 µg Se/g) of 727±28 days - mean±SEM obtained in five different experiments (Fig.
1). The half life for methyl mercury in the brown shrimps fed chicken muscle with a low selenium content
(0.06 µg Se/g) was approximately twice as high (Fig. 1). Exposure to selenite, seleno-cystin and selenomethioneine (but not selenate) in the food decreased the half life methyl mercury in the brown shrimps (Fig.
2
1). For selenium concentrations ≤ 0.79 µg Se/g, half lives could be described: t ½ = 1226 – 1229*[Se] (r =
2
0.81; p = 0.002) and for [Se] ≥ 0.79: t ½ = 226 – 3.95*[Se] (r = 0.64; p = 0.01) (Fig. 1).
Figure 1: Selenium concentrations in food and half lives for methyl mercury in the brown shrimp
Crangon crangon. Half lives for methyl mercury in the experiments where shrimps were fed diets with
various selenium concentrations: Artificially low selenium concentration (□), background selenium
concentrations of mussels (●: mean±SEM for 5 experiments), food amended with selenite (○), seleno-cystin
(▼) or seleno-methionine (∆). The lines show linear regressions for selenium concentrations ≤ 0.79 µg Se/g
(red line) and [Se] ≥ 0.79 (blue line).
Retention of methyl mercury in copepods A. tonsa fed uncontaminated algae could be described by a two
0.0574*t
0.0125*t
+ 20*e
(with t in hours). Thus, 80% of the
compartment model: % methyl mercury retained = 80*e
methyl mercury was eliminated from a compartment with a half life of 12.1 hours and 20% from a
compartment with a half life of 55.6 hours. Retention of methyl mercury in copepods A. tonsa fed selenium
exposed algae could also be described by a two compartment model: % methyl mercury retained =
0.0867*t
0.0180*t
+ 28*e
(72% of the methyl mercury eliminated from a compartment with a half life of 8.0
72*e
hours and 28% from a compartment with a half life of 38.5 hours). The difference between the control and
selenium exposed group was statistically significant, although the effect was not nearly as pronounced as in
the brown shrimp.
Elimination of methyl mercury from the shore crab C. maenas was very slow and no effect of selenium in the
food could be identified
4. Conclusions
The finding that there is a negative correlation between low, environmentally realistic selenium
concentrations in the the food and the half life for methyl mercury in brown shrimps indicates that selenium
may play an active role for the biokinetics of methyl mercury in the environment. This is corroborated by
similar findings in zebrafish Danio rerio [1] where small increases in the selenium concentration in the food
also lead to a dose-dependent decrease in the retention of methyl mercury. The potential significance of the
somewhat more limited effect of selenium at the lowest trophic levels (the copepods) and lack of effect in the
shore crab needs more detailed elucidation.
5. References
[1] Bjerregaard P, Fjordside S, Hansen MG, Petrova MB. 2011. Dietary selenium reduces retention of methyl
mercury in freshwater fish. Environ. Sci. Technol. 45:9793-9798.
Acknowledgement – This investigation was supported by grants from the Danish Natural Science Research
Council
Mercury Pollution in China: Releases, Use and Impacts
Thorjørn Larssen1
1
Norwegian Institute for Water Research, Gaustadalleen 21, 0349 Oslo.
1. Introduction
China currently has the world’s largest intentional consumption as well as unintentional environmental
release of mercury (Hg). Atmospheric emissions has been estimated to about 700 tons annually, accounting
for one third of the global anthropogenic emission. There are also large (but not quantified) releases to local
soil and water environments. The intentional use og Hg in industrial processes and consumer products has
been estimated at 1000 tons annually, roughly half of the global total [1].
Mercury is released to the environment by a wide range of sectors, including key industries such as mining,
power generation, non-ferrous metal production, and the cement and chemical industries. The industrial use
of mercury in China has caused severe pollution incidents in the past. Today, as a result of past practices,
high mercury levels are found in water, soil and rice near abandoned mercury mining and smelting areas.
The presentation gives an overview of the major issues regarding China’s Hg pollution issues, including
releases, intentional use, environmental concentrations as well as human exposure.
2.1. Unintentional Mercury Releases
Coal Burning: Coal combustion accounts for approximately half of the man-made atmospheric Hg
emissions in China. China is the largest consumer of coal in the world, with coal accounting for almost 75%
of its energy production (coal-fired power plants and industrial boilers). The power plants sector is the largest
consumer of coal, but the industrial boiler sector is a larger source of atmospheric mercury emissions due to
less air pollution control in this sector. The total mercury emission from power plants was estimated to be
123 tonnes in 2007 and 214 tonnes for the industrial boilers [1]. Although the coal based energy production
in China will continue to increase, there is a potential for reduced Hg emissions : There are considerable cobenefits in terms of reduced mercury emissions from control measures for other pollutants (sulfur, nitrogen,
particulate matter), as mercury is removed from the flue gas by the air pollution control devices.
Non-Ferrous Metals Smelting: The non-ferrous metal smelting industry is a major sector for mercury
releases as Hg present in the ores is mobilized during the smelting process. Modern non-ferrous metals
smelters collect most of the mercury and either sell it or dispose it, but at more backward units large
quantities of Hg are released to the environment. (as wastewater, solid waste or the atmospheric emissions).
Non-ferrous metal smelting in China still involves a large number of small- and medium-sized plants. Many
use out-dated technology and hence have large mercury releases to the environment. There is a great
potential for reducing the releases of Hg to the environment from this sector, as technologies for collecting
Hg from the waste streams are readily available.
Cement Production: Close to half of the world’s cement is produced in China . Mercury is a trace element in
the raw feedstock materials, and in the fuels (mostly coal), making the cement industry an important mercury
pollution source. Cement demand will keep growing in the near future as development goals continue to be
pursued. To some extent modernization of the sector and air pollution control measure s targeting
atmospheric particulate matter emissions may reduce mercury emissions. The sector will still continue to be
an important sector for mercury releases.
2.2. Intentional use of Mercury
VCM/PVC Industry: Production of VCM for PVC in China is based on a process using coal as feedstock
(rather than oil or gas), which requires Hg as a catalyst. Currently the sector uses about 1000 tonnes of Hg
as catalyst annually and is by far the sector with the largest intentional Hg use. Currently there are no Hgfree alternative catalysts available (although development testing is ongoing). As a result, large amounts of
waste mercury catalysts, mercury-containing active coal, mercury-containing HCl, and mercury-containing
alkaline agents are generated. With the exception of the used catalyst, these are rarely recycled for technical
and economic reasons. The fate of Hg in these waste streams is uncertain. The future of Hg releases from
this sector largely depends on the development of Hg free alternatives. The sector is forecasted to continue
to grow rapidly and is hence a key issue in future management of Hg pollution.
Mercury-Added Products: The most important mercury-added products in china are medical devices
(thermometers and blood pressure measuring devices), fluorescent lamps and batteries. The total annual
use in the sector is uncertain, but may be around 300 tonnes annually. A major problem is the management
and disposal of wastes from the production processes as well as from end users’ disposal. Currently in China
most mercury-added products are sent to landfills along with municipal solid waste. China is serving a large
portion of the global market with these products. Therefore, legislation emerging in several countries
demanding lower Hg content will contribute to force the Chinese industry to develop alternatives with less or
no mercury. It is therefore likely that the Hg use in this sector, and in the longer term subsequent
environmental releases, may decrease in the future.
Mercury Mining Industry: China is one of only two countries in the world that still produces primary mercury
(i.e., from mining). Most of China’s mercury production is probably used domestically, though data on the
domestic trade, imports, and exports of mercury are not available. The mercury used in industry in China is
supplied by several sources: mining, imports, and recycling. A legacy of closed mines is a major problem for
China; there are dangers for local populations not just from mercury pollution but also from mine tailing pond
collapses
2.3. Mercury in the Environment and Human Exposure
The Hg levels in atmosphere and water bodies in large parts of China are highly elevated. Atmospheric Hg
concentrations show seasonal variation due to heating by coal in winter months. Remote areas in China also
showed elevated Hg levels in atmosphere and water bodies compared with other background areas in the
world. Large river estuaries, e.g. Pearl River Delta, Yangtze River Delta and Wuli Estuary, are heavily
impacted by upstream industrial sources. Those areas are also heavily populated and usually host large
scale industries [2].
Regarding fish, as a common source of human mercury exposure, data indicate that most fish contain
relatively low Hg concentrations. Fish species common in Chinese people’s diet usually are at have low
trophic levels, often farmed fish and fast growing species with low potential for bioaccumulation. There are
limited data on Hg levels in wild, older fish in China (as they are hard to find). The locations with high Hg
concentrations in fish are all strongly impacted by local Hg pollution [2].
Rice is also identified as a potential exposure pathway for population that are living near contaminated sites
and having rice growing locally as their staple food at the same time [3]. Studies showed that rice rather than
fish is the main source of MeHg exposure to Chinese in inland, although exposure levels in general are low.
Hg levels in most Chinese rice samples are well below the consumption advisory limit except a few heavily
polluted sites, e.g. mining areas.
Most Chinese people have low exposure of Hg, based on hair concentration data reported in the in literature.
However, there are data from certain areas with either a diet largely based on wild fish (Zhoushan fishermen)
or living in near vicinity of Hg mines and related contaminated sites (Wuchuan, Wanshan) showing
worryingly high exposure to Hg.
3. Conclusions
Mercury in China is a multi-faceted and complex issue, with a wide range of Hg releasing sectors, large scale
use for several purposes elevated levels in several environmental compartments. At the same time there
seems to be relatively low exposure and risk to general population at present, but a small groups of the
population at risk.
4. References
[1] The China Council for International Cooperation on Environment and Development, 2011. Special Policy
Study on Mercury Management in China.
[2] Lin, Y., Vogt, R.D., Larssen, T., 2011. Mercury in the Chinese environment - a review. In Review.
[3] Zhang, H., Feng, X., Larssen, T., Qiu, G. and Vogt, R.D., 2010. In Inland China, Rice, rather than Fish is
the major Pathway for Methylmercury Exposure. Environ. Health Perspectives. doi:10.1289/ehp.1001915.
Ecological Risk Assessment of Pesticides: Linking Non-Target
Arthropod Testing with Protection Goals (ESCORT 3)
Anne Alix1, Frank Bakker2, Katie Barrett3, Carsten A. Brühl4, Mike Coulson5, Simon Hoy6,
Jean Pierre Jansen7, Paul Jepson8, Gavin Lewis9, Paul Neumann10, Dirk Süßenbach11,
Peter van Vliet12
1
2
3
4
5
Dow AgroSciences, UK; Mitox, NL, HLS, UK; University Koblenz-Landau, Germany; Syngenta, UK;
6
7
8
9
CRD, UK; Agricultural Research Center, Gembloux, Belgium; Oregon State University, USA; JSC
10
11
International, UK; Bayer CropScience AG, Germany; Federal Environment Agency (UBA), Germany;
12
CTgB, NL.
Email contact: [email protected]
1. Introduction
The ESCORT 3 workshop (ESCORT: European Standard Characteristics Of beneficials Regulatory Testing)
dealt with questions of the protection of “Non-Target Arthropods” in the context of the use of plant protection
products in agriculture. It was the third ESCORT workshop that addressed this question.
The ESCORT 3 meeting was held as a review and update of the previous meeting outputs based on current
science. It also considered new issues and open points that had arisen in the interim period. The
proceedings of this workshop will be finalized in the 2012 and this presentation aims at offering the audience
an outline of these proceedings.
The workshop Organising Committee collated a number of questions arising from the peer review and
authorisation processes for plant protection products (PPPs), considering questions raised by risk assessors
from regulatory offices and industry performing the evaluation of PPP, as well as questions raised during the
[1]
public consultation (EFSA, 2009) on the existing Guidance Documents on Aquatic and Terrestrial
Ecotoxicology (SANCO/3268/2001 and SANCO/10329/2002). Based on these questions the Organising
Committee put together a programme of discussion topics that were addressed at the workshop in plenary
sessions alternating with work in sub-groups. This allowed for in-depth discussions on each of the four
areas identified by the Organising Committee: a) Level of protection and testing scheme; b) Off-crop
environment; c) Recovery; d) Field studies.
Approximately 60 participants registered for the workshop coming from authorities, the private sector, and
academia. The participants of the workshop were assigned to one of the four sub-groups based on their
knowledge and expertise, and regular plenary sessions gave participants the opportunity to comment on all
areas under discussion. An opening plenary session provided background information with presentations
from invited speakers.
3.1. Level of Protection and Testing Scheme
[2]
[3]
The Directive 91/414/EEC (15 July 1991) and the new Regulation (EC, No 1107/2009, 21 October 2009)
both require that products and their uses are assessed with regard to their potential effects on NTAs. In
evaluating potential approaches for structuring protection goals for these NTAs several aspects were
considered: a) spatial distribution of NTAs in the agricultural environment; b) time scale referred to for the
risk assessment; c) ecological function of the NTAs; d) life history strategies. On this basis the following
protection goals were developed differentiating between the in-field and the off-field area:
•
in the in-field area, the protection goal was identified as the maintenance of relevant functions;
•
in the off-field area, the protection goal was identified as the maintenance of NTA biodiversity.
The following functions were identified as relevant for the in-field area: a) pollination; b) control of pest
species; c) food source for wildlife and d) soil function (e.g., nutrient cycling by detritivores and
coprophagous species). With regard to the testing package and process that is needed to ensure that these
protection goals can be achieved, the information and recommendations contained in current guidance
documents were still considered appropriate.
3.2. Off-crop Environment
Off-crop in the context of arable uses of PPPs was defined at the workshop as: The area where the PPP is
not intentionally directly applied. In the context of higher tier testing, it was pointed out that phytophagous
NTA species may occur in greater overall abundance or diversity in off-crop habitats compared to in-crop
habitats. Currently it is not possible to rule out off-crop exposure, therefore the workshop considered, drift
reduction, to be a priority need, and as such, drift reduction should be promoted amongst growers, spray
machinery operators and policy makers. Mitigation measures in the context of the off-crop risk assessment
might include for example cropped or un-cropped buffer zones, or the use of low drift technology.
3.3. Recovery
When products exert some effects on NTA populations, experiments and risk assessments should evaluate
the recovery, or potential for recovery, of these populations. Recovery was defined at the workshop as: “The
return of populations, communities or functional groups to levels that would be reached without the specific
stressor.” The current risk assessment scheme for PPPs considers that effects on populations are
acceptable for the in-field area. In relation to this, as explained in ESCORT 2 it is accepted for the in-crop
area that the application of these products may result in effects above the threshold value of 50% if
“recovery” or at least the “potential for recovery” is demonstrated within one year.
The following endpoints were defined as appropriate for the following scales and areas:
•
At the landscape-level the recovery of populations is the relevant endpoint.
•
In off-crop areas, recovery of the communities (e.g., assemblage of arthropod species and their
abundance living in a grassy margin) is the relevant endpoint.
•
In in-crop areas, the recovery of ecosystem functions (e.g., pollination, pest control) assessed for
appropriate functional groups (e.g., pollinators and beneficial arthropods), is the relevant endpoint.
Considering field studies for off-crop risk assessment, no effect or only transient effects (e.g., de Jong et al.
[4]
2010) were considered acceptable, and therefore measuring long-term recovery is not applicable. In future,
recovery may be better predicted by modelling approaches and could be used for impact and risk
assessment for different agronomic practices and for extrapolation purposes.
3.4. Field Studies
Field studies can be performed if lower tier studies indicate a risk to NTAs from the use of a PPP. As effects
may have to be assessed at arthropod community level during periods that can be longer than one year,
intensive sampling on large scale plots has to be performed, especially for in-crop studies. Currently winter
wheat and apple orchards are used as model cops for arable crops and orchards, respectively. These
surrogate crops are still considered as appropriate, but for example the use of winter wheat for leafy arable
crops may need further investigation and data compilation. Results of field studies suggest that in Southern
Europe the number of species is higher than in Northern Europe, but that the abundance of the species is
lower than in Northern Europe. Current data suggest that there are no major differences in the response of
communities between North and South. Recovery trends may also vary between Northern and Southern
climatic regions.
Species Sensitivity Distributions (SSDs) which are currently used in other areas like the aquatic risk
assessment can be used for the NTA risk assessment but the actual use may need further evaluation.
Finally, the question of indirect effects from herbicides that might have no direct toxicity to NTAs but might
affect host plants of NTAs in the off-crop, was discussed. The workshop concluded that this question is
related to the protection of non-target plants and that these effects would be covered by an appropriate
protection of non-target plants.
4. Conclusions
The proceedings of the ESCORT 3 workshop aim at updating the recommendations and guidance for the
risk assessment for NTAs, based on the current knowledge on related science and on regulatory evolution.
The information and recommendations proposed in this document are to be used along with the
[5]
recommendation previously published in the ESCORT 1 (Barrett et al. 1994) and ESCORT 2 (Candolfi et
[6]
al. 2001) guidance documents.
The following four areas a) Level of protection and testing scheme; b) Off-crop environment; c) Recovery;
and d) Field studies were discussed in detail and the conclusions further developed the positions established
in ESCORT 1 and ESCORT 2
5. References
[1] EFSA. 2009. EFSA Journal 7 (12): 1438, 139 pp.
[2] EC. 2006. Directive 91/414/EEC, Council Directive of 15 July 1991 concerning the placing of plant
protection products on the market (91/414/EEC), Official Journal of the European Union. L 0414:
01.08.2006.
[3] EC. 2009. Regulation (EC) No 1107/2009 of the European Parliament and of the Council of 21 October
2009, concerning the placing of plant protection products on the market and repealing Council Directives
79/117/EEC and 91/414/EEC. Journal of the European Union, L 309/1, 24.11.2009.
[4] de Jong FMW, Bakker FM, Brown K, Jilesen CJTJ, Posthuma-Doodeman CJAM, Smit CE, van der Steen
JJM, van Eekelen GMA. 2010. Guidance for summarising and evaluating field studies with non-target
arthropods. RIVM report number 601712006/2010. ISBN/EAN: 978-90-6960-245-5.
[5] Barrett K, Grandy N, Harrison EG, Hassan S, Oomen P, editors. 1994. Guidance document on regulatory
testing procedures for pesticides with non-target arthropods, in ESCORT 1 workshop (European
Standard Characteristics of non-target Arthropod Regulatory Testing), Wageningen, The Netherlands.
SETAC Publication.
[6] Candolfi MP, Barrett KL, Campbell PJ, Forster R, Grandy N, Huet MC, Lewis G, Oomen PA, Schmuck R,
and Vogt H, editors. 2000. Guidance document on regulatory testing and risk assessment procedures for
plant protection products with non-target arthropods. ESCORT 2 workshop (European Standard
Characteristics of non-target Arthropod Regulatory Testing), Wageningen, The Netherlands.
Development of OECD Guidance on the Conduct and
Evaluation of Toxicity Tests for Endocrine Disrupting
Chemicals
Peter Matthiessen1, Gary Ankley2, Anne Gourmelon3, Laurence Musset3, and Jenny
Odum4
1
Consultant Ecotoxicologist, Old School House, Brow Edge, Backbarrow, Ulverston, Cumbria LA12 8QX, UK
2
United States Environmental Protection Agency, National Health and Environmental Effects Research
Laboratory, 6201 Congdon Boulevard, Duluth, Minnesota 55804, USA
3
Organisation for Economic Co-operation and Development, 2, rue André-Pascal, 75775 Paris CEDEX 16,
France
4
Regulatory Science Associates, PO Box 9346, Dunoon, Argyll PA23 7WR, UK
1. Introduction
It was recognised in the mid-1980s that certain chemicals and natural substances are able to interfere with
endocrine systems in wildlife and sometimes cause impacts on such vital processes as development and
reproduction. Research into these endocrine disrupting chemicals (EDCs) is still very active, but it was
realised over a decade ago that existing chemical hazard and risk assessment procedures were inadequate
for detecting these substances and fully characterising their likely effects in either wildlife or humans.
Before new regulations were made, the lack of harmonised procedures for detecting, and assessing the
effects of, EDCs was not a problem, but two jurisdictions (the United States – US - and the European Union EU) have now enacted legislation which requires that many chemicals must be assessed for their endocrine
disrupting properties, and it is likely that most other jurisdictions will follow suit. A pilot screening programme
is already operating in the US, but global screening and testing with mutual acceptance of data (MAD) was
not able to proceed until a suite of suitable internationally standardised methods became available.
In response to this concern, the Organisation for Economic Co-operation and Development (OECD) in 1996
set up its special activity on Endocrine Disrupter Testing and Assessment (EDTA) to develop and validate
new internationally standardised screens with some diagnostic capability for potential EDCs, as well as tests
with apical endpoints sensitive to EDCs. The EDTA programme was also tasked with the harmonisation of
approaches to hazard characterisation of this disparate group of chemicals. With the assistance of its
Validation Management Groups, OECD has now published a series of review papers and validation reports
[1], as well as standardised OECD Testing Guidelines (TG) for EDCs [2], and will shortly be completing a
Guidance Document (GD) on Standardised Test Guidelines for Evaluating Chemicals for Endocrine
Disruption [3].
The purpose of this paper is to briefly describe the EDC-sensitive TGs which have been, or are being,
developed by OECD, and to indicate the nature of the guidance on test interpretation in the GD.
New items for inclusion in the EDTA work programme must be proposed by member countries of OECD and
agreed by consensus of the National Coordinators of the Test Guidelines Programme. A lead country or
countries then usually prepare a Detailed Review Paper on the proposed testing method and its scientific
justification. If the review concludes that a new testing method is feasible and sufficiently underpinned by
existing research, the lead country will produce a draft TG which is then subjected to a rigorous validation
programme by the relevant Validation Management Group (VMG), for example, the VMG for Ecotoxicity
Tests (VMG-eco). Validation follows a carefully prescribed process [4] which usually includes a series of
international inter-laboratory trials to ensure that the method is fully optimised for its stated purpose, is
adequately reproducible, is statistically robust, and uses the minimum number of test organisms consistent
with its objectives. The final TG will only be published when these hurdles have been overcome and member
countries have unanimously agreed that validation has been successfully completed.
Member countries may also propose the development of Guidance Documents. In the case of the GD on
Standardised Test Guidelines for Evaluating Chemicals for Endocrine Disruption, this arose from the fact that
the various TGs for EDCs are relatively new procedures with which most chemical companies and regulatory
authorities are unfamiliar. As with draft TGs, a draft GD must be approved by consensus of OECD member
countries before final publication.
3.1. The EDTA Conceptual Framework
In order to assist assay development, EDTA initially published a 5-level conceptual framework (CF) of assay
types which it was expected would be populated with validated assays as the TG programme moved forward
[5]. This CF was recently updated (in draft), but some assays at all the levels have now either been
published or are in development. The levels range from data gathering (Level 1), through in vitro assays
(Level 2), to Levels 3-5 covering in vivo assays of increasing complexity, and at each of the in vivo levels
there are assays covering ecotoxicity as well as mammalian toxicity. It is important to note that the CF is not
a hazard testing scheme necessarily to be followed linearly from Levels 1 to 5, and testing at any level could
in theory take place at any time. At present, assays at all levels have diagnostic capability for, and/or
sensitivity to, one or more of estrogens, androgens, thyroid disrupters and steroidogenesis disrupters (socalled EATS modalities)
3.2. In vitro assays
In vitro assays will be an important component of any hazard assessment scheme for potential EDCs
because they will often provide an early alert for endocrine disrupting properties. At present, there are 2
published OECD TGs for in vitro assays: TG 456 for detecting substances with the potential to interfere with
steroidogenesis, and TG 455 for detecting interaction with the estrogen receptor. In vitro assays for other
endocrine modalities are being developed, and it is expected that, in due course, some of these assays will
also include a degree of metabolic competence.
3.3
In vivo assays
The OECD CF includes 15 in vivo mammalian assays with sensitivity towards EATS modalities, 4 in vivo fish
assays (TG 229, 230 and 234; GD 148), also with EATS sensitivity; and one in vivo amphibian assay for
thyroid-active substances (TG 231). In addition, OECD has developed, or is developing, a range of lifecycle
or partial lifecycle assays with birds, fish, amphibians and invertebrates which are able to measure the apical
effects of EDCs without generally being able to diagnose causality.
3.4 Guidance Document on Standardised Test Guidelines for Evaluating Chemicals for
Endocrine Disruption
This GD covers the published endocrine-related TGs, as well as those at an advanced stage of
development. It describes the conclusions which can be validly drawn from a particular assay result, set
against the weight of evidence of existing information. It will rarely be possible to conclude that a substance
is an EDC solely on the basis of a single assay, so the GD will provide much-needed assistance to chemical
companies and regulators alike. The GD also gives advice about a possible further stage of testing which
might be undertaken to increase the weight of available evidence.
4. Conclusions
OECD has developed a comprehensive framework which can be used for evaluating chemicals for
endocrine disrupting properties. It can already be used for in vitro and in vivo screening of chemicals with
EATS modalities, and assays for exploring the apical in vivo consequences of these modalities in vertebrates
are expected to be complete within the next few years.
5. References
[1] http://www.oecd.org/document/46/0,3746,en_2649_37465_47830318_1_1_1_37465,00.html.
[2]http://www.oecd.org/document/40/0,3746,en_2649_34377_37051368_1_1_1_1,00.html#Obtaining_Test_
Guidelines.
[3] http://www.oecd.org/document/12/0,3746,en_2649_37465_1898188_1_1_1_37465,00.html.
[4] OECD (2005). Guidance Document on the Validation and International Acceptance of New or Updated
Test Methods for Hazard Assessment. OECD Series on Testing and Assessment. Organisation for
Economic Co-operation and Development, Paris. 96 pp.
[5] http://www.oecd.org/dataoecd/54/55/48788319.pdf.
Acknowledgement - The authors thank the member countries of OECD for supporting the EDTA programme.
ECETOC Ecotoxicological assessment of endocrine disrupting
chemicals
1
2
3
4
5
A Weyers , L Weltje , JR Wheeler M Galay Burgos and M Gross
Bayer CropScience, Ecotoxicology 6620. D-40789 Monheim, Germany
2
BASF SE, Crop Protection – Ecotoxicology, D-67117 Limburgerhof, Germany,
3
Syngenta, Jealott’s Hill, Bracknell, Berkshire, RG42 6EY; UK
4
ECETOC, 4 avenue E. Van Nieuwenhuyse, B-1160 Brussels, Belgium
5
wca environment limited , Faringdon SN7 7YR, UK
1
1. Introduction
The new European Regulation on plant protection products (regulation (EC) No 1107/2009) and biocidal
products (revision to Directive 98/8/EC), as well as the regulation concerning chemicals (Regulation (EC) No.
1907/2006 ‘REACH’) only support the marketing and use of chemical products on the basis that they do not
induce endocrine disruption in humans or wildlife species. In the absence of agreed guidance on how to
identify and evaluate endocrine activity and disruption within these pieces of legislation an ECETOC task
force was formed to provide scientific criteria that may be used within the context of these three legislative
texts. This presentation focuses on wildlife species.
The synthesis of first ECETOC technical report [1] and associated workshop [2] was published by Bars et al.
(2011) [3]. Specific scientific criteria for the determination of endocrine disrupting properties that integrate
information from both regulatory (eco)toxicity studies and mechanistic/screening studies were proposed, but
it was recognised that the concept needed further refinement. Such guidance was developed by an
ECETOC task force and discussed at a workshop of invited regulatory, academic and industry scientists (WR
21 Risk Assessment of Endocrine Disrupting Chemicals 9-10 May 2011, Florence). Following input from the
workshop, ECETOC updated the guidance by taking on board some of the comments and recommendations
made.
For ecotoxicological assessments the key considerations include specificity and potency, but also extend to
considerations regarding population relevance and negligible exposure.
3.1. Specificity
The assessment of specificity is conducted at two levels. The first is to determine whether a substance
meets the agreed definition of an endocrine disrupter, i.e. causing an adverse effect, secondary
(consequent) to changes in endocrine function. A substance should only be considered of high concern
when the endocrine mediated effect occurs at concentrations lower than those that cause other significant
toxicity. The second level assesses if the most sensitive effect within one study or organism is accounted for
in a risk assessment by more sensitive non-endocrine endpoints observed in other taxonomic groups. Risk
assessment then allows for a margin of safety that sufficiently covers endocrine specific effects. If the
adverse effects are considered not specific at this stage, the substance proceeds to a standard risk
assessment based on the (lower) non-endocrine endpoint.
3.2. Potency
If the adverse effects are specific, then the potency of the substance must be considered. The substance
proceeds with a risk assessment based on the endocrine endpoint with an assessment factor based on
potency, unless exposure is negligible and no risk assessment is required. The environmental assessment
deviates at this stage from the assessment for human health, i.e. all substances proceed to a risk
assessment and there is no initial hazard-based screening or exclusion. This is because there are no readily
available threshold values for ecotoxicology in existing legislation (such as the CLP regulations) that are
suitable.
Therefore, the endocrine-mediated NOEC/NO(A)EL may be compared with other endpoints, e.g. by
assessing the magnitude of the Acute to Chronic Ratio (ACR), comparing the potency of the substance to a
reference compound (e.g. the natural ligand of interest, such as 17β estradiol for oestrogen receptor
mediated effects), consideration of the duration of exposure that is required for an adverse effect to be
induced, as well as the number of species in which the adverse effect is elicited.
3.3. Population relevance
In contrast to the human health assessment, the protection goal of environmental risk assessments is the
protection of populations rather than individuals [4]. This difference in protection goals between human and
environmental risk assessments is particularly important to the determination of endocrine disrupting
properties of chemicals. It requires specific consideration of adversity in relation to the population relevance
of endpoints measured in ecotoxicological studies.
Population relevant effects are those that affect population growth or dynamics. For example: age at first
reproduction, size of a reproductive event, frequency of reproductive events, duration of reproductive period,
viability of young and sex ratio. Clearly, some of these effects are population relevant and also diagnostic of
endocrine modulation (e.g. sex ratio in fish in the absence of sex-dependent mortality). However, some
effects are known to be responsive (and even sensitive) to, but not necessarily diagnostic of, endocrine
modulation (e.g. fecundity, which can be affected by general toxicity). In such circumstances supporting
information within a test or information from other in vivo testing tiers will be required to link the population
relevant effect to an endocrine mechanism.
3.4. Negligible exposure
A substance, which is considered to have endocrine disrupting properties, can be approved if exposure of
wildlife species to that substance under realistic proposed conditions of use is negligible. There are currently
no specified criteria for “negligible exposure of wildlife species” to plant protection products. Based on the
wording in the regulation, it is evident that negligible exposure must fall somewhere between “no exposure”
(i.e. nominal concentrations of 0, or less than the limit of detection/limit of quantification) and a concentration
representing an acceptable or low risk. Passing the standard risk assessment trigger is apparently not
sufficient.
Consideration of negligible exposure should focus on exposure of the organism group for which endocrine
disruption has been demonstrated in earlier stages of the assessment. For substances with potential
endocrine disrupting properties, the margin of safety should be considered. Alternatively, it could be possible
to use an additional assessment factor on endocrine endpoints to derive “negligible exposure” from
“acceptable exposure/low risk”.
4. Conclusions
In this paper ECETOC proposes refinements to the original ECETOC guidance [3]. The concept of potency
is proposed in order to bring the hazard-based cut-off criterion for endocrine disruptors in line with scientific
principles, so that potent endocrine disrupters will be regulated more stringently than the less potent. For
ecotoxicological assessments, further development is needed for the criteria proposed to allow them to be
fully operational in a regulatory context. ECETOC hopes that the criteria proposed in this paper will
contribute to the ongoing development of regulatory guidance under the relevant legislations.
5. References
[1] ECETOC, 2009a. Guidance on identifying endocrine disrupting effects. Technical Report No. 106.
Brussels. ISSN-0773-8072-106. Available at: http://bit.ly/ciOZCd
[2] ECETOC, 2009b. Workshop: Guidance on identifying endocrine disrupting effects, 29-30 June 2009,
Barcelona. Workshop Report No. 16. Brussels. Available at: http://bit.ly/agV83u
[3] Bars, R., Broeckaert, F., Fegert, I., Gross, M., Hallmark, N., Kedwards, T., Lewis, D., O’Hagan, S.,
Panter, H.G., Weltje, L., Weyers, A., Wheeler, J.R., Galay-Burgos, M., 2011. Science based guidance for the
assessment of endocrine disrupting properties of chemicals. Reg. Toxicol. Pharmacol. 59, 37-46.
[4] EFSA (European Food Safety Authority), 2010a. "Scientific Opinion on the development of specific
protection goal options for environmental risk assessment of pesticides, in particular in relation to the
revision of the Guidance Documents on Aquatic and Terrestrial Ecotoxicology (SANCO/3268/2001 and
SANCO/10329/2002)1." EFSA Journal 8(10): 1-55.
Acknowledgement - The authors thank the invited participants at the workshop held in Florence on May 910, 2011 for their contributions, which the task force has used to refine its original guidance.
Environmental Quality Criteria (EQC): A comparison of
methods under different regulatory regimes
Topic: F12 - Guidance documents for environmental risk assessment (ERA): needs, developments and progress
Keywords: REACH, Methods for environmental risk assessment (ERA), Environmental Quality Criteria
(EQC), Environmental Quality Standards (EQS)
Presentation preference: oral presentation
1
2
3
3
Christiane Heiss , Silke Kleihauer , Udo Hommen , Kerstin Hund-Rinke , Martin Führ
1
2
2
Federal Environment Agency (UBA), Dessau/Germany; Society for Institutional Analysis (sofia), Darmstadt;
3
Fraunhofer-IME, Schmallenberg
1. Introduction
Environmental quality criteria (EQC) form the basis for legally binding environmental quality standards
(EQS). Compulsory EQS in media legislation are for example:
−
The environmental quality standards under the Water Framework Directive 2000/60/EC (WFD), Art. 2
(24 and 35), Art. 16(7) and Annex IX WFD);
−
The European air quality provisions refer – in the context of the UNECE Convention on Long-range
Transboundary Air Pollution – to a set of different EQS, such as limit values, critical values or target
values, Art. 2 (3, 5, 6 and 9) and Art. 12 et subs. CAFE-Directive 2008/50/EC);
−
German national precautionary values for soil according to the German regulation for the protection of
soil (BBodSchV, 1999); since an EU approach to soil protection still has to be developed.
The EQC’s binding force and their scope of application depend on the legal framework they are associated
1
with. Aside the mandatory EQS for industrial installations[ ] and media legislation, the EU chemicals
legislation refer to EQC (Predicted No-Effect Concentration, PNEC) to indicate that “adequate control” of
substance related risks can be assumed.
The EQCs are derived according to different European and national legal frameworks. Most of them are
accompanied by (technical) guidance documents (TGD/GD). Within European chemical and water regimes
the (T)GDs are currently streamlined at European level. This harmonisation of methods is likely to influence
both, the other sectors of environmental policy on the European level as well as the existing national risk
assessment schemes.
The Federal Environment Agency induced a project comparing the methods and procedures under the
European REACH Regulation with the approaches to national established guidelines for water, soil and the
UNECE concept for critical loads and critical levels to protect terrestrial ecosystems against air pollutants.
2
The comparison included the legal frameworks and procedural mechanisms for quality assurance.[ ]
The methodical analysis focuses on hazardous substances (according to the CLP Regulation EC/1272/2008)
3
and their effects to the environment. EQC for metals are considered as special cases[ ] and are discussed
separately for water and soil. PBT- /vPvB- and substances causing equivalent concern (REACH Art 57, d-f)
are not included in the assessment, because according to REACH a quantified risk characterization is not
possible for PBT- and PBT-like substances.
The methodological approach contained literature work, case studies and interviews with officials to
document the experience and expertise of authorities and to evaluate the transparency of EQC derivation.
The REACH Guidance Documents (GD) served as reference method to discriminate concordances and
differences. Two main .procedural “stations” have to be distinguished (figure 1):
1. Retrieval and Evaluation of all available information [with several “steps” under REACH] to define the
“data set for the assessment”.
2. Derivation/Identification of the EQS [REACH: PNEC].
The basic procedural elements concur in all analyzed
sectors, but some variations have to be considered.
Water: The highest level of concordance has been
identified between REACH and the European water
legislation. No fundamental discrepancies can be found;
neither in the steps to define the “data set for the
assessment” nor in the derivation of the EQC by applying
“assessment factors” laid down in the guidance
documents. However, regional aspects are taken into
account in the water field, while REACH is aiming at
uniform European EQC.
Soil: As far as soil protection is concerned the German
approach proved to be highly compatible in the first
station; differences occur in the definition of the test media
(standardized soil). The derivation of the EQS differs to
some extent since the German approach takes
characteristics of the medium soil (background values)
into account.
Air: For the atmospheric compartment the REACH
guidance documents up to now do not contain any
clarification as to how a PNEC-air shall be derived. In the field of air quality legislation a case by case
approach is dominant which is based to a high degree on monitoring data and expert judgement. A
universally valid methodological guidance has not been established yet.
4. Conclusions
The conclusion of the study is that all data assembled in “Station 1” are transferable to other legislative
frameworks. But also the derivation of EQC follows similar (soil) if not identical (water) methodological
considerations. From this perspective no fundamental reservations can be made against the mutual
acceptance of the derived EQC.
However the remaining differences have to be taken into account; e.g. validity of data and tolerable grade of
uncertainties may be assessed differently by companies which register a substance under the REACH
regime and water authorities and thus different assessment factors may be used.
The test results for soil may differ due to the fact that REACH refers to “standardized soil” while national
German approach refers to regional situations and takes into account different types of soil. In this context
the question how to assess the bioavailability and the way to consider the background contamination has not
been solved satisfactorily. For metals a specific guidance document under reach is addressing these issues
to some extent.
A case by case assessment, as applied in the derivation of air quality standard is compatible with the
REACH approach, given that this is based on broad range of valid (field) data.
In an overall perspective it can be stated that the international debate is leading to a higher level of
concordance. A comparative approach – based on a step-by-step assessment of the various procedural
elements – is able to underpin this in a specific way and thus allowing to identify the remaining differences.
In cases where those differences are of minor importance the EQS may be used to fill the gaps in other
sectoral legislation, such as provisions for industrial installations (IED 2010/75/EU), water or air quality.
5. References
[1] In the context of industrial installations 'environmental quality standard` is defined as “the set of requirements
which must be fulfilled at a given time by a given environment or particular part thereof, as set out in Community
legislation” (Art. 2(7) IPPC 1996/61/EC = Art. 3(6) IED 2010/75/EU).
[2] Kleihauer/Führ/Hommen/Hund-Rinke 2011: Bestimmung von stoffbezogenen Umweltqualitätskriterien – ein
Methodenvergleich von nationalen und internationalen Bewertungsgrundlagen. Im Auftrag des
Umweltbundesamtes (FKZ 363 01 260).
[3] ECHA-Guidance on information requirements and chemical safety assessment. Appendix R.7.13-2 (July 2008):
Environmental risk assessment for metals and metal compounds, http://guidance.echa.europa.eu/docs/guidance_document/information_requirements_r7_13_2_en.pdf ( last visited: 25.11.2011).
Representativeness of Eisenia fetida for the environmental risk
assessment of pesticides to soil organisms
Michiel Daam1, Sara Leitão1, Maria José Cerejeira1 and J. Paulo Sousa2
1
2
ISA - Technical University of Lisbon, Tapada da Ajuda, 1349-017 Lisbon, Portugal
IMAR-CMA, Dep. of Life Sciences, University of Coimbra, P3004-517 Coimbra, Portugal
1. Introduction
Current pesticide risk assessments for soil invertebrates in the EU are largely based on routine testing of
earthworms, though it has been questioned whether earthworms accurately represent the sensitivity of the
wide range of soil invertebrates in terrestrial ecosystems [1-3]. For example, after reviewing laboratory
studies into the effects of pesticides on soil invertebrates, Frampton et al. [2] concluded that the standard
test earthworm Eisenia fetida sensu lato (E. fetida and E. andrei) was the least sensitive species to
insecticides based on acute mortality (i.e., LC50 values). Soil arthropods (e.g., the standard collembolan test
species Folsomia candida) appeared to be more sensitive to compounds with a broad range of (especially
insecticidal) toxic modes of action, indicating that soil arthropods should also be tested routinely in regulatory
risk assessments [2].
Due to limitations in data availability, Frampton et al. [2] could only construct SSDs for 11 (two herbicides,
two fungicides and seven insecticides) out of the total of 250 pesticides for which toxicity data was available.
Furthermore, only acute mortality data (i.e., LC50) sufficed to construct SSDs and these could also not be
constructed for individual taxonomic groups (e.g., Collembola, Lumbricidae and Nematoda) separately.
The aim of the present study was to evaluate the sensitivity of E. fetida relative to other soil invertebrates for
a greater number of compounds and endpoints using (an adapted version of) the relative tolerance (Trel)
approach as used by Wogram and Liess [4] to compare sensitivity of aquatic macroinvertebrates with that of
Daphnia magna. This enabled comparing toxicity thresholds for main terrestrial taxonomic groups and
pesticidal types of action (insecticides, fungicides, herbicides, and other) separately.
Toxicity data were obtained from the US Environmental Protection Agency (US-EPA) ECOTOX database
(http://cfpub.epa.gov/ecotox/). To enable a comparison of threshold values from different compounds, the
threshold concentrations had to be ‘‘normalised’’. This was done by transforming these concentrations to
relative tolerance (Trel) values by dividing them by the (geometric mean of) threshold value(s) of E. fetida
sensu lato. A Trel of one thus indicates a relative tolerance equal to that of E. fetida. For species more
sensitive than E. fetida, Trel is less than one and for less sensitive species it is greater than one.
To evaluate whether uncertainty factors currently applied in the EU to the threshold values of E. Fetida
sufficed to protect all other taxa included in the analyses, ‘‘Trel PNECs’’ were calculated by dividing toxicity
values of non-standard test species for the different compounds by their corresponding PNEC values. In
accordance with the ERA procedure in the EU, these PNECs were calculated by dividing the acute and
chronic toxicity data for E. fetida with 10 and 5, respectively. In addition, Trel PNECs were calculated by
considering the sensitivity of both E. fetida and F. candida, i.e. by using the lowest toxicity value of these
organisms.
3.1. Sensitivity of E. fetida sensu lato compared to other soil invertebrates
In Fig. 1, the sensitivity of soil invertebrates by taxonomic group are compared with that of E. fetida sensu
lato. The greater and lower sensitivities of collembolans to insecticides and fungicides, respectively, as also
noted by Frampton et al. [2], were confirmed. The SSDs also revealed that isopods were more sensitive to
insecticides, and nematodes to fungicides, as compared to E. fetida (Fig. 3; Table 2). Since SSDs could only
be constructed for a limited number of taxonomic-compound group combinations, 95% confidence intervals
(CI) of Trel values from these combinations were calculated. These additional analysis also indicated
significant (i.e., the value 1 is not covered by the 95% CI) greater sensitivity of Acari to insecticides, and
nematodes to fungicides. Despite this overall greater sensitivity of arthropods and worm-like taxa to
respectively insecticides and fungicides, data availability is greater for the combinations arthopods-fungicides
and worm-like taxa-insecticides (Fig. 1), implying an overall poor selection of test compound (or test species)
in the soil toxicity assays included in the database.
Potentially Affected Fraction (PAF)
Acari
Collembola
Enchytraeidae
Lumbricidae
Isopoda
Insecticides
Coleoptera
Nematoda
Fungicides
1.0
1.0
0.8
0.8
0.6
0.6
0.4
0.4
0.2
0.2
0.0
0.00001 0.0001 0.001
0.01
0.1
1
10
100
1000
0.0
0.001
Herbicides
1.0
0.8
0.8
0.6
0.6
0.4
0.4
0.2
0.2
0.0
0.00001 0.0001 0.001
0.0
0.01
0.1
1
0.1
1
10
100
1000
Only
E. fetida
10000
10
100
1000
Fungicides
Insecticides
Herbicides
Other compounds
1.0
0.01
0.01
Fungicides
E. fetida &
F candida
0.1
1
10
Insecticides
Herbicides
100
Relative tolerance (Trel)
Figure 1. Species sensitivity distributions (SSD)
comparing the sensitivity of different taxonomic groups to
insecticides, fungicides, herbicides and other compounds
with that of E. fetida. The vertical dashed line at Trel = 1
indicates the sensitivity of E. fetida sensu lato. A Trel < 1
and a Trel > 1 indicate a greater and lower sensitivity
relative to E. fetida sensu lato, respectively.
0.000001
0.001
1
1000
1000000
Trel PNEC
Figure 2. Protectiveness of predicted no effect
concentrations (PNEC) for E. fetida sensu lato alone, and
in combination with Folsomia candida, for other test
organisms included in the database. Taxonomic groups
were grouped in arthropods (black dots) and annelids and
nematodes (open diamonds). A Trel PNEC < 1 indicates
that the corresponding PNEC value(s) for the standard
test species considered is/are not sufficiently protective,
whereas a Trel PNEC > 1 indicates that the PNEC value(s)
for the standard test species considered is/are sufficiently
protective.
3.2. Implications for the terrestrial risk assessment of toxic compounds
Risk assessments (PNEC calculations) based on only E. fetida do not fully protect a great number of other
test organisms, whereas this is not the case when also including F. candida (Fig. 2). Hence, these analyses
confirm the recommendation made by Frampton et al. [2] to include F. candida in (first-tier) regulatory risk
assessments.
Especially for arthropods few Trel PNECs could be calculated, and was limited to a maximum of three
values: Acari (1), Coleoptera (2), Collembola other than F. candida (3), and Isopoda (3). Furthermore,
various Trel PNEC values lay close to 1, especially for fungicides (Fig. 2), for which three Trel PNEC values
between 1 and 2 were obtained for three different nematode taxa. Furthermore, a Trel PNEC of 0.96 was
calculated for the enchytraeid E. crypticus exposed to manganese sulphate.
Also considering that several Trel < 1 were obtained
be questionable whether sole testing of both E. fetida
the range of other potentially sensitive taxa. For the
organisms has previously been recommended (e.g.,
pesticides.
for Acari, Isopoda and Nematoda (Fig. 1), it may thus
and F. candida for the first-tier risk assessment covers
same reason, a battery of tests using a range of test
[4]), as is the case in the aquatic risk assessment of
4. References
[1] Daam MA, Leitão S, Cerejeira MJ, Sousa JP. 2011. Comparing the sensitivity of soil invertebrates to
pesticides with that of Eisenia fetida. Chemosphere 80:1040-1047.
[2] Frampton GK, Jänsch S, Scott-Fordsmand JJ, Römbke J, Van den Brink PJ. 2006. Effects of pesticides
on soil invertebrates in laboratory studies: a review and analysis using species sensitivity distributions.
Environ. Toxicol. Chem. 25:2480–2489.
[3] Römbke J, Jänsch S, Didden W. 2005. The use of earthworms in ecological soil classification and
assessment concepts. Ecotoxicol. Environ. Saf. 62:249–265.
[4] Wogram J, Liess M. 2001. Rank ordering of macroinvertebrate species sensitivity to toxic compounds by
comparison with that of Daphnia magna. Bull. Environ. Contam. Toxicol. 67:360–367.
Acknowledgement - This study was funded by the Portuguese government through a post doctoral research
position for the first author (program:Ciência 2007; reference: C2007-ISA/CEER/ECOTOX), a PhD
scholarship for the second author (reference: SFRH/BD/42306/2007) and research project
HERBITOXBIOAS (reference PTDC/AMB/64230/2006).
Information requirements under the Biocidal Products
Regulation and their implications for environmental risk
assessment, authorities and applicants
Jan Weber1, Tobias Porsbring1, Erik Van De Plassche2
1
European Commission - DG Joint Research Centre, Institute for Health and Consumer Protection; Via E.
Fermi 2749, I-21027 Ispra, Italy
2
European Chemicals Agency; Annankatu 18, FI-00121 Helsinki, Finland
1. Introduction
As of 2013, the current regulatory framework, the Biocidal Product Directive (BPD), is going to be replaced
by the Biocidal Products Regulation (BPR). The upcoming regulation will encompass a wide range of
regulatory changes. Union-wide authorisation of products, the obligation to substitute active substances with
less-hazardous equivalent alternatives, as well as an exposure-based data waiving represent major changes
that will affect both authorities and applicants. A further major change concerns the amount of quantitative
and qualitative information to be submitted by applicants for both active substances approval and biocidal
products authorisation.
Materials and methods
Current (BPD) and future (BPR) information requirements are compared. For the current data requirements,
the authors distinguish between the strictly legally binding ones, as listed in the BPD text, and the actual
applied requirements, as listed and explained in the Technical Notes for Guidance (TNsG).
The related study costs are estimated, based on study prices from a number of providers within the
European Union. The changes in information requirements are discussed in regards to the environmental
risk assessment, practical implications for the evaluating Member States competent authorities, and the
practical/economic implications for applicants.
The Biocidal Products Regulation requires more compulsory information and exhibits a more extensive list of
case-dependent additional information requirements when compared to the BPD. This could represent a
major challenge for the applicant, but has to be weighed against the aim to improve hazard and risk
assessments in a resource-efficient manner. On the other hand, for a number of endpoints only the wording
has changed and the actual study requirements are comparable to the current ones. The fact that no test
methods have been derived yet for endocrine disrupting properties as well as for nano materials, that might
be possible co-constituents in future biocidal products, results in uncertainties of this study.
3. References
[1] Directive 98/8/EC of the European Parliament and of the Council of 16 february 1998
concerning
placing of biocidal products on the market.
[2] Technical Guidance Document in support of the Directive 98/8/EC concerning the placing of biocidal
; and
DG biocidalprod
products on the market; guidance on data requirements for active substances
JRC, February 2008.
[2] Manual of Technical Agreements of the Biocides Technical Meeting (MOTA); DG JRC, September 2011.
[3] Proposal for a Regulation of the European Parliament and of the Council concerning the making available
on the market and use of biocidal products – Outcome of the European Parliament's second
reading
, (S
European
trasbourg,
Commission,
16 to 19 January
January
2012)
2011.
[4] The sources of the study costs price lists are confidential.
Proposal for a harmonized assessment of the mixture
ecotoxicity of biocidal products
Anja Kehrer1, Rolf Altenburger2, Thomas Backhaus3, Michael Faust4, Daniel Frein1,
Caroline Riedhammer1 and Beatrice Schwarz-Schulz1
1
Federal Environment Agency, Woerlitzer Platz 1, 06844 Dessau-Rosslau, Germany
UFZ Helmholtz Centre for Environmental Research, Dept Bioanalytische Ökotoxikologie
Permoserstr. 15, 04318 Leipzig
3
Dept of Plant and Env. Sciences, University of Gothenburg,
Carls Skottsbergs Gata 22B, Box 461, 40530 Göteborg, Sweden
4
Faust & Backhaus Environmental Consulting GbRBITZ (Bremer Innovations- und Technologie-Zentrum)
Fahrenheitstrasse 1, 28359 Bremen
2
1. Introduction
Biocidal products are typically mixtures of one or more active substances as well as further ingredients. With
the inclusion of the first substances in Annex I of the Biocidal Products Directive (BPD) 98/8/EC the
authorisation of the corresponding biocidal products at national level is underway and European Member
States (MS) are facing mutual recognitions.
For the authorisation of biocidal products an extensive environmental risk assessment (ERA) is required in
accordance with the BPD, i.e. not only every active substance in the product has to be subject of an
environmental risk assessment, but also substances of concern, i.e., substances leading to classification of
the product or having PBT, endocrine or CMR properties, have to be evaluated separately. In addition
estimation of mixture toxicity of the ingredients is required.
As it is well accepted that mixtures of substances usually elicit a different toxicity than the isolated
substances itself and additive effects up to synergistic effects are possible, it was agreed between the MS at
the Biocides Technical Meeting I / 2011, that the mixture toxicity of the components of biocidal products has
to be considered during the ERA and that the concept of Concentration addition is a suitable method for that.
However, until now no harmonized concept for the assessment of the mixture toxicity of biocidal products
was developed. This means that every MS has its own concept for considering the mixture toxicity during
ERA. But this is problematic due to the mutual recognition of authorisations among the Member States.
Therefore the German Federal Environment Agency (Umweltbundesamt, UBA) proposes a tiered approach
for the assessment of biocidal products based on the available data for the single product components which
also considers the existing approaches of the other MS as well as possible synergistic effects between the
product components.
The aim of the approach presented is to harmonize the assessment of the mixture toxicity of complex
biocidal products among the MS as well as to assess the mixture ecotoxicity of products and, where
relevant, of ecologically relevant mixtures, and at the same time relieve the data requirements for the
applicants as well as additional animal experiments.
2. Suggestion for a tiered approach for the authorisation of biocidal products
The proposed concept acknowledges that the initial available data for the different product components
might be quite different at the start of the assessment. In particular it is currently unclear which data might be
available for product components other than the active substance. Hence the outlined strategy provides
several starting points for beginning with the assessment, depending on the data situation (tier 1 - 4, see
below). This serves the purpose of keeping the data demands as low as possible, while at the same time
ensuring the adequate protection of the environment.
An initial assessment is already possible with a very limited set of data, using an adequately precautionary
approach. In case no indications for a reason for concern are detected, no further testing or data evaluation
is required. If there are potential reasons for concern, additional data can be supplied in order to allow a
more detailed assessment.
The approach uses the sum of PEC/PNECs in a first tier, followed by a Toxic Unit based approximation of
the joint toxicity of the relevant compounds. An important step of the proposed approach is the adequate
identification of product components which are relevant for the mixture assessment. These relevant
compounds are all compounds present in the product that are biologically active and present at a sufficiently
high concentration (“substances of concern”, according to Article 2 of the BPD). If such information is not at
hand, the only risk assessment option is the direct testing of the product. However, a list of criteria on how to
define “relevance” needs to be developed.
2.1 Possible starting points
The proposed concept provides four starting points, each based on a different data situation:
Tier 1: The minimum data that is requested consists of (i) knowledge on the amount of all relevant
compounds in the product, (ii) their relative risk ranking, and (iii) the PEC/PNEC ratio of the most risky
compound, i.e. no precise risk information is requested for all compounds. Such information is only
requested for the compound with the highest risk (PEC/PNEC ratio). For all other compounds, it only needs
to be demonstrated that their risk (PEC/PNEC ratio) is equal or smaller than the risk of the most risky
compound. This allows a convenient way to use censored toxicity data (e.g. from limit tests) and
accommodates for fluctuations in the chemical composition of the biocide product.
Under these circumstances, the initial assessment is based on the assumption that all compounds are as
risky as the most risky compound (worst case estimation, depending on the assumption that the risk ranking
is correct). The final risk of the product is then simply calculated as:
 PEC 
RQProduct = n × SynF × 

 PNEC  max
where n is the number of relevant compounds in the mixture, while the factor SynF accounts for possible
synergistic interactions.
Tier 2: If the PEC/PNEC ratio of all compounds is at hand (but not the actual data for the different trophic
levels, because e.g. data were at least partly collected from previous studies), the risk is calculated as the
sum of all PEC/PNEC ratios:
RQProduct = SynF × ∑
PEC
PNEC
Again, an additional factor SynF safeguards against the possibility of unknown, synergistic interactions.
Tier 3: If at least the base set of ecotoxicological data, i.e. acute toxicity studies using algae, daphnia and
fish, is at hand for all relevant compounds, the summation of toxic units offers a more realistic approach than
summing up PEC/PNEC ratios. The risk of the mixture is then calculated as:
RQProduct = SynF × AF × max(STU )
where AF is the Assessment Factor that is inherent in the PNEC calculation (i.e. 1000 for base-set data),
SynF is again the factor that safeguards against possible synergistic interactions, and max(STU) is the
maximum sum of toxic unit from the three trophic levels. It should be pointed out, that the use of Toxic Units
poses a special challenge in case of unbalanced data situations, when data beyond the base set are
available for some, but not all, relevant components (see discussion below).
Tier 4: Obviously, the direct biotesting of the biocide product is another possibility for risk assessment. It
should be pointed out here that product data are suited only for products that are directly entering the
environment. Otherwise, the mixture that results from the use of the product (e.g. leachates in the case of
wood preservatives), or is predicted to occur (e.g. based on MAMPEC-modeling in the case of antifoulants)
has to be tested.
It might be argued that Independent Action is more appropriate to assess the toxicity of a mixture of nonsimilarly acting compounds, which might be the majority of biocidal products. However, the inadequate use
of IA implies the risk of a toxicity underestimation and hence the applicant would need to actually prove that
IA adequately describes the toxicity of the assessed biocide product. This is only possible by comparing the
measured toxicity of the product with the IA-prediction for each considered trophic level. It would therefore be
easier and less resource demanding, to simply test the product. However, IA-based assessments might
require attention in case of biocidal products that are not entering the environment directly (in which case
product testing would not entirely reflect the environmental risk of the product).
Prioritisation of biocidal substances
for environmental monitoring
Heinz Ruedel1, Burkhard Knopf1, Stefanie Jaeger2, Stefanie Wieck2, Ingrid Noeh2
1
Fraunhofer IME - Institute for Molecular Biology and Applied Ecology, Auf dem Aberg 1, 57392
Schmallenberg, Germany
2
Federal Environment Agency (Umweltbundesamt), Wörlitzer Platz 1, 06844 Dessau-Rosslau, Germany
1. Introduction
The European Biocidal Product Directive (BPD, 98/8/EC [1]) causes a change of the use of biocidal active
substances in EU member states. The placing on the market of a number of substances has already stopped
or they will be withdrawn soon as consequence of non-inclusion decisions. Additionally, the use of certain
biocides will be restricted by risk mitigation schemes. The expected result of both measures should be a
decrease of discharges into the environment of affected biocides. This hypothesis may be proven by an
environmental monitoring in appropriate compartments (e.g. surface water, soil, biota). Such an approach
might also provide information on the necessity of regulatory improvements (e.g., improved risk mitigation
measures). Therefore, a project was initiated by the Federal Environment Agency of Germany to develop a
concept for the selection of biocides for such a monitoring.
First, the current status of biocide monitoring in Germany was investigated. A questionnaire was sent to
authorities involved in routine monitoring (e.g., surface waters for the Water Framework Directive, WFD) as
well as to research groups which run projects in this field (in total 75 addressees). Additionally a literature
research on monitoring results was performed and a database with about 500 monitoring data was compiled
(partly already aggregated data).
As further objective of the study a concept is suggested to prioritise biocidal substances for an environmental
monitoring. In a first step compounds are evaluated for emission characteristics. Since no consumption
figures are available for biocides usage in Germany, the evaluation is performed by considering the intended
use in BPD product types. This assessment is mainly based on a previous study [2]. Additionally, the number
of registered biocidal products which contain the respective compound in Germany is considered as well as
the amount produced or imported EU-wide (classified as high production or low production volume or none
of both). In cases compounds are also authorized as plant protection products this potential contribution is
considered as score, too.
The second step of the prioritisation scheme covers potential effects. Scores are applied for certain classes
of aquatic PNEC values, PEC/PNEC assessment results for biocide specific application scenarios, bioconcentration factors (BCF, to address risks of secondary poisoning) and toxicity (e.g., T+ classification).
The scores from both steps are combined and used to prioritize compounds. In a final step it was evaluated
in which environmental compartment a compound should be investigated (e.g. water, sediment, biota, soil).
This evaluation is based again on use patterns (product type specific emissions to the respective
compartment) and substance specific properties relevant for the partition into the respective compartment
(e.g. biodegradability/persistence, adsorption coefficient K oc , BCF, vapour pressure). In cases where
biocides form stable metabolites these should be assessed in the same way as the parent.
The procedure was tested with a set of 80 biocides which are either already authorised biocides (BPA Annex
I) or candidates (biocidal substances currently in the BPA review programme). Substance-specific data were
retrieved from BPA substance assessment reports (if available). The compounds in this set were mainly not
readily biodegradable.
3.1. Status quo of biocide monitoring in Germany
Evaluation of about 25 detailed answers revealed that the current monitoring mainly covers surface waters
(and partly ground water). Often biocidal compounds are monitored due to legal requirements (priority
substances according to the WFD or the German Surface Water Ordinance). However, most of the biocides
currently covered are those which are (or were until recently) also used as plant protection products (e.g.,
isoproturon, dichlorvos, diuron). Substances investigated which are solely applied as biocides are, e.g.,
cybutryne or clorophene.
Some monitoring data were also available for sewage treatment plants effluents and sludge. Only a few data
were reported for biocide levels in soils. Time series studies with aquatic biota were reported for
methyltriclosan (transformation product of triclosan) and tributyltin compounds (former biocide/antifoulant)
which were detected in fish [3].
3.2. Results from the prioritisation of biocides
About 20 substances from the test data set which are currently not used as plant protection products in
Germany were assessed as probably relevant for an environmental monitoring (e.g., difethialone, dichlofluanid, flocoumafen, methyltriclosan). For these compounds it was assessed which compartment the
monitoring should address. Lists were generated for surface waters (water phase, sediment/suspended
particulate matter, aquatic biota), terrestrial environment (soil, groundwater, biota), STP (effluents, sludge)
and atmosphere. A separate evaluation was made for substances which are also used as plant protection
products.
The plausibility of the prioritisation is discussed with regard to the compiled monitoring data as well as to
prioritisation results from other studies. For example, triclosan was detected often in surface waters and also
ranked high in the here generated list of prioritised compounds for the water compartment. High scores for
monitoring in biota received, e.g., flocoumafen (rodenticide, PBT classified) and methyltriclosan. The latter
result was also consistent with monitoring data [3].
4. Conclusions
The suggested prioritisation approach allows the selection of biocides for environmental monitoring.
Emission related assessment is performed via the intended product types. Further necessary data are easily
retrievable from either biocide assessment reports (e.g. for biocides already authorized) or may be received
from literature or by QSAR estimations.
The prioritisation scheme is recommended as tool within a future monitoring concept to follow environmental
impacts of biocides and possible changes induced by implementation of the BPA (reduction of use
categories, risk mitigation measures, marketing stop after non-inclusion decisions, increased use of certain
substances which may substitute no longer marketed biocides). For this purpose the data base for the
prioritisation has to be broadened to cover beside the already authorized biocides all of the about 250
compounds which are currently under review for the BPA.
However, to allow an assessment of the influence of BPA implementation the monitoring has to focus on
those compounds which are solely used as biocides (and not, e.g., as plant protection products or pharmaceuticals) to allow changes to be correlated to biocide use.
5. Referenced
[1] Directive 98/8/EC of the European Parliament and of the Council of 16 February 1998 concerning the
placing of biocidal products on the market. Official Journal of the European Communities L 123/1-63,
24.4.98.
[1] COWI. 2009. Assessment of different options to address risks from the use phase of biocides. Final
report on behalf of the European Commission Environment Directorate-General, March 2009. COWI A/S,
Kongens Lyngby, Denmark.
[2] Böhmer W, Rüdel H, Wenzel A, Schröter-Kermani C. 2004. Retrospective Monitoring of Triclosan and
Methyl-triclosan in Fish: Results from the German Environmental Specimen Bank. Organohalogen Comp
66:1516-1521.
Acknowledgement - The authors thank the colleagues from German and Swiss monitoring authorities and
research groups which participated in the survey on biocide monitoring and answered the questionnaire and
provided partly comprehensive information and data sets.
Pyrethroids: New contaminants in human breast milk
C. Corcellas1, M.L. Feo1, J.P.M. Torres2, O. Malm2, W. Ocampo-Duque3, E. Eljarrat1*, D.
Barceló1,4
1
Dep. of Environmental Chemistry, IDAEA, CSIC, Jordi Girona 18-26, 08034 Barcelona,Spain
2
Lab. of Radioisotopes, Biophysics Institute, Rio de Janeiro, Brazil
3
Faculty of Engineering, Pontificia Universidad Javeriana, Cali, Colombia
4
Catalan Institute for Water Research (ICRA), Parc Científic i Tecnològic de la Universitat de Girona, Girona,
Spain.
*
1. Introduction
Pyrethroids are synthesized chiral derivates of pyrethrins, which are natural insecticides that are produced
by certain species of chrysanthemum (Chrysanthemum cinerariaefolium). In the last decades, they have
increasingly replaced organochlorine pesticides due to their relatively lower mammalian toxicity, selective
insecticide activity and lower environmental persistence. Thus, they are applied in urban area primarily for
structural pest control, in agricultural areas on crops such as almonds, alfalfa, cotton, lettuce, pistachios, and
peaches, and in the home in pet sprays and shampoos.
Recently some new works point out that its massive agricultural use and domestic one could make
pyrethroids considered as pseudopersistent organic pollutant. This means that they may be constantly
present in environment not because of its stability, but given that its continuous dumping. Moreover, some
studies show that they are toxic, at specific doses on diverse organisms, rats included. As a result of that
and as long as World Health Organization marked them as potentially carcinogenic substances, interest for
this family compound is increased and some works are come out, mostly about its presence on different
kinds of food and natural environment. Nevertheless, studies in breast milk are rarely [1].
The objective of this work has been to study the presence of 12 different pyrethroids in human breast milk
samples collected from different areas, including European (Spain, 6 samples) and South-American (Brazil,
17 samples, and Colombia, 27 samples) countries, with urban and rural areas where different pyrethroids
are used.
2. Method
Analytical methodology includes a liquid extraction from freeze-dry samples with an organic solvent
(hexane:dichloromethane, 2:1) and sonication assistant, followed by a SPE cleanup with alumina and C18
cartridges in tandem. Finally, solvent is evaporated and the sample is reconstituted with ethyl acetate to be
analyzed by GC coupled with MS-MS, following the published method by Feo et al.[2].
1.
Found levels
Results indicate the presence of different pyrethroids in human breast milk from different places, with the
higher contribution for cypermethrin (with concentration levels up to 16 ng/g lw), cyhalothrin (up to 8 ng/g
dw), bifenthrin (up to 7 ng/g lw) and permethrin (up to 5 ng/g lw).
Morover, different patterns were observed depending on the country, so the typical commercial mixtures
clearly influence the pattern. The Figure 1 shows the different pattern for each country depending on the
rural or urban behaivour of the zone. In order to that, in that figure it is exposed the average contribution of
each pyrethroid to the total sum of pyrethroids in milk.
100%
CYPERMETHRIN
80%
PERMETHRIN
60%
ESFENVALERATE/
FENVALERATE
DELTAMETHRIN
/TRALOMETHRIN
l-CYHALOTRIN
40%
20%
BIFENTHRIN
Spain
(urban)
Colombia
(rural)
Colombia
(urban)
Brazil
(rural)
Brazil
(urban)
0%
TETRAMETHRIN
Figure 1: Distribution of pyrethroids in breast milk
In addition, a relationship between total pyrethroids concentration and number of children has been also
founded. In figure 2 it is revealed this tendency on Colombia samples. An analogous behavior was exposed
with Brazilian ones.
18
16
14
y = 12.056e-0.1742x
12
R2 = 0.9964
10
8
6
4
2
0
1st
2nd
3rd
4th
8th
Figure 2: Concentration (ng/g (lw) ) of pyrethroids versus number of children
As long as pyrethroid molecules have different estereoisomers, the isomerspecific enrichment in human
breast milk was also studied and checked with environmental data from the literature. With a few data on the
isomeric characterisation of the environmental samples that could be found, it is no clear the isomer
selectivity in human body.
3.2. Estimated Daily Intake (EDI)
The daily ingestion rate (EDI) of each pyrethroid was estimated. The maximum nursing infant dietary intake
-1
-1
ranged from 0.16 to 4.60 μg (kg of body weight) day , for deltamethrin and cypermethrin respectively.
These values were always lower than ADI (Acceptable Daily Intake) WHO recommends [3].
4. Conclusions
•
Pyrethroids are present in human breast milk.
•
Different patterns were observed depending on the country, so the typical commercial mixtures
clearly influence the pattern.
•
An exponential decay of the concentration levels of pyrethroids in the milk with increasing number of
children was observed.
•
It is very difficult to draw a firm conclusion on the isomer selective accumulation of pyrethroids
without more data, so this is a point to research more accurately.
•
According to the WHO criteria, the measured values are safe for human health.
5. References
[1] M.L. Feo, E. Eljarrat, D. Barceló, 2011. Rapid Commun in Mass Spectrom.25, 869-876.
[2] M.L. Feo, E. Eljarrat, M.N. Manaca, C. Dobaño, D. Barceló, J. Sunyer, P.L. Alonso, C. Menéndez, J.O.
Grimalt, 2011. Environ. Int. Doi:10.1016/j.envint.2011.08.008..
[3] WHO and UNICEF, 2005. World Malaria Report 2005. WHO/HTM/MAL/2005.1102. Geneva: World
Health Organization and United Nations Children’s Fund.Author AB, Author CD, Author EF. 1998. Title of
article. J Agric Food Chem 22:11-33.
Evaluation of Pesticides in Food – A Dynamic Multicrop Model
Ronnie Juraske1, Peter Fantke2, Assumpció Antón3, Raphaël Charles4 and Olivier Jolliet5,6
1
ETH Zurich, Schafmattstrasse 6, CH-8093 Zurich, Switzerland
University of Stuttgart, Hessbrühlstrasse 49a, 70565 Stuttgart, Germany
3
Institute of Agriculture and Food Research and Technology, 08348 Cabrils, Barcelona, Spain
4
Agroscope Changins-Wädenswil, Route de Duillier, P.O. Box 1012, CH-1260 Nyon 1, Switzerland
5
University of Michigan, 109 S. Observatory, Ann Arbor, Michigan 48109-2029, United States
6
Quantis, EPFL Science Park (PSE-D), CH-1015 Lausanne, Switzerland
2
1. Introduction
Human intake of pesticide residues via ingestion of processed food plays an important role for evaluating
current agricultural practice. However, human health impacts of pesticides are still poorly represented in
existing assessments, since almost exclusively effects from diffuse emissions are considered, thereby
disregarding ingestion exposure from residues in field crops after direct pesticide application [1]. For
pesticide residues in food crops and related impacts, steady-state is usually not reached during the short
time period from substance application to ultimate crop harvest, hence, the evolution of residues needs to be
assessed dynamically [2]. In addition, pesticide uptake and translocation mechanisms vary considerably
between crop species and may indicate significant differences in related health impacts [3]. None of the
existing models is able to contrast various crops consumed by humans. Hence, a new modeling approach
was designed to account for pesticide residues in various food crops as source for human pesticide intake.
We developed a new dynamic plant uptake model as fully described in [4] to determine human intake of
pesticide residues via consumption of food crops. The model is based on a flexible set of interconnected
compartments customized to six crops (wheat, rice, tomato, apple, lettuce and potato), thereby accounting
for the major fraction of worldwide human vegetal consumption. Modeled residues are compared with
measured concentrations of 12 different pesticide-crop combinations. Furthermore, the functioning of the
underlying dynamical system is analyzed in detail to estimate uncertainties related to input variables and to
parameterize the complex system for use in spatial modeling approaches [5]. Finally, human intake fractions
are combined with effect information distinguished between cancer and non-cancer effects to characterize
human health impacts. When combined with substance-specific pesticide application statistics, we can
estimate absolute impacts per considered land use area.
3.1. Residues in crops and human intake fractions
Measurements and model estimates correspond well with total crop-specific residual errors ranging between
a factor of 1.5 for lettuce and a factor 19 for rice with an overall residual error a factor of 4.5 over all 12
substance-crop combinations. Higher prediction accuracy is observed in crops where harvested parts are
directly exposed to the applied pesticide (apple, lettuce, tomato, and wheat), whereas crops in which
pesticides have to pass additional media like paddy water (rice) or soil (potato) showed higher uncertainties.
Intake fractions were calculated per unit mass of applied pesticide for 121 substances applied to six crops.
−2
−8
For all crops but potato, intake fractions are usually in the range of 10 and 10 kg intake per kg applied for
typical times between application and harvest. Highest intake fractions are observed for insecticides and
fungicides applied to tomato and lettuce. Intake fractions obtained after direct application were 1 to 5 orders
of magnitude higher than intake fractions estimated for indirect emissions, i.e. fractions lost to air and
freshwater during application (Figure 1). Intake fractions of a pesticide can vary between 4 and 14 orders of
magnitude for fenoxaprop-p and flufenacet, respectively, when applied to different crops at recommended
doses and harvested at typical times after applications. In contrast, intake fractions between all pesticides
applied to the same crop can vary between 3 (potato) and 16 (wheat) orders of magnitude, thereby indicating
that substance properties are almost as influential on iF as crop properties.
The main factors influencing the fate behavior of pesticides are the half-life in plants and on plant surfaces,
the residence time in soil as well as the time between pesticide application and harvest.
Figure 1: Intake fractions for 121 pesticides applied to 6 food crops at different times to harvest ∆t and for steady state [1].
3.2. Implications for current agricultural practice
The combination of large variations in intake fractions with lower variations in effect factors (up to 3.5 orders
of magnitude between pesticides) leads to an overall variation in characterization factors of almost 17 orders
of magnitude. Combining this with variability in applied pesticide mass of 4 orders of magnitude tends to
reduce the overall variability on the impact score to 13 orders of magnitude, suggesting that some of the
pesticides applied at low dose tend to have rather high toxicity potentials. Toxicity potentials can be reduced
up to 99% by defining adequate scenarios for pesticide substitution as demonstrated by an example for
different pesticide combinations applied to wheat against specific pests. However, whenever developing
substitution scenarios, we strongly recommend considering aspects related to pesticide authorization, since
a substance may be authorized for use on particular crops in some countries, but not in others, because of
e.g. decreased susceptibility of target pests (resistance) to certain pesticides.
A simplified model based on the most influential input variables enables the prediction of residues within a
factor of 10 of those calculated with the complex model. This simplified model is adequate to assess direct
residues for multimedia models used for risk and impact assessment and, hence, enables the user to
calculate direct pesticide residues by only providing a very limited set of input information.
4. Conclusions
We demonstrate the importance of dynamic pesticide assessment and distinguish between various food
crops. By identifying the combined influence of crop characteristics, application times and substance
properties we emphasize to choose appropriate times to apply pesticides and that diffuse emission pathways
may be significant for early application, but strongly underestimate human intake for late application before
harvest. Highest impacts are expected via consumption of herbaceous crops and fruit trees with usually high
intake fractions and consumption, while roots and tubers only contribute little due to very low intake fractions.
Substitution scenarios enable us to reduce impacts on human health by choosing alternative pesticides with
similar ability to control unwanted pests, but with lower toxicity. However, substitution must be discussed
separately within each pesticide target class; for example, insecticides can only be substituted by other
insecticides targeting the same insect pests.
5. References
[1] Fantke P, Juraske R, Antón A, Friedrich R, Jolliet O. 2011. Dynamic multicrop model to characterize
impacts of pesticides in food. Environ. Sci. Technol. 45: 8842-8849.
[2] Rein A, Legind CN, Trapp S. 2011. New concepts for dynamic plant uptake models. SAR QSAR Environ.
Res. 22: 191-215.
[3] Trapp S, Kulhánek A. 2006. In: Phytoremediation and Rhizoremediation; Mackova M, Dowling D, Macek
T. (Eds.). Dordrecht: Springer Press. pp 285-300.
[4] Fantke P, Charles R, de Alencastro LF, Friedrich R, Jolliet O. 2011. Plant uptake of pesticides and human
health: Dynamic modeling of residues in wheat and ingestion intake. Chemosphere 85: 1639-1647.
[5] Fantke P, Wieland P, Wannaz C, Friedrich R, Jolliet O. 2012. Dynamics of pesticide uptake into plants:
From system functioning to parsimonious modeling. Environ. Modell. Softw. (submitted).

part 1 - SETAC Europe 22nd Annual Meeting / 6th SETAC World

Transcription

Similar documents

Silencer and Non-silencer Raincaps

Mercury History

NOM-010-STPS

automotive crate engines

Managing mercury and other heavy metals

July 1973 - Warren Astronomical Society

Camp Manitoqua Aquatic Center - Iprm Construction Management

MERCURY PRINT PRODUCTIONS, INC.

Wet Deposition of Mercury In The US

Setting up Your Mercury Network Account