The Impact of the New Chemicals Policy
on Health and the Environment
Final Report
June 2003
prepared for the European Commission
Environment Directorate-General
RPA and
BRE Environment
THE IMPACT OF THE NEW CHEMICALS POLICY ON
HEALTH AND THE ENVIRONMENT
Final Report – June 2003
prepared for
European Commission – Environment Directorate-General
by
Risk & Policy Analysts Limited,
Farthing Green House, 1 Beccles Road, Loddon, Norfolk, NR14 6LT, UK
Tel: +44 1508 528465 Fax: +44 1508 520758
Email: [email protected]
BRE Environment,
Garston, Watford, WD25 9XX, UK
Tel: +44 1923 664862 Fax: +44 1923 664609
Email: [email protected]
RPA REPORT – ASSURED QUALITY
Project: Ref/Title
J427/REACH
Approach:
In accordance with Contract, Steering Group Meeting,
formal comments and associated discussions
Report Status:
Final Report
Meg Postle, Director RPA
Jan Vernon, Business Development Director RPA
Anthony Foottit, Consultant RPA
Tobe Nwaogu, Researcher RPA
Prepared by:
Dave Brooke, BRE
Mike Crooks, BRE
Approved for issue by:
Meg Postle, Director
Date:
16 June 2003
This report, if printed by RPA, is printed on 100% recycled, chlorine-free paper.
RPA & BRE
TABLE OF CONTENTS
EXECUTIVE SUMMARY
i
1.
INTRODUCTION
1.1
1.2
1.3
Background to the Study
Study Aims and Approach
Organisation of Report
2.
EXISTING LEGISLATION AND KEY COMPONENTS OF REACH
2.1
2.2
2.3
The Current Regulatory System
The Proposed White Paper Requirements
The Proposed Data and Testing Regime for REACH
3.
THE CASE STUDY ANALYSIS
3.1
3.2
3.3
Case Study Selection
The Approach to the Analysis
Evaluation of REACH Impact
1
1
2
3
6
9
13
15
20
4.
THE CASE STUDY FINDINGS
4.1
4.2
4.3
Introduction
Base Assessment
Evaluation of the Dossiers
5.
POTENTIAL WIDER IMPACTS OF REACH
5.1
5.2
5.3
5.4
Historical Environmental and Human Health Damages Avoided
The Wider Chemicals Context
Estimates of Substances Having Hazardous Properties
Implications in the Context of this Study
21
21
23
33
39
41
43
6.
CONCLUSIONS
6.1
6.2
6.3
The Study Approach
The Case Study Conclusions
The Wider Impacts of REACH
45
46
48
7.
REFERENCES
51
ANNEX 1: ALTERNATIVE TESTING REGIMES FOR REACH
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Impact of the New Chemicals Policy on Health and Environment
ANNEXES 2-5 - THE CASE STUDIES
CASE STUDY 1: NONYLPHENOLS
1.
INTRODUCTION
1.1
1.2
Background to the Case Study
Format of Case Study
2.
THE EU MARKET PROFILE
2.1
2.2
Uses and Trends
Sectoral Descriptions of Use
3.
ENVIRONMENTAL AND HUMAN HEALTH IMPACTS
3.1
3.2
3.3
4.
4.1
4.2
4.3
4.4
4.5
Introduction
Development of Concerns and Damages
Environmental Damages
Basic Assumptions
Basic Data
Exposure
Risk Assessment
Risk Management
5.
THE REACH DOSSIER CONSIDERED
5.1
5.2
5.3
The Evaluation Approach
Comparison with ESR Risk Approach
Control of Identified Risks
6.
REFERENCES
1-1
1-2
1-3
1-3
1-7
1-8
1-13
THE REACH DOSSIER
1-21
1-21
1-25
1-27
1-27
1-29
1-29
1-35
1-41
CASE STUDY 2: SHORT CHAIN CHLORINATED PARAFFINS
1.
INTRODUCTION
1.1
1.2
Background to the Case Study
Format of Case Study
2.
MARKET PROFILE
2.1
Uses and Trends
3.
ENVIRONMENTAL AND HUMAN HEALTH IMPACTS
3.1
3.2
3.3
4.
4.1
4.2
4.3
4.4
4.5
Introduction
Development of Environmental and Health Concerns
Key Properties and Presence in the Environment
Introduction
Base Data
Exposure
Risk Assessment
Risk Management Recommendations
5.
THE REACH DOSSIER CONSIDERED
5.1
5.2
5.3
The Evaluation Approach
The ESR Risk Assessment
Historical Damage Costs Avoided
2-1
2-1
2-3
2-7
2-7
2-14
THE REACH DOSSIER
2-17
2-18
2-20
2-21
2-22
2-25
2-25
2-29
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5.4
Substitution Issues
6.
REFERENCES
2-31
2-33
CASE STUDY 3: TETRACHLOROETHYLENE
1.
INTRODUCTION
1.1
1.2
Background to the Case Study
Format of Case Study
2.
MARKET PROFILE
2.1
Uses and Trends
3.
ENVIRONMENTAL AND HUMAN HEALTH IMPACTS
3.1
3.2
3.3
Introduction
Development of Environmental and Health Concerns
Key Properties and Presence in the Environment
4.
THE REACH DOSSIER
4.1
4.2
4.3
4.4
4.5
4.6
Overview
Base Data
Environmental Exposure
Environmental Risk Assessment
Human Health Exposure and Risk Assessment
Risk Management Recommendations
5.
THE REACH DOSSIER CONSIDERED
5.1
5.2
5.3
The Evaluation Approach
The ESR Risk Assessment
Historical Damage Costs Avoided
6.
REFERENCES
3-1
3-1
3-3
3-7
3-7
3-12
3-17
3-18
3-21
3-22
3-23
3-23
3-25
3-25
3-28
3-35
CASE STUDY 4: TRIBUTYLTIN
1.
INTRODUCTION
1.1
1.2
Background to the Case Study
Format of Case Study
4-1
4-1
4-2
2.
MARKET PROFILE
2.1
2.2
2.3
TBT in General
TBT in Antifouling Paints
TBT in Wood Preservatives
3.
ENVIRONMENTAL AND HUMAN HEALTH IMPACTS
3.1
3.2
3.3
3.4
3.5
3.6
Introduction
Development of Environmental and Health Concerns
The Need for Harmonised Controls
On-Going Regulation
Key Properties and Presence in the Environment
Substitutes
4.
THE REACH DOSSIER
4.1
4.2
Basic Assumptions
Base Data
4-3
4-3
4-4
4-5
4-5
4-8
4-10
4-12
4-13
4-19
4-19
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Impact of the New Chemicals Policy on Health and Environment
5.
THE REACH DOSSIER CONSIDERED
5.1
5.2
Control of Risks through REACH versus Current Measures
Historical Damage Costs Avoided
6.
REFERENCES
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4-23
4-23
4-25
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Executive Summary
1.
Study Aims and Approach
The European Commission’s White Paper (COM(2001)88 final) sets out a strategy
for a future Community Policy for Chemicals. The White Paper proposes that in the
future new and existing substances should be regulated under the same procedures
and within a single system called REACH (Registration, Evaluation, Authorisation of
CHemicals). The aim of the new strategy is to ensure a high level of protection for
human health and the environment, while ensuring the efficient functioning of the
internal market, and stimulating innovation and competitiveness in the chemical
industry. This is to be achieved by placing an increased responsibility upon industry
to provide data on the properties and uses of chemicals, and in particular existing
chemicals.
This report sets out the findings of a study that examines the potential impacts of
REACH, in terms of the types of environmental and wider public health benefits that
it may help to achieve. The aim of the study was to illustrate how a proactive
approach towards chemicals legislation, i.e. the REACH system, may improve the
environment, and public health in particular, by preventing the accumulation of
potential pollutants until their effects are well known.
The approach adopted to the study involved examination of four case study chemicals
whose uses were prohibited or restricted following observed negative impacts on
health and/or environment, or whose uses are in the process of being restricted
following the outcome of their risk assessment under the current legislative
arrangements.
2.
The Case Studies
The purpose of the case studies was to test the hypothesis that REACH can and is
likely to deliver environmental and general public safety benefits. In order to test this
hypothesis, a series of criteria were established to provide the basis for selecting the
case study substances. These criteria required:
•
•
•
•
•
•
varying restrictions and/or controls on the use of the substances;
the representation of a variety of risks and end-points;
coverage of a range of different types of applications;
consideration of chemicals having different properties of concern;
delays in action being taken where risks were known to exist; and
issues arising with the substitution of the substance by other potentially equally or
more damaging chemicals.
An additional key consideration was the degree to which information on both use and
damages was readily available. This factor alone reduced the set of chemicals that
could be considered in adequate detail to either those that have led to significant
levels of environment or human health damages on a large geographic scale, those
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Impact of the New Chemicals Policy on Health and Environment
where direct linkages with damages are easily demonstrated, or to those which have
been assessed under the EU Existing Substances Regulation (793/73/EEC).
The case studies cannot, however, be considered representative of the estimated
30,000 chemicals placed on the market in the EU at over one tonne per year per
manufacturer or importer. Instead, they are examples of the kinds of substances
which REACH is expected to identify as requiring action. They are substances which
have the kinds of uses which mean that, if the substance has hazardous properties,
there is a greater likelihood of producing effects on humans or the environment.
Furthermore, because the aim of the study was to compare what actions would be
taken under REACH with those taken under the existing framework, it was necessary
to choose substances where action is being taken.
The final list of chemicals for evaluation was agreed with the Commission and
comprises the following substances: Nonylphenol (NP); Short chain chlorinated
paraffins (SCCPs); Tributyltin (TBT); and Tetrachloroethylene (Perc).
3.
The Case Study Analysis
The case study analysis involved an examination of the damages that have arisen over
time due to the failure to control the risks associated with a given substance and the
preparation of a REACH dossier. The case studies have attempted to identify whether
REACH would:
•
•
•
require the same level of test data as required under ESR or other regulatory
regimes;
identify the same endpoints and risk compartments as those identified
(historically) and controlled by the existing legislative arrangements; and
if so, whether the risk reduction measures recommended by this retrospective
application are likely to be similar to those implemented at present.
A REACH dossier, effectively involving the retrospective application of REACH, has
been prepared for each of the case study chemicals. This required a series of
assumptions on: production levels and associated uses; the level of test data available
at the time of dossier creation; what substance-tailored testing would be undertaken;
what assumptions would be made concerning exposure; and how industry might
respond to conclusions concerning environmental risks or risks to man via the
environment.
For NPs, SCCPs and TBT no significant differences arise between the end-points
identified as having unacceptable risks. Only in the case of tetrachloroethylene is
there a significant difference, but it is likely that evaluation by a Competent Authority
would require the further testing necessary to resolve this difference. In terms of test
requirements, few significant differences were identified between what would be
required under ESR and REACH. Level 1 and Level 2 tests were identified as being
necessary by the REACH dossiers. The difference that did arise for
tetrachloroethylene was in relation to testing of degradation products and impacts on
plants from atmospheric releases (testing for which is not yet standard under ESR).
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In terms of risk reduction, the retrospective application of REACH indicated that it
would speed up the rate at which additional test data was produced compared to the
existing situation for non-priority list substances. Another key benefit is the increased
availability of toxicological data on substitutes, which may avoid the use of
environmentally damaging substitutes illustrated by the SCCPs and TBT case studies.
Furthermore, Authorisation and Accelerated Risk Management should ensure that
concerted action is taken more rapidly at the EU level, based on a common
community position.
The case studies found that, in response to the REACH dossiers prepared for each of
the case study chemicals, risk reduction measures would have been adopted. For NPs
and SCCPs, for example, it is assumed that these measures would have been similar to
what has been implemented under ESR. Thus, the costs faced by industry in either
adopting alternative processing methods or substitute chemicals would be similar1.
The key differences would be that the costs would have been incurred earlier in time
and may have related to different volumes (lower or higher) and uses of the
substances. The costs of risk reduction may not, therefore, have been any lower than
those now being incurred.
The case studies conclude that the risks associated with all of the case study chemicals
could have been controlled earlier had the testing, risk assessment and authorisation
requirements of REACH been implemented earlier. Test data available in the 1980s
had already highlighted risk issues. This suggests that damages from the use of each
of the case study chemicals could have (and most probably would have) been reduced
earlier. Table 1 provides an overview of the damages that have arisen from these four
chemicals.
4.
The Wider Impacts of REACH
Four key advantages of REACH over the current system can be identified:
•
•
•
•
1
by assessing the properties of substances and thereby making information
available more quickly, it has the potential to identify a hazard before (substantial)
damage occurs, rather than waiting for monitoring (which is slow and underfunded) to provide evidence of harm;
by providing data in a systematic manner, it enables risks to be assessed
rigorously, allowing effective risk management measures to be identified;
the availability of information on risks enables industry (chemicals manufacturers
and downstream users) to take voluntary action in response to stakeholder
pressure and/or their own policies; and
it provides a basis for quicker regulatory action for the most hazardous substances
(through ARM and authorisation).
The costs of risk reduction have not been re-examined in this study. We have assumed that the costs of
risk reduction would remain similar to those being incurred under ESR or other legislation. In reality
the costs may have varied owing to differences in usage over time, the possibility for industry to put
forward its own measures rather than responding to those proposed by Rapporteurs and the
Commission, and a range of other factors.
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Impact of the New Chemicals Policy on Health and Environment
Table 1: Summary of Historic Damages by Case Study
Case study
Damages
NP
25% to 58% of sewage treatment plants releasing ecologically significant
•
levels of NP/Es into the environment
elevated levels in sewage sludge, preventing land spreading and thus
•
increasing costs of disposal
25% of EU rivers having levels of NP/E that are regularly in excess of the no
•
effect concentration
70% of EU rivers having levels likely to exceed the predicted no effect
•
concentration under low flow
over 50% of observations in freshwaters, marine waters, rivers and lake
•
sediments exceeding the predicted no effect concentrations in affected areas
SCCPs
very bioaccumulative and very toxic substance to aquatic organisms, which
•
may cause long term adverse effects in the aquatic environment
possible involvement in long range transport, as they have been detected in
•
areas and regions remote from any notable sources
detection in higher predatory animals and human breast milk, which may
•
produce irreversible effects in humans (e.g. cancer)
Tetrachloropotential carcinogenic effects on workers through occupation exposure
•
contamination of numerous groundwater resources with example costs of
ethylene
•
remediation varying from €4 to €30 million per waterbody
potential carcinogenic effects on the general population through
•
contamination of drinking water supplies
TBT
geographically widespread impacts on commercially harvested shell fisheries
•
- estimated at €150 million alone at Arcachon Bay, France
documented imposex impacts in as many as 150 species of marine snails,
•
with the exact number of organisms affected unknown
shell deformity effects and larval mortality in aquatic organisms
•
clean-up cost to harbour and port authorities
•
The case studies highlight the fact that, for the chemicals concerned, there was
awareness of their potential impacts long before regulatory action was taken.
However, the information was often incomplete and considerable further data
collection and risk assessment work, taking place over a long period of time, was
necessary before there was agreement on the need for action. In some cases, the
hazards were only identified once environmental damage had occurred, as in the case
of the imposex impacts on dog whelks from TBT. In other cases, such as SCCPs, it
was the widespread distribution of the substance in the environment that led to
recognition of the associated risks.
Had more rigorous testing and risk assessment requirements for existing substances
been introduced in 1981, alongside the requirements placed on new substances,
information to provide the basis for risk management would have been available
sooner and damages to the environment and man could have been reduced. This
argument holds even though our knowledge and expertise concerning the impacts of
chemicals has increased considerably since the mid 1990s through the ESR priority
list programme (and other related work at the international level). Indeed, one could
further argue, that there would have been a speeding up in the development of that
knowledge and expertise.
ESR is a slow and costly process. As additional existing chemicals are subjected to
the more rigorous testing and risk assessment regime established for priority list
substances under ESR, an increasing number are being found to cause damage to the
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environment and public health. For the bulk of chemicals that fall outside the priority
list process only limited testing and risk assessment data are available under the
current regime. Furthermore, within the marketplace, it is often very difficult to
ascertain which chemicals are used in which products and in what quantities. As a
consequence, it would appear inevitable that there may be significant, as yet
undetermined, risks associated with hazardous chemicals placed on the market, which
are not currently subject to rigorous regulation.
Even though the case studies may represent ‘worst case’ scenarios, they also highlight
that there are clear benefits to society of avoiding such damage costs in the future.
Furthermore, other research undertaken indicates that hundreds of substances may be
found to require some form of control in the future. While one might expect the
damage costs for any one substance currently lacking data to be lower than those
highlighted above, the sum of all such damage costs could prove to be significant.
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Impact of the New Chemicals Policy on Health and Environment
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1.
INTRODUCTION
1.1
Background to the Study
On 13 February 2001, the European Commission adopted a White Paper
(COM(2001)88 final) setting out its strategy for a future Community Policy For
Chemicals. The White Paper proposes that in the future new and existing substances
should be regulated under the same procedures and within a single system. The
current system for regulating new substances should be revised and made more
effective and efficient, with the revised obligations being extended to existing
substances. This revised system is called REACH (Registration, Evaluation,
Authorisation of CHemicals).
The aim of the new strategy is to ensure a high level of protection for human health
and the environment, while ensuring the efficient functioning of the internal market,
and stimulating innovation and competitiveness in the chemical industry. This is to
be achieved by placing an increased responsibility upon industry to provide data on
substances, in particular existing substances. The strategy advocates the provision of
earlier and more comprehensive information on substances to downstream users.
Moreover, it would place a requirement upon downstream users to notify the
authorities of uses not originally envisaged by the manufacturer (‘unintentional’ uses)
and to undertake assessments of the risks associated with those uses.
In order to help decision-makers adopt informed positions on this proposed major
legislation, it is important to understand the potential impacts of the legislative
proposals prior to their adoption. Past studies have examined, for example, the
business impacts of the strategy and the potential occupational health impacts of
increased information on chemical properties.
This report sets out the findings of a study that examines the potential impacts of
REACH, in terms of the types of wider public health and environmental
improvements that it may help to achieve.
1.2
Study Aims and Approach
The aim of the study was to illustrate how a proactive approach towards chemicals
legislation, i.e. the REACH system, may improve environmental and public health, in
particular by preventing the accumulation of potential pollutants until their effects are
well known. This was to be achieved by identifying and analysing example
chemicals. These examples were to be taken from the groups of chemicals whose
uses were prohibited or restricted following observed negative impacts on health
and/or environment and or whose uses are in the process of being restricted following
the outcome of their risk assessment under the current legislative arrangements.
The basic hypotheses tested in the case studies was that:
•
the provision of substance tailored testing information on chemicals properties
will allow the swift identification of any risks of possible concern;
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Impact of the New Chemicals Policy on Health and Environment
•
this information will then enable manufacturers and downstream users to respond
by taking (or proposing) suitable actions to reduce those risks to acceptable levels;
where the information indicates that a substance is of very high concern, or where
risk reduction is required at the Community level, then appropriate controls can be
established by the authorities more quickly;
•
such action can be taken before (significant) damage to public health or the
environment occurs; and/or
•
where there is the potential for damage to be currently taking place, this will be
predicted before it would otherwise have been under the current legislative
arrangements, with action taken earlier than would be under the current regime.
In order to test these hypotheses, a set of four case study chemicals were identified
through consultation with the Commission, Member State Competent Authorities and
the European Chemicals Bureau. Once the case study chemicals had been selected,
they were analysed to establish the detrimental impacts on health and environment
caused by the substances over time. This analysis relied on an examination of how
concern about the chemical developed over time, and in some cases the valuation of
impacts at particular sites to illustrate the potential magnitude of damages. The final
step was then to compare the impacts that have arisen from a chemical’s use to date
(as a proxy for impacts under the current legislative arrangements) with the likely
impacts on health and environment had the new REACH chemicals strategy been in
place.
1.3
Organisation of Report
The findings of the study are presented in the remainder of this report:
•
Section 2 provides an overview of the proposed requirements under REACH and
how these have been interpreted for the purposes of this study and compares these
with the requirements of existing legislation;
•
Section 3 presents the case study chemicals and the reasons for their selection,
together with a description of the approach taken to the analysis;
•
Section 4 discusses the REACH dossier developed for each of the case studies,
and compares this to what has occurred under the existing regime; while
•
Section 5 reviews the damages caused by the case study chemicals and draws out
the more general lessons concerning the potential benefits of REACH; and
•
Section 6 provides our conclusions from the study.
More detailed discussion of the case studies are presented in the Annexes, with a
separate Annex provided for each case study.
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2.
EXISTING LEGISLATION AND KEY COMPONENTS OF REACH
2.1
The Current Regulatory System
2.1.1
Introduction
Under the Sixth Amendment to Directive 67/548/EEC, new chemical substances are
subject to a notification regime that requires testing and risk assessment prior to their
marketing in volumes above 10 kg. The testing requirements are tiered according to
the volume of the substance to be placed on the market. As the quantity increases,
more in-depth testing and risk assessments are required. The testing package for
substances marketed above 1 tonne per year (t/y) is referred to as the Base Set.
Testing packages that apply to new substances marketed in volumes over 100 t/y and
over 1,000 t/y are referred to as Level 1 and Level 2 respectively. Comprehensive
data are therefore available on substances placed on the market since September 1981;
in total, around 2,400 new substances have been notified under this regime.
Existing substances, which form the bulk of substances on the EU market (around
30,000 produced at over one tonne per year per manufacturer or importer), are not
subject to the same testing and risk assessment requirements. Instead, requirements
for the classification and labelling of existing substances relate only to data that are
already available. In most cases, this information is not comprehensive.
The Existing Substances Regulation (ESR – Council Regulation (EEC) 793/93 on the
evaluation and control of existing substances) provides for the testing, risk assessment
and risk management of existing substances giving rise to concern. Under this
Regulation, 141 existing substances have been prioritised for comprehensive risk
assessment. In the 10 years since this Regulation was adopted, draft risk assessments
have been completed for 96 substances and conclusions have been agreed for 64 of
these. For the remaining substances, risk assessment work is continuing. For the
non-prioritised existing substances, which form the bulk of those on the EU market,
only limited testing and risk assessment data are available. This may not only mean
that potential risks are unrecognised, but also that the legislation to address risks to
health and the environment, which often relies on the classification of substances,
may not be operating as effectively as it should. Likewise, it means that downstream
users do not have information on the risks that substances may pose to workers, the
environment or consumers, which may affect their choices on the substances used in
processes and end-products.
The requirements under these two Directives, and the differences between them, are
discussed further below. In addition to these, the relevance of the Technical Guidance
Document in support of the risk assessment process for new notified substances and
for existing substances and the OECD Existing Substances programme are discussed
briefly.
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Impact of the New Chemicals Policy on Health and Environment
2.1.2
Directive 67/548/EEC
The Sixth Amendment to Directive 67/548/EEC introduced a system of pre-market
notification for new substances. A detailed notification file must be submitted to a
national competent authority that includes a range of information on the physicochemical, toxicological and ecotoxicological properties of the substance. The extent of
data to be submitted varies with the quantity of the substance to be placed on the market.
Box 2.1 sets out the different levels of information required under this Directive.
Box 2.1: Information Requirements Under Directive 67/4548/EEC
Base Set Dossier Annex Viia (Substances> 1 Ton Per Annum)
Identity: of the notifier and the manufacturer, functions and desired effects of the substance,
estimated quantities on the market
Chemical identity: IUPAC name, structure formula, CAS-number, purity and composition of
impurities, different spectra, determination methods.
Exposure in production/compounding process: exposure estimates related to workplace and
environment
Physicochemical properties: state, granulometry, melting and boiling point, relative density,
vapour pressure, surface tension, water solubility, PoW, flash point, flammability, explosive and
oxidising properties.
Toxicological properties: acute toxicity (oral, dermal and/or by inhalation), skin and eye
irritation, sensitisation, mutagenicity, subacute toxicity (one but relevant route), assessment of
the toxicokinetic behaviour, screening for reprotoxicity
Ecotoxicological properties: acute fish and Daphnia toxicity, growth-inhibition test on algae,
bacterial inhibition, biodegradation, hydrolysis, absorption and desorption screening test.
Recommendations and precautions: precautions at use, storage and transport, recommended
methods for disposal or destruction
Proposals: proposal for classification and labelling, proposal for a safety data sheet.
Annex Viii, Level 1, Additional Testing Data (> 100 Tons Per Annum)
Physicochemical properties: further studies dependent from Annex VII results
Toxicological properties: fertility (one species, one generation, most appropriate route; when
equivocal findings, 2nd generation is required), teratology study (required when positive
indications in fertility study or when not examined), sub-chronic and/or chronic study (required
when positive results in subacute study), additional mutagenesis and/or screening for
carcinogenesis (strategy in Annex V dependent from Annex VII results), basic toxicokinetic
information
Ecotoxicological properties: prolonged Daphnia study, higher plant toxicity, terrestrial toxicity on
earthworms, further fish toxicity studies, bioaccumulation (preferably fish), supplementary
degradation studies when not degradable, further adsorption/desorption studies dependent on
Annex VII results
Annex Viii, Level 2, Additional Testing Data (> 1000 Tons Per Annum)
Toxicological properties: chronic toxicity, carcinogenicity, 3-generation fertility, developmental
toxicity, teratology (other species), toxicokinetic studies, organ/system toxicity
Ecotoxicological properties: additional bioaccumulation, degradation, mobility and
adsorption/desorption, further fish toxicity, avian toxicity, toxicity other organisms
Notification files are reviewed by a competent authority, which prepares a risk
assessment based on the dossier information. There are four possible conclusions
from the risk assessment; they range from ‘no further information about the dangers
of the substance’ is needed to immediate ‘recommendations for risk reduction’. The
latter recommendation could require restrictions on the marketing and use of the
substance in accordance with Directive 76/769/EEC. Once the notification file is
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accepted by the Competent Authority, the substance may be placed on the market
throughout the EU.
Within this process, Competent Authorities have interpreted the requirements of the
Directive according to the need to provide clarification in specific cases. For example,
in relation to test strategies, and more specifically in relation to strategies for the Annex
VIII, Level 1 and Level 2 additional testing, interpretations have been published which
act as a guide to the type of substance-tailored testing required as part of notifications
for substances having different properties.
2.1.3
Regulation (EEC) 793/93
The Existing Substances Regulation ((EEC) 793/93) sets out the requirements for the
provision of data on existing substances. These are significantly different from those
for new substances, and it is these differences that REACH is trying to address. The
data requirements under ESR generally relate to existing information only, with
manufacturers and importers expected to make all reasonable efforts to obtain existing
data on the end-points listed in Box 2.2. Where such data do not exist, manufacturers
and importers are not bound to carry out any further tests in order to provide such
data.
Furthermore, there is no requirement to carry out a risk assessment concerning the use
of that substance, unless a substance is entered onto one of the priority lists. If a
substance is entered onto a priority list, a Competent Authority becomes responsible
for undertaking the risk assessment and proposing any appropriate risk management
measures. As noted above, priority lists currently account for only a small proportion
of chemical substances on the EU market (i.e. 141 out of 30,000 substances produced
at over one tonne per year per manufacturer or importer).
Once a substance has been entered onto one of the priority lists, manufacturers/
importers are required to submit all relevant information and corresponding study
reports for risk assessment of the substance concerned to the European Chemicals
Bureau (ECB), within six months of publication of the list. The manufacturers and
importers who have submitted such information are then obliged to carry out the
testing necessary to obtain any missing data and to provide the test results and test
reports in order to complete the data requirements of Annex VII A to Directive
67/548/EEC. Derogations from these requirements can be requested, for example
when a particular physico-chemical property is not relevant for a substance, or where
data from a higher level test already exists.
In addition, for high production volume substances, further information may be
requested where this is considered necessary on the basis of the risk assessment being
prepared for the substance. Manufacturers and importers are obliged to carry out the
testing necessary to obtain the specified information.
For low volume substances, the Commission in consultation with the Member States
determines the cases in which it is necessary to request manufacturers and importers
to submit additional information for risk assessment purposes. However, there is no
obligation to conduct further testing for that purpose, unless the decision is made
following specified procedures.
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Box 2.2: Information Requirements under the Existing Substances Regulation
Hedset Dossier Annex III (> 1000 Tons Per Annum )
Identity of the submitter and co-submitters, use pattern of the substance, quantities marketed or
imported
Chemical identity: Name, synonyms, molecular and structural formula, CAS-number, purity and
composition of impurities
Physicochemical properties: state, melting and boiling point, relative density, vapour pressure, water
solubility, PoW, flash point, (auto) flammability, explosive and oxidising properties, other available
data.
Environmental fate and pathways: Stability in water and soil, photodegradation, monitoring data,
distribution among and between environmental compartments, biodegradation
Toxicological properties: acute toxicity (oral, dermal, by inhalation or other route), corrosivity,
skin and eye irritation, sensitisation, mutagenicity in vitro and in vivo, subacute toxicity,
carcinogenicity, reprotoxicity, other relevant information, experience with human exposure
Ecotoxicological properties: toxicity to fish, to Daphnia and to other aquatic invertebrates, growthinhibition test on algae, bacterial inhibition, toxicity to terrestrial organisms and to soil dwelling
organisms
C&L: (provisional) classification and labelling.
Hedset Dossier Annex IV (10 - 1000 Tons Per Annum )
Identity of the submitter and co-submitters, use pattern of the substance, quantities marketed or
imported
Chemical identity: Name, synonyms, molecular and structural formula, CAS-number, purity and
composition of impurities
Physicochemical properties: state
C&L: (provisional) classification and labelling.
2.1.4
TGD on Risk Assessment and OECD Existing Substances Programme
The Technical Guidance Document developed to support risk assessments for both
new and existing substances also contains guidance on additional testing. It
elaborates detailed testing strategies for human toxicity and environmental toxicity.
The OECD Existing Substances programme is also relevant. This programme is based
on the Screening Information Data Set (SIDS), which constitutes the minimum data set
needed to carry out an initial assessment of the substance. The SIDS is similar to
Annex VIIA, though it excludes some physical-chemical data regarding flammability
and explosivity properties, irritation and sensitisation but includes an additional
reproductive toxicity screening test (OECD TG 421).
2.2
The Proposed White Paper Requirements
2.2.1
Overview
The REACH system has been proposed as a way of addressing the lack of information
on the potential risks posed by the majority of chemicals on the EU market under the
current legislative framework. The aim is to ensure that equivalent information is
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available on both new and existing substances. The REACH system will comprise
three different elements:
a)
Registration of basic information on chemicals (existing and new) exceeding
a production or import volume of 1 t/y, including test data and preliminary risk
assessments, as well as proposals for classification and labelling, safety data
sheets and proposals for risk management;
b)
Evaluation by authorities of the registered information for all substances
exceeding a production or import volume of 100 t/y and for lower volume
substances where there is concern; and
c)
Authorisation of substances considered to be of very high concern.
Authorisation will be granted for specific uses of such substances only where
it is justified either in terms of the well controlled nature of their use or on
socio-economic grounds.
Manufacturers and importers will be required to notify Competent Authorities of their
intention to produce/import a substance in volumes greater than 1 t/y and to submit a
registration dossier. This is an increase in the threshold that currently applies to new
substances (from 10 kg) and is a new requirement for existing substances.
Competent Authorities will be responsible for the evaluation of registration dossiers.
In the case of substances produced in volumes greater than 100 t/y, the evaluation will
include consideration of the information and the strategy for substance-tailored testing
(relating to Level 1 and Level 2 tests) submitted by industry. The authorities will then
agree with industry an appropriate course of action with regard to any further testing,
risk assessment and risk management requirements. For substances produced in
quantities below 100 t/y, spot checks and computerised screening will be undertaken.
Authorisation will apply to any substance that has hazardous properties giving rise to
very high concern (with there being no volume thresholds). The White Paper
identifies carcinogenic, mutagenic and reprotoxic substances (CMRs categories 1 and
2) and persistent organic pollutants (POPs) as definitely requiring authorisation. It
also highlights the potential for inclusion of persistent, bioaccumulative and toxic
substances (PBTs) and very persistent and very bioaccumulative substances (vPvBs).
There have also been proposals that sensitisers should be included. Exemptions from
authorisation may be given for uses of the substances that do not give rise to concerns,
and continued use may be permitted on the grounds of socio-economic justification,
provided that adequate safety measures are taken.
In parallel to authorisation, an accelerated risk management process is envisaged for
those substances that do not demonstrate the properties that would trigger
authorisation but for which restrictions are required because they pose unacceptable
risks. This process will draw upon the test and exposure data that will become
available as a result of the registration process and the preliminary risk assessments
prepared by industry. In these cases, the greater availability of hazard and exposure
data should make it possible for the type of comprehensive risk assessment that is
carried out for priority listed substances under the current system to be replaced by
more targeted assessments.
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2.2.2
The REACH Dossier
To meet the requirement for equally comprehensive data on new and existing
substances, manufacturers and importers will have to submit a registration dossier,
providing information on:
i)
ii)
production quantities;
the properties of the substance (including test data on physico-chemical,
toxicological and ecotoxicological properties);
iii) intended uses and estimated human and environmental exposures for these;
iv) proposals for classification and labelling of the substance;
v) a safety data sheet;
vi) a preliminary risk assessment covering intended uses; and
vii) proposed risk management measures.
As well as a summary of relevant test data, the dossier will include a preliminary risk
assessment, an indication of any necessary changes in the classification and labelling
of the substance and an indication of any risk management measures that are required
in relation to its manufacture or use. The provision of this information will enable
appropriate action to be taken to control any risks to man or the environment, where
this may include regulations to be introduced as a result of the Authorisation of a
substance, or it’s going through Accelerated Risk Management (ARM).
Or action may be taken voluntarily by the manufacturer of the substance, or by
downstream users, to limit the risks of concern. Indeed, for substances produced in
tonnages below 100t/y, a greater expectation may exist in terms of actions voluntarily
adopted by industry. For these substances, the same level of evaluation by Competent
Authorities is not envisaged, suggesting that it is less likely that such substances will
be highlighted for ARM (although one would expect companies to notify authorities
of substances of ‘very high concern’).
For the higher volume substances, where evaluation of dossiers by Competent
Authorities identifies a range of different uses and potential risks requiring action, one
would expect the substance to be called in for a Community assessment through
ARM. Furthermore, when a substance has been identified as having properties of
high concern, then the dossier will automatically form part of any submission to the
Authorisation process.
A tiered approach is proposed for the submission of dossiers for existing substances.
(Dossiers for new substances will be submitted during notification, as under the
current system). Dossiers for substances produced in higher volumes are to be
submitted earlier than dossiers for lower volume substances, although there is likely
to be sufficient flexibility to allow for the earlier registration of substances produced
in lower volumes particularly if they are substances of concern, having either proven
or suspected hazardous properties.
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2.2.3
Obligations on Downstream Users in Dossier Preparation
The White Paper also proposes that a series of obligations be placed on downstream
users within the above system:
•
•
•
downstream users should inform the authorities of any use that has not been
envisaged by a manufacturer or importer and which is not addressed by the
preliminary risk assessment (these are referred to as ‘unintended uses’);
downstream users should inform the authorities of management measures
different from those reported by the manufacturers or importers; and
downstream users may be required to perform testing, where uses differ from
those originally envisaged by manufacturers or importers and the exposure
patterns also differ substantially from those evaluated by them.
The essential aim of these obligations is to ensure that all uses of chemicals are
covered either within the main dossiers submitted by manufacturers or in ‘postcard’
notifications made by downstream users. The need to register an unintended use
through a postcard notification may arise either because a manufacturer/importer
decides not to support a particular use, or because a downstream user does not wish to
release commercially sensitive information on how it uses a substance.
2.3
The Proposed Data and Testing Regime for REACH
2.3.1
Underlying Principles
The White Paper proposes that the scope of the substance tailored testing regime is
based on the requirements set out in Annex VIII of Directive 67/548/EEC, with the
following general testing regime recommended for new and existing substances:
•
Substances produced/imported in quantities between 1 – 10 t: data on the physicochemical, toxicological and ecotoxicological properties of the substance; testing
should generally be limited to in vitro methods;
•
Substances produced/imported in quantities between 10 – 100 t: ‘base set’ testing
according to Annex VII A of Directive 67/548/EEC. Waiving of testing will be
acceptable on due justification and, in particular, for existing substances;
•
Substances produced/imported in quantities between 100 – 1000 t: ‘Level 1’
testing (substance-tailored testing for long-term effects). The scope of the
additional testing will be based on the requirements set out in Annex VIII of
Directive 67/548/EEC. Guidelines, including decision trees for testing, will be
developed to allow the tailoring of testing according to the results of the available
information, physico-chemical properties, use and exposure to the substance;
•
Substances produced/imported in quantities above 1000 t: ‘Level 2’ testing
(further substance-tailored testing for long-term effects). The scope of the
additional testing will be based on the requirements set out in Annex VIII of
Directive 67/548/EEC. Guidelines, including decision trees for the testing
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strategy, will be developed to allow the tailoring of testing according to the results
of the available information, physico-chemical properties, use and exposure to the
substance.
The White Paper recognises the criticism that the current regime for new substances
does not take differences in exposure sufficiently into account. It recognises the need
for the new system to be more flexible, indicating that guidelines will be developed
on how to carry out substance tailored testing, according to the results of the available
information, physico-chemical properties, use and exposure to the substance.
The aim is to develop a system that allows the waiving of tests or the extension of
testing as appropriate, on the basis of exposure. However, detailed information on
exposure may not be available. As a result, more generic assumptions may have to be
used. For the higher volume substances, therefore, any Level 1 and Level 2 testing
programme will need to be determined by the Competent Authority in consultation
with the manufacturer or importer during the evaluation process.
2.3.2
General Testing Considerations
As part of its work on the development of its more detailed proposals, the
Commission established a series of Working Groups that examined particular issues
raised by the White Paper. One of these focused on the detail of the testing regime
that would apply. This Working Group operated on the principle that “the testing
regime proposals should be built as a system to obtain all the information relevant for
protection of human and environment in an efficient, flexible, human and economic
manner” (TRE/TS01/04/004 REV 1 produced by the Working Group on Testing,
Registration and Evaluation).
Rather than setting out lists of tests that are required, it is proposed instead that the
testing regimes should be based on lists of information required. This type of
approach better recognises that there may be different ways of providing information
(in vitro methods, animal testing, (Q)SAR and read-across to structural analogues,
etc) and that those submitting dossiers should be able to choose the most appropriate
method for the substance of concern.
In relation to substance-tailored testing, the key elements were indicated as being:
•
•
•
•
the use of existing data;
provision of scientific justification for the test strategy (chemistry of the
substance, data on analogues, toxicokinetic and toxicodynamic data, etc);
recognition of the technical non-feasibility of undertaking some tests because of
physico-chemical properties; and
the need to take exposure into account in developing the test strategy.
The underlying aim is not only to ensure that testing requirements are not prescriptive,
but also that the reasons for not providing information are adequately explained and
well justified.
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Two different approaches to building a stepwise system for substance tailored testing
were identified by the Working Group:
•
a list of basic information requirements (BIR) with possible waiving of tests based
on a substance’s properties and sound justification in relation to use and exposure
patterns; and
•
a list of minimum information requirements (MIR), with additional test
requirements depending on the substance’s properties and use and exposure
patterns.
The emphasis of these approaches is different, but it has been argued by industry that
they should theoretically reach the same level of testing requirements (although the
Business Impact Assessment (RPA and Statistics Sweden, 2002) found that the two
implied different levels of testing and associated costs). In particular, it is argued that
the choice between these two approaches will have little influence on the testing
programmes carried out for the higher production volume substances (tonnages
greater than 100 t/y). The so-called Level 1 and Level 2 requirements established for
these substances will be agreed by Competent Authorities and submitters (rather than
be determined by the dossier submitter alone).
For either approach, the following principles were considered of particular
importance:
•
existing information, including all experimental and human data, should be used
as far as possible to obtain the relevant information. Available human data should
be analysed regardless of the substance tonnage;
•
animal tests should only be conducted when their outcome is expected to be
relevant to the risk assessment and where the data requirement cannot be satisfied
by a validated in vitro test method or by other means;
•
testing strategies designed to provide guidance on the systematic and stepwise
gathering of information are presented in the TGD or in OECD guidelines. They
should be used as a tool, in combination with expert judgement, to determine the
need for testing. Any testing strategy should be reconsidered when new data
become available, including exposure related data; and
•
it is paramount that information related to use categories and exposure is available
at the earliest step of the registration process in order to adapt the strategy
accordingly.
The two testing regimes proposed for BIR and MIR, based on the above principles,
are set out in Annex 1 of this report (note that a third testing regime was identified by
the Working Group, which is a variation on MIR).
2.3.3
Testing Option Examined
For the purposes of this study, we have assumed that Option I as set out in Annex 1
(BIR) provides the basis for the testing regime adopted under REACH. The main
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Impact of the New Chemicals Policy on Health and Environment
reason for this decision is that this regime has starting requirements that are more
closely aligned with those of the Dangerous Substances Directive. It is also more
likely to be in line with the types of data required for priority listed substances
undergoing risk assessment under ESR (793/93 (EEC)).
As it has been argued that, in theory, adopting either option should result in the same
data being provided, we have highlighted in the case study dossiers when different
data may have been presented where this is relevant.
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3.
THE CASE STUDY ANALYSIS
3.1
Case Study Selection
3.1.1 The Selection Criteria
The criteria used to select the case study chemicals required that they illustrated:
adoption of different types of restrictions and/or controls on the use of the
substances;
risks associated with different end-points of concern;
risks arising from a range of different types of applications;
risk issues arising from different properties of concern;
delays in action being taken where risks were known to exist; and
issues arising with the substitution of the substance by other potentially equally or
more damaging chemicals.
•
•
•
•
•
•
As a starting point, the case study chemicals were drawn from the set of chemicals
whose uses are being or have been prohibited or restricted. Box 3.1 provides an
indication of the source lists from which chemicals were selected for the purposes of
this study. Reference to these lists provides a means of testing the ability of the
REACH process to identify potential concerns for particular uses of a substance. This
includes consideration of the likely importance of registrations submitted by
downstream users for unintended uses; for example, where a substance being
regulated under one of the lists given in Box 3.1 has been found to have an
unintended use which has subsequently been controlled. It should also help illustrate
how risk management under REACH might differ from the current system.
Box 3.1: Source Lists from European Legislation, Legislative Procedures and International
Agreements
•
•
•
•
•
Substances covered by the Marketing and Use Directive (76/769/EEC);
List I and list II substances under the Dangerous Substances Directive (76/464/EEC);
Priority substances that have undergone risk assessment (or where risk assessment is nearly
complete) under the Existing Substances Regulation (793/93/EEC);
Substances for which controls have been introduced/proposed under the Convention for the
Protection of the Marine Environment of the North East Atlantic (OSPAR Convention); and
United Nations (UN) International Programme on Chemical Safety (IPCS).
In order to assess the value of the substance tailored testing regime, it was important
that the case studies reflected a range of different environmental and public health risk
issues (in terms of exposure pathways and endpoints). This includes risks to specific
environmental compartments and exposure routes as well as a number of
compartments. This would help not only in determining whether REACH would
identify these, but also in reflecting the range of damages (to the environment and
man via the environment) that may be avoided in the future.
In addition, it was considered important that the examples should also cover a range
of use sectors and applications. This included case studies examining risks associated
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Impact of the New Chemicals Policy on Health and Environment
with production of the substance, use as a chemical intermediate, professional use
across sectors having point source and dispersive emissions, and consumer uses.
There are also examples where the need for restrictions and prohibitions was in clear,
with this being the case for a number of persistent, bioaccumulative and toxic
substances. One of the key motivations for the new strategy has been the delay in
introducing regulation, even though an awareness of risks has existed for a long
period of time. Thus, although the case studies needed to include such examples, it
was also important to consider examples where the advantages of controls (compared
to their drawbacks) have been less clear cut.
Another possible advantage of the REACH system lies in improving risk reduction
decision making in relation to the suitability of substitutes. Under the existing
system, the lack of toxicity data on substitutes often constrains comparison of the
advantages and drawbacks of restricting the use of a substance to the consideration of
probable hazard. Under REACH, because data are to be produced for all substances
over the same time period (and associated testing data are to be gathered), this
problem should be significantly reduced, if not eliminated. Thus, the case studies
should examine at least one substance where substitution has been found to be an
issue.
3.1.2
Availability of Information
In addition to these criteria for the selection of examples, there are other practical
considerations concerning the suitability of candidate substances for further analysis
in the study.
In order to prepare a hypothetical dossier for each of the case study chemicals,
information is required on the market as a whole for the substance and on use of the
substance by sector and application type. This data feeds into assumptions on
exposure pathways to the environment. Ideally, such information would be available
for production activities and for a range of use categories, to allow flexibility in
choosing what will be covered by a particular hypothetical dossier. In this regard, the
data generated for risk assessments of priority substances under ESR is
comprehensive for any given substance, while this is not the case for other
legislation/international agreements.
Data availability is also important in relation to toxicity, observed and predicted
effects, associated damages, etc. These data are essential not only for preparing the
dossier, but also for assessing the historical and on-going impacts, including long term
effects, on health and/or environmental compartments for each substance. Again, this
has resulted in a bias towards substances that have been assessed under the priority
lists of ESR, or those which have been singled out within the scientific and academic
literature for their damaging effects.
Furthermore, the extent to which environmental impacts or impacts on general public
safety can be quantified will, to a large extent, be guided by the nature of the available
information. For this study, the aim is to focus on the damages associated with
normal use of the substance, as opposed to accidental discharges from, for example, a
major accident at an installation. The degree to which such quantification can be
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undertaken is limited though for the majority of chemicals, due to a lack of
documented data on the magnitude of effects and difficulties in separating out the
influence a given chemical from other factors. As a result, the information required to
allow quantification exists mainly for those high profile chemicals that have caused
significant levels of damage in the past.
Two additional criteria were added as a result of responses received from consultees.
These relate to the ability of REACH to highlight risk issues arising from the
breakdown products of a substance rather than the substance itself, and the possibility
that dossiers may not highlight a risk issue depending on how it is scoped and
developed.
3.1.3
The Case Study Chemicals
Several substances were originally proposed for consideration by consultees2. Those
substances not selected were eliminated because: one of the case study chemicals
provided a better example, the risks related to mis-use rather than accepted use, the
substance was complicated in terms of the related family and risk issues (e.g.
chromates), or the issues raised would detract from the main purpose of the study.
The final list of chemicals for evaluation was agreed with the Commission was as
follows:
•
•
•
•
Nonylphenol (NP);
Short chain chlorinated paraffins (SCCPs);
Tributyltin (TBT); and
Tetrachloroethylene.
Table 3.1 overleaf provides a brief overview of how these chemicals meet the criteria
set out above.
3.2
The Approach to the Analysis
3.2.1
Overview of Approach
There were three main areas of investigation within the case study analysis:
2
•
the first concerns the damages that have arisen over time due to the failure of
action to control the risks associated with a given substance;
•
the second concerns the type of dossier that is likely to have been produced under
a ‘retrospective’ REACH for each of the substances and how this compares to
what would happen under the current regime; and
Substances suggested but not taken forward include: acrylamide; asbestos, benzene, PCBs,
acrylonitrile, trichlorobenzene, decabromodiphenyl ethers, cumine, MTBE, and chromates.
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Table 3.1: Candidates for Final Selection
Criteria
Substance
Risks are fairly obvious and damages TBT and Tetrachloroethylene
fairly tangible/significant
or
Risks are less obvious and less tangible SCCPs and NPs
Varying restrictions and/or controls on All four chemicals
the use of the substances
Coverage of a range of different types NP and NP ethoxylates (NPEs) together are used by 20
of applications
industry sectors
Other case studies all include different types of
applications: marine anti-fouling paints, dry-cleaning and
solvents and processing additive
A variety of risks and end-points Case studies cover: freshwater, marine, sediment, worker
should be represented
safety and man via the environment together with other
issues (such as long-range transport)
Different properties of concern
Case study substances range from full PBT to only meeting
a sub-set of these criteria and a category 3 carcinogen
Delays in taking action to reduce risks Applies most obviously to TBT and NPs but also to
which may have aggravated damages
Tetrachloroethylene
Risk management has resulted in
substitution, which has subsequently
been found to pose potential risk issues
Additional Criteria
An unintended use or breakdown
product was discovered and found to
present risks
Substance was placed on the priority
list and action is proposed, but
substance tailored testing under
REACH may not identify a problem
•
3.2.2
TBT (replaced by cuprous oxides)
SCCPs (replaced by Medium chain length chlorinated
paraffins with associated issues)
Nonylphenol ethoxylates which breakdown into NP in the
environment
Tetrachloroethylene and trichloroacetic acid breakdown
products
Nonylphenols and their use in nonylphenol ethoxylates
TBT owing to the specificity of the risk end-point
the third is the types of actions that manufacturers and downstream users would be
most likely to have taken in response to any unacceptable risk conclusions arising
from REACH, or whether the substance would be subject to Authorisation or
ARM.
Assessment of Historic Damages
The process that has been adopted to assessing the damages caused by the four
substances has involved the following steps:
1) reviewing the scientific and academic literature to identify when research on
different hazardous properties began and when concern started to arise;
2) making chronological links between the scientific research and the introduction of
either voluntary or regulatory measures aimed at reducing risks to the
environment and/or to public health;
3) collating monitoring data (where available) to illustrate the possible scale of
environmental damages; and
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4) analysing the history of testing and risk management activities in relation to
properties of concern (persistence, bioaccumulation and toxicity, carcinogenicity,
etc.) and developing conclusions on the avoidable damages.
The review of the scientific and academic literature is not intended to be
comprehensive. Instead, the aim was to provide a snapshot of the research activities
that were being undertaken at different points in time. This illustrates how concern
over a particular substance has developed and provides an indication of when there
was sufficient information to indicate the need for risk management.
Linking the emergence of risk management activities in practice to information
provision also acts as a signal of the awareness of and concern over potential
environmental risks. It also highlights where there have been delays in taking action
to minimise or protect against environmental damages or risks to public health (in
terms of man via the environment).
Monitoring data, where available, were compared to either emission limits set in
current regulations or to predicted no effect concentrations (PNECs) to provide an
indication of the possible scale of actual environmental damages. Where there is time
series monitoring data (i.e. data taken for the same site over a period of years), it can
also help clarify the degree to which voluntary or early regulatory measures affected
the potential scale of environmental damages.
In terms of properties of concern, the analysis of avoidable damages focused on a
comparison of the data available at different points in time to the various PBT criteria,
CMR properties, and in relation to other factors such as damages to actual physical
resources. The criteria adopted in establishing whether or not a substance is PBT are
based on those included in the TGD. These are set out in Table 3.2. For the CMR
properties, the criterion was that of whether or not a substance has more recently been
formally categorised in this regard and in what category.
Table 3.3: Assumed Criteria for PBT
Property
Criterion
Persistence
half life >60 days in marine waters and >40 days in estuarine and fresh
waters or >180 days in marine and >120 days in fresh and estuarine
sediments; also a proposed extension to the terrestrial compartment for
half-life in soil >120 days
Bioaccumulative
Bioaccumulation Factor (BCF) >2000
Toxicity
Chronic No Observable Effects Concentration (NOEC) <0.01 mg/l or
CMR or endocrine disrupting effects
very Persistent
half life >60 days in marine, estuarine or fresh waters and >180 days in
marine, estuarine or fresh water sediments; note proposed extension to the
terrestrial compartment for half-life in soil >180 days
very Bioaccumulative
BCF >5000
Source: TGD
3.2.3
Preparation of Hypothetical Dossiers
Developing dossiers for each of the case study chemicals effectively involves the
retrospective application of REACH, to a point in time prior to their being listed as a
priority substance under ESR or otherwise restricted. As it is necessary to project
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Impact of the New Chemicals Policy on Health and Environment
backwards in time, these dossiers are hypothetical in nature, requiring that a series of
assumptions are made concerning:
•
•
•
•
•
production levels and associated uses for a particular manufacturer or consortia
submitting the dossier;
the level of information available to the manufacturer or consortia at the time of
dossier creation;
the substance-tailored testing that would be undertaken by the
manufacturer/consortia for completion of the dossier (in line with Testing Option I
as presented in Annex 1);
the assumptions that would be made concerning exposure and hence the
conclusions that would be reached regarding potential risks; and
initial industry proposals for risk management for any conclusions concerning
environmental risks or risks to man via the environment.
Because REACH is based on manufacturer specific dossiers for identified uses, we
have used historical data on production, the number of manufacturers, types of use
and the number of users to determine the level of testing that would be required (in
terms of base set, Level 1 and Level 2). This permits examination of whether or not
REACH is likely to identify the full range of risks of concern that have been
identified through other processes (such as ESR) and whether it is likely to result in
suitable and sufficient measures for the protection of human health and the
environment.
Data for the base set test requirements was drawn from original IUCLID data sets (in
other words pre-ESR where relevant) for individual manufacturers where possible.
Where this was not possible, we drew on the combined data sets provided by
manufacturers to the ESR process. From an examination of the base set data (for the
appropriate tonnage band), we identified what further tests under Levels 1 and 2 (as
appropriate) would be assumed appropriate for the key endpoints according to the
guidelines set by the Working Group (and for each dossier) reported in Section 2. For
each substance, this results in a series of hypotheses concerning what further tests
would be undertaken at Level 1 and Level 2.
Actual test results were then pulled from the available testing data sets and fed into
the EUSES and EASE models to prepare the risk assessment. In addition,
assumptions were made concerning exposure scenarios. Where exposure data was
readily available (i.e. from the ESR risk assessments3), this was used in the
hypothetical risk assessments prepared here. Where such data were not readily
available, TGD default values were used to estimate exposure. The risk assessment
process set out in the TGD was then applied to determine whether unacceptable risks
arise for any of the uses considered in the hypothetical dossiers created for the
purposes of this study.
Once the risk assessment results were available for each dossier, we assessed whether
the substance would be identified as requiring Authorisation or as a likely candidate
for Accelerated Risk Management. Where this was not the case, we developed
3
Note that BRE prepared the ESR risk assessments for SCCPs, NPs, and Tetrachloroethylene.
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RPA & BRE
scenarios concerning what measures industry would be likely to adopt in order to
reduce risks. These scenarios are use specific and are based on actual industry
responses in the past (e.g. voluntary agreements, product withdrawal from certain
uses, etc.).
3.2.4
The REACH Dossier Compared
The key output from the above work is a comparison of the risk conclusions and risk
management activities resulting from the preparation of a REACH dossier and the
actual outcome of ESR or other risk management activities. In particular, the aim was
to highlight whether they would differ to any significant degree. In theory, had
REACH been in place sooner, it would have identified the case study substances as
being of concern and resulted in some form of risk reduction measures sooner.
As such, the analysis has attempted to identify whether REACH would:
•
•
•
•
require the same level of test data as required under ESR or other regulatory
regimes;
raise any concerns for the example substances and, if so, for which endpoints and
risk compartments;
identify the same endpoints and risk compartments as those identified
(historically) and controlled by the existing legislative arrangements; and
if so, whether the risk reduction measures recommended by this retrospective
application are likely to be similar to those implemented at present.
It is noteworthy that REACH is not being introduced out of concern that existing
procedures, such as ESR, are not sufficiently robust (once a substance has become the
subject of attention). Rather, it is the slow rate at which substances for which data are
lacking can be processed through the procedures such as the ESR priority lists that is
the concern.
As a result, REACH is unlikely to identify any additional risks or risk reduction
measures beyond those that would eventually be put in place for priority list
substances under ESR. Where the risks identified by a retrospective REACH are
similar to those identified under current legislation, it can be assumed that the same
risk reduction options would be applicable.
Under REACH, however, the selection of appropriate substitutes will be facilitated by
the fact that all substances will be undergoing registration (essentially)
simultaneously. Owing to the greater amount of test data that will be available, the
analysis of substitutes is likely to be more thorough and reliable than is possible under
current legislative arrangements. Thus, when examining any differences between
recommendations from a retrospective application of REACH and the types of risk
reduction proposed/implemented under existing provisions, the study has considered
whether any actual issues concerning substitutes have arisen.
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Impact of the New Chemicals Policy on Health and Environment
3.3
Evaluation of Damages Avoided for Case Study Chemicals
A key concept underlying REACH is that it should be able to identify substances/uses
of concern and recommend suitable controls earlier than would otherwise occur under
the existing system. This can be tested, by combining the hypothetical REACH
dossiers with the chronology of scientific investigation and concern with regard to
each of the substances. Through such a comparison, it is possible to highlight the
environmental and public health damages that might have been avoided had REACH
been in place sooner.
For the example substances, REACH is unlikely to recommend anything additional to
the risk reduction measures that have already been implemented, except with regard
to substitution issues. This means that, where the types of risks and risk reduction
measures identified from a retrospective application of REACH are similar to those
identified by ESR, for example, one can conclude that had REACH been in place
earlier damages would have been avoided. The converse is also true. If the
retrospective application of REACH suggests that it may have failed to account for all
risks, then this would suggests that REACH would have failed to prevent all of the
observed environmental or public health damages.
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4.
THE CASE STUDY FINDINGS
4.1
Introduction
The case studies are presented in full in the annexes to this report as follows:
•
•
•
•
4.2
Annex 2 – Case Study 1:
Annex 3 – Case Study 2:
Annex 4 – Case Study 3:
Annex 5 – Case Study 4:
Nonylphenols (NPs);
Short Chain Chlorinated Paraffins (SCCPs);
Tetrachloroethylene (Perc); and
Tributyltins (TBTs).
Base Assessment
The starting point for each of hypothetical manufacturers’ dossiers under the REACH
scenario is a set of base assumptions concerning the relevant production volume, who
is submitting the dossier and for what uses. The base assumptions for each of the
dossiers are summarised in Table 4.1 overleaf.
The dossiers vary in scope, in terms of the completeness of the uses covered. For
example, the nonylphenol dossier does not consider the potential for releases of NP
from the use of nonylphenol ethoxylates (NPE), as was the case in the ESR risk
assessment. This is because NPEs are substances in their own right and, thus, would
be addressed separately by REACH. It is possible that the manufacturers of NP could
consider that their responsibility for the substance ends when it is turned into a
different substance. The producers of NPE would then have to address the releases
associated with the uses of their substances. In an attempt to reflect this type of
situation for REACH, the dossier starts by considering the production of NP and NP
derivatives alone and then explores whether REACH would make the important step
of identifying the relationship between sources of NP in the environment and the use
of NPEs4.
For SCCPs, it has been assumed that the manufacturer produces formulations for
leatherworking processes and, as such, omits entries for:
•
•
•
•
•
metal working (formulation);
metal working (use);
rubber formulations;
paints and sealing compounds; and
textile applications.
Similarly, the TBT dossier focuses on use in anti-fouling paints and wood
preservatives, although the conclusions would apply to most uses. In contrast, the
tetrachloroethylene case study relates to all current uses.
4
Note that if in this case, some of the producers of NP also made the ethoxylates so a close connection
between the two assessments would be expected. Furthermore, if the TGD acts as the basis for the risk
assessments, then breakdown products should be considered.
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Impact of the New Chemicals Policy on Health and Environment
Table 4.1: Applications, Tonnage Bands and Key Issues to be Examined
Dossier
Applications
Tonnage Band and Testing Regime
NPE/NP: Dossier
• Production
prepared by four NP
>1000t/y
• NPEO production
manufacturers as a
• NP/formaldehyde resins
consortium covering
Subject to substance tailored testing
• TNPP production
uses of NP.
requirements up to Level 2 testing (i.e.
• Epoxy resin production
Testing Dossiers A, B, C and D)
• Stabiliser production
Dossier
SCCPs:
prepared by a single
manufacturer
covering uses of
SCCPs.
Tetrachloroethylene:
Dossier prepared by a
consortium of all
manufacturers
Tributyltin: Dossier
prepared by a single
manufacturer of TBT
anti-foulings.
• Phenolic oximes
• Production
• Leather formulation
• Leather
fat
(processing)
>100 -1000t/y
liquors
• Production/intermediate
• Dry cleaning
• Metal cleaning
• Vessels operating on inland
waterways
• Vessels frequently operating
in inshore waters/harbours
(e.g.
servicing/dredging
vessels, tugs, pilot vessels,
ferries)
• Deep sea vessels
• Dry dock operations
• Private vessels
Subject to substance tailored testing
requirements up to Level 1 testing (i.e.
Testing Dossiers A, B and C)
>1000t/y
Subject to substance tailored testing
requirements up to Level 2 testing (i.e.
Testing Dossiers A, B, C and D)
>100 -1000t/y
Subject to substance tailored testing
requirements up to Level 1 testing (i.e.
Testing Dossiers A, B and C)
In developing the hypothetical dossiers, the following assumptions have been made:
•
•
•
•
the data available in the IUCLID submitted to the European Chemicals Bureau
following the introduction of ESR, but before priority listing under ESR, were
available to the manufacturers at the start of dossier preparation;
any further substance tailored testing that is necessary to complete a dossier must
be undertaken in line with the requirements set out for Basic Information
Requirements (BIR) for tonnage bands as set out in Section 2;
where site specific release data are not available, default data from the TGD and
within the EUSES model are applied; exposures for workers and consumers are
estimated using specific data or the methods in the TGD; and
EUSES provides the basis for reaching conclusions as to whether or not
unacceptable risks result from a particular application or sector.
For TBT the data have been taken from the IUCLID submission on the substance,
supplemented with a small amount of data from other sources. There was sufficient
information on measured levels in the environment to prepare the dossier to represent
the period before restrictions on the use of TBTO in anti-fouling paints were
introduced.
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RPA & BRE
4.3
Evaluation of the Dossiers
4.3.1
Overview
The full conclusions of each of the four manufacturers’ dossiers can be found in
Section 4 of the appropriate case study (see Annexes 2 to 5). Table 4.2 provides a
summary of the conclusions from each of the four dossiers concerning the activities
and applications that present a risk (by endpoint). The risk management measures
assumed to be proposed by manufacturers in their dossier submissions are also
provided in the table.
4.3.2
Case Study 1: Nonylphenols
Comparison of REACH Dossier Conclusions with ESR
Because it has been assumed that NP ethoxylates would come under a separate
dossier, this REACH dossier identifies a smaller set of activities as posing risks than
did the ESR risk assessment. Assessing the main uses of NP separately shows that
for most of these no control is required. It is when these uses are combined with the
background emissions of NP from the NP ethoxylates (which contribute 94% of the
total continental burden) that all uses become a concern. Although the ESR risk
assessment reached conclusion (iii) for almost all of the end-points, the contribution
of NP ethoxylates to these conclusions was taken into account when the ESR risk
reduction strategy was developed for both NPs and NPEs. Risk reduction measures
were identified for all applications of NPEs, but only for a few uses of NP.
This difference in conclusions between the REACH dossier and the ESR risk
assessment relates to the following processes and activities:
•
•
•
NP production;
epoxy resin production; and
production of phenolic oximes.
The key question raised by this case study is whether the necessary linkages would be
made between the production and use of NPEs and the fact that they degrade to NP in
the environment, subsequently posing unacceptable risks to the aquatic environment
and potentially the terrestrial environment and from secondary poisoning. If it is
assumed that manufacturers of NP would also be involved in any assessment of the
formulation and use of NPEs (either because they formulate NPE-based products for
downstream users or have an interest in preparing such a dossier), then such a linkage
between emissions of NPEs and NPs in the environment would be made.
Information that NPEs can break down to NP was in the general literature. This may
not be the case for other chemicals, however. An important issues for REACH then is
whether the link between substances and their decomposition products would be
made, where the link is less clear. If not, guidance may need to attempt to address
this issue more robustly.
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Impact of the New Chemicals Policy on Health and Environment
Table 4.2: Conclusions of the Dossiers regarding Risks of Activities
Life cycle step
Water
Sediment
Soil
Secondary
Poisoning
Nonylphenols
Production
NPEO production
NP/formaldehyde resins
TNPP production
Epoxy resin production
Stabiliser production
Phenolic oximes
SCCPs
Leather formulation
Risk Red.
Risk Red.
Risk Red.
Risk Red.
Further monitoring: to improve emission for formulation and processing activities.
Risk Red.
Risk Red.
Risk Red.
Risk Red.
Page 24
Further testing: as the risk characterisation for sediment and soil is based on the
equilibrium partitioning method, and has an extra safety factor of 10, further testing
on sediment and soil organisms would be likely to refine the assessment.
Emissions control: required at formulation sites giving rise to PEC values above the
PNEC values.
Leather fat liquors
(processing)
Regional
Further monitoring of discharges to water at NPEO production sites and downstream
user (formaldehyde resins and stabiliser production) locations to refine estimates of
losses to receiving environments.
As appropriate, additional emissions control technology to be employed to ensure
that emissions are below ecologically significant levels.
Risk Red.
Risk Red.
Tetrachloroethylene
Production/intermediate
Dry cleaning
Metal cleaning
Summary of Risk Management Measures Proposed by Manufacturers
Risk Red.
Risk Red.
Risk Red.
Voluntary Phase Out: in smaller leather processing facilities where emissions
control technology may not be cost-effective. Substitute with either non-chlorinated
processing fat liquor agents or longer chain length chlorinated paraffins (LCCPs: C18
- C20 (liquid)) depending on the results of the LCCP dossier). (Note that MCCPs
would not be chosen as the substitute given the conclusions of their dossier with
regard to use in leather fat liquors).
No further testing is proposed.
No further risk reduction is proposed as it is assumed that tetrachloroethylene is
disposed of properly, as per the controls on disposal already in place.
RPA & BRE
Tributyltins
TBT would be classified as a PBT and the substance would immediately be called in
Vessels operating on
Risk Red.
Risk Red.
for Authorisation and subsequent controls and bans on the use substance in the full
inland waterways
range of applications. (Consideration of what measures would be proposed by
Vessels frequently
manufacturers and users is, hence, immaterial)
operating in inshore
waters/harbours (e.g.
Risk Red.
Risk Red.
servicing/dredging
vessels, tugs, pilot
vessels, ferries)
Deep sea vessels
Risk Red.
Risk Red.
Dry dock operations
Risk Red.
Risk Red.
Private vessels
Risk Red.
Risk Red.
Key: Risk Red. = risk reduction required - = endpoint not considered blank = no risk control identified
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Impact of the New Chemicals Policy on Health and Environment
Control of Identified Risks
Under ESR, the fact that the magnitude of the predicted environmental risks varied
considerably by industry sector led to the most stringent measures being targeted at
those sectors that contribute most to the continental burden. The strategy was based
on a stepped approach, aimed at ensuring that the environmental benefits were gained
in a cost-effective manner by first reducing the continental burden and then
addressing any remaining risks at the local level.
Table 4.3 below summarises the main proposals of the ESR Risk Reduction Strategy
(RPA, 2000) for each of the sectors and the risk management measures it is assumed
would be proposed under REACH in relation to a NP-only dossier. It must be made
clear that the proposals under the REACH dossier would not result in less stringent
level of protection nor does the reduced set of measures under REACH necessarily
mean that unnecessary risk reduction measures by industry would have been avoided5.
The differences arise because of the difference in the scope of the assessments; thus,
one would expect a complementary NPEs REACH dossier to result in proposals for
the remaining measures to be implemented.
Table 4.3: Proposed Risk Reduction Measures – REACH versus ESR
Recommended
Measure
Marketing and
use restrictions
REACH Proposals: NP only
Integrated
Pollution
Prevention and
Control (IPPC)
Production of NPE
Production of phenol/formaldehyde
resins
Production of other plastic stabilisers
Environmental
Quality
Standards/Limit
Values
ESR Risk Reduction Strategy:
NPs and NPEs
Metal working
Pulp, paper and board
Cosmetics and personal care products
Industrial and institutional cleaning
Textile processing
Leather processing
Agriculture (biocidal products, in
particular in teat dips)
Production of NPE
Captive use
Production of phenol/formaldehyde
resins
Production of other plastic stabilisers
Emulsion polymerisation
Formulation for other uses
Production of epoxy resins
Production of phenolic oximes
Paints (production, domestic use and
industrial use)
Civil and mechanical engineering
Electronic/electrical engineering
Mineral oils and fuel
Photographic industry
Source: RPA (2000) and case studies
5
For those categories of use where no risk management under IPPC was identified under the Dossier,
the ESR process identified optional controls because of the linkage with NPE-related background
concentrations. The exception was for TNPP where no risk was identified under ESR, but the option of
instigating controls was included anyway.
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RPA & BRE
Assuming that the link were made between NPEs and their NP decomposition
products in a separate NPE dossier (which is unlikely), then it is likely that
manufacturers and downstream users would argue for voluntary reductions in the key
dispersive uses of NPEs. These would be in place of the prescriptive marketing and
use restrictions that have been recommended under the ESR risk reduction strategy
(as voluntary measures were not considered to provide a sufficiently effective risk
management measure).
However, it is also likely that a NPE dossier would be called in for Accelerated Risk
Management, given the context and timescale of the problems and the concerns
surrounding this substance (described in more detail in Section 3 of the case study).
It may also be the case that risk management under ARM will provide a more robust
mechanism than the ESR process, since all uses would have to be declared. Initially,
a large number of the uses of NPEs were not considered in the risk assessment as they
had not yet been identified. During the course of preparing the risk reduction strategy
these uses were identified and entered into the risk assessment.
However, a significant proportion of NPE usage remains unaccounted for – allocated
to ‘miscellaneous other uses’. This should not occur under REACH, as downstream
users would be required to submit postcard notifications, including risk assessments,
of such uses.
Overall, it is likely that, had REACH been in place earlier, it would have identified
risks and recommended risk management measures much earlier. Most of the data
used in preparing the REACH dossier were available in the early to mid-1980s; where
they were not, substance tailored testing under REACH would have filled the
remaining gaps.
4.3.3 Case Study 2: SCCPs
Comparison of REACH Dossier Conclusions with ESR
The risk assessment carried out under ESR considered not only the risks associated
with the production of SCCPs and their use in leather processing industry as in the
dossier produced here, but also all of the other downstream uses of SCCPS (as
discussed in Section 2 of the case study). Risks were identified for the aquatic
compartment and for secondary poisoning for:
•
•
the formulation (aquatic only) and use of metal working fluids; and
the formulation and use of SCCPs in leather processing.
For most scenarios for sediment and soil, a conclusion (i) was reached, with this
indicating that further information on emissions was required and that testing on
sediment and soil organisms was needed. However, the risk reduction measures
required as a result of the conclusion (iii) findings for the aquatic compartment were
expected to also impact on the risk assessments for sediment and soil. It was
therefore concluded that further monitoring and testing work should await the
outcome of risk reduction proposals.
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Impact of the New Chemicals Policy on Health and Environment
The conclusions of the REACH dossier are very similar for those life cycle scenarios
that it covers. Possible risks are indicated for the aquatic, sediment and soil
compartments for formulation and use in leather processing activities. The key
difference is that there is no risk from secondary poisoning from these uses in the
REACH dossier. This is due to the lower bioconcentration factor used in the
assessment. The BCF values used in the ESR assessment came from studies that were
internal to one industry producer, and it was assumed for the purposes of this
hypothetical dossier that they would not be made available (in the first instance) to the
particular company submitting this dossier. They may be made available later as part
of information sharing, or they may not become available if the company holding that
information does not also wish to register a dossier for SCCPs.
The suggested recommendation within the dossier, that more information on
emissions be sought for the aquatic compartment, was also reached in the ESR
assessment at an early stage, but no better information was provided and so a
conclusion (iii) was reached for the aquatic compartment. If it were assumed that no
further information would be provided for REACH, then a conclusion (iii)
(unacceptable risks) would also be obtained for these endpoints, as the PNEC cannot
be revised upwards.
Control of Identified Risks
One of the outcomes of the ESR risk assessment was the classification and labelling
of SCCPs as being dangerous for the environment (R50/53). The second outcome
was the preparation of a risk reduction strategy. The risk reduction strategy for
leather processing considered a range of different options for managing the risks
associated with the use of SCCPs, and recommended (RPA, 1997):
•
•
Classification and labelling of SCCPs as dangerous for the environment; and
Marketing and use restrictions under Directive 76/769/EEC.
The use of marketing and use restrictions was recommended not only because it was
deemed to be the most effective means of controlling risks to the environment, but
also because the leather processing industry was already moving away from the use of
SCCPs and indicated that the costs of a ban would not have a significant effect on
those companies using SCCPs.
These recommendations have since been
implemented in Directive 2002/45/EEC, which bans the use of SCCPs in both leather
processing and metalworking and leather finishing from late 2003. The Directive also
requires that the European Commission reviews all remaining uses of SCCPs by the
end of 2003.
It is of note that it took five years from the production of the risk reduction strategy to
the introduction of this Directive.
The main difference between the ESR risk reduction strategy and what the
manufacturer might recommend as risk management in the REACH dossier is likely
to be the degree to which use restrictions would be considered. Following current EU
Guidance, SCCPs would now be considered a borderline (in relation to toxicity) PBT.
Thus, the starting point for risk management under REACH may lie with
manufacturers rather than the substance going immediately to authorisation. In this
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RPA & BRE
case, though because the leather processing industry was already moving away from
their use (due in part to historic concerns), this sector would probably have preferred
to switch to substitutes rather than adopt additional emissions control.
More importantly, the case study indicates that had a REACH dossier been prepared
earlier in time, it is likely to have reached the same conclusions with regard to risks to
the environment and the need for some action to be taken. This is despite the fact that
a more limited data set was used for some of the end-points in the hypothetical dossier
produced here. One would also expect a similar dossier to conclude that the use of
SCCPs in metalworking fluids presented risks to the aquatic environment and
sediment.
However, while REACH would have introduced similar controls as ESR, its ability to
ensure suitable substitution is much enhanced over ESR. At the time that the ESR
risk reduction strategies were produced for both leather processing and metalworking,
MCCPs and LCCPs were both considered to pose lower risks than SCCPs (on the
basis of readily available information). As a result, many leather processors and
metalworking facilities may have shifted to the use of these other CPs. In the case of
MCCPs, this is unlikely to have resulted in a significant reduction in risks to the
environment as the draft ESR risk assessment has found unacceptable risks for both of
these sectors (Environment Agency, 2000). These conclusions are strengthened if
Directive 2002/45/EC has led to an increased use in MCCPs.
Since REACH will ensure that information will be available on the risks posed by the
substitutes, downstream users will be able to take better informed decisions when
considering substitutes or alternative processing methods. The result should be a
faster reduction in risks to the environment and man under REACH compared with
ESR.
4.3.4
Case Study 3: Tetrachloroethylene
Comparison of REACH Dossier Conclusions with ESR
The Draft Environmental Risk Assessment for tetrachloroethylene (perc) was issued
in March 2002 by the UK, as rapporteur, on behalf of the EU. The Draft Health Risk
Assessment was issued prior to this, but is currently being revised.
The ESR risk assessment found no risks from tetrachloroethylene production or use
for surface water, sediment, waste water treatment plants or soils. Questions were
raised during the discussions on the assessment about the possible effects of
tetrachloroethylene on plants exposed through the air, and about possible effects of
breakdown products produced through the degradation of tetrachloroethylene in air.
This is still under investigation. Some member states believe that the available
evidence is sufficient to reach a conclusion of risk for this endpoint; other member
states believe that further study is required. These studies are still in progress and
effectively a conclusion (i) currently applies.
The REACH dossier comes to the same conclusions for surface water, sediment,
waste water treatment plants or soils. The other two issues do not emerge directly
from the data requirements. The discussion on degradation in air in the dossier does
Page 29
Impact of the New Chemicals Policy on Health and Environment
include some comments on the formation of the specific breakdown product,
trichloroacetic acid or TCA, which is the subject of the investigation under ESR in
relation to the terrestrial environment (TCA levels in soil have been identified as
posing a risk in some local scenarios).
The information requirements indicated so far, however, for the BIR Dossiers under
REACH do not appear to require a detailed consideration of potential breakdown
products. Such a consideration may have been included if there was evidence that
biodegradation led to the production of a stable product at a high rate (yield), but in
this case the rate of TCA production is relatively low (a few percent).
In relation to effects on plants through atmospheric releases, the BIR Dossier
requirements under REACH do not include any mention of testing by this route. At
the time of the ESR assessment, there were a small number of references in the
literature to possible effects, but these were not well reported or convincing, and the
industry position was that they were not scientifically valid. Under such
circumstances, it is unlikely that the submitter of the REACH Dossier would have
pursued this aspect, or even considered it. Although no strategy for this has yet been
devised, this is seen as an issue particularly for volatile substances which may be
released to air in quantity.
The Draft ESR risk assessment for human health is not currently publicly available, so
no comparison can be made here between it and the conclusions of the hypothetical
industry dossier. However, tetrachloroethylene is to be classified as a category 3
carcinogen and is a candidate for the 29th ATP. There is also some concern for
reproductive toxicity. Taken together, these may affect its use both in an occupational
setting and within consumer products.
The human health assessment is also understood to address the risks to man from
exposure to tetrachloroethylene in groundwater; however, this relates to supplies
which meet EU limits and not contaminated aquifers. The latter are assumed not to be
used as drinking water supply sources.
Control of Identified Risks
No risk reduction strategy has yet been prepared for tetrachloroethylene under ESR.
The preparation of a strategy is not likely to take place until the further information
identified by the environmental risk assessment is available, allowing for firmer
conclusions to be reached with regard to trichloroacetic acid and risks to the terrestrial
environment and possible effects of tetrachloroethylene on plants exposed through the
air. However, given the conclusions reached to date, risk reduction in relation to the
environment would only address emissions to air from its use in chemical synthesis.
A similar need for risk reduction does not arise in the hypothetical REACH dossier
prepared for this case study.
Given that the dossier has been produced to represent a production volume of greater
than 1,000 t/y, it would be subject to evaluation by a Competent Authority under
REACH. This would provide an opportunity for the Competent Authority to raise
questions concerning trichloroacetic acid, effects on plants exposed through the air,
and carcinogenicity and reproductive toxicity. Assuming that such issues are raised, it
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RPA & BRE
is likely that the Competent Authority may request further information be provided or
propose Accelerated Risk Management. This case study also highlights that it may be
important for Competent Authorities to be able to exchange views in a forum such as
the current Technical Meetings for ESR.
Assuming that the Competent Authority does request further information or
Accelerated Risk Management, there would be no significant difference in the
outcome for the environment between REACH and ESR in terms of the robustness of
controls. The further testing that has been required under ESR highlights the fact that
the current regime places no duty on manufacturers of a substance to undertake new
testing in response to risk issues.
If REACH had been implemented sooner, manufacturers would probably have been
obliged to undertake the further testing now being sought. Although substances
produced in tonnages below 100 t/y will not be subject to evaluation by Competent
Authorities, REACH effectively places a duty of care on manufacturers with regard to
any potential risks arising from the release of a chemical to the environment. This
aspect alone may result in companies undertaking the additional testing necessary to
clarify potential uncertainties and associated risks.
The following points demonstrate that there was sufficient concern and evidence to
suggest that these uncertainties would have been addressed earlier under REACH:
4.3.5
•
although research started in the 1980s on the potential impacts of atmospheric
releases of tetrachloroethylene on plants, little further testing was undertaken
subsequently to validate these findings;
•
the possibility of carcinogenic effects was raised in the 1970s; and
•
starting in 1976, a number of different regulatory initiatives were introduced to
reduce releases of tetrachloroethylene to the environment; these were first
introduced in relation to surface and groundwaters, and then in relation to
atmospheric emissions.
Case Study 4: Tributyltins
Unlike the other case studies, there is no single risk assessment document or
mechanism for developing a risk reduction strategy for the use of Tributyltins within
marine anti-fouling paints (in particular) with which to compare the results.
However, given that TBT would be classified as a PBT, it is unlikely that
Authorisation through REACH would propose any controls that are less stringent than
the existing controls. Secondly, it is possible that Authorisation under REACH would
propose wider controls than those currently in place.
The key difference between the hypothetical situation of REACH being in place
versus the ‘real life’ situation (where it was not) is likely to be the speed at which a
suite of controls would be proposed and implemented compared to the more
piecemeal approach that has occurred. One of the reasons for this piecemeal approach
is the international dimension of the problem and of shipping in general. This has
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Impact of the New Chemicals Policy on Health and Environment
complicated both risk management and enforcement issues. In this respect, reducing
the risks of shipping originating/registered outside European under REACH would
still require international agreement and action from the IMO.
However, if REACH had been in place earlier, it would have facilitated a unified
European mechanism and position. At the very least, one would have expected a
speeding up of action to reduce the problem in continental European inland waters
and, probably, inshore and offshore waters. By virtue of more rapid pan-European
action, clearly an important shipping origin and destination, it follows that there could
have been reduced inputs of TBT to international waters and other non-EU waters.
In addition to the probability that, had REACH been in place sooner, risks and risk
reduction measures would have been put in place faster, there is the issue of
substitutes. The consideration of what comprises a suitable and safe alternative to
TBT has also caused both a slowdown in regulatory controls on TBT itself and a shift
in some cases to poor substitutes from a toxicological perspective. Because REACH
will require testing and risk assessment of all existing chemicals, this means that full
dossiers for the substitutes and associated toxicological and fate data would be
available for alternative (chemical based) anti-fouling paints. This may have
increased the speed at which a final decision and substitution was made.
4.3.6
Summary
Table 4.5 provides a summary of the performance of REACH compared to the
existing regime(s) when considering its ability to recognise and control risks quickly
and effectively.
Table 4.5: Comparison of the REACH Dossier with Existing Regimes
Differences in Risk
Performance of REACH versus Existing
Assessment Conclusions
Regime
No significant difference
Enhanced through better use data availability,
Nonylphenols
toxicological data on substitutes and, possibly
implementation.
Partially dependent on REACH linking
substances and decomposition products –
which is considered likely
No significant difference
Enhanced through more rapid toxicological
SCCPs
data gathering to identify uncertainties. Much
enhanced by the availability of data to prevent
unsuitable substitutions that have occurred.
REACH identifies same
Probably enhanced by addressing the large
Tetraoutcomes but there is an issue
number of uncertainties in environmental and
chloroethylene
concerning breakdown
human toxicity that have existed for some time.
products and plant toxicity
through air. It is considered
Dependant on the competent authority
that further testing would be
requesting more test data to address
requested by competent
decomposition products and toxicity to plants.
authority. Assuming this, there
is no significant difference.
No difference
Enhanced through more concerted action being
TBT
taken more rapidly based on a common
community position. In addition, substitution
with more suitable alternatives would have
been more certain.
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5.
THE POTENTIAL WIDER IMPACTS OF REACH
5.1
Historical Environmental and Human Health Damages Avoided
5.1.1 Introduction
Based on the conclusions of the REACH dossiers presented in Section 4, it can be
argued that the risks and consequent damages, associated with all of the case study
chemicals could have, and probably would have, been controlled or avoided had
REACH been implemented earlier. This is particularly true if the requirements placed
on the notification of new chemicals in 1981 under the Sixth Amendment to the
dangerous substances directive had been extended to include existing substances.
In addition to the REACH dossiers, the case studies present an overview of the scale
and extent of damages caused by each of the case study chemicals, the time period
over which these damages occurred and the concerns that were raised over this time
period. Table 5.1 summarises the chronology of events from impacts first being
observed (through studies still considered robust), the first regulatory or voluntary
industry response by Member State or industry association and the first EU response.
The aim of this table is to illustrate the potential length of time over which damages
will have been occurring prior to action being taken under the current system.
In all of these cases, the data required to trigger concern within a REACH-style
dossier was available in the 1980s (at least) to indicate that risk reduction action might
be required or that further testing was necessary. It is the generation of data on
hazardous properties combined with the preparation of risk assessments that is key to
REACH delivering significant environmental and public health benefits.
Table 5.1: Table showing the year in which environmental and health impacts were first
observed for case study chemicals & the year of the initial regulatory and EU regulatory
responses.
1st Regulatory
1st EU
1st Environmental
1st Health
Chemicals
or Voluntary
impact observed
impact observed
Response
Response
1971 (r)
1972 (MS)
Nonylphenols
1970 (r)
1995 (RAR)
SCCPs
1975 (r)
1975 (r)
1991 (IND)
1995 (RAR)
Tetrachloroethylene
1975 (r)
1975 (r)
1987 (MS)
1990 (DIR)
Tributyltin
1976 (d)
1970 (d)
1982 (MS)
1991 (DIR)
NB: These dates are based on documented evidence - undocumented environmental and health
impacts may thus have been observed prior to these dates.
*(r) : Research shows possibility of impact *(d) : Damage occurs to draw attention to impacts
*(RAR): Risk Assessment Report * (DIR): Directive *(MS): Member State *(IND): Industry
It is of note that the case studies also assume that the risk reduction measures adopted
in response to the REACH dossiers (and in particular for NPs and SCCPs) would have
been similar to what has been (is being) implemented under ESR or other legal
instruments. Thus, the costs faced by industry in either adopting alternative
processing methods or substitute chemicals would be similar. The key differences
had REACH been in place sooner is that the costs of control would have been
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Impact of the New Chemicals Policy on Health and Environment
incurred earlier in time and may have related to different volumes (lower or higher)
and uses of the substances. The costs of risk reduction may not therefore have been
any lower than those now being incurred. The damages, however, could have been
reduced significantly.
5.1.2
Summary of Damages Associated with the Case Study Chemicals
Nonylphenols
Both the REACH dossier and the ESR risk assessment identify risks to the aquatic
and terrestrial environment arising from the production and use of NPs (and NPEs).
Although these risks cannot be translated to concrete examples of impacts on
populations of particular fish (although a link has been made in one study) or other
species, this does not mean that damages to the aquatic and terrestrial ecology have
not occurred.
The detailed case study (See Annex 2) highlights that the use of NP/E could have
resulted in an estimated 25% to 58% of sewage treatment plants releasing ecologically
significant levels of NP/Es in effluent. Elevated levels in sewage sludge have also
been a source of contamination. Consideration and comparison of measured levels
from the literature reveals that:
•
•
•
for freshwaters, 52% of observations exceeded the predicted no effect
concentration by a factor of between 1.2 and 1,091 times;
for marine waters, 87% of observations exceeded the predicted no effect
concentration by factors of between 1.3 and 10.3 times; and
86% of river and lake sediments exceeded the predicted no effect concentration by
factors of between 1.3 and 191.
These data are drawn from observations in the literature and represent levels at
locations where one might expect to find NP/Es. As a result, they cannot be directly
extrapolated to determine, for example, that 52% of rivers have levels of NP/E
exceeding the predicted no effect concentrations. In addition, there has been no
representative sampling of rivers in the EU. However, extrapolating from US data,
(as a best estimate), it is estimated that the use of NP/E could have resulted in 25% of
EU rivers having levels of NP/E that are regularly in excess of the predicted no effect
concentration. Furthermore, 70% of EU rivers could have exceeded the predicted no
effect concentration under low flow conditions. The full ecological implications of
such elevated levels are not known.
Considering that most of the data on the effects of Nonylphenols were available in the
early to mid-1980s, it can be suggested that under REACH, the risks from
nonylphenol use would have been identified much earlier, and risk management
measures introduced sooner. Where data were not available, substance tailored
testing under REACH would have filled the remaining gaps. The elevated levels
reported above for freshwaters, marine waters, and river and lake sediments would not
have occurred, or would at the least, been much reduced from those observed today.
Furthermore, control of these risks is still not fully in place although industry has
itself moved away from the use of NPEs. However, as a priority hazardous substance
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under the Water Framework Directive, action will have to be taken to address any
levels in excess of the currently proposed 0.33µg/l Environmental Quality Standard.
It is only by 2009 that measures may be drawn up to tackle remaining discharges to
the aquatic environment and any associated contamination of sediments. It is not until
then that the costs of clean-up will be realised.
More immediately, however, costs may be incurred in meeting limits placed on the
concentration of NP/Es in sewage sludge spread onto land. Based on UK data, an
estimated 13,200 tonnes of NP/Es end up in sewage sludge in the EU. Samples taken
at numerous sewage treatment works in the EU have found levels of NP/Es in sludge
at concentrations well above the proposed limit of 450 mg/kg currently being
proposed for the Sludge Directive. Much lower limits currently apply in Sweden and
Denmark, at 50 mg/kg and 10 mg/kg respectively. As sludge containing NP/Es above
these limits cannot be spread to land, wastewater treatment plant operators will face
the increased costs of landfilling or incinerating these sludges. These increased
disposal costs may become significant at a local level, as the difference in costs
between land spreading and incineration are estimated at between €150 to €190 per
tonne.
Furthermore, because not all uses of NP/Es will be banned under the proposed risk
reduction strategy, these increased disposal costs might be expected to be realised by
an increased number of wastewater treatment plant operators in Europe (assuming the
proposed EU limits become a legal requirement).
SCCPs
The SCCPs case study (See Annex 3) highlights the fact that investigations into the
toxicity, biodegradation and bioconcentration properties of SCCPs started in the
1970s, gaining significant momentum in the 1980s. High levels of SCCPs were
subsequently detected in seabirds (eggs), herons, guillemots, herring gulls, grey seal,
sheep and other mammals.
The present state of knowledge based on over 25 years of SCCPs research shows that:
•
•
•
•
SCCPs are very bioaccumulative with whole body bioconcentration factors of up
to 7,500 found in fish;
SCCPs are very toxic to aquatic organisms and may cause long term adverse
effects in the aquatic environment;
SCCPs may be involved in long range transport, as they have been detected in
areas and regions remote from any notable sources; and
SCCPs have been detected in higher predatory animals and human breast milk,
and may produce irreversible effects in humans (e.g. cancer).
These conclusions have been drawn from research carried out by various
governments. No definite links have been made, however, between the presence of
SCCPS and a specific environmental or human health impact (although it is a
suspected carcinogen). One of the reasons for this is the difficulty involved in
establishing direct pollutant-effect linkages, given the effects of other contaminants
and environmental factors. Levels measured at certain locations associated with
SCCP production and use have, however, been found to be above the predicted no
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Impact of the New Chemicals Policy on Health and Environment
effect level. Furthermore, SCCPS may now meet the criteria for marine PBTs
(although they are borderline toxic).
Taken together, this suggests that had REACH been in place earlier, the
environmental impacts arising from the use of SCCPs could have been minimised
considerably (i.e. by at least 10 years, assuming action had been taken in 1992 rather
than 2002), and on-going impacts minimised. In particular, levels of SCCPs found in
the Arctic and other locations distant from sites of use may be significantly lower.
Furthermore, it could be argued that the testing required to demonstrate conclusively
whether or not SCCPs are persistent in the atmosphere would have been completed by
now, rather than still being the subject of debate at the international level.
As with NPs, it may take years for the full damage costs arising from the use of these
substances in the applications of concern to be realised. For example, a study by
Stevens et al (2003) found particularly high concentrations of SCCPs and MCCPs
within sewage sludge (ranging between 7 to 200 mg/kg dm and 30 to 9700 mg/kg).
Although no limits are currently proposed for chlorinated paraffins within sludge, the
authors note the potential for concern (particularly as many uses will not be restricted
under Directive 2002/45/EEC).
Tetrachloroethylene
An examination of the chronology of research, presented in the case study (See Annex
4), would suggest that human health concerns from tetrachloroethylene use first arose
over 30 years ago. Epidemiological studies concluded in the late 1980s and early
1990s indicated that exposure presented an increased, albeit inexplicable, risk of
developing various forms of cancer.
Based on this time scale of over 20 years, had there been further testing requirements
placed on manufacturers (i.e. through a system such as REACH), there may be less
uncertainty remaining today as to the carcinogenic and reproductive toxicity effects of
occupational exposure to tetrachloroethylene. If tetrachloroethylene is found to be a
category 1 or 2 carcinogen in the future (as a result of further testing), then the lack of
data may have resulted in an increased number of cancers within the EU worker
population.
The damages caused by tetrachloroethylene contamination of groundwater sources
have been significant. Remediation costs have been reported as ranging from €4
million - €30 million for a particular site as shown in the case study. Some of these
damages could have been avoided if tetrachloroethylene had been regulated as a List I
substance earlier. Indeed, tetrachloroethylene could be discharged direct to
groundwater under a consent system until 1990, when it became a List I substance
(along with trichloroethylene and other similar solvents) and guidance on the use and
disposal was formally introduced in the EU under Directive 90/415/EEC.
Given the insolubility and persistence of tetrachloroethylene in groundwaters, one
could argue that under REACH, tetrachloroethylene may have been categorised
sooner as a List I substance and guidelines on its’ use and disposal introduced earlier.
This would probably have promoted an acceleration in the voluntary decline in the
use of tetrachloroethylene in sectors such as dry cleaning, which accounted for a
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significant number of groundwater pollution incidents. Given the magnitude of the
potential damage costs associated with the loss of groundwater resources as drinking
water supply sources across the EU, even a short acceleration in its becoming a List I
substance could have resulted in significant savings in resource costs.
For example, at the lower remediation costs quoted in the tetrachloroethylene case
study, avoidance of future contamination of only 400 drinking water supply sources6
with a difficult to treat substance could save society some €2 billion in remediation
and related costs.
The importance of reducing the time frame over which such damages occur is
highlighted by a much publicised legal case between a leather tannery and a water
company in the UK (ENDS, 1999):
•
In 1976, Water Company X acquired land and abstraction rights to a borehole to
supply drinking water to at least 50,000 people. The same year,
tetrachloroethylene was declared a List II substance by the Commission;
•
In 1980, the EC Directive 80/68/EEC resulted in drinking water being tested for
tetrachloroethylene. Water Company X discovers levels far in excess of
acceptable limits;
•
In 1983, the contamination was finally traced to the borehole acquired in 1976,
which was then shut down. Water Company X incurred costs of €1.5 million
providing an alternative source of water for the residents, and the borehole was
abandoned;
•
In 1993, the water company lost a court case to recoup damage costs, based on the
ruling that the tannery responsible for the contamination, could not have
reasonably foreseen the damage caused. This ruling of ‘unforeseeable damage’
was 15 years after tetrachloroethylene was declared a List II substance; and
•
In 1999, 16 years after the leakage, tetrachloroethylene has been discovered at
concentrations of 15,000 mg/l in plumes rising from the groundwater.
The financial costs faced by this company are in addition to any health impacts (as
tetrachloroethylene is a suspected carcinogen) that may have arisen within the general
public from the use of contaminated drinking water supplies. For example, scientists
in the US have estimated an increased risk of 1 in 1 million of an individual
contracting cancer when concentrations of tetrachloroethylene in drinking water
exceed 1 µg/l, while the EU drinking water limit is 5 µg/l.
However, tetrachloroethylene is just one of many pollutants that have resulted in the
abandonment or the need for expensive remediation works of both groundwater
sources and land. The UK Environment Agency has estimated that some 130
groundwater supply sources have been affected to abandonment by various point
6
Solvent contamination accounted for 250 cases in the UK alone in 1996 (Environment Agency, 996). If a
difficult to treat substance is found in only 20% of cases, this gives 50 such cases in the UK alone.
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Impact of the New Chemicals Policy on Health and Environment
sources of pollution, with a further 370 at risk. Not all of these cases of past pollution
can be linked to inadequate information on the hazardous properties of chemicals.
Yet, even if only a sub-set are, when scaled up to the EU level, the implied damage
costs are significant. It is contamination issues such as these that current proposals
for Article 4 of the Water Framework Directive are trying to address. One of the aims
is to encourage Member States to restore contaminated groundwaters, even though
this may impose significant costs on society, owing to the scarcity of water resources
more generally for use by both man and the environment.
Tributyltins
The use of TBT-based antifouling paints has been identified with impacts across a
range of endpoints. The case study (See Annex 5) highlights the following damages
associated with the use of these paints:
•
•
•
•
the reduction in shellfish stocks on a widespread geographic basis around the
world;
the documented discovery of imposex in as many as 150 species of marine snails,
with the exact number of organisms affected unknown;
shell deformity effects and larval mortality in aquatic organisms; and
corresponding financial losses suffered by the aquaculture industry and costs
imposed on the harbour authorities.
The TBT case study illustrates the potential magnitude of the economic damages that
can arise from the widespread, dispersive use of a PBT chemical. Unfortunately, the
data required to estimate impacts of TBT on shellfish harvests in estuaries across the
EU are lacking, with reliable figures only available for Arcachon Bay, France. In this
case, oyster harvests decreased from production levels of 10,000 to 15,000 tonnes per
year in the mid-1970s to only 3,000 tonnes in 1981 (Ruiz et al, 1996; Evans, 2000).
The financial costs to shellfish farmers can be estimated as ranging between €14
million to €26 million per annum (at current prices), equating to a minimum of €140
million over the 10 year period of a serious decline in oyster harvests. These figures
are just for one estuary, but population-level effects have been widely documented
throughout the EU and elsewhere (e.g. the US and Japan). Thus, one could expect
considerable damage costs to have arisen to shellfishery operators in other estuaries
throughout the EU, with significant impacts documented in Irish, UK, Dutch, French
and Mediterranean waters.
TBT could be regarded as an unusual case, in that concerns arose early in its use and
led to some restrictions at the regional/national level. In addition, its impacts on
molluscs reflect a highly sensitive, chemical specific phenomenon (Santillo et al,
2002). However, it serves to show the potential implications that continued
widespread use of PBT or vPvB substances could have on the environment (and man
via the environment).
In the case of TBT, however, evidence of imposex along shipping lanes and in
proportion to the density of shipping traffic allowed a more speedy and conclusive
linkage of the chemical to its impacts. Indeed, where such direct linkages have been
made for other chemicals, the estimated financial impacts can be equally high. For
example, the Japanese Ministry of the Environment has estimated that for the period
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from 1956 when Minamata Disease (from methylmercury contamination of fish and
shellfish) was first diagnosed to March 2001, 2,955 people had been diagnosed as
having the disease and approximately 144 billion yen had been paid in compensation
by the responsible companies (Ministry of the Environment, 2002).
5.1.3
How Representative are these Case Studies?
In trying to understand the magnitude of the current level of damages that may be
occurring to the environment or man via the environment because of the current
unavailability of data on the majority of existing substances, it is important to
consider how representative the four case studies are. They were selected to reflect a
range of criteria concerning chemical properties, types of uses, identified risks,
regulatory action and substitution issues.
The aim of these criteria was, in part, to ensure that the case studies were more rather
than less representative of the types of risk issues that have arisen in the past and that
are likely to arise in the future. However, in order to ensure that the case studies were
truly representative would require consideration of a much larger number of criteria
(e.g. based on properties of substances, aspects of their use pattern, routes of
exposure, mechanisms of toxicity etc), and there would almost certainly be
exceptions.
Instead, the case studies can be considered as examples of the kinds of substances
which REACH is expected to identify as requiring action. They are mostly substances
which have the kinds of use which mean that, if the substance has hazardous
properties, there is a greater likelihood of producing effects on humans or the
environment. The aim of the study was to compare what actions would be taken under
REACH with those taken under the existing framework, making it necessary to
choose substances where action is being taken.
All of the case studies are high production volume (HPV) chemicals, however, and
one of the reasons why such substances are used in higher tonnages is that there are
no readily substitutable low cost alternatives; to some extent then, most high tonnage
substances will be individual. In this respect, it is worth noting that although
substances were selected for the priority lists under ESR for specific reasons, the risks
identified through the assessments were in many cases in different areas to those for
which they were selected (as discussed further below). An example is the effects on
plants for tetrachloroethylene. Thus, even the priority substances could be considered
as random choices. Also there are no specific hazardous properties which are
‘required’ in a high tonnage substance, so there is no obvious reason to expect that
higher tonnage substances are inherently more hazardous than those used in lower
tonnages. The difference relates to their potential for exposure.
The general paucity of data on existing chemicals, however, makes it difficult to draw
further conclusions concerning the representativeness of the case studies, and thus the
scale of the problem to be addressed by REACH.
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Impact of the New Chemicals Policy on Health and Environment
5.2
The Wider Chemicals Context
The implications of the limited information that is currently available on existing
chemical substances has been the subject of studies within the EU and the US. The
availability of data on chemicals is considered in this section, to provide a better
understanding of the potential magnitude of the future damages that might be avoided
through the implementation of REACH.
5.2.1
Data Availability within the EU
Under ESR, companies are obliged to make every reasonable effort to obtain
information on the HPV chemicals that they produce or import. The data provided by
industry are entered into the IUCLID database to allow the wider exchange and use of
the data (for example, in priority listing substances under ESR).
In 1999, the European Chemicals Bureau (ECB - Allanou et al, 1999) examined the
availability of data on HPV chemicals within the EU and found that:
•
•
Base set data: 31% have data for environment end-points, 22% for human health
end-points, and only 14% for both environment and human health end-points;
Full data set: 5% have data for all environment end-points, 12% for all human
health end-points, and only 3% for all environment and human health end-points.
The study concludes that there are considerable data gaps in relation to both
environmental and human health end-points. Indeed, some 15% have no data at all,
and only 14% have a complete Base set dossier (although this lack of data is partially
explained by the difficulty of testing for some substances – such as petroleum
streams).
Approximately 98% of the 2,465 HPVs included on IUCLID were found to have
some entry in the sections for classification. Of these, around 700 have no R-phrases,
i.e. they are not classified (either because the data indicate no classification, or
because there is insufficient data). Approximately 70% of the HPVs are classified.
The ECB web-site indicates that around 2550 existing substances and around 700 new
substances are included on Annex 1 of the Dangerous Substances directive, i.e. they
are classified with one or more R phrases. This is described as covering some 7000
substances, which means that approximately 50% of the substances considered have
been given some form of classification.
5.2.2 US EPA Data Availability Study
In 1998, the US EPA’s Office of Pollution Prevention and Toxics carried out an
analysis of test data availability for over 2,800 organic HPV chemicals produced in or
imported into the US. The study found that no basic toxicity information – neither
human health nor environmental – is publicly available for 43% of the HPVs
manufactured in the US7. Furthermore, a full set of data (where this relates to OECD
7
In comparing the figures from the US with those for the EU, it is important to note that the US HPV list
does not include petroleum substances, metals and UVCBs (Unknown or Variable composition,
Complex reaction products or Biological material).
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SIDS) is publicly available for only 8% of these chemicals (including those being
assessed at the time under the OECD Chemicals programme).
The study also included a search for data availability for chemicals appearing on the
EPA’s Toxic Release Inventory (TRI), which would be expected to be relatively well
tested. It was noted that around 20% of the TRI HPV chemicals were missing two or
more of the six basic SIDS tests (although 74% of the 91 high release TRI HPVs have
the full SIDS dossier available). More importantly, the EPA note that the majority of
HPVs lack the basic information needed to determine whether they should be listed
on the TRI or not; 46% of the non-TRI chemicals have no data available, with less
than 4% having the full set of basic tests (EPA, 1998).
The EPA also found that of the 193 HPVs that have permissible exposure limits
(PELs), only 52% had basic screening tests for all four of the human health end-points
considered in SIDS. Furthermore, the bulk of HPVs without PELs lack even the
minimal test data needed to support development of PELs for the protection of
workers (US EPA, 1998).
5.2.3
Member States
This latter finding by the US EPA is supported by reports produced by Member State
Competent Authorities, such as the UK Health and Safety Executive (HSE) which
noted that occupational exposure limits have only been set in the UK for 517 of the
total 30,000 substances placed on the market within the EU (HSE, 2002). Gaps in
scientific knowledge on potential health effects and resource constraints affect the
development and validation of new limits. In a consultation document on revising the
Occupational Exposure Limits framework in the UK, the HSE note that many of the
existing occupational exposure standards may not be soundly based; for most of these
substances there are inadequate data to support the setting of a health protective limit
(HSE, 2002).
5.3
Estimates of Substances Having Hazardous Properties
5.3.1
Overview
The EU and US EPA studies summarised above concentrated on the availability of
test data for the HPV chemicals. These are however estimated to account for less than
9% of the 30,000 existing substances currently placed on the market in the EU at over
one tonne per manufacturer or producer. While the precise number of the hazardous
chemicals in use is not known, a number of estimates have been made of the potential
numbers of existing substances having one or more hazardous properties.
5.3.2
Screening and Other Estimates
The Danish EPA (2001) has subjected 47,000 organic compounds from EINECS
(which had not previously been classified) to analysis using QSARs8. Their results
8
Quantitative Structure-Activity Relationships (QSARs) are computer models used to predict hazardous
properties in the absence of measured data.
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Impact of the New Chemicals Policy on Health and Environment
indicate that over 20,600 (about 44% of those analysed) could be classified as having
one or more hazardous properties (where these were taken as acute oral toxicity,
sensitisation, mutagenicity, carcinogenicity and danger to the aquatic environment).
Of note is that nearly 7,000 chemicals (nearly 15% of those analysed) would be
classified as N; R50/53 or N; R51/539.
Screening exercises have also been undertaken in relation to the PBT and vPvB
criteria, as presented in the revised Technical Guidance Document (EC, 2002).
Screening of the data contained in the IUCLID database for 2,000 HPV chemicals
(some of which may no longer be produced at these volumes) by the UK’s
Environment Agency on behalf of the UK Stakeholder Chemicals Stakeholder Forum
(2003) identified 32 PBT and 35 vPvB chemicals (based on the data available and
using the revised TGD criteria (Chemicals Stakeholder Forum, 2002).
Similarly, screening by the Danish EPA (2001) using QSARs predicted that roughly
2% (i.e. 2,000) of the 100,000+ EINECS listed substances would classify as being
either PBT or vPvB. This suggests that when smaller production volumes are taken
into consideration, the end number may well be significantly higher than the 70 to 80
HPV substances which have been identified to date as being PBT or vPvB chemicals
(particularly when intermediates are included).
In addition, 850 substances are currently classified as CMR (Categories 1 and 2)
under Directive 67/548/EEC. The White Paper (2002) adopted a working assumption
that a further 500 may be identified through future testing, with these estimates based
on a review of existing data, experience with ESR and other programmes such as the
OECD HPV programme. It should be recognised that this was a working assumption,
and the figure could be lower or higher. There are also currently some 90 respiratory
sensitisers listed under Annex I of 67/548/EEC, with a further 16 skin sensitisers
listed and which have also had limits placed on their concentration within
preparations (to below 1%). A further 400 substances have been classified as skin
sensitisers, with an unknown number expected to be identified through REACH.
5.3.3
Business Impact Assessment Estimates
In order to estimate testing costs for the Business Impact Assessment (RPA and
Statistics Sweden, 2002), figures generated by the ECB on the level of post Base Set
testing for human health and the environment required for priority substances going
through ESR were adopted. These data may over- or under-predict what will be
required under REACH; the number of substances that will require post-base set
testing is unknown.
As noted above, the ESR substances are ‘data rich’ compared to other HPVs and to a
greater degree to the non-HPVs. Thus, they may require less additional testing than
other existing substances which are ‘data poor’. However, the ESR substances have
9
N; R50/53: dangerous for the environment; very toxic to aquatic organisms, may cause long term
adverse effects in the aquatic environment; and N; R51/53: dangerous for the environment; toxic to
aquatic organisms, may cause long term adverse effects in the aquatic environment.
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been identified in part on the basis of historical risk concerns (but also on other
screening activities), and thus may also represent a set of substances requiring more
than average levels of test data.
The ECB data were, therefore, combined in the Business Impact Assessment with
industry responses on the level of test data currently available for substances falling
under the different production volumes and scenarios to derive the number of
chemicals that would be registered under REACH. Estimates were derived of the
number of substances likely to require substance-tailored testing at Levels 1 and 2
because of their potentially hazardous properties.
Under Scenario 3 (that representing best estimates), it was assumed that 2,000
substances produced in volumes greater than 100 tonnes per year would require Level
1 testing for either environmental or human health end-points, with a further 1,200
requiring testing for both sets of end-points. For Level 2 testing, roughly 1,000
substances would require testing for either environment or human health, with a
further 600 requiring testing for both environment and health end-points. Taken
together, these estimates suggest that some 600 of the substances currently placed on
the market are likely to be identified as having properties of concern. As noted above,
this could be either an over- or an under-estimate as it is based on what was required
for ESR substances, which may be of higher risk than most substances.
5.4
Implications in the Context of this Study
The above discussion has highlighted the potential significance of the lack of data on
chemical properties. It effectively means that, for the majority of chemicals,
regulators, downstream users and consumers lack the information required to
determine whether a given use is ‘safe’ in risk terms. This is highlighted by the
concerns raised by the US EPA evaluation of data availability, by the UK HSE report
and the Danish EPA analysis amongst others. This problem is further compounded by
the conclusion that, even for data rich substances, our a priori knowledge of the
potential risks arising from the use of particular chemicals is poor, particularly in
relation to human health (and in particular man via the environment and workers)
(RIVM, 2002).
Although one cannot extrapolate from the damages caused by the case study
chemicals to the other 30,000 in order to estimate the level of unknown damages
currently being caused, they do provide indicators of the types of damages that should
be reduced or avoided in the future as a result of REACH. From the preceding
sections, four key advantages of REACH over the current system can be identified:
•
•
by assessing the properties of substances, and thereby making information
available more quickly, it has the potential to identify a hazard before (substantial)
damage occurs, rather than waiting for monitoring (which is slow and underfunded) to provide evidence of harm;
by providing data in a systematic manner, it enables risks to be assessed
rigorously, allowing effective risk management measures to be identified;
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Impact of the New Chemicals Policy on Health and Environment
•
•
Page 44
the availability of information on risks enables industry (chemicals manufacturers
and downstream users) to take voluntary action in response to stakeholder
pressure and/or their own policies; and
it provides a basis for quicker regulatory action for the most hazardous substances
(through ARM and authorisation).
RPA & BRE
6.
CONCLUSIONS
6.1
The Study Approach
The aim of this study has been to illustrate how a proactive approach towards
chemicals legislation, i.e. the REACH system, will improve the environment and
public health in particular, by preventing the accumulation of potential pollutants until
their effects are well known. This aim has been translated into the following
hypotheses which have been tested through the use of case studies:
•
the provision of substance tailored testing information on chemicals properties
will allow the swift identification of any risks of possible concern; this will occur
more quickly than would take place through more traditional monitoring
activities;
•
by providing systematic data, it enables the risk to be assessed rigorously,
allowing effective risk management measures to be identified; and
•
this information will enable manufacturers and downstream users to respond by
taking (or proposing) suitable actions to reduce risks to acceptable levels or to
eliminate them; where the information indicates that a substance is of very high
concern, or where risk reduction is required at the Community level, then
appropriate controls can be implemented more quickly by the authorities.
In order to test this hypothesis, four case study chemicals were selected based on
criteria concerning current levels of controls on the use of the substances, the risks of
concern, usage patterns, regulatory delays and issues arising from substitution. Just as
importantly, the choice of the case studies was constrained by data availability. This
alone reduced the set of chemicals that could be considered in adequate detail to those
that have led to significant levels of environment or human health damages on a large
geographic scale, those where direct linkages with damages are easily demonstrated,
or to those which have been assessed under the EU Existing Substances Regulation
(793/73/EEC).
The case studies cannot be considered representative, however, of the estimated
30,000 chemicals currently placed on the market in the EU at over one tonne per
manufacturer or producer. Instead, they are examples of the kinds of substances
which REACH is expected to identify as requiring action. They are substances which
have the kinds of use which mean that, if the substance has hazardous properties,
there is a greater likelihood of producing effects on humans or the environment.
Furthermore, because the aim of the study was to compare what actions would be
taken under REACH with those taken under the existing framework, it was necessary
to choose substances where action is being taken.
The case study analysis has had three main strands of investigation:
•
the first strand relates to the damages that have arisen over time due to the failure
of action to be taken sooner to control the risks associated with a given substance;
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Impact of the New Chemicals Policy on Health and Environment
6.2
•
the second concerns the type of dossier that is likely to have been produced under
a ‘retrospective’ REACH for each of the substances and how this compares to
current assessments of risk; and
•
the third is the types of actions that manufacturers and downstream users would be
most likely to have taken in response to any unacceptable risk conclusions arising
from REACH, or whether the substance would be subject to Authorisation or
ARM.
The Case Study Conclusions
REACH dossiers were prepared for each of the case study dossiers, with this being a
key part of the REACH process. We then considered how the REACH process would
compare to the conclusions that have been reached under the existing regime. This
has allowed us to identify whether REACH would:
•
•
•
•
require the same level of test data as required under ESR or other regulatory
regimes;
identify the same endpoints and risk compartments as those identified
(historically) and controlled by the existing legislative arrangements;
if so, whether the risk reduction measures recommended by this retrospective
application are likely to be similar to those implemented at present; and
lead to action being taken sooner than under the current system and hence reduce
levels of environmental damage and risk to man via the environment.
For NPs, SCCPs and TBT no significant differences arise between the end-points
identified as having unacceptable risks. Only in the case of tetrachloroethylene is
there a significant difference, but it is likely that evaluation by a Competent Authority
would require the further testing necessary to resolve this difference. In terms of test
requirements, few significant differences were identified between what would be
required under ESR and REACH. Level 1 and Level 2 tests were identified as being
necessary by the REACH dossiers.
The difference that did arise for
tetrachloroethylene was in relation to testing of degradation products and impacts on
plants from atmospheric releases (testing for which is not yet standard under ESR).
In terms of risk reduction, the retrospective application of REACH indicated that it
would speed up the rate at which additional test data was produced compared to the
existing situation for non-priority list substances. Another key benefit is the increased
availability of toxicological data on substitutes, with the fact that this may avoid the
use of environmentally damaging substitutes illustrated by the SCCPs and TBT case
studies. Furthermore, Authorisation and Accelerated Risk Management should ensure
that concerted action is taken more rapidly at the EU level, based on a common
community position.
The case studies found that, in response to the REACH dossiers prepared for each of
the case study chemicals, risk reduction measures would have been adopted. For NPs
and SCCPs, for example, it is assumed that these measures would have been similar to
what has been implemented under ESR. Thus, the costs faced by industry in either
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adopting alternative processing methods or substitute chemicals would be similar10.
The key differences would be that the costs would have been incurred earlier in time
and may have related to different volumes (lower or higher) and uses of the
substances. The costs of risk reduction may not, therefore, have been any lower than
those now being incurred.
The case studies conclude that the risks associated with all of the case study chemicals
could have been controlled earlier had the testing, risk assessment and authorisation
requirements of REACH been implemented earlier. This suggests that damages from
the use of each of the case study chemicals could have (and most probably would
have) been reduced earlier. Table 6.1 provides an overview of the damages that have
arisen from these four chemicals.
Table 6.1: Summary of Historic Damages by Case Study
Case study
Damages
NP
25% to 58% of sewage treatment plants releasing ecologically significant
•
levels of NP/Es into the environment
elevated levels in sewage sludge, preventing land spreading and thus
•
increasing costs of disposal
25% of EU rivers could have levels of NP/E that are regularly in excess of
•
the no effect concentration
70% of EU rivers could have levels exceeding the predicted no effect
•
concentration under low flow
over 50% of observations in freshwaters, marine waters, rivers and lake
•
sediments exceeding the predicted no effect concentrations in affected
areas
SCCPs
very bioaccumulative and very toxic substance to aquatic organisms,
•
which may cause long term adverse effects in the aquatic environment
possible involvement in long range transport, as they have been detected in
•
areas and regions remote from any notable sources
detection in higher predatory animals and human breast milk, which may
•
produce irreversible effects in humans (e.g. cancer)
Tetrachloroethylene • potential carcinogenic effects on workers through occupation exposure
contamination of numerous groundwater resources with example costs of
•
remediation varying from €4 to €30 million per waterbody
potential carcinogenic effects on the general population through
•
contamination of drinking water supplies
TBT
geographically widespread impacts on commercially harvested shell
•
fisheries - estimated at a minimum of €140 million alone at Arcachon Bay,
France
documented imposex impacts in as many as 150 species of marine snails,
•
with the exact number of organisms affected unknown
shell deformity effects and larval mortality in aquatic organisms
•
clean-up cost to harbour and port authorities
•
10
The costs of risk reduction have not been re-examined in this study. We have assumed that the costs of
risk reduction would remain similar to those being incurred under ESR or other legislation. In reality
the costs may have varied owing to differences in usage over time, the possibility for industry to put
forward its own measures rather than responding to those proposed by Rapporteurs and the
Commission, and a range of other factors.
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Impact of the New Chemicals Policy on Health and Environment
6.3
The Wider Impacts of REACH
Four key advantages of REACH over the current system can be identified:
•
•
•
•
by assessing the properties of substances and thereby making information
available more quickly, it has the potential to identify a hazard before (substantial)
damage occurs, rather than waiting for monitoring (which is slow and underfunded) to provide evidence of harm;
by providing data in a systematic manner, it enables risks to be assessed
rigorously, allowing effective risk management measures to be identified;
the availability of information on risks enables industry (chemicals manufacturers
and downstream users) to take voluntary action in response to stakeholder
pressure and/or their own policies; and
it provides a basis for quicker regulatory action for the most hazardous substances
(through ARM and authorisation).
The case studies highlight the fact that, for the chemicals concerned, there was
awareness of their potential impacts long before regulatory action was taken.
However, the information was often incomplete and considerable further data
collection and risk assessment work, taking place over a long period of time, was
necessary before there was agreement on the need for action. In some cases, the
hazards were only identified once environmental damage had occurred, as in the case
of the imposex impacts on dog whelks from TBT. In other cases, such as SCCPs, it
was the widespread distribution of the substance in the environment that led to
recognition of the associated risks.
Had more rigorous testing and risk assessment requirements for existing substances
been introduced in 1981, alongside the requirements placed on new substances,
information to provide the basis for risk management would have been available
sooner and damages to the environment and man could have been reduced. This
argument holds even though our knowledge and expertise concerning the impacts of
chemicals has increased considerably since the mid 1990s through the ESR priority
list programme (and other related work at the international level). Indeed, one could
further argue, that there would have been a speeding up in the development of that
knowledge and expertise.
ESR is a slow and costly process. As additional existing chemicals are subjected to
the more rigorous testing and risk assessment regime established for priority list
substances under ESR, an increasing number are being found to cause damage to the
environment and public health. For the bulk of chemicals that fall outside the priority
list process only limited testing and risk assessment data are available under the
current regime. Furthermore, within the marketplace, it is often very difficult to
ascertain which chemicals are used in which products and in what quantities. As a
consequence, it would appear inevitable that there may be significant, as yet
undetermined, risks associated with hazardous chemicals placed on the market, which
are not currently subject to rigorous regulation.
Even though the case studies may represent ‘worst case’ scenarios, they also highlight
that there are clear benefits to society of avoiding such damage costs in the future.
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Furthermore, research undertaken elsewhere indicates that hundreds of substances
may be found to require some form of control in the future. While one might expect
the damage costs for any one substance currently lacking data to be lower than those
highlighted above, the sum of all such damage costs could prove to be significant.
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Impact of the New Chemicals Policy on Health and Environment
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7.
REFERENCES
Allanou R et al (1999): Public Availability of Data on EU High Production Volume
Chemicals, EUR 18996 EN, European Chemicals Bureau, European Commission
Joint Research Centre.
Bernson V (1999): Chemicals Control Regulations vs Enlisting Marketplace Competition
and Burden of Proof, paper to KEMI board, dated 19 April 1999.
Chemicals Stakeholder Forum (2003): The Stakeholder Forum’s PBT and vPvB Criteria,
paper discussed at the 11th Meeting on 11 March 2003 (and associated papers
discussed at earlier meetings, with particular reference to those of 11 June and 10
September 2002).
Danish EPA (2001): Report on the Advisory List for Self-classification of Dangerous
Substances, Environmental Project No. 636, Copenhagen, Miljøstyrelsen.
Danish EPA (2001a): Identification of Potential PBTs and vPvBs by use of QSARs, draft
report dated 1 November 2001.
ENDS (1999): Sixteen Years on, Agency Discovers New Risks at
Eastern Counties Leather,
ENDS Report, Vol 293, June 1999.
European Commission (2002): Technical Guidance Document on Risk Assessment, draft
copy dated May 2002.
Health & Safety Commission (2002): Discussion Document on Occupational Exposure
Limits (OEL) Framework, Discussion Document, HSE, London.
Joint Research Centre (2003): Report of JRC Expert Group on Chemical Intermediates,
draft report dated 23 January 2003.
RIVM (2002): Evaluation of EU Risk Assessments Existing Chemicals (EC Regulation
793/93), report prepared for the Dutch Ministry VROM.
RPA & Statistics Sweden (2002): Business Impact Assessment of EU Chemicals Strategy,
report for DG Enterprise, dated June 2002.
Stevens J et al (2003): PAHs, PCBs, PCNs, Organochlorine Pesticides, Synthetic Musks
and Polychlorinated n-Alkanes in UK Sewage Sludge: Survey Results and
Implications, Environmental Science Technology, 37, 462-467.
US EPA (1998): Chemical Hazard Data Availability Study, EPA’s Office of Pollution
Prevention and Toxics, Washington DC.
White Paper Working Groups (2002): Substances of Very High Concern, report by
Subgroup 1 working on the Chemicals Strategy, dated 20 February 2002.
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Impact of the New Chemicals Policy on Health and Environment
WWF/EEB (2003): A New Chemicals Policy in Europe - New Opportunities for
Industry, discussion paper dated January 2003.
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ANNEX 1:
ALTERNATIVE TESTING REGIMES FOR REACH
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Impact of the New Chemicals Policy on Health and Environment
A.
OPTION I: BASIC INFORMATION REQUIREMENTS
A1.1 Introduction
In addition to the principles laid out more generally for substance tailored testing
under REACH, the Working Group convened by the Commission also listed
principles specific to each of the options. Those underlying the testing regime for
BIR are as follows (TRE/TS01/04/004 REV 1).
•
When it is not technically possible or when it does not appear scientifically
necessary to give information, the reasons shall be clearly stated and be subject to
acceptance by the authorities. Arguments related to exposure can also be
considered when relevant. They should be supported by reliable estimates or
measurements of human or environment exposure.
•
Whereas testing will normally follow the tonnage triggers as described below, the
risk assessment may indicate that further information is required before the
quantities reach the tonnage threshold.
•
Accepted testing strategies designed to provide guidance on the systematic and
stepwise gathering of information should be used as a tool, in combination with
expert judgement, to determine the need for testing. Any testing strategy should be
reconsidered when new data become available, including exposure related data.
Against this background, data requirements for four dossiers – one for each of the
tonnage bands – are set out below.
A1.2 Dossier A for Quantities 1 – 10 t/y
Box A1.1 sets out the information requirements for Dossier A. As can be seen from
Box A1.1, the starting point of the proposal is not far from a VIIB dossier
(67/548/EC) and is consistent with the White Paper which recommends that the gap in
knowledge about the intrinsic properties for existing substances should be closed to
ensure that equivalent information to that on new substances is available (although for
new substances the VIIB is currently connected to a lower tonnage). In vitro methods
are not available for all end-points, requiring that in vivo methods are used until in
vitro methods are either developed or gain regulatory acceptance.
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Box A1.1 : Basic Information Requirements - Dossier A (1 – 10 t/y)
Spectra (UV, IR, NMR or mass spectrum)
Methods of detection and determination, including known analytical methods to allow determination
of human and environmental exposure.
Physico-chemicals properties
Melting point
Boiling point
Relative density
Water solubility
Vapour pressure
Surface tension on a case by case basis
Partition coefficient
Flash point (liquid) or flammability (solid or gas)
Self-ignition temperature
Explosive properties
Oxidizing properties
Granulometry on a case by case basis
Toxicity
Acute toxicity (by oral route for substances other than gas or by inhalation for gas and volatile
substance)
Skin irritation
Eye irritation
Skin sensitisation
Mutagenicity : at least one bacteriological test. A positive test should be followed by further testing
according the accepted strategy
Repeated dose toxicity (a 28-day study when justified by regular and/or frequent exposure, oral route
unless contra-indicated)
Ecotoxicity
Daphnia acute toxicity
Degradation (biotic) including bacterial inhibition
Growth inhibition test on algae if justified by exposure
Some of the data, such as that on explosive properties, oxidizing properties and
granulometry, are only required depending on the structure of the substance or in
particular situations (e.g. for granulometry, for solids of small particle size and when
there is exposure by inhalation).
In addition, other comments are made by the Working Group on information
requirements for this dossier relating to acute toxicity, skin and eye irritation,
sensitisation, mutagenicity and repeated dose toxicity. For acute toxicity and skin and
eye irritation, the importance of using non-animal test methods where possible is
stressed (and of speeding up the acceptance of these); as is the need to give priority to
developing an in vitro method for sensitisation and the possibility of using structureactivity relationships and ‘read across’ to avoid animal testing.
The Working Group also stressed the importance of testing for mutagenicity as part of
this dossier, as it is a toxicological effect that may be exerted at all dose levels.
Information about mutagenicity is considered essential to being able to predict
possible carcinogenic or reproductive effects and prioritising substances. The Ames
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Impact of the New Chemicals Policy on Health and Environment
test is recognised as a widely accepted screening test and it is inexpensive, so should
be included within these dossiers.
Repeated dose toxicity is more difficult. In this case, the approach should be
substance tailored, and the need for information decided on a case by case basis
(although further guidance may be required in the future).
A1.3 Dossier B for Quantities 10 – 100 t/y
Dossier B effectively builds on Dossier A, by adding additional information. Thus,
the comments that applied to Dossier A also apply here. Further comments made by
the Working Group again relate to the need to develop further test methods (in vitro
tests for reprotoxicity screening and a fish egg test or fish cell test for fish acute
toxicity) and the availability of guidance on some end-points in the TGD (i.e. predict
toxicokinetics in the absence of experimental data).
Box A1.2: Basic Information Requirements - Dossier B (10 – 100 t/y)
Spectra (UV, IR, NMR or mass spectrum)
Methods of detection and determination, including known analytical methods to allow determination
of human and environmental exposure
Physico-chemicals properties
Melting point
Boiling point
Relative density
Water solubility
Surface tension on a case by case basis
Partition coefficient
Flash point (liquid) or flammability (solid or gas)
Self-ignition temperature
Explosive properties
Oxidizing properties
Granulometry on a case by case basis
Toxicity
Acute toxicity (by oral route for substances other gas or by inhalation for gas and volatile substances
supplemented by another route when justified by exposure)
Skin irritation
Eye irritation
Skin sensitisation
Repeated dose toxicity (a 28-day study by oral route unless contra-indicated)
Mutagenicity: the substance should be examined in two in vitro tests, one bacteriological and one non
bacteriological. A positive test should be followed by further testing according the accepted strategy.
Reproductive/developmental toxicity screening (may be combined with repeated dose toxicity)
Predicted toxicokinetics
Ecotoxicity
Daphnia acute toxicity
Degradation (biotic) including bacterial inhibition
Growth inhibition test on algae
Fish acute toxicity
Adsorption/desorption screening test
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A1.4 Dossier C for Quantities 100 – 1000 t/y
All of the data considered for Dossier B are also relevant to Dossier C. Box A1.3 sets
out the additional information that should be considered.
Box A1.3: Basic Information Requirements - Dossier C (100 – 1000 t/y)
Toxicity
Toxicokinetic information: in vitro skin absorption, in vitro metabolism studies (will be useful to
decide further testing)
Respiratory sensitisation (as soon as a test is available)
Fertility (a two-generation study, preferably on rats)
Developmental toxicity (one species, preferably rat)
Subchronic toxicity (a 90-day study) an/or chronic toxicity if available data show the need for
following the accepted strategy
Additional mutagenicity tests as prescribed in the accepted testing strategy
Ecotoxicity
Daphnia reproduction test
Test on higher plant
Acute toxicity on earthworms
Further toxicity studies with fish
Test for species accumulation (one species, preferably fish)
Supplementary degradation studies
Further studies on absorption/desorption
A1.5 Dossier D for quantities >1000 t/y
All of the data considered for Dossier C are also relevant to Dossier D, with Box A1.4
listing the additional information that should be considered.
Box A1.4: Basic Information Requirements - Dossier D (> 1000 t/y)
Toxicity
Toxicokinetic study including biotransformation (ADME)
Developmental toxicity on a second species unless clear adverse effects have been shown in the first
developmental study following the accepted strategy
Chronic toxicity/carcinogenicity unless available data show no indication of concern following the
accepted strategy
Ecotoxicity
Additional test for accumulation , degradation, mobility and absorption/desorption
Further toxicity studies with fish
Toxicity studies with birds
Additional toxicity studies with other organisms
A1.6 Exposure-Based Waiving of Tests
Within the overall testing regime, the potential for waiving particular tests on the
basis of exposure is highlighted. The waiving of tests is indicated as only being valid
if production and use result in no or very low emissions, or in only rare or occasional
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Impact of the New Chemicals Policy on Health and Environment
‘short’ emissions to the workplace and would result in negligible consumer and
indirect exposures (with intermediates potentially falling into these categories).
In these circumstances, the required test package would take the form of a reduced or
minimum set of information. These minimum information requirements are listed in
Box A1.5 for each of the tonnage bands.
Box A1.5: Exposure Based Waiving of Tests and Minimum Information
Tonnage Band
Minimum Information Requirements (MIR)
Dossier A less:
- subacute toxicity
1 – 10 t/y
- daphnia acute toxicity
- Growth inhibition test on algae
Dossier B less:
- subacute toxicity
10 – 100 t/y
- reproductive/developmental screening toxicity
- acute fish toxicity
- adsorption/desorption
100 – 1000 t/y
Dossier B
Dossier B and additional information on reprotoxicity
>1000 t/y
A.2
Option II: Minimum Information Requirements
A2.1
Introduction
The Minimum Information Requirements testing regimes are based on the concept of
providing risk-adequate information - structured by tonnage - for each use, exposure
and the already known (inherent) properties of a substance. No differentiation is
made as to whether this information can be split into what has to be delivered during
the registration step or can be provided at a later stage (e.g. after evaluation), or
whether it is provided all in one step. This aspect is a significant variation from
Option I above, where it is assumed that all of the necessary information is provided
as part of registration. In addition, in this case, it is argued that it should be up to the
submitter of the registration dossier whether or not information extra to the
requirements set out below is submitted as part of registration. Again this is a
different approach to that under Option I, which requires all existing information to be
taken into account, particularly in relation to human health effects.
More generally, it is argued that available existing information should be used rather
than generating new data when it is of an acceptable standard; for example, data on
physico-chemical endpoints that have been generated by reliable tests should be
accepted even when these have not been conducted according to GLP. In addition, it
is argued that, where possible, bridging of information, expert judgement, (Q)SAR,
and/or alternative test methods should be used to avoid unnecessary animal testing.
A further principle underlying the regime is based on the White Paper. This is that
testing should only be required if the additional information generated could have a
consequence on risk management measures already in place. So, for example:
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•
•
•
•
when skin contact can be excluded because of other toxic properties which result
in worker safety designed to avoid repeated exposure, there is no need to test for
skin sensitisation;
if carcinogenic properties are already known (classified as a Category 1 or
Category 2) and measures are implemented which exclude exposure, then testing
for reprotoxicity should not be necessary as generally, no more stringent measures
would result. The same principle applies to substances classified as mutagens;
if a property like sensitisation is not known but assumed and the substance is
classified and handled according to this property, then testing of this property
should not be necessary; and
if different substances are handled exclusively together and for one substance a
property like carcinogenicity is already known and measures are implemented
which exclude exposure to all substances, then testing of this property should not
be necessary for each substance separately.
Proponents of this regime also argue that if there is no relevant exposure then there
should be no requirement to provide test data above minimum requirements. Relevant
in this case is interpreted as meaning above limits or thresholds which are already
accepted in existing legislation, with “no relevant exposure” then being below such
limits. The examples given relate to thresholds. If a threshold which has been
defined by an EU-wide or internationally accepted OEL-value is not exceeded, then
there is no need for additional testing; e.g. use of a preparation with a maximum of
0.1 % of a substance: no additional requirements due to this use. If a substance has no
pathway to the environmental compartment, no additional ecological data beyond
basic information is necessary.
It is further proposed that there is no generation of additional data above minimum
requirements if there is no bioavailability in general, or bioavailability in specific
cases. With regard to the first case, an example is given of a substance that has a
molecular weight > 750 g/mol and the diameter of a molecule is > 950 pm (although
it is noted that these are just proposals and that the chemical structure would also have
to be taken into account). In the second case of bioavailability…….?? only be
required in specific cases, it is argued that, for example, no testing for inhalative
bioavailability should be required if: the vapour pressure of the substance is < 0.1 Pa
(20°C), particle size diameter is > 10 µm; no aerosol formation is foreseeable; or (for
exposure indirectly via e.g. wastewater) Henry-constant < 1 Pa m3/mol.
Taken together, these proposals for this Option could result in less testing than carried
out under Option I. However, as noted earlier, in theory, significant differences
should not arise if Option I is appropriately substance tailored.
A2.2 Minimum Information Requirements for 1-10 t/y
Box A2.1 sets out the information required under this Option for the lowest tonnage
band. Based on the arguments put forward above, additional requirements are also
specified where use of the substance includes professional or consumer use and thus
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Impact of the New Chemicals Policy on Health and Environment
there may be greater potential for exposure.
information should be provided as follows:
•
•
•
Where such use occurs, further
for professional use: information about sensitisation potential when there is a
foreseeable frequent exposure (without protective equipment) and for highly
reactive substances;
for consumer use: information about sensitisation potential for highly reactive
substance; and
for consumer and professional use: information about mutagenicity (Ames test) is
proposed.
Box A2.1: Minimum Information Requirements (1 - 10 t/y)
Melting Point
Boiling Point
Relative Density
Vapour Pressure
Partition Coefficient
Water Solubility
Flash Point
Granulometry
Acute Toxicity (1 Route)
Skin Irritation
Eye Irritation
Biodegradation
Acute Ecotoxicity (Preferably Daphnia)
depending on exposure and physical state
A2.3 Minimum Information Requirements for 10-100 t/y
For the next highest tonnage band, information is required for the endpoints set out in
Box A2.2. That information that is in addition to what is required for the lowest
tonnage band is highlighted in bold.
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Box A2.2: Minimum Information Requirements (10 – 100 t/y)
Melting Point
Boiling Point
Relative Density
Vapour Pressure
Partition Coefficient
Water Solubility
Flash Point
Flammability
Explosive Properties
Self-Ignition Temperature
Oxidising Properties
Granulometry
Acute Toxicity (1 Route)
Skin Irritation
Eye Irritiation
Sensitisation
Ames Test
28 Days Repeated Dose Study
Acute Ecotoxicity (1 or 2 Species)
Biodegradation
depending on substance-specific
properties
depending on exposure and physical state
depending on use categories
In this case, if information on exposure is not available (or not given with the
registration file) beyond categorisation into industrial, professional or consumer use,
then the following information should also be provided:
•
•
•
for professional and consumer use: information about sensitisation and
mutagenicity;
for professional and consumer use: additional mammalian toxicity (28 day study)
and a screening study on reprotoxicity; and
for wide dispersive use (professional and consumer use) and in case of pathways
into the environment: a second ecotoxicity test (preferably algae).
It is also argued that the testing strategy needs to be flexible enough that it is the
detailed exposure patterns and inherent properties that determine whether more or less
information is acceptable. Two examples are given in this regard. The first relates to
a case of frequent consumer exposure and a positive Ames test, with this indicating
that a 2nd test on mutagenicity should be undertaken. The second relates to a case of
professional use with effective exposure control information about sensitisation. In
this case, it is argued that mutagenicity, mammalian toxicity (28 day study) and a
screening on reprotoxicity should not generally be required.
A2.4 Minimum Information Requirements for 100-1000 t/y and >1000 t/y
For existing substances produced in tonnages between 100 and 1000 t/y and over
1000 t/y, information requirements would be the same. The full set of information
required is set out in Box A2.3. In comparing these requirements to those for
substances produced between 10 and 100 10/y, the only differences is in relation to
acute ecotoxicity and the potential need for studies across three species.
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Impact of the New Chemicals Policy on Health and Environment
Box A2.3: Minimum Information Requirements (100 – 1000 t/y and >1000 t/y)
Melting Point
Boiling Point
Relative Density
Vapour Pressure
Partition Coefficient
Water Solubility
Flash Point
Flammability
Explosive Properties
Self-Ignition Temperature
Oxidising Properties
Granulometry
Acute Toxicity (1 Route)
Skin Irritation
Eye Irritation
Sensitisation
Ames Test
28 Days Repeated Dose Study
Acute Ecotoxicity (1 - 3 Species)
Biodegradation
depending on substance-specific
properties
depending on exposure and physical state
depending on use categories
Again it is stressed that the strategy should be flexible and relate to exposure patterns
and the inherent properties of a substance. Where information on exposure is not
available beyond categorisation into industrial, professional or consumer use, then the
following information should also be provided:
•
•
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for professional and consumer use: information about mammalian toxicity (28 day
study) and a screening study on reprotoxicity; and
for wide dispersive use (professional and consumer use) and in case of pathways
into the environment: in total 3 ecotoxicity tests (3 species) are proposed.
RPA & BRE
CASE STUDY 1:
NONYLPHENOL (NP)
Case Study 1: Nonylphenols
RPA & BRE
1.
INTRODUCTION
1.1
Background to the Case Study
‘Nonylphenol’ refers to a large number of isomeric compounds of the general formula
C6H4(OH)C9H19. The type and extent of branching of the Nonylphenols (NPs) depend on
the production process and the feedstock used in production. Although many NP isomers
have discrete CAS numbers, the second priority list identifies only two; these were chosen
by NP manufacturers because they are the most representative of the product as they
make it.
NP is used almost exclusively as an intermediate in the production of various NP
derivatives. Releases of NP from these production processes are very low. As a result,
very little NP enters into the environment directly. Rather, the primary source of NP in
the environment is considered to be nonylphenol ethoxylates (NPEs)1, a family of nonionic surfactants which can break down into NP after being released into the environment
during their production, their formulation into various other products, and the use of such
products. NPEs are part of the alkylphenol ethoxylate (APE) group of non-ionic
surfactants and represent some 90% of the APEs used in the EU (by tonnage with
octylphenols and their ethoxylates being the other more common AP/Es).
NP is a priority substance under the Existing Substances Regulation (Council Regulation
No 793/93), with nearly 80,000 tonnes being used in Europe in 1997. The risk assessment
under ESR concluded that use of NP and NPE posed risks to the aquatic environment. In
response, the risk reduction strategy under ESR, recommended that comprehensive phaseouts under the Marketing and Use Directive (76/769/EEC) should be applied to those
industry sectors and uses that contribute most to regional concentrations and/or for which
alternatives to NP and NP ethoxylates (together referred to as NP/E) are known to be
available. This risk reduction strategy has not yet been implemented fully.
NP/Es were chosen as a case study chemical for a number of different reasons. The
principal reason is the examination of how REACH is likely to deal with the complex
problems that NP/Es presented to ESR. These complex problems surround the fact that
the principal source of NP in the environment (94%) is the decomposition of NPE based
products from other uses, that would be covered by a different dossier under REACH.
Other reasons for selecting NPs are as follows:
•
•
1
their use first started about 40 years ago (possibly more) and they have been used in
an extremely wide range of different applications;
the case study highlights the types of damages that could be avoided in relation to
chemicals that, although highlighted as a priority, are not linked to pollution incidents
and, thus, relate to less obvious environmental and health impacts;
NPEs are also referred to as nonoxynol, ethononylphenol, polyoxyethylene nonylphenol ether and
nonylphenoxypoly (ethyleneoxy) ethanol.
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Case Study 1: Nonylphenols
•
•
1.2
ecotoxicological effects formed the initial basis of concern, with possible oestrogenic
effects later becoming an issue; and
their use has been the focus of a wide range of voluntary and regulatory initiatives,
with these including PARCOM and OSAPR initiatives.
Format of Case Study
The case study has been organised as follows:
•
Section 2 presents an overview of the market profile for NP/Es as it stood at the time
of the ESR Risk Assessment;
•
Section 3 sets out an overview of the timescale of environmental and human health
concerns associated with NPs in the environment and the regulatory and voluntary
initiatives that have been adopted in relation to NPs up to their assessment under ESR.
Conclusions concerning environmental and human health damages are presented;
•
Section 4 presents the hypothetical REACH Dossier developed for this case study;
and
•
Section 5 discusses this hypothetical REACH Dossier in the light of the findings of the
ESR Risk Assessment and Risk Reduction Strategy and the historical context of
environmental and human health concerns.
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RPA & BRE
2.
The EU Market Profile
2.1
Uses and Trends
In 1997, nearly 80,000 tonnes of NP were used in the EU, almost exclusively as an
intermediate in the production of other chemicals. Some 60% of this (corresponding to
around 47,000 tonnes) was used to make NPEs, with the remainder used to produce other
NP derivatives.
Table 2.1 provides an indication of the use patterns of NP/Es based on 1997 data (the
most recent for the EU as a whole). As shown, industrial and institutional cleaning was
one of the main users of NP/Es despite voluntary agreements being in place since the
1970s. Other key sectors of use were use as an intermediate within the chemicals
industry, use in emulsion polymerisation, textile and leather processing, use in veterinary
medicines and pesticides within the agricultural sector, and use in paint products.
It is of note that, as illustrated in Table 2.1, some 9% of use was within ‘other niche
markets’ with a significant proportion of this being in personal care products (shampoos,
make-up, etc.). A further 7% of use was unaccounted for at the time of preparing both
the risk assessment and the risk reduction strategy.
Table 2.1: Use of NP/Es at the EU Level
Types of Use
Use at EU Level
Volume (tonnes)
As percentage of EU Use
Industrial and institutional cleaning
23,000
30
Emulsion polymerisation
9,000
12
Textile auxiliaries
8,000
10
Intermediate/Captive use
7,000
9
Leather auxiliaries
6,000
8
Agriculture
5,000
6
Paints
4,000
5
Metal industry
2,000
3
Pulp and Paper
1,000
1
7,000 (12,000)
9
Total known use
72,000
93
Unaccounted for use
5,600
7
Total EU Use
77,600
100
Other niche markets
Source: RPA (2000) and RPA (2001)
2.2
Sectoral Descriptions of Use
A description of how NP/Es are used in these various sectors is provided in Table 2.2, in
descending order of importance in terms of tonnage consumed (based on RPA, 2000).
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Case Study 1: Nonylphenols
Table 2.2: Function of NP/Es on an Industry Basis (in decreasing order of annual NP/E tonnage used)
Industry
Function of NP/Es
Notes
Laundries; floor and surface cleaning in Includes ‘industrial and institutional’ cleaning and
Industrial
buildings; vehicle cleaners; anti-static domestic products; also covers releases from NPE&Institutional
cleaners; metal cleaning
Cleaning
based detergents used in some other sectors (e.g.
electronics/electrical engineering)
Emulsion
Added to acrylic esters used for specialist Act as dispersants and aid the stability of the
formulation; present (see also ‘civil & mechanical
Polymerisation
coatings, adhesives and fibre bonding
engineering’ under ‘other niche markets’)
Used as processing aids in formulation of End applications for polymer dispersions include
emulsion polymers, including polyvinyl paints, paper, inks, adhesives, carpet backings,
textiles and leather finishing
acetates and acrylic acids
Textile Auxiliaries
Captive Use (use
by the chemical
industry in
synthesis of other
chemicals)
Potentially used in polymerisation
reactions to make polymer solutions that
are used for wastewater treatment
Main use is wool scouring (removing
natural oils from wool); also for fibre
lubrication, dye levelling and flocking
(a)
synthesis of nonylphenol ether
sulphates
(b)
synthesis of nonylphenol ether
phosphates
Leather
Auxiliaries
Thought to be used in the wet degreasing
of hides in the leather industry
Agriculture
(a) pesticides
(b) veterinary medicines (principally in
teat dips for treating mastitis; also in sheep
dips)
Paints
Used in the preparation of the paint resin
(polyvinylacetate) and also as a paint
mixture stabiliser
Metal Industry
Metal cleaning processes (iron and steel
manufacture);
steel
phosphating,
electronics cleaning (for metal contacts)
and cleaning of metal products prior to
storage; formulation and usage of cutting
and drilling oils
Felt
conditioner/cleaner
(woollen/
synthetic drying machine that needs
periodic cleaning); defoamers (these are
dripped into the wet end of paper
manufacturing to reduce foaming); wire
cleaner; descaler; system cleaner; retention
aid; mould inhibitor; tissue softener; delignification of wood.
Pulp and Paper
Page 1-4
APEs used in wastewater treatment are thought to
account for 3-4% APE exposure to the
environment in the EU
Being phased out of wool scouring in the UK
(a) used as an emulsifier for styrene and other
monomers (probably low impact), as emulsifier in
agrochemicals, and additive to special types of
concrete
(b) normally used as agrochemicals or in the
emulsion polymerisation process; may also be used
in I&I cleaning products
New information from the leather industry
indicates that almost half of NPE usage attributed
to them is exported for use outside the EU
Used as wetting agents, dispersants and emulsifiers
in pesticides; run-off from the soil surface and
leaching are not significant sources of water
contamination because NP/Es are strongly bound
to soil; teat and sheep dips eventually applied to
land as sewage sludge
Other possible uses of NPE in the coatings industry
include the formulation of inks for laser jet printers
and the formulation of ‘blanket wash’ chemicals
for use with lithographic printers; NPEs used in
water-based paints
Use of detergents for cleaning in the metal
working industry is considered under I&I cleaning
RPA & BRE
Table 2.2 (cont): Function of NP/Es on an Industry Basis (in decreasing order of annual NP/E tonnage
used)
Industry
Function of NPEs
Notes
Other niche markets:
* Civil and
Possible uses include manufacture of wall May also be used as an air-entraining admixture in
Mechanical
construction materials, road surface cement, but this is a small usage; releases from
Engineering
materials, and also in cleaning of metals production of plastics and use of NP-based
etc; may also be in some plastic materials additives related to civil and mechanical
used in construction, particularly if engineering is considered elsewhere in the risk
assessment
produced via emulsion polymerisation
* Electronics/
Used in fluxes in the manufacture of
Electrical
printed circuit boards, in dyes to identify
Engineering
cracks in printed circuit boards and as a
component of chemicals baths used in the
etching of circuit boards
* Mineral oil and Nonylphenol ethoxylate phosphate esters The manufacture and blending of additives
used as additives in lubricating oil (used in packages are thought to be main sources of
Fuel Industry
military
gearboxes);
nonylphenol environmental release for this industry, where the
ethoxylate esters (prevent aggregation of risk assessment indicates that NP/Es are mostly
metal fragments in engine boxes; reduce burnt off during end use
the impact of water contamination); NPE
and NP used in blending of fuel additive In fuels, detergents are used to clean engines
packages; used either in a lubricant or in a internally as a means of meeting vehicle emission
targets
fuel oil
* Photographic
In products intended for home use by Regulations require that commercial photo
Film
amateur photographers; for photo developers do not discharge to sewer; largest users
developers who develop film for amateur of photo chemicals pre-treat their waste, then
photographers; in some professional discharge to sewers; small/medium scale users
products; also reported to be used in x-ray generally have waste removed from the site and
incinerated, although some residue from wash
film
tanks is discharged directly; home hobbyists
discharge to sewer
** Personal Care Cosmetics, spermicides, shampoos, Used as a surfactant in cosmetics
shower gels, shaving foams, etc.
** Public
Non-agricultural pesticides; vehicle and These products were part of the category ‘Public
Domain
office cleaning products; correction fluids, domain’ which was largely, but not entirely,
inks and other office products
replaced by the category ‘I&I’
* Tonnage for these sectors are aggregated as part of ‘other niche markets’ in industry usage data; however, for the
purpose of calculating the sector-specific NP burden, they are treated independently in the risk assessment.
** These are also sub-sectors of ‘other niche markets’, but neither their tonnage nor the associated NP burden is
treated independently.
It is important at this stage to emphasise the distinction between how NP/Es are used as
indicated in Table 2.2. A number of the uses relate to NPs only and their use as a
chemical intermediate. In other case, use relates to the manufacture of other products,
such as use in textile processing, while in others it relates to use in products, such as use in
paints and personal domestic products that are then used by the end consumer.
There is a certain element of overlap between these categorisations. For example, NPEs
are used in production of emulsion polymers (and are present in small amounts in the final
product). These emulsion polymers may then be used in coatings (e.g. paints) that are
applied to textiles.
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Case Study 1: Nonylphenols
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RPA & BRE
3.
ENVIRONMENTAL AND HUMAN HEALTH IMPACTS
3.1
Introduction
Historically, in addition to the uses set out in Table 2.2, a major use of NPE surfactants
was in domestic cleaning agents. This use category was phased out generally by 1995
under the Paris Commission (PARCOM) Recommendation 92/8. A number of individual
Member States took earlier action:
•
•
•
•
•
Sweden took action as early as 1972 when the use of NPEs ceased in household
cleaning products;
the UK secured a voluntary agreement for phase out in domestic cleaning products in
1976;
in 1986, Germany decided to phase out NPEs in domestic products;
Denmark enacted a voluntary agreement with industry in 1987; and
the Netherlands had a voluntary agreement to phase out use in household cleaning
agents by 1988.
Clearly, there was sufficient concern over the impacts of NPE for industry to take
voluntary action to phase out their use as early as 1972 (Sweden). It is thought that this
initial action was taken because of the magnitude of use, the characteristics of use which
would increase transport and exposure (i.e. its use as a surfactant and carriage in water),
and because of its toxicological profile.
NP was not classified in the European Union until 2001. The current classification is
summarised in Box 3.1. This indicates that NP is considered to be very toxic to aquatic
organisms; consideration of its persistence also indicates that it meets EU criteria for both
freshwaters and sediment. It does not meet the criterion for bioaccumulation, however.
More recently, there has been concern over the mounting evidence that NP is an
oestrogenic compound.
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Case Study 1: Nonylphenols
Box 3.1: Classification and Labelling of NP
The classification and labelling of NP is listed in Annex I to Directive 67/548/EEC (28th Adaptation to
Technical Progress; January 2001), as follows:
Classification: Harmful if swallowed
Causes burns
Very toxic to aquatic organisms, may cause long-term adverse effects in the aquatic
environment
Labelling: Causes burns
Very toxic to aquatic organisms, may cause long-term adverse effects in the aquatic
environment
Keep locked up and out of the reach of children
In case of contact with eyes, rinse immediately with plenty of water and
seek medical advice
Wear suitable protective clothing, gloves and eye/face protection
In case of accident or if you feel unwell seek medical advice immediately
(show the label where possible)
This material and its container must be disposed of as hazardous waste
Avoid release to the environment. Refer to special instructions/safety data sheets
3.2
Development of Concerns and Damages
3.2.1 Pre 1970
Before 1970, NPE was widely used as a surfactant in a number of applications, including
use in the domestic cleaning and laundry sector. At this time, few concerns were
expressed in the literature concerning potential problems with its use. The earliest studies
on the effects of NP/Es found in the literature were conducted by Knaak et al (1966) (in
the case of toxicokinetics in rats) and Smyth et al (1969), who exposed a group of six rats
to an unquantified “concentrated vapour” of NP for four hours and found that there were
no deaths. It is unclear whether the motivation for undertaking these studies was as part
of a general testing procedure, or whether the tests were undertaken because there was a
concern over NP/Es at this point in time. The former appears more likely.
3.2.2 1970 - 1979
Concern over the use of NP/Es rose steadily over the period of the 1970s. Box 3.2
provides an overview of significant events and studies in this period. During this decade,
there was growing awareness that NPE formed NP as a decomposition product and that
this occurred in municipal waste waters. In the same period, the effects on marine species
(cod) were investigated. Early awareness and concern over this evidence may have been
one of the factors in the development of the Swedish control measures in 1972 and the UK
voluntary agreement to phase out use of NPEs in domestic cleaning agents in 1976,
although the literature is not clear on this point. An early indication of human skin
sensitisation is reported in Japan in 1979.
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RPA & BRE
Box 3.2: Developments in the 1970s
1970
Swisher (1970) reviewed the biodegradation of surfactants including NPEs and made conclusions
concerning the breakdown of NPE to NP. Swisher’s conclusions are consistent with those that
appear in the 2002 European Risk Assessment.
1971
Swedmark et al (1971) examined the biological effects of surface active agents on marine animals
(cod) and defined a 15 day LC50 of 0.1 mg/l.
1972
Swedish Voluntary Agreement to phase out domestic use.
1976
Gaffney (1976) observed biodegradation of NP in municipal wastewaters which contained NP and
so may have been adapted. No biodegradation was observed in tests with domestic wastewater
which was not adapted.
1979
UK establish a voluntary agreement to phase out the use of NPEs in domestic cleaning products.
Ikeda et al (1979) reported two isolated case reports of leucoderma on the hands and forearms
among Japanese workers exposed to alkaline detergents containing polyethylene alkylphenylether
which the authors speculated might be caused by free octylphenol or NP (which were also found
in the detergents).
3.2.3 1980-1989
Developments over this period are provided chronologically in Box 3.3. There was an
increase in the number of studies on the environmental concentrations of NP/Es, the
transport, toxicokinetics, toxicity and biodegradation of NPE. There was also a steady
increase in evidence that NP/Es could present a problem because of the toxicological
profile of NP and the nature of the main use of NPEs as a cleaning/emulsifying agent (i.e.
used with water, and associated transport issues).
In the mid to late 1980s, concerns in the US grew, Germany began a process of voluntary
withdrawal from certain applications (beginning with domestic and industrial laundry
detergents and cleansers) and Switzerland banned the use of NPEs (and OPEs) in washing
agents and auxiliaries. By the end of the decade, there was evidence of phytotoxic effects.
Box 3.3: Developments in the 1980s
Circa
Ahel et al and co-workers began to report on concentrations of NP in the Glatt river in
1980
Switzerland (the most recent work has shown a significant decrease in surface water
concentrations from the early 1980s).
1981
In Canada, Sundaram & Szeto (1981) examined the degradation of NP in stream and pond under
simulated field conditions
1982
Kravetz et al (1982) looked at the biodegradation of radiolabelled NPnEO (n=9) during
wastewater treatment.
1984
Nethercott & Lawrence (1984) provide evidence that Nonoxynol-6 found in an industrial
waterless hand cleanser induced allergic contact dermatitis on the upper extremities of a uranium
mill maintenance worker.
1985
US OTS asks the public for unpublished information about chemicals including 4-nonylphenol
(para-, ortho-, and mixed) for Chemical Hazard Information Profiles (CHIPs).
1986
Manufacturers and processors of NP ethoxylates entered into a German voluntary agreement to
phase-out the use of alkylphenol ethoxylates (NP and diisobutylphenol ethoxylates) in domestic
laundry detergents and cleansers as well as for detergents used in commercial laundry (by the end
of 1986), and in aerosol-filled cleansers and disinfectant cleansers (from November 1987). They
also agreed to look into possible substitution of NP ethoxylates in industrial uses (wetting agents
and detergents in the textile industry by January 1989 and use in leather and fur, paper, textiles
and industrial cleaners by January 1992 (BUA, 1988).
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Case Study 1: Nonylphenols
Box 3.3 ctd: Developments in the 1980s
1987
In Switzerland the use of octylphenol ethoxylates and NP ethoxylates in washing agents and
washing auxiliary substances was banned in September 1987.
Schaffner et al (1987) reported the concentration of NP in groundwater near the River Glatt,
Switzerland.
1989
The US Chemical Manufacturers Association organizes a panel to assess the environmental
impact of alkylphenols and ethoxylates. The panel includes members from the chemical industry
and planned to work with US EPA in determining how best to proceed with surveys and analyses
of the chemicals’ presence in water at various sites.
Prasad (1989) studied the effects of NP on the macrophtyes Lemna minor L. and Salvinia
molesta. Inhibition of frond production was noticed after 2 days at 0.5 mg/l, 2.5 mg/l and 5 mg/l
NP and photosynthetic activity was curtailed after 4 days. Reductions in growth were observed in
lower concentrations of NP (0.125-0.5 mg/l) and bleaching, chlorosis and mortality observed at
NP concentrations of 0.5-2.5 mg/l.
3.2.4 1990 – 2001
The 1990s is the period where concerns were heightened, partly by further toxicological
studies, but also because of the possible link made with oestrogenic effects.
In 1992, PARCOM Recommendation 92/8 required signatory countries to achieve a phase
out of NPEs in domestic detergents by 1995 and in all detergent applications by 2000. NP
was included on the EU Second Priority List under the EU’s Existing Substances
Regulation (793/93/EEC) by the middle of the decade and production of an EU Risk
Assessment on NP under ESR began.
The US began its considerations of NP/E under the US Toxic Substances Control Act
(TSCA) and published its RM-1 document for para-NP in 1996. It concluded that “risks
do not appear widespread, but that there are some impacted areas where aquatic areas
could be affected”. This conclusion differed from that of the EU risk assessment process
which, even in the early drafts of 1996, identified concerns for a number of use categories.
The process of developing an EU risk reduction strategy was begun on this basis.
Growing evidence and concern over the presence and possible toxic and oestrogenic
effects of NP/Es resulted in Sweden further tightening limit values on concentrations in
sewage sludges applied to agricultural soils and in Denmark implementing limits. NP/E
was included on a list of 15 chemicals and groups of chemicals for control through
OSPAR in October 1997.
By the end of this period, the Final EU Risk Assessment had been published (EC, 2002)
and the EU risk reduction strategy (RPA, 2001) was submitted by the UK to the
Commission. Given the lack of conclusive proof of oestrogenic effects at concentrations
below ‘conventional’ toxic thresholds, the identified risks and control measures were
based on the toxicological profile of NP derived from standard ecotoxicity tests. This
called for widespread controls on the marketing and use of the substance in the EU and
these measures are still being taken forward by the Commission. In the US, conversely,
the use of different values to derive No Observed Effect Levels (NOELs) in the RM-1
Page 1-10
RPA & BRE
report have meant that NP/Es are not (currently) regarded with the same level of priority
as in the EU.
Box 3.4: Developments in the 1990s
1990
A well-conducted in vitro mammalian cell gene mutation test proved negative (Hüls, 1990).
1991
The first evidence of oestrogenic activity was reported by Soto et al (1991) who reported that the
release of the oestrogenic antioxidant p-nonylphenol from the polystyrene centrifuge tubes had
induced both cell proliferation and progesterone receptors in human oestrogen sensitive MCF
breast tumour cells and also triggered mitotic activity in rat endometrium. The authors noted that,
not only may this lead to spurious results, but that these compounds may be potentially harmful to
the reproductive function of exposed humans and to the general environment.
1992
Germany found that the target of a complete phase out in the area of washing and cleaning agents
by 1992 was not achieved.
PARCOM Recommendation 92/8 required signatory countries to achieve the phase out of NPEs
in domestic detergents by 1995 and in all detergent applications by 2000.
1993
1994
1995
Initiative Umweltrelevante Altstoffe study on rats (1992) provides first mammalian evidence that
NP exposure over several generations can cause minor perturbations in the reproductive system of
offspring, namely slight changes in the oestrous cycle length, the timing of vaginal opening and
possibly also in ovarian weight and sperm/spermatid count, although functional changes in
reproduction were not induced at the dose levels tested.
Jobling & Sumpter (1993) tested oestrogenic activity of NP/Es in rainbow trout (Oncorhynchus
mykiss) using vitellogenin response (a yolk protein normally produced in response to oestrogen in
female trout). They found changes (albeit slight) in the oestrous cycle length, timing of vaginal
opening, ovarian weight and sperm/spermatid count. The effects on the oestrous cycle were seen
in both the F generations and the timing of vaginal opening was influenced in all three generations.
White et al (1994) reported that NP can stimulate vitellogenin secretion, in vitro, at concentrations
of 0.2 mg/l and above in hepatocytes from rainbow trout (Oncorhynchus mykiss). The authors
also found that NP showed competitive displacement of oestrogen from its receptor site in
rainbow trout (Oncorhynchus mykiss).
Purdom et al (1994) placed cages containing rainbow trout in the effluent from sewage-treatment
works. The study showed that plasma vitellogenin concentrations rose rapidly and very markedly
(500 to 100,000-fold, depending on site) when trout were maintained in the effluent. The identity
of the oestrogenic substance was unknown but the authors hypothesised that one of the two most
likely possibilities was alkylphenol-ethoxylates (APE), originating from the biodegradation of
surfactants and detergents during sewage treatment.
NP and ethoxylates included on EU Second Priority List under Directive 793/93/EEC.
In Sweden use of NPEs in cleaning agents was found to have reduced by 70-80% during the
period 1990-1995 as a result of both administrative actions and voluntary actions from industry.
The “Bund-/Länderausschuß für Umweltchemikalien” (BLAU, 1995) reviewed the available
information on NP concentrations in the environment in Germany. NP concentrations in sludge
from domestic and industrial wastewater treatment plants in Brandenburg (Eastern Germany) were
determined between October 1993-May 1994. The concentrations were in the range of <1 to 214
mg/kg dry weight in domestic wastewater treatment plants and in the range of <1 to 39 mg/kg dry
weight in industrial wastewater treatment plants.
Harries et al (1995) found elevated levels of blood vitellogenin in rainbow trout (Oncorhynchus
mykiss) exposed in vivo to NP for 3 weeks. Levels of blood vitellogenin were found to be
significantly elevated at concentrations of 20.3 µg/l (a ten-fold increase over controls) and 54.3
µg/l (a 1,000 fold increase over controls).
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Case Study 1: Nonylphenols
Box 3.4: Developments in the 1990s
Concern about the suggested association between environmental oestrogens, endocrine disrupters
& declining human reproductive health prompts the UK Government to review current
knowledge.
1996
Following a recommendation under the US Toxic Substances Control Act (TSCA) that NPEs be
tested, the US EPA invokes the TSCA Section 8(a) Preliminary Assessment Information Rule
(PAIR) and the TSCA Section 8(d) Health and Safety Reporting Rule requiring manufacturers
and importers to report production, use and exposure-related information and any unpublished
health and safety data.
USEPA publishes RM-1 document for NP which concludes that “risks do not appear widespread,
but that there are some impacted areas where aquatic areas could be affected”.
First draft of the EU Risk Assessment was completed, concluding that there was a need for risk
reduction in a number of applications. The process of producing a risk reduction strategy began
and identified a number of (low volume) uses not previously identified in the Risk Assessment.
Ahel et al (1996) studied the infiltration of nonylphenolic compounds from river water to
groundwater in the Glatt River region of Switzerland and found evidence that breakthrough of NP
into the aquifer from river water may occur.
Jobling et al (1996) exposed two-year-old male rainbow trout (Oncorhynchus mykiss) to NP in a
flow through system for 3 weeks. Histological examination of the testes showed that fish exposed
to NP had a significantly higher proportion of spermatogonia type A than controls. A significant
stimulation of blood vitellogenin levels was also seen later in the fish development.
Colerangle & Roy (1996) studied the influence of NP on growth and cell proliferation and of the
mammary gland in rats of the Nobel strain (particularly sensitive to oestrogenic activity) using
non-standard methods. This study suggested that NP at dose levels of 0.05 and 35.6 mg/kg/day
increases growth and proliferation activity in a dose-related manner in the mammary gland (Odum,
1999, duplicated the study more robustly and failed to confirm the observation of such activity
following NP exposure at similar dose levels).
In 1996/97, the British Association for Cleaning Specialities (BACS) and the Soap and Detergent
Industry Association (SDIA) reached a voluntary agreement to remove all alkylphenol ethoxylates
from industrial and institutional detergent in 1998. The agreement did not cover solvent
degreasers.
1997
Sweden was expected to call for a ban at the Paris Commission meeting.
An initial priority list of 15 chemicals and groups of chemicals for control of pollution in the North
Sea agreed at OSPAR in October 1997. Substances on the list include NPEs & related
substances.
In Denmark limit values for NP in sludge to be applied to farmland were set. From 1 July 1997
the limit value for NP and NP ethoxylates (with 1 or 2 ethoxylate groups) in soil was 50 mg/kg dw
(this limit value was due to be reduced on the 30 June 2000 to 10 mg/kg dw).
In Sweden, the recommended limit value for NP in sludge for agricultural use was reduced to 50
mg/kg dw (from 100 mg/kg dw) in 1997.
Based on voluntary commitments, the use of alkylphenol ethoxylates in detergents and cleaning
agents in Germany was found to have reduced by about 85% from 1986 to 1997.
Baldwin et al, (1997) find that NP is capable of significantly perturbing components of androgen
metabolism in daphnids at concentrations of ≤25 µg/l. The reproductive chronic value derived
was 71µg/l and this concentration was estimated to reduce the elimination of testosterone by
approximately 50%. The results indicate that NP can cause effects on steroid hormone
metabolism that may contribute to its reproductive toxicity.
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RPA & BRE
Box 3.4: Developments in the 1990s
Shurin & Dodson (1997) The daily production of deformed live offspring/adult was found to be
related to NP exposure in Daphnia galeata mandotae as a clear dose-response curve. The
deformities were seen in 11% of live young at a NP concentration of 10 µg/l, and only animals that
were prenatally exposed to NP exhibited deformity.
1998
Final Draft of the EU Risk Assessment and Stage 4 Risk Reduction Strategy published (the final
Risk Reduction Strategy is published in 2000 and the Final Risk Assessment is published in
2002).
UK Environment Agency summarised a review of current scientific evidence on endocrine
system-disrupting chemicals and proposed a priority list of substances including octyl-and NP and
their ethoxylates.
The National Institute of Environmental Health Sciences in the US concluded studies on the effect
of NP on the reproductive systems of rats and found that although sperm density and numbers
decline at doses of 650 ppm and above, fertility was not affected.
3.3
Key Properties and Presence in the Environment
3.3.1 Introduction
According to the most recent classification, NP/Es are “very toxic to aquatic organisms
and may cause long-term adverse effects in the aquatic environment”. While there is an
abundance of toxicological and kinetics data that suggest NP/E has the potential to cause
environmental damage, there are no examples in the literature that directly implicate the
presence of NP/E as the cause of any observed damages.
This does not necessarily mean that NP/E has not been responsible for any environmental
damages. Rather there are, as yet, no studies or examples of observed damages to, for
example, freshwater or marine ecosystems that have as yet been able to isolate NP/E from
the range of pollutants and pressures that act to inhibit ecosystem function, thus
demonstrating a causal link to damages. It is therefore not possible to draw on observed
effects in the field for the purposes of this case.
However, whilst there is a paucity of data on actual observed impacts, there are reliable
documented measurements of levels of NP/E in the various media which permit some
assessment of the likely scope, extent and duration of impacts.
3.3.2 Human Health Concerns
There is limited human epidemiological data on the known hazardous properties of NPs in
man. Most of the information on the toxicokinetics of NP is based on a small number of
limited rat and human studies, supported by data relating to other phenols with close
structural relationship to NPs.
Studies have shown NP to be moderately toxic by the oral route. On ingestion, NPs are
rapidly and extensively absorbed from the gastrointestinal tract, from where they distribute
throughout the whole body, with the highest concentrations in fatty tissues. There is,
however, insufficient data to make specific conclusions on the bioaccumulative potential
Page 1-13
Case Study 1: Nonylphenols
of NPs. NPs are poorly absorbed across the skin, although liquid NP can be corrosive to
the skin, depending on its potency (which might vary according to source and
composition). NP liquid is also a severe eye irritant (EC, 2002).
The main toxic effects associated with exposure are thus: acute toxicity, corrosivity,
repeated dose toxicity and effects on the reproductive system. Current information on the
carcinogenic potential of NPs indicate that they are unlikely to be mutagenic and as such,
the cancer risks are quite low (EC, 2002). NP has been shown to have oestrogenic activity
in a number of in vitro and in vivo assays, although studies are inconclusive.
The main occupational health concerns, are associated with manufacture of NPs, the use
of NPs as an intermediate and the use of speciality paints. Apart from the use of speciality
paints, NP is always likely to be processed in ‘closed’ plant, so that exposure is likely to
only occur when the plant is breached (EC, 2002).
3.3.3 Presence in the Environment
As indicated in Boxes 3.2 to 3.4, a number of studies have examined the presence of NP/E
in various media, effluents and wastes. Table 3.1 provides a summary table of measured
levels in these various media, including data on:
•
•
•
•
•
•
freshwaters;
freshwater sediments;
marine/estuarine waters;
effluent/sludges from industrial and domestic waste water treatment plants (WWTP);
groundwaters; and
agricultural soils.
As Table 3.1 shows, the data cover a relatively long time period beginning in the early
1980s up until the late 1990s, and show a wide variation in measured levels depending on
the media and the location of the measurement. It should be noted that the values
included in the table are taken from a range of different studies. In some cases they
represent large numbers of samples, in other cases they form studies with single samples or
small numbers of samples. Hence, they do not all have the same significance in terms of
environmental occurrence.
In relation to ecological effects of such measured levels and what they indicate about the
magnitude of environmental damages generally, the risk assessment defines the predicted
no effect concentration (PNEC), which is based on standard toxicity data on no observed
effect concentrations (NOECs) with the application of a safety factor.
To aid interpretation of the data, Table 3.1 also provides the ratios of measured levels to
values derived from the risk assessment and elsewhere, with those values exceeding the
PNEC marked in bold. The following values have been defined to provide the distinction
between effect and no effect:
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RPA & BRE
•
•
•
•
•
•
PNEC water (algae) = 0.33 µg/l;
WWTP = 3.3 µg/l (based on a standard dilution factor of 10)2;
PNEC Soil (reproduction earthworms) = 0.3 mg/kg;
PNEC Oral (secondary poisoning) = 10 mg/kg (food);
PNEC Sediment (derived from water in the RA) = 0.039 mg/kg; and
Sludges = 50 mg/kg (based on Swedish and Danish limit values).
A simple comparison of the number of measurements exceeding the predicted no effect
concentrations/limit values indicates the following:
2
•
Freshwaters: 52% of the values in the Table exceed the PNEC value, with the level
of exceedance varying from 1.2 to 1,091 times the PNEC value. It should be noted
that both ends of this range were measured in the same river in the same year, with
only the location varying. As might be expected, given that sources of NP/E in
freshwaters are, in the main, from point sources, this demonstrates that the location of
measurement is an important factor in the levels found and, therefore, that some
stretches are likely to be more affected by elevated concentrations than others;
•
Marine/Brackish Waters: 87% of the values in the Table exceed the PNEC, with
values in the range of 1.5 to 10.3 times the predicted no effect concentration. As with
freshwaters, the values at the extremes of this range relate to the same estuary, in the
same year, but in different concentrations, again highlighting the importance of the
location of measurement in estimating environmental damages;
•
River and Lake Sediments: 86% of the values in the Table exceed the predicted no
effect concentration by factors of between 1.3 and 191;
•
Tissue: There are no values in the Table above the PNEC for secondary poisoning in
the risk assessment. There are, however, few field measurements on which to base an
assessment of damages via this route;
•
WWTP Effluents: based on a dilution factor of 10, 58% of the values for treated
effluents from WWTP show levels likely to result in concentrations in freshwaters
above the predicted no effect concentration. If the dilution factor is increased to 100,
25% of the values are still likely to result in ecologically significant levels in receiving
waters;
•
WWTP Sludges: 82% of the values for NP/E in sludges are greater than the 50mg/kg
limits for use on agricultural land set down by Sweden and Denmark. Levels of
exceedance of this value range between 1.5 and 30 times; and
•
Soils and Sludge Amended Soils: There are no observations of agricultural soil
levels above the PNEC for soil set out in the RA.
This value is the concentration in effluent from waste water treatment plants (WWTPs) which if diluted in
surface water by a factor of 10 would equal the PNEC for surface water. It is not the PNEC for effects on
micro-organisms in WWTPs (which is 9.5 mg/l in the risk assessment).
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Case Study 1: Nonylphenols
Table 3.1: Measured Levels of NP in Different Media
Measured in
Fresh Waters
River
Receiving WWTP
River
River
River
River
River, Industrial
River
River
River
River
River (background)
River (Downstream of
WWTP)
River
River
Lake (back ground)
Lake (1km from operation)
River
River
River
River
River
River
River
River
River
River
River
River
River Abstraction Point
River
Marine/Brackish Waters
Estuarine
Estuarine
Estuarine
Estuarine
Estuarine
Estuarine
Estuarine
Estuarine
River and Lake Sediments
Lagoon
Lake
River
River
River
River, Suspended
Lake (1km from WWTP)
Tissue
Duck
Fish
Effluents
Secondary
Page 1-16
Average or Median
Value
Observed:
PNEC/Limit
Where
Year
Ref.
Glatt River, Switzerland
WWTP
Glatt River, Switzerland
Glatt River, Switzerland
Glatt River, Switzerland
Glatt River, Switzerland
River Sava, Croatia
River Main
River Main
River Main
Hessian Rivers
Bavarian Rivers
1981
1985
1985
1985
1987
1997
1991
1989
1990
1991
1995
1997
1.80
3
2.70
22.65
4.10
0.18
0.55
0.04
0.05
0.12
bd
0.05
µg/l
µg/l
µg/l
µg/l
µg/l
µg/l
µg/l
µg/l
µg/l
µg/l
µg/l
µg/l
5.5
9.1
8.2
68.6
12.4
0.5
1.7
0.1
0.2
0.4
0.1
1
3
2
2
21
16
4
10
10
10
14
26
Bavarian Rivers
1997
0.25
µg/l
0.8
26
Thur River
Thine River
Eastern Finnish Lake
Eastern Finnish Lake
River Aire, UK
River Dart, UK
River Thames, UK
River Aire, UK
River Aire, UK
River Aire, UK
River Aire, UK
River Aire, UK
River Thames, UK
River Lea, UK
River Wye, UK
River Ouse, UK
River Arun, UK
Representative Rivers USA
1996
1996
1995
1995
1995
1995
1995
1995
1995
1995
1995
1995
1995
1995
1995
1992
0.17
0.04
0.01
0.45
90.80
0.03
0.23
0.20
0.40
4.80
360
4.30
1.55
6.25
1.45
2.95
0.20
0.12
µg/l
µg/l
µg/l
µg/l
µg/l
µg/l
µg/l
µg/l
µg/l
µg/l
µg/l
µg/l
µg/l
µg/l
µg/l
µg/l
µg/l
µg/l
0.5
0.1
0.0
1.4
275.2
0.1
0.7
0.6
1.2
14.5
1,090.9
13.0
4.7
18.9
4.4
8.9
0.6
0.4
10
10
22
22
9
9
11
23
23
23
23
23
9
9
9
9
9
19
Tees Estuary, UK
River Tees, UK
River Tees, UK
River Tees, UK
River Tees, UK
River Tees, UK
River Tees, UK
Mersey, UK
1995
1995
1995
1995
1995
1995
1995
1995
2.65
3.40
0.90
0.50
0.50
0.80
3.40
0.20
µg/l
µg/l
µg/l
µg/l
µg/l
µg/l
µg/l
µg/l
8.0
10.3
2.7
1.5
1.5
2.4
10.3
0.6
9
23
23
23
23
23
23
9
Venice, Lagoon
Lake Constance
River Main
River Main
Glatt River, Switzerland
Main and Hessian Rivers
Eastern Finnish Lake
1990
1991
1991
1991
1994
1994
1996
0.01
0.05
7.44
6.90
3.06
0.80
0.54
mg/kg
mg/kg
mg/kg
mg/kg
mg/kg
mg/kg
mg/kg
0.4
1.3
190.8
177.0
78.5
20.5
13.7
17
8
10
10
6
10
22
Glatt River, Switzerland
Glatt River, Switzerland
1993
1993
0.60
0.80
mg/kg dwt
mg/kg dwt
0.1
0.1
5
5
WWTP
1981
40
µg/l
12.1
1
RPA & BRE
Table 3.1: Measured Levels of NP in Different Media
Measured in
Where
Secondary
WWTP
Effluent
WWTP
Anaerobic
WWTP
Effluent Discharge
Glatt River, Switzerland
Primary
WWTP
Secondary
WWTP
Effluent
WWTP, Industrial, Hessian
Effluent
WWTP, Industrial
Effluent
WYYP, River Dart, UK
Effluent to Sea
Tanker Washing Effluent, UK
Primary
WWTP, Domestic
Secondary
WWTP, Domestic
Effluent
Light vehicle washing
Effluent
Heavy vehicle washing
Sludges from Waste Water Treatment Plants
Anaerobic
WWTP
Activated
WWTP
Mixed Prim and Sec
WWTP
Activated
WWTP
Aerobic
WWTP
Anaerobic
WWTP
Activated
WWTP
Aerobic
WWTP
Anaerobic
WWTP
Digested
WWTP
Raw
WWTP
Sludge
WWTP, Domestic, Hessian
WWTP, Domestic, Brandenburg
Sludge
(Eastern Germany)
WWTP, Industrial, Brandenburg
Sludge
(Eastern Germany)
Sludge
Agricultural Grade Sludge
Aerobic
WWTP
Anaerobic
WWTP
Soils and Sludge Amended Soils
No Sludge
Soil
Sludge Amended, Day 322 Soil
Sludge Amended, Initial
Soil
Sludge Amended, Initial
Soil
1:
2:
3:
4:
5:
6:
7:
Ahel et al (1981)
Ahel et al (1985)
Ahel and Giger (1985)
Ahel et al (1991)
Ahel et al (1993)
Ahel et al (1994)
Ahel (1996)
8: After EU RA - 2002
9: Blackburn and Waldock (1995)
10: BLAU (1995)
11: Britnell (1995)
12: Brunner et al (1988)
13: Danish EPA
14: Fooken et al (1995)
Year
1981
1985
1985
1985
1988
1988
1995
1995
1995
1995
1995
1995
1996
1996
Average or Median
Value
26 µg/l
5 µg/l
467 µg/l
3 µg/l
15 µg/l
2.70 µg/l
1.70 µg/l
330 µg/l
1.20 µg/l
27 µg/l
6.70 µg/l
1.55 µg/l
600 µg/l
430 µg/l
Observed:
PNEC/Limit
7.9
1.5
141.5
0.9
4.5
0.8
0.5
100.0
0.4
8.2
2.0
0.5
181.8
130.3
Ref.
1
3
3
2
12
12
14
9
9
9
9
9
20
20
1984
1984
1984
1985
1985
1985
1988
1988
1988
1988
1988
1995
1010
120
90
128
280
1000
74
385
1550
1500
190
25
mg/kg
mg/kg
mg/kg
mg/kg
mg/kg
mg/kg
mg/kg
mg/kg
mg/kg
mg/kg
mg/kg
mg/kg
20.2
2.4
1.8
2.6
5.6
20.0
1.5
7.7
31.0
30.0
3.8
0.5
15
15
15
3
15
15
12
12
12
12
12
14
1995
107
mg/kg
2.1
10
1995
19.50
mg/kg
0.4
10
1995
1996
1996
0.03
88
705
mg/kg
mg/kg
mg/kg
0.0
1.8
14.1
13
25
25
1992
1992
bd
0.46
1
4.70
mg/kg
mg/kg
mg/kg
mg/kg
0.0
0.0
0.1
18
18
13
18
1992
15: Giger et al (1984)
16: Giger et al (1997)
17: Marcomini et al (1990)
18: Maromini et al (1992)
19: Naylor et al, 1992a; Radian
Corporation, 1990
20: Praxéus (1996)
21: Schaffner et al (1987)
22: Suoanttila (1996)
23: UK EA (1995)
24: Warhurst (1995)
25: Williams and Varineau
(1996)
26: Zellner and Kalbfus
(1997)
In order to gauge the level of environmental damage that has been (and is being) caused
by elevated levels of NP/E in the EU, it is necessary to consider the levels of exposure and
extent of contamination of media. The measurements provided in Table 3.1, however,
only provide a list of the values that can be found in the literature and, within this, values
from locations where one might expect to be able to find higher levels of NP/Es. For
these reasons, the sample of observations is unlikely to be representative. However,
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Case Study 1: Nonylphenols
extrapolating the percentages reported above could suggest that, for example, 52% of
freshwaters in the EU have levels over the predicted no effect concentration. As noted,
this is likely to be an overexaggeration of the situation.
In the EU, there has been no study comparable to that undertaken in the US where
samples were taken from a cohort of 30 rivers that were representative of the range of
rivers in the US. Comparing the measurements found in this representative sample with
the EU PNECs indicates that 7% of rivers in the US had mean measured levels higher than
the EU PNEC, 17% had calculated harmonic mean concentrations higher than the PNEC
and 70% of rivers would exceed the PNEC under low flow conditions.
If a broad assumption is made that the structure and pressures operating in US rivers is
similar to European continental rivers it can be tentatively be suggested that 25% of EU
rivers have levels of NP/E that are regularly in excess of the predicted no effect
concentration and that 70% could have levels which exceed the predicted no effect
concentrations under low flow conditions.
Based on the data in Table 3.1, and assuming that samples are more representative of
WWTPs in the EU, it can be estimated very tentatively that, historically, 25% to 58% of
sewage treatment plants have discharged ecologically significant levels of NP/E to
receiving waters.
3.3.4 Key Properties
Degradation and Persistence
The data available indicate that NP undergoes biodegradation in water, sediment and soil
systems. The results from standard biodegradation tests are variable but indicate that NP
is probably inherently biodegradable. In terms of this variability, the risk assessment
reports a possible explanation for some of the inconsistencies found in the various tests as
being due to the toxicity of NP to micro-organisms at the concentrations used in some of
the tests. A second factor that seems to be important in the biodegradation of NP is that
microorganisms need a period of adaptation. Another factor is that the NP supplied is a
mixture of compounds with differing degrees of branching/isomers in the nonyl chain and
it may be expected that some of the components of the NP mixture would degrade faster
than others.
Although NP is probably inherently biodegradable, it is not considered to biodegrade
readily. Significant biodegradation was seen in ready biodegradability tests when adapted
micro-organisms were used. A rate constant equivalent to a half-life for biodegradation in
surface waters of 150 days has been established. The estimated half life of mineralization
in soil used in the risk assessment is 300 days and 3,000 days in sediment (estimated by the
methods in the TGD). The rate constants and half-lives estimated for NP in surface water,
sediment and soil are thought to be representative of a realistic worst case for
mineralisation of NP. In some situations, particularly where well-adapted microorganisms are present, the actual half-life for NP in surface water and soil may be less than
these values. In contrast, in other situations the actual half-lives could be longer than
estimated here, given that the overall degradation rate of NP in the environment will
Page 1-18
RPA & BRE
depend on the factors mentioned above, such as the possibility of minor amounts of more
persistent NP isomers being present in the product or the absence of suitably adapted
micro-organisms.
Based on these data, and combined with exposure data, it is clear that the continued
presence of NP/E in the media set out in Table 3.1 is variable but, potentially, fairly longterm depending on the media. Using the observations in Table 3.1, the maximum
measured exceedances and assuming limited exchange between media and immediate
cessation of inputs, the following timescales provide an indication of the persistence of NP
in the environment at ecologically significant levels:
•
Freshwaters: applying the half life for mineralization of 150 days, degradation to
levels below the predicted no effect concentration could take up to five years;
•
Marine/Brackish Waters: applying the half life for mineralization of 150 days,
degradation to levels below the predicted no effect concentration could take one to
two years; and
•
River and Lake Sediments: applying the half life for mineralization of 3,000 days,
degradation to levels below the predicted no effect concentration could take seven to
eight years.
In terms of the persistence in organisms and the potential to bioconcentrate, it is clear
from the available data that NP bioconcentrates to a significant extent in aquatic species,
with BCFs (on a fresh weight basis) of up to 1,300 in fish. However, this value may
overestimate the BCF as it would include any metabolites of NP as well; more reliable
values with a mean of 741 have been measured, which are of a similar order of magnitude.
Bioconcentration factors of around 2,000-3,000 have been measured in mussels. The
BCF calculated from the log Kow of 4.48, using the TGD guidance is 1,280, which agrees
well with the measured values. This calculated value of 1,280 was used in the risk
assessment.
Whilst NP has the potential to bioconcentrate, biomagnification is not expected to occur.
Nevertheless, historical (and existing) use of NP/Es is likely to have resulted in elevated
and ecologically significant levels of NP in the environment (in particularly the aquatic
environment) in a significant proportion of rivers and these concentrations can be expected
to stay at significant levels for a number of years after the (eventual) cessation in the use of
the substance.
3.3.5 Environmental Concerns
There are no direct observations of effects in the environment which can be directly
attributed to NP. Purdom et al (1996) did identify NP/Es as one of the two likely causes
of effects seen on fish in the effluent from waste water treatment plants, but were not able
to say definitively that this was the cause. The assessment of effects therefore has to rely
on the results of laboratory tests and the use of safety factors. Based on this, a
comparison with measured levels shows that effects in the environment would be
expected. Thus, it is concluded that the elevated levels are or have been responsible for
Page 1-19
Case Study 1: Nonylphenols
adverse effects on ecosystem functioning and health because of both direct toxicity to
populations of organisms and indirect effects because of reductions and variations in the
size and stability of populations at different trophic levels (with subsequent changes to the
nature of ecosystems).
Thus the environmental damages associated with NP/E can be summarised as follows:
•
historical and existing use of NP/Es has resulted in ecologically significant levels,
particularly in the aquatic environment, although it remains unclear what the resultant
effects on ecosystem function and health are likely to be;
•
if a broad assumption is made that the structure of, and pressures operating in, US
rivers are similar to European continental rivers it can be tentatively suggested that
25% of EU rivers have levels of NP/E that are regularly in excess of the predicted no
effect concentration and 70% have levels which exceed the predicted no effect
concentrations under low flow conditions. Measured levels suggest an absolute
maximum of 1,091 times the predicted no effect concentration, but a more realistic
figure may be in the range of 10 to 100 times the predicted no effect concentration
(based on measured levels largely from the mid 1980s to the mid 1990s);
•
if there were zero emissions, recovery of affected aquatic systems (including the
sediment) could take as long as eight years, but possibly more or less depending on
individual situation and historical contamination;
•
contamination of marine waters and sediments is likely to be less widespread but with
similar recovery times in severely affected areas;
•
it appears likely that, historically, 25% to 58% of sewage treatment plants have been
releasing ecologically significant levels of NP/Es; and
•
there will have been some contamination of soils through the application of sewage
sludges.
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RPA & BRE
4.
THE REACH DOSSIER
4.1
Basic Assumptions
As noted earlier, NP is produced and used in quantities of 50,000 – 100,000 tonnes in the
EU. In addition, there are only four manufacturers within the EU. For the purposes of
this case study, it has been assumed that the four manufacturers have joined together and
formed a consortium in order to prepare the REACH dossier. The dossier must fulfil the
requirements for a substance produced in excess of 1,000 t/y.
In developing this hypothetical dossier, the following assumptions have been made:
•
the data that were available in IUCLID submitted to the European Chemicals Bureau
are assumed to have been available to the manufacturers at the start of dossier
preparation;
•
any further substance tailored testing must be undertaken in line with the requirements
set out for Dossier D (See Annex 1);
•
where site specific release data are not available, default data from the TGD and
within the EUSES model is applied; exposures for workers and consumers are
estimated using specific data or the methods in the TGD (including the EASE model);
and
•
EUSES provides the basis for reaching conclusions as to whether or not unacceptable
risks result from a particular application or sector.
The remainder of this section sets out the key results for the hypothetical dossier compiled
on NPs. No details of the underlying studies are included (see the full ESR risk
assessment for further information on the underlying studies).
4.2
Base Data
4.2.1 Identity of Substance
The information presented in this dossier represents the substance produced and sold
commercially as NP. There are a number of possible CAS numbers which relate to this
substance. The main one is 84852-15-3, which has the substance name 4-NP (branched).
Other numbers relate to more specific forms of the nonyl group, and components of these
may be present in the commercial product.
Methods of detection are available for water, soil, sediments and biota. Detection limits
vary. For water, the lowest in recent studies is 1 ng/l, but older studies more commonly
have limits of 0.1 µg/l or sometimes higher. In soil, levels down to 0.003 mg/kg have
been reported.
Page 1-21
Case Study 1: Nonylphenols
4.2.2 Physico-chemical Data
The basic physico-chemical data are presented in Table 4.1.
Table 4.1: Physico-chemical Data
Physical state (at ntp):
Clear to pale yellow viscous liquid with a slight phenolic odour
Melting point:
circa -8°C (may vary according to production process)
Relative density:
0.95 at 20°C
Vapour pressure:
circa 0.3 Pa at 25°C
Octanol-water partition
4.48 (Log Kow)
coefficient:
Water solubility:
6 mg/l at 20°C (may be pH dependant)
Flash point:
141-155°C
Flammability:
circa 370°C
Oxidising properties:
Not applicable
Viscosity:
2,500 mPa s at 20°C
4.2.3 Ecotoxicity Data
Aquatic Toxicity
Table 4.2 below summarises the lowest valid results for each species among the data set3.
Note that almost all of these data were included in the IUCLID dossier available at the
start of the ESR risk assessment. Given the availability of these data, it is assumed here
that no further testing is required with regard to aquatic toxicity. Indeed, there are a
number of other test results for fish, invertebrates and algae which support these values,
while showing effects at higher concentrations. A PNEC value of 0.33µg/l is estimated
from the available data.
In addition, test data indicated that NP was weakly estrogenic in an in vitro test with fish
hepatocytes. (Note that this was the only comment in the environmental part of the
IUCLID on this aspect.)
Terrestrial Toxicity
The data that are assumed to be readily available for terrestrial plants and terrestrial
invertebrates are summarised in Tables 4.3 and 4.4.
None of these data were included in the IUCLID dossier, even though all but the last
could have been. As the IUCLID dossier did include comments about NP reaching soil as
a result of NPE use, it seems reasonable to assume that under the new system the
company might have looked harder for information in relation to this, and possibly to have
tested worms as these are specifically mentioned for Dossier C.
3
Note, of these studies, the only one not available at the time of the IUCLID submission is the Kopf (1997)
study.
Page 1-22
RPA & BRE
If it was assumed instead that the worm result was not available for inclusion in the dossier
at the start, this would suggest that further testing would have to be undertaken.
The PNEC for the terrestrial compartment is 0.3 mg/kg wet weight.
Table 4.2: Aquatic Toxicity Data
Trophic level
Species
Freshwater fish
Fathead minnow
Pimephales promelas
Saltwater fish
Sheepshead minnow
Cyprinodon variegatus
Ceriodaphnia dubia
Freshwater
invertebrates
Daphnia magna
End point
Concentration
(mg/l)
0.128
0.0074
Reference
Validity
Brooke (1993a)
Ward & Boeri
(1991b)
Ward & Boeri
(1990d)
England (1995)
Valid
Valid
0.085
0.024
Brooke (1993a)
Comber et al
(1993)
Valid
Valid
96hr EC50
96hr LC50
0.0207
0.043
Valid
Valid
28 day NOEClength
0.0039
Brooke et al (1993)
Ward & Boeri
(1990c)
Ward & Boeri
(1991c)
Ward & Boeri
(1990b)
Kopf (1997)
96hr LC50
33 day
NOECsurvival
96hr LC50
0.31
96hr EC50
7 day
NOECreproduction
48hr EC50
21 day
NOECsurviving
0.069
0.0887
Valid
Valid
offspring
Saltwater
invertebrates
Fresh water
algae
Hyalella azteca
Mysidopsis bahia
Selenastrum
capricornutum
Scenedesmus
subspicatus
96hr EC50(Cell
0.41
growth)
72hr EC50 (Biomass)
72hr EC10 (Biomass)
72hr EC50 (Growth
rate)
0.0563
0.0033
0.323
0.0251
Valid
Valid
Valid
72hr EC10 (Growth
rate)
Saltwater algae
Skeletonema costatum
96hr EC50(Cell
0.027
growth)
Mesocosm
study
20 day NOEC
20 day LOEC
Table 4.3: Toxicity to Terrestrial Plants
Species
Test substance
Soil type
Lettuce (Lactuca
sativa)
Sorghum (Sorghum
bicolor)
Sunflower
(Helianthrus rodeo)
Soya (Glycine max)
4-nonylphenol
NP
Agricultural
loam
Grit/loam
soil
0.005
0.023
Ward & Boeri
(1990a)
Liber et al (1999)
Endpoint and effect concentration
(wet weight)
7 day EC50 (Growth) 559 mg/kg
14 day EC50 (Growth) 625 mg/kg
21 day NOEC (Growth) 100 mg/kg
21 day EC50 (Growth) 1,000 mg/kg
21 day NOEC (Growth) 100 mg/kg
21 day EC50 (Growth) 1,000 mg/kg
21 day NOEC (Growth) 100 mg/kg
21 day EC25 (Growth) 1,000 mg/kg
Valid
Use
with
Care
Reference
Hulzebos et al
(1993)
Windeatt &
Tapp (1987)
Page 1-23
Case Study 1: Nonylphenols
Table 4.4: Toxicity to Terrestrial Invertebrates
Species
Test substance
Soil type
Springtails
(Folsomia
fimetaria)
Earthworms
(Apporec-todea
calignosa)
NP
sandy soil
4-nonylphenol in
sludge
NP
LUFA soil
NP
LUFA soil
Endpoint and effect concentration
(wet weight)
21 day EC10 (Reproduction) 27 mg/kg
21 day EC50 (Reproduction) 39 mg/kg
21 day EC10 (Reproduction) 48 mg/kg
21 day EC50 (Reproduction) 59 mg/kg
21 day EC10 (Reproduction) 24 mg/kg
21 day EC50 (Reproduction) 66 mg/kg
21 day EC10 (Mortality) 75 mg/kg
21 day EC50 (Mortality) 151 mg/kg
21 day EC10 (Mortality) > 40 mg/kg
21 day EC50 (Growth) 23.9 mg/kg
21 day EC10 (Reproduction) 3.44 mg/kg
21 day EC50 (Reproduction) 13.7 mg/kg
Reference
Holm
Krogh et
al (1996)
Sediment Toxicity
Limited sediment toxicity data are available, with a 14-day MATC of 26 mg/kg
determined for Chironomus tentans in exposures with sediment. The lack of data means
that the equilibrium partitioning method will be used to estimate the PNEC for sediment.
Avian Toxicity
No data are available. A NOAEL of 15 mg/kg body weight is available from mammalian
studies for reproductive effects. This converts to a PNEC of 10 mg/kg in food using the
methods of the Technical Guidance document.
4.2.4 Environmental Fate
Biodegradation
Standard ready biodegradability tests show that NP is not readily biodegradable. A ready
test in which acclimated sludge was used showed 78% removal after 40 days. Studies on
degradation in surface waters and soil are also available. Half lives of 150 days in surface
water and 300 days in soil are consistent with the observed results. NP is considered to be
inherently biodegradable in waste water treatment plants4.
The degradation of NP ethoxylates, the main products made from NP, has also been
studied. These tend to adsorb to sludge in waste water treatment plants, and reach the
anaerobic clarification stage. Here the ethoxylate chain is removed, leaving NP which is
stable under these conditions. From this, NP can reach soil through the application of
sludge, or be returned to the clarification plant with the surplus water and hence be
released to surface water5.
4
More recent studies show that NP can reach the pass level in a ready test but not within the 10-day window,
hence it would still be classed as inherently biodegradable.
5
This comment was included in the IUCLID. There was no attempt to estimate the yield of NP from this. NPEs
also degrade under aerobic conditions, with the gradual removal of the ethoxylate chain, but this tends to stop
Page 1-24
RPA & BRE
In terms of microbial inhibition, the readily available data were for Pseudomonas, where
an EC10 of >10 mg/l was determined. In developing this dossier, we have assumed that,
in reviewing the data, the manufacturer decided that this was not good enough test data
for preparing a REACH dossier, and a further inhibition test was carried out. A second
test following OECD protocols was undertaken and, in this case, an EC50 of 950 mg/l was
determined in an activated sludge respiration test.
Adsorption/Desorption
For adsorption/desorption, experiments have been carried out for three different soil types.
These determined Koc values as 4.35-5.69 (log values). The relevant test was carried out
to a standard beyond the basic information tests indicated as being required in Dossier B
and can be reconciled to the prediction from log Kow. As a result, the manufacturer
decided that no further testing needed to be undertaken in relation to adsorption or
desorption.
Accumulation
A range of test data is available for bioconcentration. The bioconcentration factor in fish
has been determined in a number of experiments, with values ranging from 220 – 741 on a
whole fish basis. Higher values have been observed in mussels, but also more variable
values, ranging from 10 to 3200.
4.3
Exposure
4.3.1 Overview
As part of the dossier, an exposure assessment is required in order to assess the risks
associated with production and use of NPs. In this case, the assessment was carried out
in relation to the production and major use of the substance, so those aspects directly
related to the producers and their initial (major) customers. This exposure assessment has
been prepared so as to represent all four companies involved in the consortium.
The main uses of NPs addressed in the dossier are:
•
•
•
the production of NPEs (NP ethoxylates);
the production of resins, plastics and stabilisers; and
the more minor uses such as the production of phenolic oximes.
with one to two ethoxylate units still in place, and the terminal group may be oxidised to a carboxylic acid.
Information on this degradation path was to some degree available at the time of the IUCLID. These products
are also considered to have some endocrine-disrupting ability.
Page 1-25
Case Study 1: Nonylphenols
4.3.2 Monitoring Data
With regard to the production of NPEs, concentrations in the receiving water at one site
were 0.54 – 3.02 µg/l. Concentrations at other sites were calculated from measurements
on the levels in effluent and the local dilution. The estimated concentrations were 0.26
mg/l, 0.3 mg/l, 1.49 µg/l, 1.36 µg/l. Three other sites had no emissions to water.
Concentrations in receiving waters have been measured at one site producing NP-based
resins, plastics and stabilisers, with levels of <0.2 µg/l. Concentrations in effluents at three
other sites have also been measured, and these have been used with local information to
estimate the surface water concentrations. These are <0.02, 0, and <0.15 µg/l.
A survey of concentrations reported in the literature indicates that general background
levels have reduced in recent years, possibly due to the removal of NPEs from domestic
cleaning products. Current background levels in countries which have applied restrictions
are probably up to 0.2µg/l. Higher concentrations are measured near to industrial sources,
water treatment plants etc. NP has been measured in sewage sludges.
4.3.3 Estimated Exposures
The specific data for sites included above has been combined with default assumptions
from the Technical Guidance Document to produce estimated concentrations in the
environment. These relate to the three main use areas indicated above. The resulting PECs
are presented in Table 4.4.
Table 4.5: Estimated Exposures
Life cycle step
Production
NPEO production
NP/formaldehyde resins
TNPP production
Epoxy resin production
Stabiliser production
Phenolic oximes
Regional
Page 1-26
Site
A
B
C
D
B
C1+2
C3
D1
D2
E,F,G
PEC water
(µg/l)
<0.2
<0.08
<0.078
1.7-3.2
260
300
1.55
1.42
1.7-2.7
0.11
2.84
0.063
0.0059
PEC sediment
(µg/kg)
23.5
9.4
9.2
200-355
30,500
35,200
182
167
200-317
13
333
7.4
PEC soil
(mg/kg)
0.24
15.5
18
1.27
1.27
159
0.003
0.17
3x10-6
RPA & BRE
4.4
Risk Assessment
The predicted exposure concentrations above have been compared with the PNEC values
derived earlier. The results are presented in Table 4.5 as conclusion (iii), risk, or
conclusion (ii), no risk, rather than as PEC/PNEC ratios. An assessment of risk from
secondary poisoning is included, based on the concentrations in water and soil from above
and using bioconcentration factors as estimated using the Technical Guidance methods.
Ratios for micro-organisms in waste water treatment plants, not shown, indicate no risk
for any life cycle step.
Table 4.6: Identification of Risks
Life cycle step
Production
NPEO production
Site
Water
Soil
A
B
C
D
B
C1+2
C3
D1
D2
E,F,G
(ii)
(ii)
(ii)
(ii)
(iii)
(iii)
(iii)
(iii)
(iii)
(ii)
(iii)
(ii)
(ii)
(iii)
(ii)
(ii)
(ii)
(ii)
(ii)
(ii)
(ii)
(iii)
(iii)
(iii)
(iii)
(ii)
(ii)
(ii)
(ii)
(ii)
(ii)
(ii)
NP/formaldehyde resins
TNPP production
Epoxy resin production
Stabiliser production
Phenolic oximes
Regional
4.5
Secondary
poisoning
(ii)
(ii)
(ii)
(ii)
(ii)
(iii)
(iii)
(iii)
(iii)
(ii)
(ii)
(ii)
(ii)
(ii)
(ii)
(ii)
Risk Management
4.5.1 Conclusions from the Risk Assessment
The risk assessment concludes that there is a need for risk reduction for the following
processes and applications of NP:
•
•
•
NPEO production (on the basis of risks to water, soil and secondary poisoning);
NP/Formaldehyde resins (on the basis of risks to the water environment); and
Stabiliser production (on the basis of risks to the water environment).
No risk management is required for the other processes and applications, namely:
•
•
•
•
NP production;
TNPP production
Epoxy resin production; and
Phenolic oximes.
Page 1-27
Case Study 1: Nonylphenols
4.5.2 Recommended Further Testing or Risk Assessment Activities
Based on the findings, further monitoring of discharges to water at NPEO production sites
and downstream user (formaldehyde resins and stabiliser production) locations should be
undertaken to refine estimates of losses to receiving environments.
4.5.3 Further Risk Management Measures
Depending on the conclusions concerning further emissions monitoring, further risk
management activities may be required. In the event that further measures are required, it
is recommended that additional emissions control technology is employed to ensure that
emissions are below ecologically significant levels.
In this regard, one or more of the following control technologies and methods could be
employed:
•
•
•
•
Page 1-28
the installation of non-contact vacuum systems to reduce releases of NP via steam
ejectors, cutting releases by an estimated 51%;
the introduction of new in-process cleaning technology to reduce emissions of NPEs
by 33%;
reductions in emissions by almost 60% through waste minimisation and good
housekeeping initiatives; and
reductions in emissions by around 90% over three years using a combination of waste
minimisation, containment and minor process adjustments.
RPA & BRE
5.
THE REACH DOSSIER CONSIDERED
5.1
The Evaluation Approach
The aim of developing the hypothetical dossiers is to provide a basis for comparing what
might have happened had REACH been introduced earlier with what happened under the
existing regime. In order to do this, we discuss below whether REACH would:
•
•
•
•
•
5.2
require the same level of test data as required under ESR or other regulatory regimes;
raise any concerns for the example substance and, if so, for which endpoints and risk
compartments;
identify the same endpoints and risk compartments as those identified (historically) and
controlled under the existing legislative arrangements;
recommend through this retrospective application, similar risk reduction measures to
those implemented at present; and
lead to action being taken sooner than under the current system and hence reduce
levels of environmental damage and risk to man via the environment.
Comparison with ESR Risk Assessment
5.2.1 Conclusions of the ESR Risk Assessment
The risk assessment carried out under ESR identified the conclusions set out in Table 5.1
for production of NP and NP derivatives6.
Table 5.1: Conclusions of the ESR Risk Assessment for Production of NP and NP Derivatives
Life cycle step
Water
Soil
Secondary
Poisoning
Production
(ii)
(ii)
(iii)*
NPEO production
(iii)
(iii)
(iii)
NP/formaldehyde resins
(ii)
(ii)
(iii)
TNPP production
(ii)
(ii)
(ii)
Epoxy resin production
(ii)
(ii)
(iii)*
Stabiliser production
(ii)
(ii)
(iii)
Phenolic oximes
(ii)
(ii)
(iii)*
* = identified as a risk by virtue of background levels from NPE uses
As can be seen from Table 5.1, the ESR risk assessment concluded that there was a need
for risk reduction for the following processes and applications:
•
•
•
•
6
NPEO production (on the basis of risks to water, soil and secondary poisoning);
NP/formaldehyde resins (on the basis of risks to the water environment);
Stabiliser production (on the basis of risks to the water environment);
NP production (on the basis of risks to the water environment);
Note that the ESR Risk Assessment covered a range of applications of NPEs. This is discussed later.
Page 1-29
Case Study 1: Nonylphenols
•
•
Phenolic oximes (on the basis of risks to the water environment); and
Epoxy resins (on the basis of risks to the water environment).
No requirement for risk reduction was identified for TNPP Production.
5.2.2 Comparison of REACH Dossier Conclusions with ESR
The conclusions reached by the hypothetical REACH Dossier for the production of NP
and NP derivatives differ from those found in the ESR risk assessment for the following
processes and activities:
•
•
•
NP production;
Epoxy resin production; and
Phenolic oximes.
For all of these processes and activities, the ESR risk assessment identified that risk
reduction was required. However, in all cases, the reason that the ESR risk assessment
reached these conclusions was because of the background concentrations in the
environment resulting from the use of NPEs.
In the environmental risk assessment carried out under ESR, the endpoints (ecosystems)
considered are the primary environmental ‘compartments’ (aquatic, terrestrial and
atmospheric), as well as effects relevant to the food chain (secondary poisoning7). Impacts
on each of these four endpoints were assessed independently for each phase in the NP lifecycle. These phases are:
•
•
•
•
NP production;
production of NPE and other NP derivatives;
formulation of NPE-based products; and
use of NPE-based products in each of the identified industry sectors.
The full conclusions reached for all phases are summarised in Table 5.2 overleaf. For the
aquatic environment, the conclusions of the risk assessment were that unacceptable risks
(conclusion (iii)) arise from all industry sectors which use NP/E with the exception of
TNPP production. For the ‘terrestrial’ and ‘secondary poisoning’ endpoints, the
conclusion concerning unacceptable risks (conclusion (iii)) applies to fewer sectors. For
the atmospheric compartment, no unacceptable risks were found for any of the sectors.
Thus, some sectors require risk reduction for only one endpoint, while others require risk
reduction for two or three endpoints. Of all the endpoints, the aquatic is the most
sensitive in that it has the lowest NP concentration threshold to trigger adverse effects.
The conclusions from the REACH assessment are that there are risks for the aquatic
compartment from NPE production, NP/formaldehyde resin production and production of
stabilisers. NPE production also gives rise to risks for the terrestrial compartment and
7
This includes bioconcentration, bioaccumulation and biomagnification.
Page 1-30
RPA & BRE
Table 5.2: Conclusions of the Environmental Risk Assessment*
Life Cycle
Industry Sector
Risk to
Stage
Aquatic
Environment
NP Production NP production
(iii)
Production
NPE
(iii)
of NP
Phenol/formaldehyde resins
(iii)
Derivatives
TNPP
(ii)
Phenolic oximes
(iii)
Epoxy resins
(iii)
Other plastic stabilisers
(iii)
Formulation of Formulation (excluding paints)
(iii)
NPE-based
Paints
(iii)
Products
Use of NPEI&I
(iii)
based
Emulsion polymerisation
(iii)
Products
Textile auxiliaries
(iii)
*
Risk to
Terrestrial
Environment
(ii)
(iii)
(ii)
(ii)
(ii)
(ii)
(ii)
(iii)
(iii)
Risk of
Secondary
Poisoning
(ii)
(iii)
(ii)
(ii)
(ii)
(ii)
(ii)
(iii)
(ii)
(iii)
(iii)
(iii)
(iii)
(iii)
(iii)
(ii)
(iii)
(ii)
(iii)
(ii)
(ii)
(iii)
(ii)
(iii)
not given
not given
(ii)
(iii)
(iii)
Captive use
(iii)
Leather auxiliaries
(iii)
Agriculture (pesticides)
(iii)
Agriculture (veterinary care)
(iii)
Paints
(iii)
Metal industry (extraction)
(iii)
Pulp and paper
(iii)
Other niche markets
Civil and Mechanical Eng.
(iii)
(iii)
(ii)
Electronics/Electrical Eng.
(iii)
(iii)
(ii)
Mineral Oil and Fuel Industry
not given
not given
(iii)
Photography (small scale)
(ii)
(ii)
(iii)
Photography (large scale)
(iii)
(iii)
(ii)
Other
(iii)
not given
not given
The table excludes assessment of the risk to the atmosphere as the risk assessment reached conclusion (ii) –
no unacceptable risks - overall for the atmospheric compartment.
The risk assessment notes that Conclusion (iii) – unacceptable risks - was reached for these sectors only
because the background regional PEC was added to the local PEC.
for secondary poisoning; the other uses are not a risk to these compartments. There are
no risks at the regional level. The main differences between the ESR and REACH
assessments relate to the aquatic compartment, and are due to the lower regional
background concentration estimated in the REACH dossier (as it does not include any
contribution from the breakdown of NPEs).
Because the hypothetical dossier prepared for this case study was (realistically) limited to
the production of NP, NPEs and other NP based derivatives, it did not highlight any issues
associated with the formulation and use of NPE based products. Instead, it is assumed
that the formulation and use of NPE products would be the subject of a second dossier.
The key question here, then, is whether the necessary linkages would be made between the
use of NPEs and the fact that they degrade to NP in the environment, subsequently posing
unacceptable risks to the aquatic environment and potentially the terrestrial environment
and from secondary poisoning. If it is further assumed that manufacturers of NP would
Page 1-31
Case Study 1: Nonylphenols
also be involved in any assessment of the formulation and use of NPEs (either because
they formulate NPE-based products for downstream users or have an interest in preparing
such a dossier), then one assumes that the linkage between emissions of NPEs and NPs in
the environment would be made.
In the case of NPs, this seems reasonably likely as information that NPEs can break down
to NP was in the general literature. One of the important issues for REACH, however, is
whether this would be the case for other substances where the link between substances
and their decomposition products is less clear and, if not, whether guidance should attempt
to address this issue more robustly.
A further consideration for REACH and NPs is whether there would be a single dossier
for NPEs or whether there would be several, as it may depend on whether they are listed
as one substance on EINECS or several.
5.2.3 Detailed Consideration of Hazardous Effects and Routes of Exposure between the
Assessments
Aquatic Compartment
The risk assessment reviewed standard toxic effects on the aquatic environment (fish,
aquatic invertebrates and algae) as well as bioaccumulation of NP (see ‘secondary
poisoning’ below). It was found that the standard toxicity effects occur at lower
concentrations than significant effects associated with bioaccumulation, so standard toxic
effects were used as the basis for deriving the ‘predicted no effect concentration’ for water
(PNECwater) of 0.33 g/l. The regional ‘predicted environmental concentration’ (PECwater)
is 0.60 g/l. Based on background regional concentrations alone, the PEC/PNEC ratio will
always be greater than one. Thus, the production, formulation or use of any product
containing NP or its derivatives will automatically result in Conclusion (iii). The only
exception to this is TNPP production, as the risk assessment concludes that the two TNPP
production sites in the EU contribute nothing to the local (nor, therefore, to the regional)
concentrations.
The toxicity data set used in the REACH dossier is the same as that for ESR, so the same
PNEC is derived. The background concentration resulting only from the use of NP, which
is what is calculated in the REACH dossier, is significantly lower than that in the ESR
assessment, and as a result is not a risk in itself. Hence some of the use areas indicated as a
risk in the ESR assessment (production, epoxy resins, oximes) are not shown as a risk in
the REACH dossier.
Terrestrial Compartment
Toxicity tests of NP on terrestrial plants show effects on growth, while tests on terrestrial
invertebrates show impacts on reproduction and mortality. The PNECsoil of 0.3 mg/kg is
based on the most sensitive of these test subjects. The PEC varies according to industry
sector, exceeding PNEC where discharges to sewer (and hence levels in sludge) are
particularly high.
Page 1-32
RPA & BRE
According to the risk assessment, the PEC for soil is primarily a result of NP/E in sewage
sludge applied to land. The quantity of NP/E in sewage sludge is a direct result of the
many industrial uses of NPE-based products and, potentially, its use as a flocculant in
sewage treatment processes.
The PNEC and the conclusions from the REACH dossier are the same as those for the
ESR assessment.
Secondary Poisoning
The risk assessment also considers secondary poisoning, an effect on higher organisms
(e.g. birds, fish-eating mammals) which can arise through their consumption of lower
organisms containing the substance (e.g. fish, daphnia). This is assessed by comparing the
concentrations in the food organisms with the effect concentrations on the higher
organisms. The risk assessment suggests that these effects will occur at concentrations of
10 g/l or higher, a figure well above the PNECwater and the regional PECwater. Thus, the
reduction of concentrations to below the PNECwater should provide adequate protection.
The conclusions from the REACH dossier are the same as those for the ESR assessment.
5.2.4 Contribution of Each Industry to Risk Levels
Table 5.3 overleaf, which is based on the ESR risk assessment, shows the contribution to
the continental burden of NP attributed to the various industry sources of NP and NPE.
From the Table, it can be seen that the industrial and institutional cleaning products,
textile, leather and NPE production together contribute some 70% of the total burden.
The 24% of the total burden associated with ‘other niche markets’ is largely unaccounted
for, although a small part of this is attributable to the civil and mechanical engineering,
electronics/electrical engineering, mineral oil and fuel, and photographic sectors. The final
5% is distributed across the remaining industry sectors. More than 90% of the burden is
associated with final use of NPE-based products.
Whilst some sectors give rise to a relatively severe risk (those with the highest PEC/PNEC
ratio such as NPE production, production of plastic stabilisers, leather processing, textile
processing), for others the risk is more marginal and may only exist because of the
‘continental burden’ (i.e. widespread background pollution). For seven industries, the risk
assessment reached conclusion (iii) only when the regional background concentration was
added to the local concentration. These are: NP production; epoxy resin production;
phenolic oxime production; use of agricultural pesticides (but not veterinary medicines);
captive use; small photographic users (but not large users); and use of paints. Reducing
regional background concentrations would result in these no longer posing unacceptable
risks.
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Case Study 1: Nonylphenols
Table 5.3: NPs and NPEs – Usage, Contribution to Continental Environmental Burden and Risk
Ratios
% of EU
%
Usage (NP or
Continental
PEC/PNEC
NPE)
NP Burden
Direct releases of NP
NP production
n/a
0.003
<0.6 to <1.8
NPE production
60
5.82
5.91 to 1,394
NP/formaldehyde resin production
Tris (nonylphenyl)phosphite (TNPP) –
production
29
0.007
4.9 to 9.7
5
0
n/a
Epoxy resin manufacture
2
0.004
1.97
Production of other plastic stabilisers
1
0.02
11.3
Phenolic oxime production
3
0
1.79
100 (NP)
6
Subtotal
Indirect releases via NPEs
Formulation
n/a
5.79 to 39.4
c
Pesticide application
6
0.54
Captive use by chemical industry
9
0.1
1.88
Electrical engineering applications
<1
0.001
11
Industrial and institutional cleaning
30
44.7
79.7
Leather processing
8
6.09
52.4 to 255.8
Metal processing and extraction
Fuel and oil additives (manufacture and
blending)
3
1.22
427
<1
0.008
4.8 to 108
Photographic materials
Polymer production/emulsion
polymerisation
<1
0.16
2.06 to 6.45
Pulp, paper and board industry
1
1.72
50
Textile processing
13
14.7
1060
see below>>
see below>>
16.7
Paint use
5
0.04
1.8
Civil engineering
<1
0.02
94.8
Misc. other (incl. unallocated tonnage)
10
23.5
n/a
100 (NPEs)
94
n/a
n/a
n/a
1.78
Paint production
Subtotal
2 to 2.8
5.55
Background risks
Regional PEC/PNEC ratios
Source: adapted from RPA (2000)
Discharges of NP/Es from all sectors contribute to the continental burden, although some
to a greater extent than others. Some of the sectors are relatively ‘emissive’ in that a large
proportion of the quantity of NP/E used is emitted to the environment, whereas others
operate relatively closed processes, leading to low levels of emissions. Thus, the amount
released (and resulting environmental risks) is often not proportional to the quantity used.
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RPA & BRE
The REACH dossier deals only with the production and direct use of NP. Emissions to
water are dominated by those from the production of NPEs, which account for ~75% of
regional emissions and 98% of continental emissions. All other processes have relatively
small emissions. The total emissions in the REACH dossier are only ~6% of the total in
the ESR assessment.
5.3
Control of Identified Risks
5.3.1 The Risk Reduction Strategy under ESR
The finding that the magnitude of environmental risks varies considerably according to the
industry sector (as set out in Table 5.3 above) was taken into account in developing the
risk reduction strategy under ESR. Relative advantages and drawbacks of various policy
measures were considered, with the final strategy adopted on the basis that the most
stringent measures should be targeted at those sectors that contribute most to the
continental burden.
The strategy was therefore based on a stepped approach, aimed at ensuring that the
environmental benefits were gained in a cost-effective manner by, first, reducing the
continental burden and background concentration and eliminating regional concerns (i.e.
reduce PEC/PNEC to <1) and, later, concerns at the local level.
The first step involved the introduction of marketing and use restrictions on those uses
that contributed the most to the continental burden of NP/Es. The next step was then to
regulate controlled processes via the IPPC regime, with residual risks controlled through
the use of environmental quality standards under the Water Framework Directive (WFD).
Table 5.3 below summarises the main proposals of the Risk Reduction Strategy (RPA,
2000) for each of the sectors.
It is important to note that the risk reduction strategy has not yet been fully implemented.
However, on the 19 May 2003, the Council adopted Directive 2003/…/EC amending for
the 26th time Council Directive 76/769/EEC relating to restrictions on the marketing and
use of certain dangerous substances and preparations. The Directive has not yet been
published in the OJ and, as such is not yet implemented. The Directive places restrictions
on NP/E and restricts its use as a substance or constituent of preparations in
concentrations equal to or higher than 0.1% by mass for the following purposes:
•
•
•
•
•
•
industrial and institutional cleaning (except controlled closed dry cleaning/cleaning
systems where washing liquid is recycled or incinerated);
domestic cleaning;
textiles and leather processing (except processes with no release into waste water/pretreatment to remove organic fraction);
emulsifier in agricultural teat dips;
metal working (except use in closed systems where washing liquid is recycled or
incinerated);
manufacturing of pulp and paper;
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Case Study 1: Nonylphenols
cosmetic products;
other personal care products (except spermicides); and
co-formulants in pesticides and biocides.
•
•
•
5.3.2 Comparison with Risk Control through REACH
Table 5.4 provides a comparison of the risk management measures proposed in the
REACH Dossier (Section 4) and those recommended in the NPs Risk Reduction Strategy
under ESR.
Table 5.4: Proposed Risk Reduction Measures – REACH versus ESR
Recommended
Measure
Marketing and
use restrictions
Integrated
Pollution
Prevention and
Control (IPPC)
REACH Proposals: NP only
Production of NPE
Production of phenol/formaldehyde
resins
Production of other plastic stabilisers
Environmental
Quality
Standards/Limit
Values
ESR Risk Reduction Strategy:
NPs and NPEs
Metal working
Pulp, paper and board
Cosmetics and personal care products
Industrial and institutional cleaning
Textile processing
Leather processing
Agriculture (biocidal products, in
particular in teat dips)
Production of NPE
Captive use
Production of phenol/formaldehyde
resins
Production of other plastic stabilisers
Emulsion polymerisation
Formulation for other uses
Production of epoxy resins
Production of phenolic oximes
Paints (production, domestic use and
industrial use)
Civil and mechanical engineering
Electronic/electrical engineering
Mineral oils and fuel
Photographic industry
Source: RPA (2000) and case studies
As can be seen from the table, there is no substantial difference between the outcomes as
regards the risk management measures. The only difference is that, for those categories of
use where no risk management was identified under the Dossier, the ESR process
identified optional controls, because of the linkage with background concentrations
(triggering of a conclusion (iii)) on a large number of NPE uses. The exception was for
TNPP where no risk was identified under ESR, but the option of instigating controls was
included anyway.
In terms of overall environmental risk levels as applied to production of NP and NP
derivatives alone, there is no substantive difference between risk management under
REACH and ESR in terms of the robustness of recommended controls, whether or not
REACH made the link between NPE and NP and background concentrations.
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RPA & BRE
As regards the issue of the separate dossier on NPEs, the various applications of NPEs are
responsible for 94% of the total environmental burden. If the link were not made between
NPEs and their NP decomposition products under REACH (which is unlikely), the
outcome of risk management measures for NPE uses is uncertain. This emphasises the
importance of making sure decomposition products are considered robustly. In terms of a
dossier on NPEs, these have a lower toxicity to aquatic organisms than NP and they are
readily biodegradable. It is likely that that they would not present the same level of risk,
unless large quantities are discharged. So it would be their breakdown products which
would be the concern. These are more complex, in that if we are considering the
breakdown of NPEs (rather than the formation of NP) then we would probably need to
consider the 1 and 2 ethoxylate products and their carboxylic acid equivalents as well as
NP itself. These other products have been shown to have some estrogenic effect. A toxic
equivalent approach might be needed.
If the link were made in the case of NP/Es (which is likely), it is unlikely that the key
dispersive users of NPEs and those identified for marketing and use restrictions in the ESR
risk reduction strategy (namely: industrial and institutional cleaning; textile processing;
leather processing; agriculture; metal working; pulp, paper and board; cosmetics and
personal care products) would recommend anything greater than an agreement to phase
out uses over time.
Clearly, as an effective risk management measure, this is not as robust as the marketing
and use restrictions proposed by the ESR process, however, REACH may be better able
to identify the downstream uses and quantities used for the large number of uses that are
still classified under “other” in the ESR process8. Such users would be obligated under
REACH to submit postcard notifications. It is unlikely, however, that this enhanced
ability to identify additional downstream (unintended) uses would compensate for the less
robust measures (voluntary agreements) that may be proposed by manufacturers and users
under REACH. In this respect, the only way in which robust risk management of the sort
recommended under ESR would occur if NP and NPE were called in for Authorisation.
Whether NP/E would have been called in for Authorisation is debateable, but seems likely
given the context and timescale of the problems and concerns described in Section 3.
Assuming it would have been (which seems likely given the above), it is likely that the
process of risk management under REACH/ Authorisation could provide an even more
robust mechanism than the ESR process since all uses would be declared (where some are
still ‘unidentified’ in the ESR assessment) and the consideration of substitutes would be
based on more robust criteria than the hazard profiles currently employed to ‘screen’ likely
substitutes under ESR (as dossiers with a full toxicological and fate profile would be
available for these substitutes).
8
A large number of additional uses not considered in the risk assessment were identified in the course of
undertaking the risk reduction strategy under ESR.
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Case Study 1: Nonylphenols
5.3.3 Damage Costs Avoided Under REACH
Given these points, it is likely that, had REACH been in place earlier, it would have
identified risks and recommended risk management measures much earlier, considering
that most of the data were available in the early to mid-1980s. Where they were not,
substance tailored testing under REACH would have filled the remaining gaps. As such,
the following might apply:
•
historical and existing use of NP/Es would not have resulted in ecologically significant
levels, particularly in the aquatic environment;
•
if a broad assumption is made that the structure and pressures operating in US rivers is
similar to European continental rivers, it can be tentatively suggested that 25% of EU
rivers could have had levels of NP/E that are regularly in excess of the no effect
concentration and 70% could have had levels which exceed the no effect
concentrations under low flow conditions;
•
if REACH had been in place 15 years ago (allowing time for measures to fall into
place), recovery of affected aquatic systems (including the sediment) could be almost
complete by this time;
•
similarly, contamination of marine waters and sediments would not have been
widespread and affected areas would have recovered by this time;
•
the 25% to 58% of sewage treatment plants that have been releasing ecologically
significant levels of NP/Es would have ceased doing so long ago; and
•
there would have been much less contamination of soils through the application of
sewage sludges.
Based on existing data, it is difficult to derive a complete set of damage costs for NPs.
However regulatory changes for the treatment of sludges provide an insight into a part of
the costs. In 2000, the EC published the 3rd Draft of the “Working Paper on Sludge”
which, among other changes to the current legislation, proposes limit values for
concentrations of organic compounds, including NPEs. This means that wastewater
treatment plant operators will face increased disposal costs in landfilling or incinerating
sludge, in order to meet limits placed on NPE concentrations in sewage sludge spread
onto land.
Samples taken at numerous sewage treatment works in the EU have found levels of NP/Es
in sludge at concentrations well above the proposed limit of 450 mg/kg dm currently being
proposed for the Sludge Directive. This limit is quite high compared with the Danish and
Swedish limits of 10 mg/kg dm and 50mg/kg dm respectively. However, it is obvious that
a number of EU member states would have incurred significant costs in trying to meet the
limits, given the values of NPEs in sludge across the EU shown in Table 5.5.
Page 1-38
RPA & BRE
Table 5.5: Values of NP/Es in sludge from literature
Where
Year
Average value
Reference
(mg/kg dm)
Hessian
1995
25
Fooken et al (1995)
Germany
1995
107
BLAU (1995)
Germany
1995
128.2
Jobst (1995)*
Germany
1997
50 – 300
Schnaak et al (1997) *
Spain
2000
10 – 48
Castillo et al (2000) *
Spain
2000
nd – 16
Castillo et al (2000) *
Norway
1989
25 – 2298
Paulsrud et al (2000)**
Sweden
1989-91
44 – 7214
Paulsrud et al (2000)**
Denmark
1995
0.3 – 67
Torslov et al (1997) **
Denmark
1993-94
55 – 537
Torslov et al (1997) **
*
**
Source: Rogers et al, 2002; Source: Lagenkamp et al, 2001
To estimate the cost implications of this, the following calculations have been used. It is
known that:
•
•
•
approximately 77,600 tonnes/year of NP/Es are used within the European Union
(RPA, 2000), while 17,600 tonnes of NP/Es are consumed in the UK alone (CES,
1993);
of the 17,600 tonnes, an estimated 10,690 tonnes (approximately 61%) is discharged
to the sewer, while 2,950 tonnes (approximately 17%) ends in sludge;
assuming that the UK situation applies throughout Europe, the quantities of NP/Es in
sludge for the EU would be an estimated 13,200 tonnes (17% of 77,600).
Based on this 1997 data, an estimated 13,200 tonnes of NP/E end up in sewage sludge
every year in the EU. As the difference in costs between land spreading and incineration
are estimated at between €150 to €190 per tonne, the increased disposal costs may
become significant. Because not all uses of NP/Es will be banned under the proposed risk
reduction strategy, these increased disposal costs could be realised by a number of
wastewater treatment plant operators in Europe (assuming the proposed EU limits become
a legal requirement).
NPs are also a priority hazardous substance under the Water Framework Directive.
Although they appear to be in higher concentrations in sludge than in the aquatic
environment, action will have to be taken to address any levels in excess of the currently
proposed 0.33µg/l Environmental Quality Standard. It is only by 2009 that measures may
be drawn up to tackle remaining discharges to the aquatic environment, and any associated
contamination of sediments. It is not until then that the costs of clean-up will be realised.
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Case Study 1: Nonylphenols
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RPA & BRE
6.
REFERENCES
Ahel M et al (1981): Organic Micropollutants in Surface Waters of the Glatt Valley,
Switzerland in Analysis of Organic Micropollutants in Water: Proceedings of the
Symposium held in Ireland 1981 (eds: Bjørset A and Angeletti G).
Ahel M et al (1996): Behaviour of Alkylphenol Polyethoxylate Surfactants in the Aquatic
Environment -- 3. Occurrence and Elimination of their Persistent Metabolites During
Infiltration of River Water to Groundwater, Water Res, Vol 30 No 1, pp37-46.
Baldwin W et al (1997): Metabolic Androgenization of Female Daphnia magna by the
Xenoestrogen 4-Nonylphenol, Environ. Toxicol. Chem., Vol 16, pp1905-1911.
Brooke LT (1993a): Acute and Chronic Toxicity of Nonylphenol to Ten Species of Aquatic
Organisms, USEPA Draft Report, EPA Contract No 68-C1-0034.
Bund-/Länderausschuß für die Bewertung von Umweltchemikalien (BLAU) (1995): Bericht der
Arbeitsgruppe ‘Datensammlung für die Bewertung von Umweltchemikalien’, Stand
14.08.1995.
BUA (1988): Nonylphenol - BUA Report 13, GDCh-Advisory Committee on Existing
Chemicals of Environmental Relevance, January 1988.
CES (1993): Uses, Fate and Entry to the Environment of Nonylphenol Ethoxylates, Final
Report submitted to the Department of the Environment, London.
Colerangle JB & Roy D (1996): Exposure of Environmental Oestrogenic Compound
Nonylphenol to Noble Rats Alters Cell Cycle Kinetics in the Mammary Gland, Endocrine,
Vol 4, pp115-122.
Comber MHI et al (1993): The Effects of Nonylphenol on Daphnia magna, Water Research, Vol
27 No 2, pp273-276.
EC (2002): 4-Nonylphenol (Branched) and Nonylphenol, Risk Assessment Report,
European Commission. EUR 20387 EN (Pl-2, Volume 10)
England DE (1995): Chronic Toxicity of Nonylphenol to Ceriodaphnia dubia, report prepared
for the Chemical Manufacturers Association by ABC Laboratories Inc. Report #41756,
Environ. Toxicol. Chem., Vol 12, pp1079-1094.
European Commission (2000): Working Paper on Sludge – 3rd Draft, Brussels, April 2000.
Gaffney, PE (1976): Carpet and Rug Industry Case Study II: Biological Effects, J. Water Pollut.
Control Fed., Vol 48 No 12, pp2731-2737.
Harries et al (1995): As reported in Chemicals with Estrogen-like Effects, Tema Nord 1996:580,
Nordic Council of Ministers, Copenhagen, 1996.
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Case Study 1: Nonylphenols
Hüls AG (1990): In Vitro Mammalian Cell Gene Mutation test with Nonylphenol, IBR
project no. 95-86-0449-90.
Hulzebos EM et al (1993): Phytotoxicity Studies with lactuca sativa in Soil and Nutrient
Solution, Issue 12, pp21.
Ikeda, M et al (1970): Two Cases of Leucoderma, Presumably Due to Nonyl- or Octylphenols in
Synthetic Detergents, Ind Health, Vol 8, pp192-196.
Initiative Umweltrelevante Altstoffe (1992): Teratogenicity Study in Wistar Rats Treated
Orally with Nonylphenol. IBR project no. 20-04-0502/00-91.
Jobling et al (1996): Inhibition of Testicular Growth in Rainbow Trout (Oncorhynchus mykiss)
Exposed to Oestrogenic Alkylphenolic Chemicals, Environ. Toxicol. Chem., Vol 15,
pp194-202.
Jobling S & Sumpter JP (1993): Detergent Components in Sewage Effluent are Weakly
Oestrogenic to Fish: An in vitro Study Using Rainbow Trout (Oncorhynchus mykiss)
Hepatocytes, Aquat. Toxicol., Vol 27 No 3-4, pp361-372.
Lagenkamp H & Part P (2001): Organic Contaminants in Sewage Sludge for Agriculture
Use, European Commission, Joint Research Centre, Institute for Environment and
Sustainability, Soil and Waste Unit, October 2001.
Liber K et al (1999): Lethality and Bioaccumulation of 4-nonylphenol in Bluegill Sunfish in
Littoral Enclosures, Environ. Toxicol. Chem., Vol 18 No 3, pp394-400.
Kravetz L et al (1982): Ultimate Biodegradation of an Alcohol Ethoxylate and a Nonylphenol
Ethoxylate under Realistic Conditions, Soap Cosmetics Chem. Specialities, Vol 58, pp3442.
Kopf W (1997): Wirkung Endokriner Stoffe in Biotests mit Wasserogranismen, in Stoffe mit
Endokriner Wirkung in Wasser, (ed: Institut für Wasserforschung München)
Oldenbourg.
Naylor C et al (1992a): Alkylphenol Ethoxylates in the Environment, J. Am. Oil Chem. Soc.,
Vol 69 No7, pp695-703.
Nethercott JR & Lawrence MJ (1984): Allergic Contact Dermatitis due to Nonylphenol
Ethoxylate (Nonoxynol-6), Contact Dermatitis, Vol 10 No 4, pp235-239.
Odum J et al (1999): Comparative Activities of p-nonylphenol and Diethylstilbestrol in Noble
Rat Mammary Gland and Uterotrophic Assays, Reg Tox Pharmacol, not yet published.
Proceedings of a Seminar on Nonylphenol Ethoxylates and Nonylphenol, Grand
Hotel, Saltsjobaden, Sweden, 6-8th February, 1991.
Paxéus N (1996): Vehicle Washing as a Source of Organic Pollutants in Municipal Wastewater,
Water Sci Tech, Vol 3 No 6.
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RPA & BRE
Prasad R (1989): Effect of Nonylphenol Adjuvant on Macrophtyes, Adjuvants Agrochem., Vol 1,
pp51-61.
Purdom CE (1994): Estrogenic Effects of Effluents from Sewage Treatment Works, Chem. Ecol.,
Vol 8 No 4, pp275-285.
Rogers et al (2002): Review of Pollution Inventory Substances in Sludge from Sewage
Treatment Works, R&D Technical Report No P2-288, Environment Agency.
RPA (2000): Nonylphenol Risk Reduction Strategy, Final Report for the Department of the
Environment, Transport and the Regions, London.
RPA (2001): Nonylphenols, Supplementary Information in Support of the Risk Reduction
Strategy, Final Report for the Department of the Environment, Transport and the
Regions, London.
Schaffner C et al (1987): Field Studies on the Behaviour of Organic Micropollutants During
Infiltration of River Water to Ground Water, Water Sci. Technol., Vol 19, pp1195-1196.
Shurin J & Dodson S (1997): Sublethal Toxic Effects of Cyanobacteria and Nonylphenol on
Environmental Sex Determination and Development in Daphnia, Environ. Toxicol.
Chem., Vol 16.
Soto AM et al (1991): p-Nonylphenol: an Oestrogenic Xenobiotic Released from Modified
Polystyrene, Environ. Health Perspect., Vol 92, pp167-173.
Sundaram, K & Szeto, S (1981): The Dissipation of Nonylphenol in Stream and Pond Water
under Simulated Field Conditions, J. Environ. Sci. Health, Vol B16, pp767-776.
Suoanttila M (1996): Autojen maahantuontivarastojen jätevesien määrittäminen ja
bioakkumuloitumisen selvittämien GC/MS - laitteistolla. Diplomityö, Lappeenrannan
teknillinen korkeakoulu, Finland 1996. (Wastewater identification and determination of
bioaccumulation with GC/MS - Combination at an Import Terminal of Cars)
Swedmark, M et al (1971): Biological Effects of Surface Active Agents on Marine Animals,
Marine Biol., Vol 9, pp183-201.
Swisher R (1970): Surfactant Biodegradation, New York, Marcel Dekker Inc.
Ward TJ & Boeri RL (1990a): Acute Static Toxicity of Nonylphenol to the Marine Alga
(Skeletonema costatum), report prepared for Chemical Manufacturers Association by
Resource Analysts, Study No 8970-CMA.
Ward TJ & Boeri RL (1990b): Acute Static Toxicity of Nonylphenol to the Freshwater Alga
(Selenastrum capricornutum), report prepared for Chemical Manufacturers Association by
Resource Analysts, Study No 8969-CMA.
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Case Study 1: Nonylphenols
Ward TJ & Boeri RL (1990c): Acute Flow Through Toxicity of Nonylphenol to the Mysid
(Mysidopsis bahia), report prepared for Chemical Manufacturers Association by Resource
Analysts, Study No 8974-CMA.
Ward TJ & Boeri RL (1990d): Acute Flow through Toxicity of Nonylphenol to the
Sheepshead Minnow (Cyprinodon variegatus), report prepared for Chemical
Manufacturers Association by Resource Analysts, Study No 8972-CMA.
Ward TJ & Boeri RL (1991b): Early Life Stage Toxicity of Nonylphenol to the Fathead
Minnow (Pimephales promelas), report prepared for Chemical Manufacturers
Association by Resource Analysts, Study No 8979-CMA.
Ward TJ & Boeri RL (1991c): Chronic Toxicity of Nonylphenol to the Mysid (Mysidopsis
bahia), report prepared for Chemical Manufacturers Association by Resource Analysts,
Study No 8977-CMA.
White R et al (1994): Environmentally Persistent Alkylphenolic Compounds are Estrogenic,
Endocrinology, Vol 135, pp175-182.
Williamson J B & Varineau P T (1996): Nonylphenol in Biosolids and Sludges, SETAC Poster
Session P0576, November 20, 1996.
Windeatt AJ & Tapp JF (1987): The Effects of Six Chemicals on the Growth of Sorghum
bicolor, Helianthus rodeo and Glycine max, Brixham Laboratory Report BL/A/2836.
Zellner A & Kalbfus W (1997): Belastung bayerischer Gewässer durch Nonylphenole in:
Stoffe mit endokriner Wirkung in Wasser. Bayerisches Landesamt Für
Wasserwirtschaft, Institut für Wasserforschung München (ed.), Oldenbourg, MünchenWien.
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CASE STUDY 2:
SHORT CHAIN CHLORINATED PARAFFINS (SCCPS)
Case Study 2: SCCPs
RPA & BRE
1.
INTRODUCTION
1.1
Background to the Case Study
Short chain chlorinated paraffins (SCCPs) were chosen as a case study chemical for a
number of different reasons. One of the key reasons is that they were on the first
priority list of chemicals under the Existing Substances Regulation (ESR - EEC
793/93). As a result of risk assessment conclusions that their use posed risks to the
environment, they are also one of the first substances to be regulated through the ESR
process.
Other reasons for selecting SCCPs are as follows:
1.2
•
their use first began about 40 years ago as a result of their chemical stability, and
they have been used in a range of different applications;
•
the case study highlights the types of damages that could be avoided in relation to
chemicals that, although highlighted as a priority, are not linked to pollution
incidents and, thus, relate to less obvious environmental and health impacts;
•
occupational health effects and wider public health effects formed the initial basis
of concern, with environmental effects later becoming an issue;
•
their use has been the focus of a wide range of voluntary and regulatory
initiatives, with these including a PARCOM initiative; and
•
restricting the use of SCCPs raises substitution issues as the most cost-effective
substitutes are other chlorinated paraffins.
Format of Case Study
A profile of the more recent market for SCCPs is provided first (Section 2), including
a brief description of how they have been used in different applications. This is
followed in Section 3 by a historical review of how SCCPs became an issue of
concern, and when either voluntary industry or regulatory action in response to such
concerns was first initiated.
The hypothetical REACH dossier is presented in Section 4. This includes a summary
of what we assume for the dossier in terms of production volumes, uses, test data,
exposure, risk assessment conclusions, further testing and risk management
recommendations.
The dossier is then considered further in Section 5, which
compares its conclusions to the findings of the ESR process. Further hypotheses are
then made as to the damages that could have been avoided had REACH been in place
earlier.
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Case Study 2: SCCPs
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RPA & BRE
2.
MARKET PROFILE
2.1
Uses and Trends
2.1.1 Overview
There is a wide range of commercially produced chlorinated paraffins which act as
functional additives in formulations used by a number of downstream user sectors.
These products are classified according to the length of their carbon chain and the
degree of chlorination by percentage of weight, with this placing them in distinct
‘families’. Table 2.1 sets out this classification system and the four main families.
Table 2.1: Classification of Chlorinated Paraffins
C10 – C13
Chlorine Content
(% by weight)
48 – 71
C14 – C17
40 – 59
85535-85-9
C17+ (chlorparaffin waxes)
26 – 59
63449-39-8
C18 - C20 (liquid)
<20, 69-72
85535-86-0
Length of Carbon Chain
CAS number
85535-84-8
Both the length of the carbon chain and the degree of chlorination affect the properties
of the chlorinated paraffins and hence their suitability for different uses. These
factors also affect the toxicity and environmental effects associated with the different
families.
Short chain chlorinated paraffins are those with a carbon chain length between C10 –
C13. They are used as additives in a disparate range of applications, including:
•
•
•
•
•
•
•
metalworking fluids as extreme pressure additives;
leather processing as fat liquoring agents;
paints and coatings as plasticisers and/or flame retardants;
sealant and adhesive manufacture as plasticisers;
flame retardants in rubber conveyor belts;
flame retardants for textiles; and
PVC manufacture as a secondary plasticiser with flame retarding properties.
A brief description is provided below for each of these uses. However, as can be seen
from Table 2.2, the quantities consumed within the EU in the period leading up to
their assessment under ESR varied considerably by sector, with use in metalworking
fluids accounting for the largest proportion historically (table based on ERM, (1999)).
Although data are not readily available on sales to these different sectors for more
recent years (or earlier periods), the trend of declining sales of SCCPs is likely to have
continued. In addition, the importance of the different sectors is likely to have
changed since 1999. This is due to the ESR risk assessment finding unacceptable
risks arising from the use of SCCPs in the metalworking and leather sectors and the
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Case Study 2: SCCPs
Table 2.2: Historic Use Patterns for SCCPs in Western Europe (tonnes/year)
Sector
1994
1995
1997
1998
Value (€1998)
Metalworking
9,380
6,215
5,152
2,018
1,513,000
390
104
249
45
33,750
1,845
1,150
1,537
713
534,750
1,493
1,347
692
638
478,500
n/a
n/a
n/a
13
9,750
100
n/a
41
648
486,000
13,208
8,816
7,671
4,075
3,056,250
Leather Fat Liquors
Paints and Sealants
(plasticiser)
Rubber and Textiles
(flame retardant)
PVC Plasticisers
Other (including sales
to formulators)
Total
Source: ERM (1999)
introduction of Directive 2002/45/EC which places marketing and use restrictions on
these applications.
In comparison to the above figures, worldwide production of chlorinated paraffins as
a whole is estimated at around 300 kt/year, with five producers of chlorinated
paraffins currently operating within the EU. It is believed that only two of these EU
manufacturers may be producing SCCPs at the current time, although non-EU
producers may be importing them into the EU.
Overall there has been a transition from the production of SCCPs to medium chain
chlorinated paraffins (MCCPs) (C14 – C17), as a result of voluntary industry action and
the regulation of their use in metalworking and leather processing (see also Section 4).
In 1999, MCCPs were estimated to account for over 80% of total chlorinated paraffin
production in the EU (excluding imports).
2.1.2 Use in Metalworking Fluids1
Metalworking fluids have two main functions: to cool and lubricate the tool/metal
interface; and to flush away ‘chips’ of cut metal. Chlorinated paraffins are highly
valued additives to fluids used for extreme pressure metalwork owing to their ability
to react with the metal surfaces (tool/metal being worked) at a molecular level,
creating a continuous lubricating layer. This is important as, when the tool and metal
meet at high temperatures, they can weld together reducing tool life and the quality of
the finished piece. SCCPs are particularly valued in extreme pressure processes
where rapid and severe metalwork is being undertaken, as they can be chlorinated to a
high degree (and so provide greater lubrication) and yet maintain a low viscosity
(unlike other paraffins which become almost solid at high chlorine levels).
Traditionally, extreme pressure additives include phosphorus, chlorine and sulphur
compounds, (with respectively increasing working temperature ranges) although there
1
Note that this discussion is based on RPA (1997).
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RPA & BRE
is disagreement over the precise temperature ranges at which they are effective.
Under extreme pressure (EP) conditions, the lubricating oil containing these elements
is believed to combine at the molecular level with the metal surfaces (tool/metal). It
is this reaction involving the phosphide (or phosphate), chloride or sulphide which
maintains lubrication between the two surfaces at EP. The type of fluid applied to an
operation is governed by the severity of the process and the compatibility of a fluid's
EP agents with the metal and tool being worked.
2.1.3 Use in Leather Processing
In leather processing, SCCPs are used as bulking agents in fat liquors which fatten
and soften the leathers. They are relatively cheap compared to some of the
alternatives but do not convey any fat liquoring properties themselves. Their main
advantages are that they are odour free and cost-effective; although in some cases and
for some types of animal skins it has been argued that they may offer better adhesion
and greater washability in comparison to natural oils.
Within the EU, SCCPs tended to be used in lower grade fat liquoring agents, and only
in some countries. Starting in 1994, there was a general trend away from SCCPs to
long chain chlorinated paraffins (LCCPs) and MCCPs (in order of importance) as
substitutes, as well as use of a number of natural animal and vegetable oils.
2.1.4 Use in the Paint Industry
SCCPs are used as plasticisers in paint resins, providing the base for more demanding
coating applications such as marine coatings and protective coating systems for steelwork and other applications exposed to aggressive industrial environments. Their use
is mainly restricted to acrylic based coatings, and may also be for the purpose of
imparting flame retardant, water proofing and chemical resistance properties.
The key properties of chlorinated paraffins in relation to this use are their insolubility
in water, chemical inertness, and extremely low volatility. Although LCCPs are more
generally used, SCCPs may be added to paint resins to soften and produce the paint
(ERM, 1999).
2.1.5 Use in Sealants and Adhesives
Within sealants and mastics, SCCPs are used mainly as a plasticiser to control the
elasticity and the hardness of rubbers. They are used as both a plasticiser and a flame
retardant additive in adhesives. The two key reasons for using SCCPs are that they
have a low level of leachability from the sealants and have a very low volatility. This
is important as it extends the lifetime of the sealant, to as much as 20 years, allowing
them to be used in building, industrial and automotive applications.
They are used in polysulphide (together with LCCPs), polyurethanes, acrylic sealants
and other polymer sealants (ERM, 1999).
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Case Study 2: SCCPs
2.1.6 Use in the Rubber Industry
Consumption of SCCPs by the rubber industry is very small (and may not take place
any longer) and is confined to their use as a flame retardant. They are used alone or
in combination with antimony trioxide to improve the flame retardancy of rubber and
other synthetic materials.
The main application is in rubber conveyor belts, hoses and tubes used in the mining
industry, where a high flame resistance is required by current safety standards.
2.1.7 Use in the Textile Industry
The use of SCCPs in the textile industry may be for one of three purposes:
•
•
•
as a flame retardant;
to confer water resistance; or
to confer rot prevention.
Across all three types of use, the quantities involved are small, with most use as backcoatings on textiles. The key types of materials are tent and sail cloth, tarpaulins, and
in the past military clothing.
2.1.8 Use in PVC Manufacture
SCCP usage in PVC manufacture is as a secondary plasticiser, which will also confer
flame retarding properties. They are generally used in conjunction with the phthalate
plasticisers to provide the additional level of flame retardancy required due to the high
flammability of the phthalates. Use of SCCPs in this sector, however, is expected to
be low as it has generally been replaced by the use of MCCPs.
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RPA & BRE
3.
ENVIRONMENTAL AND HUMAN HEALTH IMPACTS
3.1
Introduction
In order to understand the types of environmental and public safety benefits that may
be generated by REACH through the increased availability of information on
chemicals, a historical overview is given below of concerns and actions in relation to
SCCPs. This historical overview is not meant to provide a comprehensive summary
of scientific and other research concerning SCCPs nor is it intended to question or
validate research conclusions. Instead, the aim is to illustrate when concern first
arose, the types of risk issues that have been highlighted and raised in relation to
SCCPs, and how these concerns have been addressed either voluntarily by industry or
through regulatory and other more formal risk management measures.
The aim of this section is to:
1) review the scientific and academic literature to identify when research on different
hazardous properties began and when concern started to arise;
2) make chronological links between the scientific research and the introduction of
either voluntary or regulatory measures aimed at reducing risks to the
environment or to public health;
3) present monitoring data (where available) to illustrate the possible scale of
environmental damages that have occurred as a result of SCCP use; and
4) analyse the history of testing and risk management activities in relation to
properties of concern (persistence, bioaccumulation and toxicity, carcinogenicity,
etc.) and develop conclusions on the avoidable damages.
3.2
Development of Environmental and Health Concerns
3.2.1 1970 – 1979
Although SCCPs had been used in a number of applications since the early 1960s,
very little is found in the scientific literature about research into their health and
environmental effects until the 1970s. A high proportion of the studies conducted in
the 1970s, however, remains confidential and unpublished. The published
investigations, few as they were, indicate that the effects of SCCPs on the aquatic
environment were a subject of concern to the scientific community.
Early studies investigated the uptake, bioaccumulation and toxicity of chlorinated
paraffins. Lombardo et al (1975) investigated the bioaccumulation of chlorinated
paraffins in the rainbow trout and results showed initial evidence of the potential for
bioaccumulation of SCCPs when ingested. In 1979, Bengtsson et al conducted a
similar experiment, studying the uptake and accumulation of SCCPs by bleak (a
species of fish). Results showed that the level of chlorination of SCCPs had
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Case Study 2: SCCPs
implications for its accumulation, with the lower chlorinated grades showing the
greatest uptake over the exposure period. Linden et al (1979) then conducted further
tests on bleak examining acute toxicity effects of SCCPs to brackish water organisms.
At about the same time, the United States Environmental Protection Agency (US
EPA) conducted a toxicological investigation of selected potential environmental
contaminants, which included chlorinated paraffins (Howard et al, 1975).
One of the early human studies involved occupational exposure to SCCPs. In this
study, exposed employees were patch tested with various constituents of the cutting
fluids containing chlorinated paraffins. No positive reactions were reported for any of
the constituents (Menter et al, 1975). In general, the results on toxicity and
bioaccumulation could be said to have been inconclusive but were useful indicators of
potential issues to be associated with SCCPs in future.
Table 3.1 below gives a summary of the developments in the research on SCCPs in
the 1970s as discussed above. No regulatory or voluntary industry initiatives have
been identified for this period.
Table 3.1: Research and Regulatory Developments in the 1970s
1975
Lombardo et al (1975) investigate the bioaccumulation of chlorinated paraffins in
rainbow trout. Trout were fed a diet containing 10 mg/kg food of a C12 chlorinated
paraffin for 82 days. Results showed that the concentration of chlorinated paraffins
increased during the study, reaching a level of 1.1 mg/kg tissue (18 mg/kg fat) by the
time the study was terminated, with the equilibrium point not believed to have been
reached. Potential bioaccumulation of SCCPs noted.
1975
One of the early human studies involved occupational exposure to SCCPs was
conducted by Menter et al (1975). 134 non-exposed employees and 75 exposed
employees were patch tested with various constituents of the cutting fluids coolants
containing chlorinated paraffins. No positive reactions were reported for any of the
constituents, although the authors themselves suggested that the tests were not
sufficiently stringent.
1975
USEPA conducted an investigation of selected potential environmental contaminants
which includes SCCPs.
1979
Bengtsson et al (1979) conducted an experiment studying the uptake and accumulation
of several SCCPs by bleak (Alburnus alburnus). The concentration of chlorinated
paraffin in the bleak was measured by a neutron activation analysis method and results
showed that the uptake was greatest for the lower chlorinated grades over the exposure
period.
3.2.2 1980 - 1989
Investigations on the biodegradation and bioconcentration properties of SCCPs
continued in the 1980s. Tests on the biodegradability of SCCPs were carried out
using OECD guidelines, with results showing that 98% of the chlorinated paraffin
used in the experiment remained at the end. SCCPs were not readily or inherently
biodegradable (Street et al, 1983 in EC, 2000). Madeley and Maddock (1983) then
exposed rainbow trout to measured concentrations of SCCPs to determine the
bioconcentration potential of SCCPs in fish. Whole body bioconcentration factors
(BCFs) of over 7,500 were determined in fish ,with the BCFs found to increase with
decreasing exposure concentration. Further tests showed high levels of SCCP
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RPA & BRE
accumulation in the liver and viscera of the rainbow trout after exposure to measured
concentrations.
The first significant reports of measurements of chlorinated paraffins in the
environment appeared in this period. In many cases these did not distinguish the
chlorinated paraffins as short chain, medium chain etc., but tended to report wider
ranges of composition. Early indications of environmental exposure came from data
published in 1980, which compared chlorinated paraffin (C10-C20) levels in waters in
non-industrial areas to marine waters and industrial areas in the United Kingdom.
The concentration of chlorinated paraffins in the three types of water showed that the
maximum levels of chlorinated paraffins detected in industrial areas, albeit low, were
of the order of two to five times the levels in non-industrial and marine waters
(HELCOM, 2002).
Other data published the same year showed high levels of chlorinated paraffins
(C10-C20) (up to 12,000 µg/kg) in fish and mussels from the Wyre estuary close to a
paraffin production site in England. High levels of chlorinated paraffins were also
detected in seabirds (eggs), herons, guillemots, herring gulls, grey seal and in sheep
close to a chlorinated paraffin production plant in the United Kingdom (Campbell and
McConnell, 1980). By 1986, measurements of chlorinated paraffins (unspecified
chain length) in mammals in Sweden showed even higher concentrations than
measurements in 1980. Accumulation of SCCPs in rabbit and moose muscle was also
observed (Jansen et al, 1993).
Table 3.2 below gives a summary of the developments in research on SCCPs in the
1980s as discussed above. No regulatory or voluntary industry initiatives have been
identified for this period.
3.2.3 1990 – To Date
The 1990s saw a significant increase in the reported investigations into the effects of
SCCPs on health and the environment. There were early reports of skin sensitisation
and allergic reactions associated with use of SCCP based metalworking fluids.
However, it has since been shown that SCCPs do not have the potential to be skin
sensitisers and that other constituents of the fluids or stabilisers in the chlorinated
paraffin were responsible for the alleged reactions.
Further SCCP toxicity tests, however, soon established the liver, kidney and thyroid
to be the main target organs of repeated doses of SCCPs in mice and rats. Analysis of
the acute toxicity of chlorinated paraffins with differing chain lengths on Japanese
medaka eggs showed the C10 compounds to be more toxic than the C11, C12, and C14
homologues (Fisk et al., 1999). More advanced techniques for homologue specific
analyses of SCCPs for the lake trout showed calculated bioaccumulation factors of
between four to seventy times the EU criteria for bioaccumulation (BCF ≥ 2,000).
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Case Study 2: SCCPs
Table 3.2: Research and Regulatory Developments in the 1980s
1980
Data published in 1980 compared chlorinated paraffin levels in waters in nonindustrial areas to marine waters and industrial areas in the United Kingdom. The
concentration of SCCPs in the three types of water were 0.1-0.3 µg/l, 0.1-1µg/l and
0.1-2 µg/l respectively, with the water from industrial areas showing higher SCCP
levels than the non-industrial and marine waters (HELCOM, 2002). (Note the original
measurements were of C10-C20 chlorinated paraffins, the SCCP content was estimated
as part of the ESR assessment).
Other data published the same year show levels of chlorinated paraffins (C10-C20) of
up to 200 µg/kg in fish, 100-12,000 µg/kg in mussels, and above 200 µg/kg in
mussels from the Wyre estuary close to a paraffin production site in England. Other
measurements showed levels of 50-2,000 µg/kg in seabirds (eggs), 100-1,200 µg/kg
in heron and guillemot, 200-900 µg/kg in herring gull, 50-200 µg/kg in sheep close to
a chlorinated paraffin production plant and 40-100 µg/kg in grey seal being found in
the United Kingdom (Campbell and McConnell, 1980).
1981
1983
1983
1986
1986
1987
Campbell and McConnell (1980) also measured the average levels of chlorinated
paraffins found in human foodstuffs which showed levels up to 0.3 mg/kg in dairy
products, 0.15 mg/kg in vegetable oils and derivatives, and 0.005 mg/kg in fruit and
vegetables. Levels in shellfish close to sources of discharge of up to 12 mg/kg and in
meat of up to 4.4 mg/kg on a fat weight basis (the sample contained ~2% fat) have
been measured. While these values are low compared with the PNEC value of
16mg/kg, they may have implications where bioaccumulation and bioconcentration of
SCCPs occurs along a food chain.
Measurements of chlorinated paraffin levels in ringed seal blubber in Kongsfjorden,
Norway and the grey seal blubber from the Baltic Sea between 1979 – 85 showed
levels of around 130 and 280 µg/kg chlorinated paraffins on a lipid basis respectively
(Environment Canada, 2002).
Street et al (1983) investigated the biodegradability of SCCPs. Results showed that
SCCPs are not readily or inherently biodegradable.
Madeley and Maddock (1983) investigate the bioconcentration of SCCPs in rainbow
trout exposing them to measured concentrations of SCCPs for 60 days to find out the
bioconcentration factor. Whole body bioconcentration factors (BCFs) of 1,173-7,816
were determined based on radioactivity measurements in the fish and BCFs of 5747,273 were determined based on the parent compound analysis with the BCFs found
to increase with decreasing exposure concentration. Further tests showed high levels
of SCCP accumulation in the liver and viscera of the rainbow trout after exposure to
measured concentrations.
Measurements of chlorinated paraffin levels in surface water in rivers from industrial
areas in the United Kingdom showed low levels of 0.12-1.45 µg/l SCCP.
Measurements of chlorinated paraffin (chain length not specified) levels in mammals
in Sweden showed high concentrations of chlorinated paraffins, 2,900 and 4,400
µg/kg, respectively on a lipid basis, in rabbit and moose muscle (Jansen et al, 1993).
Levels of around 0.50-1.2 µg/l are reported in two rivers in Germany. The levels
measured in Germany in 1987 were similar to those found in the United Kingdom in
1986, although the maximum levels were slightly lower (HELCOM, 2002).
SCCPs were placed on the first priority list of substances to be assessed under ESR,
with the first draft risk assessment completed in 1995 by the UK Government (which
acted as rapporteur for the substance). The risk assessment resulted in the
classification and labelling of SCCPs as being dangerous to the environment
(R50/53), and identified risks to the environment, particularly the aquatic environment
from use in metalworking fluids and leather processing. Use in the remaining sectors
was found not to pose unacceptable risks to the environment (EC, 2000).
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RPA & BRE
Meanwhile, another issue arose as concentrations of SCCPs were being measured far
away from known sources. The detection of SCCPs in Arctic air, biota and lake
sediments and in the water column around the Bermuda Islands, in the absence of
significant sources of SCCPs in these regions, served as evidence that SCCPs were
being transported. Measurements in northern Canada and Norway were recorded
(Tomy et al, 1998; Borgen et al, 2000), while in England, measurements found high
levels remote from any known sources. For example, levels of SCCPs in aquatic
organisms from various regions, remote from any notable sources, were reported.
Marine mammals were also found to contain SCCPs. Tomy et al (2000) found
SCCPs in Beluga whales from northwest Greenland and Hendrickson Island, in
walrus from northwest Greenland and in ringed seal from southwest Ellesmere Island.
Analysis of SCCP concentrations in the liver and blubber of beluga whales from the
St Lawrence River estuary by Bennie et al (2000) showed high levels of SCCPs, with
the blubber levels being comparable to total concentrations of PCB and DDT
compounds.
Further studies showed that about 95% of the SCCPs in air samples were present in
the gas phase (Peters et al, 2000). One of the fundamental themes of subsequent
research (particularly in the Arctic regions) was that long range transport was a key
factor influencing SCCP availability in the environment and, as such, environmental
and health effects were liable to occur (and be observed) far away from their sources.
Bearing in mind that the main environmental source of human exposure to SCCPs is
food and, to a lesser extent, water, the presence of SCCPs in animals ultimately has
led to concerns over the health implications for humans. Levels in food in the range
of 30 µg/kg to several thousands µg/kg have been measured in different regions.
Total chlorinated paraffins in food, fish and marine animals have been reported with
levels measured (on a fat weight basis) ranging from 62 µg/kg to 963 µg/kg in
mackerel, fish oil (herring), margarine containing fish oil, common porpoise, fin
whale, pork, cows milk and human breast milk. SCCPs were thought to make up a
very small percentage of the total in mackerel, fish oil, porpoise and fin whale, but
around 7% in human milk, 11.5% in margarine, 21% in cows milk and 30% in pork
(Greenpeace, 1995). Calculations based on research figures put the maximum
estimated human intake (ignoring contributions from inhalation) as of the order of 20
µg/kg (body weight)/day, with the major contribution coming from fish/shellfish
(Campbell and McConnell, 1980, so relates to older data). Analysis of human breast
milk from Inuit women living in communities on Hudson Strait in Northern Quebec,
Canada showed on a lipid basis, levels of 10.6-16.5 µg/kg (Stern, 1998).
Regulatory and Voluntary Industry Initiatives
The 1995 ESR risk assessment was followed by the development of a risk reduction
strategy. This identified the following possible options for control: marketing and use
restrictions; classification and labelling; a voluntary agreement and limit values on
emissions to the aquatic environment. These options were assessed in some detail,
and it was concluded that the most effective policy option would be a Europe-wide
restriction on the marketing and use of SCCP-based metalworking fluids and leather
processing agents under Directive 76/769/EEC (RPA, 1997).
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Case Study 2: SCCPs
During a workshop held earlier in 1994, Euro Chlor developed a voluntary agreement
to reduce the use of SCCPs and this was put before PARCOM (the Paris
Commission). In 1995, however, PARCOM went further and proposed a total ban on
the use of SCCPs, based on the fact that they are persistent, toxic to aquatic organisms
and bioaccumulative in certain species and environments. Less environmentally
hazardous substitutes were also thought to be available for most major applications of
SCCPs.
In response to the voluntary agreement and PARCOM decision, production of SCCPs
was stopped by the German authorities in 1995; while in 1997 the Swedish authorities
passed a law (1997/98: 145) phasing out the use of SCCPs in all sectors by 2000.
This law effectively completed the process started in 1991, which phased out the use
of SCCPs in metalworking through a voluntary agreement and set targets for other
sectors.
In 1998, SCCPs were classified as Category 3 carcinogens by the European
Commission under Directive 67/548/EEC. Four years later, Directive 2002/45/EEC
was introduced, banning the use of SCCPs in metalworking and leather processing
across Europe entirely by 6 January 2004 at the latest. As part of the risk reduction
measures developed in response to these conclusions, it was decided subsequently to
review the other uses by the end of 2002 in the light of any new relevant scientific
data.
Table 3.3 below summarises the regulatory developments in relation to SCCP use
from the 1990s to 2002 as discussed above. Key research findings are also
summarised in the table.
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RPA & BRE
Table 3.3: Research and Regulatory Developments in the 1990s
1991
Voluntary Agreement with PVC sector to phase out SCCP use agreed in Denmark.
1994
Euro Chlor proposed voluntary agreement to phase out use of SCCPs in metalworking
fluids to PARCOM (Paris Commission).
1995
1995
1995
1995
1995
1997
1998
PARCOM and Euro Chlor jointly proposed an 80% reduction in the use of SCCPs in
metalworking from 1993 levels by the end of 1996 and elimination of the use of SCCPs in
metalworking by the year 2000.
PARCOM Decision (95/1) seeking a phase-out across all sectors of SCCPs use by year
2000 was agreed and signed by member countries.
Production of SCCPs stopped in Germany.
Preparation of an EU SCCP risk assessment under the Existing Substances Regulation.
Total chlorinated paraffins in food, fish and marine animals reported with levels measured
(on a fat weight basis) of 271 µg/kg in mackerel, 62 µg/kg in fish oil (herring), 98 µg/kg in
margarine containing fish oil, 16-114 µg/kg in common porpoise, 963 µg/kg in fin whale,
69 µg/kg in pork, 74 µg/kg in cows milk and 45 µg/kg in human breast milk (Greenpeace,
1995). Levels have implications for food chain and bioaccumulative effects.
The first draft risk assessment completed by the UK under the Existing Substances
Regulation (EEC 793/93), concluded that there were risks to the aquatic organisms from
SCCP use in metalworking fluids and leather fat liquors.
Sweden passed Environment Bill phasing out use of SCCPs in all sectors by 2000.
The detection of SCCPs in Arctic air, biota & lake sediments and in the water column
around the Bermuda Islands, in the absence of significant sources of SCCPs in these
regions served as evidence that SCCPs are being transported to these regions from other
places. Measurements of up to 8.5 pg/m3 in northern Canada, 9.0 – 57 pg/m3 at Norway,
and 65 – 924 pg/m3 in Egbert, Ontario were recorded (Tomy et al, 1998; Borgen et al.
2000) & in Lancaster, England, average levels were around 320 pg/m3 (Peters et al., 2000).
Analysis of fish (carp and Lake trout) from the Lake Ontario showed 59 – 2,600 ng
SCCPs/g wet weight whole fish, while yellow perch and catfish from the Detroit River
showed 1,100 and 300 ng SCCPs /g wet weight respectively (Muir et al. 2001; Tomy et al,
1997). Other high concentrations ranging from 100-770 µg/kg wet weight were found in
arctic animals (Environment Canada, 2002).
1998
1999
1999
2001
2002
Marine mammals also found to contain SCCPs, with concentrations measured in blubber
from beluga whales and walrus ranging from 110 to 1360 ng/g wet weight (Tomy et al,
2000; Stern et al. 1998). Analysis of SCCP concentrations in beluga whales from the St
Lawrence river estuary by Bennie et al. (2000) showed the liver and blubber contained 1.1
– 59 and 6.4 – 166 microgram SCCPs/g fresh weight respectively (the blubber levels being
comparable to total concentrations of PCB and DDT compounds).
SCCPs were classified as Category 3 carcinogens by the European Commission under
Directive 67/548/EEC.
International Maritime Organisation categorised SCCPs as Severe Marine Pollutant.
Analysis of the acute toxicity of chlorinated paraffins with differing chain lengths on the
Japanese medaka (Oryzias latipes) eggs showed the C10 compounds to be more toxic than
the C11, C12, and C14 homologues with further tests indicating the acute toxic mechanism to
be narcosis (Fisk et al., 1999).
Sampling exercise in the UK targeting locations thought to use chlorinated paraffins
(Nicholls et al, 2001). Individual types (short-chain, medium-chain) identified where
possible but in many cases this not feasible. In water, SCCPs identified at only one site,
near to a metal working facility. In sediments, MCCPs dominated most samples, SCCPs
identified near a chlorinated paraffin production site and a PVC/paint site. SCCPs
tentatively identified in biota near a sealants manufacturer, CP production site, and two
metalworking sites.
Introduction of Directive 2002/45/EEC banning use of SCCPs in metalworking use and
leather processing across Europe.
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Case Study 2: SCCPs
3.3
Key Properties and Presence in the Environment
3.3.1 Presence in the Environment
There has been very little measurement of SCCPs in the aquatic environment. BRE
(1995) found only two studies reporting on concentrations in the UK, both of which
were undertaken by ICI. Consultations undertaken during the development of the
risk reduction strategy found that neither the Environment Agency for England and
Wales nor any of the water and sewerage companies had undertaken any monitoring
or measurement of SCCPs in the aquatic environment at that time (RPA, 1997).
In 1986, ICI measured levels of SCCPs at 16 UK sites, most of which were in
industrial areas. SCCPs were identified at only nine of the sites, at concentrations
varying between 0.12 and 1.45 µg/l. At four sites, the concentration was higher than
0.50 µg/l (the predicted no effect concentration (PNEC) for aquatic organisms). An
earlier study measured the concentrations of chlorinated paraffins (CPs) with a chain
length ranging from 10 to 20 (i.e. some short and some medium). This study found
that:
!
in the marine environment, half the samples had detectable amounts of CPs in
water with concentrations being in the range 0.50 to 4.0 µg/l. With respect to
sediment, CPs were found in fewer samples (although the detection limit was
much higher at 50 µg/l compared with 5 µg/l for water) with concentrations being
in the range 50 to 100 µg/kg;
!
in fresh waters remote from industry, half the samples had detectable amounts of
CPs in water with concentrations being in the range 0.50 to 1.0 g/l. With respect
to sediment, CPs were found in fewer samples with concentrations being in the
range 300 to 1,000 µg/kg; and
!
in fresh waters close to industry, almost all of the samples had detectable amounts
of CPs in water and sediment. For water, concentrations were in the range 0.50 to
6.0 µg/l while for sediment, these were in the range 1,000 to 15,000 µg/kg.
Consideration of these and other measurements taken outside the EU led BRE to
conclude that typical measured concentrations of SCCPs are 0.05 to 0.30 µg/l in
waters in areas remote from industry and 0.10 to 2.0 µg/l in areas close to industry.
Modelling of the distribution of SCCPs in the environment has indicated that releases
to water are most likely to end up in the sediment or soil and this is confirmed by the
above measurements of SCCP levels in sediments which are approximately 100 to
1,000 times those in water. However, SCCPs may be slightly mobile in the
environment and so a small fraction of releases may be transported over a wide area
away from the sources of release (as suggested by the findings reported above with
regard to levels in the Artic and other areas remote from use).
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RPA & BRE
3.3.2 Key Properties
The assessment of environmental impacts for chemicals is based on measures of
persistence, bioaccumulation and toxicity and each of these is addressed below, along
with SCCPs potential for long range transport.
•
Persistence: SCCPs are quite stable in the environment, degrading very slowly
under bacterial action and binding strongly to soils and sediment long after their
entry into the environment (Street et al, 1983 in EC, 2000). Although there is
some evidence to suggest that SCCPs can degrade to a limited extent under some
circumstances, they are not readily biodegradable or inherently biodegradable.
They therefore meet the criterion for persistence in the ESR Marine Risk
Assessment guidance. They probably also meet the criterion for very persistent
(vP). The Canadian authorities are also currently pushing for SCCPs to be
recognised by the UN ECE as a persistent pollutant.
•
Bioaccumulation: SCCPs are very bioaccumulative with bioaccumulation factors
greater than 5,000 reported in a variety of freshwater and marine organisms. Tests
to ascertain the bioconcentration factor (BCF) (a measure of the chlorinated
paraffins levels present in fish compared to those present in water) have shown
SCCPs to have whole body BCFs of up to 7,816 for the rainbow trout. Later
research showed even higher BCFs ranging from 21,000 to 141,000 for specific
homologues and individual tissues. SCCPs are thus without doubt ‘very
bioaccumulative’.
•
Toxicity: SCCPs exhibit the highest toxicity of the polychlorinated n-alkanes and
are also widely accepted to be highly toxic to aquatic invertebrates and algae.
SCCPs are of low acute toxicity to fish with no significant effects up to solubility,
however in longer-term studies SCCPs are known to produce toxic effects.
Results from experiments indicate that SCCPs are specifically toxic to some
aquatic species (e.g. Daphnia) and may be of concern in areas where higher levels
of SCCPs are present. They are currently classified as R50/R53 chemicals – very
toxic to aquatic organisms and may cause long term adverse effects in aquatic
environment, with the EU risk assessment highlighting a need for specific
protective measures for the aquatic ecosystem from SCCPs.
•
Long Range Transport: Long-range transboundary atmospheric transport is
thought to be an important aspect of the global distribution of SCCPs and
responsible for their occurrence in remote areas.
3.3.3 Human Health Concerns
The main occupational health concerns associated with the use of SCCPs arise from
their application in metalworking fluids. Where oil mists may form as part of the
metal cutting activity, respiratory disorders may develop if worker protection
measures are not implemented. Metalworking fluid-related asthma cases in the UK
numbered around 30 in the early 1990s, although many more may have gone
unreported (Anon, not dated). It is not possible to attribute a proportion of these
specifically to SCCPs, as opposed to substitute extreme pressure additives or other
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Case Study 2: SCCPs
components of the fluids. However, occupational exposure to SCCPs is not
considered to present any risks provided that an Occupational Exposure Limit (OEL)
for oil mists (5 mg/ml in the UK) is met (RPA, 1997).
More generally, the implications of SCCPs in the environment for man are currently
unclear, as indicated earlier. The EU risk assessment report, however, concluded that
there were no significant risks to man exposed to SCCPs via the environment2. Since
then, the European Commission decided in 1998 to classify SCCPs as Category 3
carcinogens, with a risk rating R40 - possible risk of irreversible effects. They have
also been classed as carcinogenic under the Canadian Environmental Protection Act
1988; in the United States, the short chain (C12), 58% chlorine product is the only
chlorinated paraffin to be classified and labelled as a carcinogen; while in Germany,
the MAK Commission has classified virtually all chlorinated paraffins as Category 3B
(i.e. suspect carcinogens).
Box 3.1 shows the current EU classification and labelling for SCCPs.
Box 3.1: Current Classification of SCCPs
The current EU classification is:
Carcinogen Category 3: R40, with the symbol Xn; and
Dangerous for the Environment, with the symbol N
Risk phrases:
R40 – possible risk of irreversible effects
R50/53 – very toxic to aquatic organisms, may cause long-term adverse effects in the aquatic
environment.
Safety phrases:
S2 – keep out of the reach of children
S24 - Avoid contact with skin
S36/37 - Wear suitable protective clothing and gloves
S60 - This material and its container must be disposed of as hazardous waste.
S61 - Avoid release to the environment. Refer to special instructions/Safety data sheets.
2
Although laboratory experiments have shown SCCPs to cause tumours in rats and mice in widely
accepted methodologies, the mechanisms which cause tumours to be formed are specific to rodents
and of no relevance for human health.
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RPA & BRE
4.
THE REACH DOSSIER
4.1
Introduction
4.1.1 Overview
Developing dossiers for each of the case study chemicals effectively involves the
retrospective application of REACH. As it is necessary to project backwards in time,
these dossiers are hypothetical in nature, requiring that a series of assumptions are
made concerning:
•
•
•
•
•
production levels and associated uses for the manufacturer/consortia submitting
the dossier;
the level of information available to the manufacturer at time of dossier creation;
the substance-tailored testing that would be undertaken for completion of the
dossier (in line with Testing Option I as presented in Section 2 of the main report);
the assumptions that would be made concerning exposure and hence the
conclusions that would be reached regarding potential risks; and
the manner in which industry would respond to any conclusions concerning
environmental risks or risks to man via the environment with regard to risk
reduction activities.
The remainder of this section sets out the key assumptions and associated results for
the hypothetical dossier compiled on SCCPs. No details of the underlying studies are
included (see the full ESR risk assessment for further information on the underlying
studies).
4.1.2 Basic Assumptions
For the purposes of this case study, this registration relates to the production of shortchain chlorinated paraffins (SCCPs) and their use in the processing of leather. The
chlorinated paraffins used in this area generally have chlorine contents of 20-40%.
Data for SCCPs with other chlorine contents have been included in the dossier to fill
gaps where there are no data specific to the chlorination range indicated.
The substance is produced by the manufacturer preparing this dossier in a quantity of
370 tonnes per year. It has therefore been assumed that the dossier should follow the
basic information requirements for Dossier B, with additional information or
comments in relation to any additional substance tailored testing requirements.
In developing this hypothetical dossier, the following assumptions have been made:
•
•
the data that were available in IUCLID submitted to the European Chemicals
Bureau are assumed to have been available to the manufacturer at the start of
dossier preparation;
any further substance tailored testing must be undertaken in line with the
requirements set out for Dossier C (See Section 2 of the main report);
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Case Study 2: SCCPs
where site specific release data are not available, default data from the TGD and
within the EUSES and EASE models are applied; and
EUSES provides the basis for reaching conclusions as to whether or not
unacceptable risks result from a particular application or sector.
•
•
4.2
Base Data
4.2.1 Identity of the Substance
The CAS number 85535-84-8 is the relevant number for this substance; it relates to
alkanes with C10-13 chains, chlorinated. As such it covers a range of chlorine contents.
There are serious difficulties in measuring SCCPs in the environment. The substance
is a mix of components, with a range of chain lengths and differing chlorine contents.
The choice of substances to use as standards is crucial, as is the way in which the
response of the standard is related to that of the substance in the sample. Methods
have been reviewed and the most reliable identified. It is possible to measure down to
0.05 µg/l for water, although 0.50 µg/l is more commonly found. In sediment the
lowest value is 5 µg/kg, with 50 µg/kg being more common.
4.2.2 Physico-chemical properties
The basic physico-chemical data are presented in Table 4.1.
Table 4.1: Physico-chemical Properties
Property
Physical state at ntp
Pour point
Chlorine content
(% wt)
49-70
49
Vapour pressure (at
40oC)
Water solubility (at 20oC)
Log octanol-water
partition
coefficient
Flash point
Autoflammability
Explosivity
Oxidising properties
Page 2-18
Remarks
-
Clear to yellowish liquid
o
-30.5 C
>200oC
Boiling point (at ntp)
Density (at 25oC)
Value
Commercial mixtures nodistinct melting point
Decomposition with release
of hydrogen chloride
49-70
1.2-1.6 g/cm3
50
0.021 Pa
59
0.15-0.47 mg/l
With partial hydrolysis
49
4.39-6.93
50
166oC
Measured by a high
performance thin layer
chromatography method
Closed cup
Not stated
Not explosive
None
Decomposes with liberation
of hydrogen chloride above
200oC
RPA & BRE
4.2.3 Ecotoxicity
Acute Daphnia Toxicity
No data are available for SCCPs that are 20-40% chlorine by weight. The closest
chlorine content for which data are available is 56%. For this SCCP, a 24 hour EC50
of 0.44 mg/l was established. Other values are available for other chlorine contents.
Degradation
No degradation was observed in a ready biodegradability test. In an inherent
degradability test, 7.4% and 16% degradation were observed at concentrations of 50
and 25 mg/l respectively. Thus, SCCPs do not meet the criteria for inherent
degradability. Aniline added to this test was degraded as normal, so these
concentrations (which are well above the solubility limit) are not toxic to microorganisms.
In a sewage treatment plant simulation test (coupled units), 93% removal was seen but
was considered to be due to adsorption to sludge.
There are some indications from non-standard studies that micro-organisms which
have been previously acclimated to lower chlorinated SCCPs may be able to degrade
them to some extent.
Growth Inhibition in Algae
For freshwater algae, an EC50 of 1.3 mg/l and a NOEC of 0.39 mg/l were determined.
The study was over 10 days, and the EC50 value is extrapolated above the highest
measured concentration achieved in the test (1.2 mg/l).
A further test had to be carried out to provide information data for a seawater based
NOEC. This was carried out for saltwater algae. An EC50 of 0.043 mg/l and a NOEC
of 0.012 mg/l were determined in a 96 hour test.
Fish Acute Toxicity
The available results all quote values which are well in excess of the solubility. It is
assumed that the substance has no acute effects at solubility.
Adsorption/Desorption
There are no data available. The log Kow value indicates a high potential for
adsorption. The sorption of the substance is likely to be well predicted by the log
Kow value, therefore no specific testing for adsorption is considered necessary. The
calculated values are used in place of the test.
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Case Study 2: SCCPs
4.2.4 Dossier C Considerations in Addition to Dossier B
Chronic Daphnia Toxicity
A NOEC of 0.05 mg/l was determined in a 21-day test using an SCCP with 20%
chlorine content. This study was carried out by the submitter in view of the acute
(short term test) toxicity results with Daphnia.
Higher Plant Test
No data located. The need for such testing will be discussed on the basis of the results
of the risk assessment.
Acute Earthworm Tests
No data located. The need for such testing will be discussed on the basis of the results
of the risk assessment.
Further Fish Studies
A 20-day NOEC of <40 µg/l has been determined for fish (for progressive loss of
motor function).
The need for any further chronic fish studies will be discussed with the authorities,
depending on whether sufficient exposure is expected.
Accumulation
Measured values of 800-1,000 have been reported in the literature. These reports
suggest that lower chlorinated substances have higher accumulation factors. The
higher end of the range of values will be used for this submission.
Supplementary Degradation Information
No further data. The data reported above on degradation is considered to be sufficient
for the purposes of this submission.
Further Sorption/Desorption
No additional comment to that above.
4.3
Exposure
Production site emissions have been monitored, and estimated releases to water are
low (<10 kg/year). Sludges from water treatment are incinerated.
No specific information on releases from formulation or use in leather is available.
Instead, the methods in the Technical Guidance Document (TGD) have been used to
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RPA & BRE
estimate potential concentrations as a result of these operations. Default values for
emissions and amounts used on sites have been used throughout, with the exception of
the amount formulated at a site which has been taken from customer information.
The resulting predicted environmental concentration (PEC) values are given in Table
4.2.
Table 4.2: PEC Values for Formulation and Use in Leather
Formulation
0.088
Sediment
(mg/kg)
69
Processing
0.018
14
74
4.2
0.037
0.35
-
Water (mg/l)
-5
Regional
2.8x10
Soil (mg/kg)
Fish (mg/kg)
364
15.5
In addition to the concentrations in fish, the concentration in worms was also
estimated to provide an assessment of secondary poisoning. However, these gave
very high values, in the order of 3 g/kg. These are not considered to be realistic. It is
therefore argued that, for the purposes of this dossier, the method used to estimate
them may not be appropriate for substances with high log Kow values. The
concentrations in worms are not considered further in this assessment.
4.4
Risk Assessment
From the aquatic toxicity data presented, there are three long term NOEC (no
observable effects concentration) values. Following the approach laid out in the
TGD, a factor of 10 is therefore applied to the lowest (the NOEC for marine algae, at
0.012 mg/l). The resulting predicted no effects concentration (PNEC) is therefore
derived as 1.2 µg/l.
There are no data for soil or sediment, so the equilibrium partitioning method is used
for these endpoints. Test data on birds are not indicated at this level but, as the
substance is accumulative, some information on this has been located. A NOEC of
166 mg/kg in food is reported for a chronic feeding study with Mallard ducks looking
at effects on reproduction. The use of an assessment factor of 10 is indicated in the
TGD, giving a PNEC of 16.6 mg/kg food.
The PEC/PNEC ratios are presented in Table 4.3. A ratio greater than unity (1.0)
indicates a risk of concern, with the end-points for which such a ratio exists
highlighted in bold.
Table 4.3: PEC/PNEC Ratios
Water
Sediment
Soil
Formulation
74
740
831
Secondary
poisoning
0.93
Processing
15
150
170
0.25
0.023
0.23
4.6
-
Regional
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Case Study 2: SCCPs
Exposures from production are low and so no risks are anticipated (with no
corresponding values presented in table 4.3). The sediment and soil ratios in the
Table have been increased by a factor of 10 as indicated in the TGD to account for
possible uptake through ingestion (as the substance has a log Kow value above 5).
4.5
Risk Management Recommendations
4.5.1 Conclusions from the Risk Assessment
SCCPs will be classified as PBTs in relation to the marine environment, and it is
recognised that this will lead to their going to Authorisation.
However, as part of any further risk assessment work, the emission estimates should
be improved for both formulation and processing. Information on releases to water
and on the fate of sludge from the waste water treatment plant would be especially
useful. As the risk characterisation for sediment and soil is based on the equilibrium
partitioning method, and has an extra safety factor of 10, testing on sediment and soil
organisms would be likely to refine the assessment.
4.5.2 Recommended Further Testing or Risk Assessment Activities
Based on the above findings, the following activities should be undertaken:
•
Emissions monitoring: further monitoring of discharges to water should be
undertaken at downstream user locations (formulation and processing) to improve
the quality of the data used within the exposure assessments. It is expected that
this will clarify whether additional emissions control technology is required at
these downstream user sites; and
•
Testing: further testing on sediment and soil organisms should be undertaken to
refine the PNEC values being used in the risk assessment.
4.5.3 Further Risk Management Measures
As SCCPs are a PBT substance, it is recognised that further risk management will be
required, however, the nature of this should take into account the conclusions of the
above monitoring and testing data. The proposed measures for adoption within
Authorisation are as set out below.
The need for action at downstream user sites that formulate SCCPs for use in leather
processing agents (fat liquors) is dependent upon the types of measures that could be
adopted by leather processing facilities. As a worst-case, smaller leather processing
sites may not be able to introduce additional emissions control technology and may
have to switch to alternative fat liquor agents.
1) If the further monitoring activities confirm the emission rates assumed in the
exposure assessment, then additional emissions controls would be required at
formulation sites giving rise to PEC values above the PNEC values should use
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RPA & BRE
continue in leather processing. However, whether such sites adopt controls will
be dependent on the measures required of leather processors.
2) For leather processing, the adoption of emission controls may not be cost-effective
for smaller facilities. In these cases, it is proposed that these downstream users
substitute SCCPS with either non-chlorinated processing fat liquor agents or
longer chain length chlorinated paraffins (LCCPs: C18 - C20 (liquid)). Note that
MCCPs do not provide an appropriate substitute given the conclusions of their
dossier with regard to use in leather fat liquors. The dossier for LCCPs is not yet
complete. However, use of LCCPs in leather fat liquors used by sites with
inadequate emissions control should only occur if the dossier for LCCPs
concludes that there is no environmental risk associated with their use by this
sector.
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Case Study 2: SCCPs
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RPA & BRE
5.
THE REACH DOSSIER CONSIDERED
5.1
The Evaluation Approach
The aim of developing the hypothetical dossiers is to provide a basis for comparing
what might have happened had REACH been introduced earlier with what happened
under the existing regime. In order to do this, we discuss below whether REACH
would:
•
•
•
•
•
5.2
require the same level of test data as required under ESR or other regulatory
regimes;
raise any concerns for the example substance and, if so, for which endpoints and
risk compartments;
identify the same endpoints and risk compartments as those identified
(historically) and controlled by the existing legislative arrangements;
recommend through this retrospective application, similar risk reduction measures
to those implemented at present; and
lead to action being taken sooner than under the current system and hence reduce
levels of environmental damage and risk to man via the environment.
The ESR Risk Assessment
5.2.1 Conclusion of the ESR Assessment
The risk assessment carried out under ESR considered not only the risks associated
with the production of SCCPs and their use in leather processing industry as in the
dossier, but also all of the other downstream uses of SCCPs discussed in Section 2.
The conclusions of the ESR assessment are summarised in Table 5.1.
Table 5.1: Conclusions of the ESR Risk Assessment
Scenario
Water
Sediment
Soil
Secondary
poisoning
Production
(ii)
(i)
(ii)
(ii)
Metalworking (formulation)
(iii)
(i)
(i)
(ii)
Metalworking (use)
(iii)
(i)
(i)
(iii)
Rubber formulations
(ii)
(i)
(ii)
(ii)
Paints and sealing compounds
(ii)
(ii)
(ii)
(ii)
Leather formulation
(iii)
(i)
(i)
(iii)
Leather fat liquors
(processing)
(iii)
(i)
(i)
(iii)
Textile applications
(ii)
(ii)
(ii)
(ii)
Regional
(ii)
(i)
(i)
(ii)
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Case Study 2: SCCPs
Risks were identified for the aquatic compartment and for secondary poisoning for:
•
•
the formulation (aquatic only) and use of metal working fluids; and
the formulation and use of SCCPs in leather.
For most scenarios and for sediment and soil, a conclusion (i) was reached, with this
indicating that further information on emissions was required and that testing on
sediment and soil organisms was needed. However, the risk reduction measures
required as a result of the conclusion (iii) findings for the aquatic compartment would
be expected to have an impact on the risk assessments for sediment and soil. It was
therefore concluded that further monitoring and testing work should await the
outcome of risk reduction proposals.
The conclusions of the REACH dossier are very similar for those life cycle scenarios
it covers, in that possible risks are indicated for the aquatic, sediment and soil
compartments for formulation and use in leather processing activities. The key
difference is that no risk from secondary poisoning from these uses is identified in the
REACH dossier. This is due to the lower bioconcentration factor used in this
assessment. The BCF values used in the ESR assessment came from studies that were
internal to one industry producer, and it was assumed for the purposes of this
hypothetical dossier that they would not be made available (in the first instance) to the
particular company submitting this dossier. They may be made available later as part
of information sharing, or they may not become available if the company holding that
information does not also wish to register a dossier for SCCPs.
The recommendations within the REACH dossier, that more information on
discharges be sought for the aquatic compartment, were also reached in the ESR
assessment at an early stage, but no better information was provided and so a
conclusion (iii) was reached for the aquatic compartment. If it were supposed that no
further information would be provided for REACH, then the conclusion from REACH
would also be (iii) for these endpoints, as the PNEC cannot be revised upwards.
5.2.2 Hazardous Effects and Routes of Exposure
Aquatic Compartment
The ESR assessment reviewed the available data on aquatic toxicity. Results were
available from long term studies on species from three trophic levels, so an
assessment factor of 10 was used to derive a PNEC of 0.50 µg/l. Emissions from the
production of SCCPs were assessed using site specific information. Releases from
formulation and use in leather were estimated using mainly default values, with some
information about the quantities used. The assessment indicated risks to surface water
and to sediment. The initial conclusion was that further information was needed,
relating to exposure and testing. No further information on exposure to the aquatic
compartment was received, and so the conclusion for surface water was revised to one
of risk needing to be reduced. For sediment, testing on sediment organisms could
refine the PNEC, as the current value is based on the equilibrium partition method,
and so a conclusion (i) was retained. Similar conclusions were reached for
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RPA & BRE
formulation and use in metalworking fluids, but these are not relevant to this REACH
dossier.
The data set for aquatic toxicity in REACH is smaller, as it was compiled by only one
manufacturer. Other data in the ESR assessment came from company confidential
studies; such studies may or may not be available to all manufacturers submitting a
dossier depending on the manner in which consortia are formed and what
requirements for information sharing are finally included within REACH.
The PNEC in the REACH dossier is 1.2µg/l; this is a little higher than that from the
ESR assessment, but the change has little effect on the conclusions (no ratios which
are above one in the ESR assessment are below one in the REACH assessment). The
risk characterisation in REACH indicates possible risks from formulation and use in
leather. The initial conclusion is that further information should be sought for
releases from these areas, and possible testing on sediment organisms. If it were
assumed that no further exposure data would be forthcoming, then the conclusion
from REACH would be the same as that from ESR – there would be a risk requiring
control for surface water from the formulation and use in leather. A possible risk to
sediment could be revised through testing, but it would be appropriate to await the
outcome of decisions on risk management for emissions to surface water first.
Terrestrial Compartment
No studies of effects on terrestrial organisms were identified for the ESR study, and
so the equilibrium partition approach was used, giving a PNEC of 0.80 mg/kg. The
risk characterisation ratios were increased by a factor of 10 to account for possible
ingestion of soil (a similar adjustment was made to the sediment ratios). The ESR
assessment concluded there was no risk to soil from production sites. Ratios were
above one for formulation and use in leather. The assessment concluded that further
information was needed, but that this need should await the outcome of the risk
reduction measures identified for the aquatic compartment, as these were likely to
have an impact on the terrestrial assessment. The same conclusion was reached for
metalworking fluids.
The conclusions for REACH are similar. The PNEC is a little different as the aquatic
PNEC is different, but this has no effect on which ratios are above one. As for ESR,
the ratios are increased by a factor of 10. Also, as for ESR, production does not pose
a risk, but formulation and use in leather indicate possible risks. The conclusion is
that further information on exposure should be obtained, with the possibility of testing
on soil organisms. It is noted that information provided to refine the aquatic exposure
assessment is also likely to have an impact on the terrestrial assessment. A particular
question for the terrestrial compartment is the fate of sludges from waste water
treatment plants receiving waste water from the leather industry.
Secondary Poisoning
A PNEC of 16 mg/kg was derived in the ESR assessment. Formulation and use in
leather indicated a risk, but production did not. A risk through this route was
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Case Study 2: SCCPs
concluded, as no further information on aquatic exposures was provided which would
have allowed the PEC in fish to be revised.
The PNEC for REACH is the same. The concentrations in fish estimated in REACH
are lower than those in the ESR assessment, as the data on bioconcentration in the
dossier is limited (other values for the ESR assessment came from company
confidential studies). The conclusion from the REACH dossier is that there is no risk
for secondary poisoning from formulation and use in leather.
5.2.3 The ESR Risk Reduction Strategy
One of the outcomes of the risk assessment was the classification and labelling of
SCCPs as being dangerous for the environment (R50/53). The second outcome was
the preparation of a risk reduction strategy (RPA, 1997). The risk reduction strategy
for leather processing considered a range of different options for managing the risks
associated with the use of SCCPs. The options were assessed in detail against the
criteria used within ESR of effectiveness, practicality, economic impact and
monitorability. The conclusions for leather working were as follows (RPA, 1997):
•
Classification and labelling of formulations containing SCCPs as dangerous for
the environment: this was not considered a feasible regulatory option at the time,
as the Directive on the classification and labelling of preparations had not yet been
introduced. Furthermore, industry had voluntarily labelled SCCPs as ‘dangerous
for the environment’, with this leading to no decline in sales.
•
Limits on emissions: at the time the risk reduction strategy was first developed,
SCCPs would have had to be classified as either a List 1 or List 2 substance under
the Framework Directive (76/464/EEC) on pollution caused by certain dangerous
substances. Although this would have been feasible, there was concern that, due
to the nature of the leather processing industry, a significant proportion of
discharges would be ineffectively controlled leading to on-going environmental
damages.
•
Marketing and use restrictions: the final option was the introduction of marketing
and use restrictions under Directive 76/769/EEC. This ended up being the
recommended strategy, not only because it was deemed to be the most effective in
controlling the risks to the environment, but also because the leather processing
industry was already moving away from the use of SCCPs and indicated that the
costs of a ban would not have a significant effect on those companies using
SCCPs.
The recommendations of this strategy have since been implemented in Directive
2002/45/EEC, which bans the use of SCCPs in both leather processing and
metalworking and leather finishing from late 2003. The Directive also requires that
the European Commission reviews all remaining uses of SCCPs by 6 January 2004.
Given that SCCPs probably meet the PBT criteria for EU marine risk assessments,
this group of substances would go to Authorisation, and be most likely to face
marketing and use restrictions given the availability of substitutes. Thus, there are
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unlikely to be significant differences as to what would be adopted as risk management
under REACH compared with ESR. Furthermore, given the historic trend in use,
manufacturers themselves may have proposed this outcome if it did not appear
feasible for emission controls to be adopted across the board by leather processors.
Because the leather processing industry was already moving away from SCCPs, the
user sector would probably have preferred to switch to substitutes rather than adopt
additional emissions control. However, a key factor in the leather processing sector’s
decision making in this regard related to the availability of substitute agents for use in
fat liquors. In particular, these were (and are) LCCPs and MCCPs. At the time, the
test data available for MCCPs and LCCPs suggested that they might be of lower
toxicity and hence pose lower risks than SCCPs.
5.2.4 General Conclusions
Overall, one can conclude that a REACH dossier prepared earlier in time is likely to
have reached the same conclusions with regard to risks to the environment and the
need for some action to be taken. This is despite the fact that a more limited data set
was used for some of the end-points in the dossier. One would also expect a similar
dossier relating to the use of SCCPs in metalworking fluids to conclude that they were
PBTs in the marine environment and these presented risks to the aquatic environment
and sediment. This could have had a more dramatic effect in terms of reducing
impacts on the environment.
Perhaps the difference under REACH is that manufacturers would be more quickly
faced with the question of whether or not they wished to undertake any further testing
required to resolve conclusion (i) situations (particularly where other tests indicate
that a substance is a PBT). One could imagine that such decisions will be taken on
the basis of a range of factors, including:
!
!
!
!
5.3
the risk management measures that they themselves will put forward to control
any risks arising to related compartments (e.g. aquatic in the case of a sediment
conclusion (i));
the value of the market for the substance in the application giving rise to the risks;
the costs of the further tests in relation to the value of sales to the relevant market
sectors; and
any concerns over the implications that risk management in relation to one use
may have for the perception of the substance and, hence, its other markets.
Historical Damage Costs Avoided
Based on the above conclusions, it is apparent that had REACH been in place sooner,
risks to the environment and potentially to man via the environment from SCCPs
could have been significantly reduced. While the damage costs associated with the
use of SCCPs are less easy to measure, the toxic, bioaccumulative and possible long
range transport effects, all result in avoidable damage costs.
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Case Study 2: SCCPs
A further consideration of the chronology of research on SCCPs (see Section 3)
highlights two stages at which definitive action could have been taken and
considerable damage costs avoided. These are when:
!
in 1980, high levels of SCCPs were being detected in seabirds (eggs), herons,
guillemots, herring gulls, grey seal, sheep and other mammals, in addition to being
found in the environment; and
!
when investigations into the toxicity and bioconcentration properties of SCCPs in
the 1970s and 1980s, found whole body bioconcentration factors (BCFs) of over
7,500 in fish, showing very bioaccumulative tendencies.
It could be argued that under REACH, the data and information that was held
unpublished through part of the 1970s and 1980s by producers of SCCPs could have
been released under the Registration stage of REACH. This combined with the need
to prepare a Chemical Safety Assessment (CSA) would have highlighted the need for
an authorisation to be sought for the use of SCCPs by downstream users. Given that
SCCPs probably meet the PBT criteria for EU marine risk assessments, they would
most likely have faced marketing and use restrictions earlier, given the availability of
substitutes.
Even in a scenario in which scientific evidence was incomplete and inconclusive,
some of the risk management initiatives taken in the 1980s and 1990s could have been
picked up and possibly come under a more integrated REACH process, given the
harmonised structure of the working relationship between national authorities under
REACH. These initiatives include the:
•
voluntary industry action first taken in relation to SCCPs in metalworking fluids
in the mid to late 1980s, and carried on through the 1990s;
•
the introduction of the Swedish Bill 90/91, which sought a phase-out of SCCP use
in the metalworking sector to be followed by a phase-out in other sectors over a
slightly longer time period; and
•
the proposals for a voluntary agreement to phase-out the use of SCCPs in
metalworking fluids put forward by Euro Chlor in 1994.
Taken together, the above suggests that had the type of regime to be introduced by
REACH been in place earlier, the environmental impacts arising from the use of
SCCPs could have been minimised considerably. In particular, levels of SCCPs
found in the Arctic and other locations distant from sites of use might be significantly
lower. This is based on the fact that REACH has the potential to react faster to the
need to minimise risks from hazardous chemicals as soon as any possible effects are
noted. It could thus be argued that the testing required to determine whether or not
SCCPs are a persistent organic pollutant in relation to long range transport would
have been completed by now, rather than being an on-going subject of debate at the
international level.
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Overall, it may take years for the full damage costs arising from the use of SCCPs in
the applications of concern to be realised. For example, a study by Stevens et al
(2003) found particularly high concentrations of SCCPs and MCCPs within sewage
sludge (ranging between 7 to 200 mg/kg dm and 30 to 9700 mg/kg). Although no
limits are currently proposed for chlorinated paraffins within sludge, the authors note
the potential for concern (particularly as many uses will not be restricted under
Directive 2002/45/EEC).
5.4
Substitution Issues
Another set of damages could be avoided under REACH, with these relating to the
damages incurred when environmentally harmful substances are used as substitutes
for other harmful substances, as was the case for SCCPs.
SCCPs were used in leather processing as bulking agents in fat liquors, particularly in
lower grade fat liquoring agents, to fatten and soften the leathers. At the time that the
risk reduction strategies were produced for leather processing and metalworking,
MCCPs and LCCPs were both considered on the basis of available scientific
information to pose lower risks than SCCPs. Although the strategies recognised that
there was some uncertainty as to whether MCCPs were more or less toxic than
SCCPs, the consensus was that they were likely to be less toxic.
This view will have resulted in many leather processors and metalworking facilities
shifting to the use of these other CPs. In the case of MCCPs, this is unlikely to have
resulted in a significant reduction in risks to the environment, as the draft ESR risk
assessment for MCCPs has concluded that (Environment Agency, 2000):
•
formulation of metal cutting fluids poses unacceptable risks to the aquatic
environment;
•
use of MCCPs in emulsifiable metal cutting/working fluids where spent fluid is
discharged to waste water presents unacceptable risks to the environment; and
•
use in leather fat liquors presents unacceptable risks to the aquatic environment.
The conclusions also hold in relation to secondary poisoning and are strengthened if
Directive 2002/45/EC has led to an increase in the use of MCCPs.
Because REACH requires the provision of data on all substances produced in volumes
over 1 t/y, this problem of substitution with other damaging substances should be
reduced in the short term and eliminated in the medium term (i.e. 10 to 12 years).
Since information will be available on the risks posed by the substitutes and of the
appropriate risk management measures required to address any risks, downstream
users will be able to take better informed decisions when selecting substitutes or
considering processing changes. The result should be an overall reduction in risks to
the environment and man.
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6.
REFERENCES
Bengtsson (1979): Structure Related Uptake of Chlorinated Paraffins in Bleaks (alburnus
Alburnus L), Ambio, Vol 8, pp121-122.
Bennie DT et al (2000): Occurrence of Chlorinated Paraffins in Beluga Whales
(Delphinapterus leucas) from the St. Lawrence River and Rainbow Trout
(Oncorhynchus mykiss) and Carp (Cyprinus carpio) from Lake Ontario, Water Qual.
Res. J. Canada, Vol 35, pp263-181.
Borgen AR et al (2000): Polychlorinated Alkanes in the Arctic Air, Organohalogen
Compounds, Vol 47, pp272-275.
Campbell I & McConnell G (1980): Chlorinated Paraffins and the Environment. 1.
Environmental Occurrence, Environ. Sci. Technol., Vol 9, pp1209-1214.
Environment Canada (2002): Short-Chain Chlorinated Paraffins (SCCP) Substance
Dossier (draft), Prepared for UNECE ad hoc Expert Group on POPs, March 2002.
European Commission (2000): European Union Risk Assessment Report. Alkanes, C1013, chloro. CAS No. 85535-84-8, EINECS No. 287-476-5. European Chemicals
Bureau, Institute for Health and Consumer Protection. 1st Priority List, Volume 4.
ERM (1999): Study on the Economic and Social Implications of Introducing
Community-wide Restrictions on the Marketing and Use of Short Chain
Chlorinated Paraffins, Draft Final Report, European Commission DGIII.
Fisk AT et al (1999): Toxicity of C10-, C11-, C12-, and C14-polychlorinated Alkanes to
Japanese Medaka (Oryzias latipes) Embryos, Environ Toxicol Chem, Vol 18,
pp2894-2902.
Greenpeace (1995): Greenpeace Zur Sache: Chlorparaffine, May 1995.
HELCOM (2002): Implementing the HELCOM Objective with regard to Hazardous
Substances. Guidance Document on Short Chained Chlorinated Paraffins (SCCP),
downloaded
from
the
HELCOM
Internet
site
(http://www.helcom.fi/land/hazardous/sccps.pdf).
Howard PH et al (1975):
Investigation of Selected Potential Environmental
Contaminants: Chlorinated Paraffins, United States Environmental Protection
Agency, Report EPA-560/2-75-007.
Jansen et al (1993): Chlorinated and Brominated Persistent Organic Compounds in
Biological Samples from the Environment, Environ. Toxicol. Chem., Vol 12, pp11631174
Linden et al (1979): The Acute Toxicity of 78 Chemicals and Pesticide Formulations Against
two Brackish Water Organisms, the Bleak (Alburnus alburnu)s and the Harpacticoid
(Nitocra spinipes), Chemosphere, Vol 11/12, pp843-851.
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Case Study 2: SCCPs
Lombardo P et al (1975): Bioaccumulation of Chlorinated Paraffin in Fish fed Chlorowax
500C, J. Assoc. Off. Anal. Chem, Vol 58, pp707-710.
Madeley JR & Maddock BG (1983): Toxicity of a Chlorinated Paraffin over 60 Days. (iv)
Chlorinated Paraffin - 58% Chlorination of Short Chain Length n-paraffins, ICI
Confidential Report BL/B/2291.
Menter P et al (1975): Patch Testing of Coolant Fractions, J.Occu Med., Vol 17 No 9,
pp565-568.
Muir D et al (2001): Short Chain Chlorinated Paraffins: Are they Persistent and
Bioaccumulative?, ACS Symposium Series, Vol 773, pp184-202.
Nicholls CR et al (2001): Levels of Short and Medium Chain Length Polychlorinated nAlkanes in Environmental Samples from Selected Industrial Areas in England and
Wales, Environ. Pollut., Vol 113, pp415-430.
Peters et al (2000): Occurrence of C10-C13 Polychlorinated n-Alkanes in the Atmosphere of
the United Kingdom, Atm Environ, Vol 34, pp3085-3090.
RPA (1997): Risk Reduction Strategy on the Use of Short-Chain Chlorinated Paraffins
in Leather Processing Final Report - Dec. 1997. Prepared for Department of the
Environment, Transport and the Regions, United Kingdom.
Schlabach, M et al (2001): Polybrominated Diphenyl Ethers and Other Persistent Organic
Pollutants in Norwegian Freshwater Fish, presented at the 11th Nordic Conference on
Mass Spectrometry, Loen, Norway, 18-21 August 2001.
Stern et al (1998): Polychlorinated n-alkanes in Aquatic Biota and Human Milk, presented at
the American Society of Mass Spectrometry and Allied Topics, 45th Annual
Conference, Palm Springs, California.
Stevens J et al (2003): PAHs, PCBs, PCNs, Organochlorine Pesticides, Synthetic Musks and
Polychlorinated n-Alkanes in UK Sewage Sludge: Survey Results and Implications,
Environmental Science Technology, Vol 37, pp462-467.
Tomy GT et al (1997): Quantifying C10-C13 Polychloroalkanes in Environmental Samples
by High Resolution Gas Chromatography/Electron Capture Negative Ion Mass
Spectrometry, Anal Chem, Vol 69, pp2762-2771.
Tomy GT et al (1998): Environmental Chemistry and Toxicology of Polychlorinated nAlkanes, Rev Environ Contam Toxicol, Vol 158, pp53-128.
Tomy GT et al (2000): Levels of C10-C13 Polychloro-n-alkanes in Marine Mammals from
the Arctic and St. Lawrence River estuary, Environ Sci Technol, Vol 34, pp16151619.
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CASE STUDY 3:
TETRACHLOROETHYLENE (PERC)
Case Study 3: Tetrachloroethylene
RPA & BRE
1.
INTRODUCTION
1.1
Background to the Case Study
Tetrachloroethylene (also known as perchloroethylene or perc) is a persistent and
toxic substance used as an intermediate in the chemicals industry, in the manufacture
of some consumer products, and in the workplace.
Perc has been selected as a case study not only for its persistence and toxicity, but
also because of its regulatory history in relation to use and, more particularly,
disposal. Other reasons for the selection of perc as a case study are based on the fact
that:
1.2
•
in contrast to the other case studies, it raises public health and worker safety
issues;
•
its historic risks to the environment are of a different nature to those covered by
the other case studies, in that the more significant damages relate to long-term
contamination of water resources;
•
a wide range of different regulatory controls have been placed on it over time; and
•
in the environment it breaks down into equally damaging substances.
Format of Case Study
A profile of the market for perc within the EU is provided first (Section 2), with this
including a brief description of its key uses. This is followed in Section 3 by a brief
overview of the health and environmental data that led to perc becoming a concern.
As part of this discussion, an overview of the various regulatory actions that have
been introduced over time to minimise risks are presented.
The hypothetical REACH dossier that has been prepared for perc is presented in
Section 4. In preparing this dossier, we have drawn upon data and experience from
preparing the ESR risk assessment for perc. The dossier is then considered further in
Section 5, which compares it (the dossier) to the findings of the ESR process. Further
discussion is provided on the damages that could have been avoided had REACH
been in place earlier.
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Case Study 3: Tetrachloroethylene
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RPA & BRE
2.
MARKET PROFILE
2.1
Uses and Trends
2.1.1 Overview
Perc is produced jointly with trichloroethylene by the TRI/PER process, which is
based on the chlorination or oxy-chlorination of the light fractions of residues from
vinyl chloride monomer manufacture. It can also be produced jointly with carbon
tetrachloride (CTC/PER process) (ECSA, not dated).
Table 2.1 shows the change in the European market for perc from 1974 to 2000. It
can be seen that there has been an almost continual decrease in the amount of perc
sold during this period, such that the total decline (1974 to 2000) is 76%, or from
290,000 tonnes in 1974 to 71,000 tonnes in 2000 (CINET, not dated). This reduction
is believed to be mainly as a result of the improvements in the dry cleaning sector,
such as: use of more efficient dry cleaning machines; an increased emphasis on recycling and improved housekeeping; and the use of enclosed systems (ECSA in
Environment Agency, 2002).
Table 2.1: European Market of Chlorinated Solvents 1974-2000 (sales only until 1992)
Year
Sales of Perc
% change
Change by decade
1974
290,000
N/A
1980
215,000
-26%
-26%
(1974 to 1980)
1981
187,000
-
1990
123,000
-34%
1991
113,000
-8%
2000
71,000
-4%
Overall change (as sales only until 1992, overall change is uncertain)
-34%
(1981 to 1990)
-37%
(1991 to 2000)
-76% (1974 to 2000)
Source: CINET (not dated), Figure 16
In 1994, perc was produced by six companies in the EU, located in Belgium, France,
Germany, Italy, Spain and the UK. Total production was estimated at 164,000 tonnes,
with production at individual sites ranging from just over 4,000 tonnes per year to
around 65,000 tonnes.
Sales within the EU are estimated at 79,000 tonnes with exports making up 56,000
tonnes. Perc is also imported into the EU from the United States and Eastern Europe,
although the amounts are thought to be negligible (Environment Agency, 2002).
The main uses of perc are (ECSA, not dated):
•
•
•
dry-cleaning;
metal cleaning and degreasing (as a substitute for 1,1,1-trichloroethane, although
trichloroethylene is a more important substitute);
chemical synthesis; and
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Case Study 3: Tetrachloroethylene
other applications, with this including use in small quantities to make paint
removers, printing inks, adhesives, special cleaning fluids, dye carriers and
silicone lubricants.
•
Table 2.2 provides estimates of the amount used in the EU in each of the above
applications (Environment Agency, 2002). The figures indicate that most perc is used
for dry cleaning (80% of sales and 38% of production volume). Exports are the
second most important ‘use’, at 34%, followed by use as a chemical intermediate.
‘Other’ uses make up only a very small proportion of total production or sales.
Table 2.2: Breakdown of Perc use in 1994 in the EU
Percentage of
Application
Percentage of Sales
production volume
Dry cleaning agent
38%
80%
Tonnes per annum
62,400
Metal cleaning agent
9%
18%
14,000
Chemical intermediate
18%
-
30,000
Exports
34%
-
56,000
Other
1%
2%
1,600
Total
100%
100%
164,000
Source: Environment Agency, 2002
2.1.2 Dry Cleaning
The process of dry cleaning involves four steps (DETR, 1999):
•
•
•
•
washing the material in hot solvent;
drying with hot air;
deodorisation of the garment; and
solvent regeneration.
Perc began replacing hydrocarbon solvents, such as white spirit, in dry cleaning about
50 years ago mainly because of its lack of flammability, ease of handling and
potential for recycling. It is the major cleaning solvent in use worldwide, with a
solvency power of 90 on the kauri-butanol scale, making it the strongest solvent in
use for commercial dry-cleaning (ECSA, 1999). Perc is also the main substitute for
1,1,1-trichloroethane and CFC113, which are controlled under the Montreal Protocol.
CFC113 was used for specialised cleaning of fabrics such as silk, fur, hide, suede and
leather, but only accounted for a small proportion of the dry cleaning market (ECSA,
1996).
There are an estimated 60,000 dry cleaning establishments in the EU. More than 90%
of these units use perc as the dry-cleaning agent, although the southern European
countries use more white spirit for dry-cleaning than those in the north (Organisation
& Environment, 1991 in Environment Agency, 2002). The remaining 10% are in the
process of switching from CFC113 to either perc or hydrocarbon systems, the latter
having been developed to provide similar cleaning performance to CFCs (DETR,
1999).
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With increased awareness of the damages that can be caused by dry cleaning agents,
there has been a general shift from open circuit machines to closed circuit machines
(DETR, 1999):
•
•
open circuit machines (OCMs) – condense the solvent through cold water with
venting to atmosphere; while
closed circuit machines (CCMs) – incorporate internal refrigerated condensers
with much reduced solvent emissions.
CCMs first became available in 1988, and the shift to their use has also been due to
the lower operating costs associated with CCMs. Operating costs for a CCM are
around 40% lower per load than for an OCM due to savings in solvent consumption,
energy, and water and residue disposal (DETR, 1999). CCMs also have the advantage
of being able to carry out cleaning and drying as a single operation (therefore not
requiring the ‘wet transfer’ of clothes and significant solvent emissions). As a result,
there is an overall decrease in the volume of perc used per machine of up to 90%
(Smith, 1995).
For example, losses of perc from dry cleaning machines have been measured at 15.5
kg per 100kg of clothes cleaned for OCMs. This reduces to 2.03 kg/100kg clothes
cleaned for CCMs (and OCMs with carbon filters). This is equivalent to 90% and
54% of total estimated perc losses from the machines (17.3 kg/100 kg for OCMs and
3.8 kg/100 kg for CCMs and OCMs with carbon filters). The remaining losses are to
water and solid waste (Organisation & Environment, 1991 in Environment Agency,
2002).
2.1.3 Metal Degreasing
Trichloroethylene is the main substitute for 1,1,1-trichloroethane in metal cleaning
and degreasing but perc is used in some small and medium-sized processes. This is in
fact the second main market for perc in the EU (see also Table 2.2). An estimated
14,000 tonnes of perc were sold to metal degreasing operations in the EU in 1994.
Since this time, a number of companies have invested in closed circuit machines,
which have reduced emissions of solvents during hot vapour degreasing. However, it
is estimated that only a small proportion (i.e. 5% in the UK) of metal degreasing
operations have these machines (Environment Agency, 2002).
Some of the perc used in metal degreasing is bought from companies recycling perc
previously used in other industries (Environment Agency, 2002).
Releases of perc from metal degreasing are estimated to be about 10% of total use,
with 90% of releases going to air, 1% to water and 9% into solid wastes (Environment
Agency, 2002).
2.1.4 Use as an Intermediate
In 1994, there were four chemical intermediate plants using perc in the EU, including
one plant which undertakes both production and processing. Two processing plants
stopped producing or using perc in 1995 (Environment Agency, 2002).
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Case Study 3: Tetrachloroethylene
Before the phase-out of CFC solvents and refrigerants such as R113, R114 and R115
under the Montreal Protocol, perc was used as a chemical feedstock for these
substances. It is now used as a raw material for production of HFCs and HCFCs,
which are CFC substitutes. The use of perc as a chemical intermediate is, however,
declining (Environment Agency, 2002).
2.1.5 Other Uses
Perc is used in small quantities to make paint removers, printing inks, adhesives,
special cleaning fluids, dye carriers and silicone lubricants. The volumes of perc used
in these sectors are minor, while releases are expected to be similar to those of dry
cleaning and metal degreasing (i.e. most perc would be released to air) (Environment
Agency, 2002).
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RPA & BRE
3.
ENVIRONMENTAL AND HUMAN HEALTH IMPACTS
3.1
Introduction
This section provides a chronology of research and regulatory activities. This
overview is not meant to provide a comprehensive summary of scientific and other
research concerning perc nor is it intended to question or validate research
conclusions. Instead, the aim is to provide a context that can be drawn upon later in
this case study analysis to illustrate whether REACH would have provided any
benefits in relation to the use of perc had it been implemented sooner.
The section:
1) reviews the scientific and academic literature to identify when research on
different hazardous properties began and when concern started to arise;
2) makes chronological links between the scientific research and the introduction of
either voluntary or regulatory measures aimed at reducing risks to the
environment, workers and public health;
3) presents monitoring data (where available) to illustrate the possible scale of
environmental damages that have occurred as a result of perc use and disposal;
and
4) analyses the history of testing and risk management activities in relation to
properties of concern (persistence, bioaccumulation and toxicity, carcinogenicity,
etc.).
3.2
Development of Environmental and Health Concerns
3.2.1 1970 -1979
Although perc was one of the most important chemicals in the dry cleaning and metal
degreasing industries in the 1970s, significant concerns about the possible
environmental and health effects of its use and disposal did not begin to arise until the
mid 1970s. Researchers investigating the already controversial health impacts of
similar solvents, trichloroethylene and 1,1,1-trichloroethane, also started to take an
interest in perc. Examples of some of the key issues researched in the 1970s included:
•
tests to establish the bioconcentration potential of perc and other similar chemicals
in fish (Neely et al, 1974);
•
investigations into the toxicity of perc on marine organisms (Pearson &
McConnell, 1975);
•
tests to establish the presence of perc in food products (McConnell et al, 1975);
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Case Study 3: Tetrachloroethylene
•
investigations into the effects resulting from the formation of carbon tetrachloride
and other dangerous breakdown products from perc in the atmosphere (Singh et
al, 1975);
•
tests to determine the degradation of perc in air and of its breakdown products
(Singh et al, 1975; Gay et al, 1976); and
•
investigations into the toxicity of perc using fathead minnows (Alexander et al,
1978).
While most of these early investigations may have been quite inconclusive, they also
highlighted sufficient concerns to result in further investigations. However, because
most of these tests focussed on two or three chemicals, further work on perc was not
pursued vigorously.
By the late 1970s, it was quite apparent that perc raised concerns, albeit uncertain, for
health and the environment. In response to these, the European Commission issued
the Directive 76/464/EEC on dangerous substances in water, identifying perc as a List
II substance. List II substances under this Directive are defined as polluting
substances of concern, the concentration of which in the environment was to be
reduced and discharges monitored by Member States using National Quality
Standards, pending further investigations into the dangers posed by the substance.
This was one of the earliest official acknowledgements of the potential hazards posed
by perc in the environment.
Box 3.1 overleaf gives an example summary of developments concerning perc from
the 1970s as discussed above.
3.2.2 1980 - 1989
By the early 1980s, the need for research into the environmental and human health
effects of perc had been more fully recognised. Konemann started investigations into
the effects of perc on the guppy in 1981, while the effects on the fathead minnows
were revisited by Walbridge et al in 1983 and Broderius & Kahl in 1985. Vonk et al
(1986) investigated the toxicity of perc to organisms such as the earthworm while
more specific studies investigating the metabolism of perc in the body were carried
out by Buben & O’Flaherty (1985).
Wider environmental effects were also being investigated. Initial suspicions were
related to ozone depletion effects but further research showed groundwater
contamination to be a substantial risk. For example, perc was observed to leach
rapidly into groundwater near sewage treatment plants in Switzerland, with no
evidence of biological transformation of perc found taking place (Schwarzenbach et
al, 1983). In 1980, the European Commission issued the Groundwater Directive
(80/68/EEC) identifying perc as a risk to groundwater. List I and II substances were
to be prevented from entering groundwater, with List II limited by a ‘consent system’.
Investigations into groundwater contamination effects by perc and related chemicals,
probably prompted by the Directive, were carried out by Fielding et al, 1981. After
further research in 1982, the Commission published a list of 129 potential Black List
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Box 3.1: Developments in the 1970s
1974
Tests to establish the bioconcentration potential of perc and other chemicals in fish are
carried out. A relatively low bioconcentration factor of 40 is reported for the rainbow
trout (Neely et al, 1974).
1975
Pearson and McConnell test the toxicity of perc to marine organisms. Initial figures
showed a 48-hour EC50 of 3.5mg/L for barnacle Eliminius modestus. Concentrations
in algae ranged from 13 – 20 µg/kg while those in mollusc bodies and organs ranged
from 0- 176µg/kg (dry body weight).
Rainwater samples collected in industrial cities in England show levels of perc up to
150 ng/L (150 ppt), while sediments from Liverpool Bay, England, were found to
contain concentrations ranging from 0.03 to 6 ppm, with most detections at the lower
limit (Pearson & McConnell 1975).
1976
1976
1978
McConnell et al detects the presence of perc in dairy products and meat in the range of
0.3-1.3 µg/kg and 0.9-5 µg/kg respectively.
Gay Jnr. et al investigates the reactions of perc in air. In one of these investigations,
perc mixtures irradiated using a smog chamber with air containing NO2 for 140
minutes, were found to have reacted forming carbon monoxide, ozone, hydrogen
chloride, triacetyl chloride and phosgene.
Singh et al, 1975 had also studied the degradation of perc in ultra zero air and found
the product yields were around 70-80% phosgene and 8% carbon tetrachloride with
carbon tetrachloride concentrations increasing after perc had been exhausted.
The European Commission issues Directive 76/464/EEC, also known as the
Dangerous Substances in Water Directive, identifying perc as a List II substance, the
concentration of which is to be reduced in the environment and its discharges
monitored using National Quality Standards, pending further investigations.
Alexander et al investigates toxicity of perc using fathead minnows. Results establish
96-hour EC50 of 18.4 mg/L (Alexander et al, 1978).
(or List I) substances which included perc (based on Directive 76/464/EEC).
Potential Black List substances are those that are highly toxic, persistent, carcinogenic
or liable to accumulate in the environment.
By the mid 1980s, certain disease linkages were also being examined, particularly in
relation to the carcinogenic and toxic properties. This prompted the United States to
commission a number of toxicological and carcinogenic studies in 1986 under the
National Toxicology Program to investigate the effects of perc. At the end of the
program, a proposal was put forward for perc to be classified as a probable human
carcinogen. Interestingly, this proposal was rejected, citing a lack of concrete
evidence. In the same year, the European Commission issued Directive 86/280/EEC
which placed limits on emissions of perc to water from industrial plants at 10 g/tonne
and 2.5 g/tonne by 1992 and 1994 respectively. In individual EU Member States,
environmental quality standards were also being established (e.g. a limit of 1 mg/l in
the freshwater environment).
Research activities continued through the late 1980s, with most investigations
pursuing the earlier leads - health (carcinogenic and toxic effects) and environmental
resource (groundwater) impacts.
Perc’s effects on plants were also scrutinised,
especially with regard to conifers, where bleaching of chlorophyll under certain
conditions was observed (Frank & Frank, 1985). Perc presence in the atmosphere for
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Case Study 3: Tetrachloroethylene
long periods and at locations remote from emission sources was demonstrated in tests.
Calculations from measurements at sites distant from any pollution sources placed the
lifetime in air of perc at 5-6 months for the northern hemisphere and 2 months for the
southern hemispheres (Class & Ballschmiter, 1987). In 1987, the London Conference
for the Protection of the North Sea agreed reductions of up to 50% for all perc
disposed of to the sea by 1995 from a 1985 baseline. Monitoring of perc levels
increased in many countries in Europe such as Finland and Germany.
Box 3.2 below summarises developments from the 1980s as discussed above.
Box 3.2 : Developments in the 1980s
1980
Directive 80/68/EEC identifies perc as one of the substances posing a risk to
groundwater.
1981
Konemann et al investigated the effects of perc on guppy
Fielding et al investigated groundwater contamination by perc.
1982
Based on Directive 76/464/EEC, a list of 129 potential Black List (or List I) substances
including perc is published. This acknowledged its potential toxic, persistent, and
carcinogenic properties.
1983
1985
1985
1986
1986
1987
1988
Effects on the fathead minnows are investigated by Walbridge et al and Broderius and
Kahl.
Perc effects on plants are investigated. Bleaching of chlorophyll from conifers under
certain conditions was observed (Frank and Frank, 1985).
Buben and O’Flaherty, 1985 investigate the effects of perc metabolism in the human
body.
National Toxicology Program launches investigation into the carcinogenic and toxic
properties of perc in the United States. Subchronic inhalation studies were carried out in
rats and mice. Liver effects were observed at 1350 mg/m3(>200 ppm) in mice and rats,
with congestion of the lungs, decreased survival and growth retardation at higher dose
levels seen in the rats.
Directive 86/280/EEC places limits on perc from industrial plants at 10 g/tonne and 2.5
g/tonne by 1992 and 1994 respectively. Environmental quality standard (EQS) of 1,000
µg/l set in the UK.
The London Conference for the Protection of the North Sea requires reductions of up to
50% for all perc disposed of into the sea by 1995 from a 1985 baseline.
Data collected from several locations in the city of Hamburg, Germany, showed ambient
air concentrations ranging from 1.8 to 70.8 µg/m3 (0.27-10.44 ppb) (Bruckmann et al
1988). The highest concentrations were detected downwind of a drycleaning facility.
Mean concentrations of perc at levels between 5.0 and 5.6 µg/m3 (0.70 ppb and 0.82 ppb)
were also found at a distance of 0.5-l.5 meters above the surface of a landfill containing
halogenated volatile organic compounds in Germany (Koenig et al. 1987).
3.2.3 1990 – Date
By the 1990s, the fact that perc posed significant environmental and health risks had
been accepted in both scientific and political circles. In 1990, Directive 90/415/EEC
(the daughter Directive to 76/464) identifying perc as a List I substance was
introduced. This included guidance on the use and disposal of perc, trichloroethylene
and similar solvents. The same year, a voluntary agreement was reached at the Third
International Conference for the protection of the North Sea at the Hague, pledging
reductions in aquatic and atmospheric perc, further to the London Conference by 1995
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RPA & BRE
and 1999 respectively. While regulatory initiatives were being taken by the
authorities, research investigations were also continuing to be commissioned to
ascertain the extent of the risks posed by perc.
In one investigation, 2,050 male and 1,924 female workers, monitored for
occupational exposure to perc, trichloroethylene and 1,1,1-trichloroethane were
followed up for incidence of cancer from 1967 to 1992. An excess of pancreatic
cancer and non-Hodgkin’s lymphoma was seen after 10 years from the first personal
measurement, with higher numbers of cancers of the stomach, liver, & prostrate
combined and the overall cancer risk increasing for a follow up period of more than
20 years for exposed workers (Anttila et al, 1995).
Another investigation studied cancer mortality from data associated with 8,163 deaths
among persons previously employed as laundry or dry cleaning workers based on the
fact that they must had been exposed to a number of organic solvents including perc
while at work (Walker et al, 1997). Using a different approach, results again showed
an excess in total cancer and oesophageal cancer mortalities in black males and in
laryngeal cancer in white males aged 15-64 years. Other effects associated with dry
cleaning workers included an elevated risk of liver and biliary tract cancer and nonHodgkin’s lymphoma. Arguments have, however, ensued as to whether the
carcinogenic effects have solely been due to perc exposure.
In another research investigation, biochemical changes in blood and urine in dry
cleaning workers of above 10 years, indicative of liver and kidney damage, were
highlighted while central nervous system symptoms such as dizziness, headaches,
sleepiness, light headedness and poor balance later became associated with exposure
to perc (Guth et al, 1997).
A considerable number of monitoring studies across Europe also started reporting that
they had detected perc in the environment, particularly in groundwater and in
leacheate. Analysis of pore water from a principal aquifer in England, the Chalk,
which has low carbon content and rapid groundwater flow in fissures, showed the
presence of perc at concentrations ranging from 0.05 to 40 mg/L at a depth of 50
meters (Lawrence et al, 1990). These concentrations exceeded the recommended
maximum acceptable concentration for perc and trichloroethylene in drinking water in
the EU of 10 µg/l. Monitoring of levels from six municipal solid waste samples from
Hamburg, Germany, also revealed levels of perc ranging from undetectable to 1.41
mg/kg (1.41 ppm) (Deipser & Stegmann, 1994).
The United Nations Economic Commission for Europe (UNECE) called for a 30%
reduction in air emissions from 1988 to 1999. Under the UNECE Gothenburg
Protocol to the Convention on Long Transboundary Air Pollution, EU Member States
agreed to reductions in emissions of VOCs by 2010, with this including perc. In
1998, perc was included within the OSPAR list of candidate substances for selection,
assessment and prioritisation as part of its review of the list of chemicals for Priority
Action.
Box 3.3 below gives a summary of key developments in relation to environmental
concerns and regulation of perc from the 1990s to date as discussed above.
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Case Study 3: Tetrachloroethylene
Box 3.3: Developments in the 1990s
1990
Directive 90/415/EEC (Daughter Directive to 76/464) identifying perc as List I substance
issued, giving guidance on the use and disposal of perc, trichloroethylene and similar
solvents.
1990
Voluntary Agreement at the 3rd International Conference for the Protection of the North
Sea at the Hague. Further to London Conference, reductions in atmospheric perc required
by 1999. Listed in Annex 1A as one of 36 substances.
1992
Voluntary scheme to develop 'Charters of Co-operation' agreed. Charters signed with
metal finishing & engineering industries to encourage recycle and reuse schemes in UK
and France.
1995
1996
1998
1999
2000
3.3
New UK Environmental Quality Standards guidance released. EQS requires continuous
improvements in perc levels in the aquatic environment.
Voluntary agreement in form of 'Charters of Co-operation' Charters signed with
distributor associations to reduce impacts of perc use in a number of European Countries
including France, UK and Belgium.
Tax on chlorinated solvents, including perc, resulting in 25% consumer price increase
introduced in Denmark.
Perc is included within the OSPAR list of candidate substances for selection, assessment
and prioritisation in order to review the list of chemicals for Priority Action.
EC Directive 99/13/EC – Solvents Emissions Directive regulating the use of perc in
industrial processes issued. Limits set for perc use in drycleaning.
Product tax introduced on VOCs to help reduce emissions to air of perc and other VOCs
in Sweden.
Key Properties and Presence in the Environment
3.3.1 Key Properties
The key properties of perc as most recently concluded in the Draft EU risk assessment
are as follows.
•
Persistence: Perc has a relatively low solubility in water and a medium-to-high
mobility in soil. It is not expected, therefore, to reside in surface environments for
more than a few days. It does, however, persist in the atmosphere for several
months and may also persist in groundwater for several months to years. These
persistent properties increase the potential for human exposure (man via the
environment) substantially. In terms of the PBT criteria, perc meets the
persistence criterion.
•
Bioaccumulation: Perc does not bioaccumulate in the aquatic food chain.
Measured bioconcentration factors (BCFs) have been found to range between 10
and 100 for perc in fish (Kenaga, 1980; Neely et al, 1974; Veith et al, 1980),
suggesting a low tendency to bioconcentrate.
Perc does not meet the
bioaccumulative criterion of the PBT assessment. Although, the presence of perc
in food stuffs may suggest bioaccumulation in plants, it is still unclear whether
accumulation takes place during growth or at some point after harvesting.
•
Toxicity: Exposure of animals to perc has been strongly linked with liver toxicity
effects. Although liver toxicity effects have been observed in humans in some
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studies, the evidence is much weaker. Common acute effects of perc in humans
are reversible neurological effects such as headache, dizziness, nausea, sleepiness,
with eye and throat irritation also reported. Chronic exposure to low levels of
perc showed subtle neurological effects, with renal effects also observed in
humans occupationally exposed to perc. In terms of environmental effects, perc is
classified as toxic to aquatic organisms. It does not, however, meet the toxicity
criterion for the PBT assessment on this basis. It may meet this criterion in
relation to its CMR properties, depending on the eventual outcome of those
discussions.
•
PBT: overall perc is not classified as a PBT substance.
3.3.2 Human Health Concerns1
Studies in humans have shown that the primary route of exposure to perc is
inhalation. On inhalation, perc is rapidly absorbed into the bloodstream, from where
it distributes readily into fatty tissues. Exposure of humans to air-borne perc has been
found to be significant in the drycleaning and metal degreasing industries.
Rapid and extensive absorption of perc has also been found to occur in cases of oral
exposure. The main toxic effect associated with acute inhalation exposure is central
nervous system depression (dizziness, headaches, sleepiness), with very high
concentrations leading to narcosis, unconsciousness and even death. Recent research
indicates that perc is irritating to both the respiratory tract and skin; although dermal
penetration appears to be generally quite low. Perc has also been found to be able to
cross placental barriers, although significant uncertainty exists regarding the risk of
spontaneous abortion in dry-cleaning workers exposed to perc (HSIA, 1999).
While there is a consensus agreement on the toxicity and carcinogenicity of perc in
rodents, the evidence for increases of cancer incidence in human studies is quite
inconsistent. One of the key contentious areas relates to whether the mechanisms
underlying the appearance of cancer tumours in rodents is of any significance in
relation to human health, and as such whether rodent data can be directly applied to
humans given the important species differences which exist. The result has been that
the need to control perc, in relation to its being a carcinogen in humans, has been
passed on to the regulatory community.
In 1995, based on the evidence of proven carcinogenicity in mice, the International
Agency for Research on Cancer (IARC, 1995) classified perc as ‘probably
carcinogenic to humans’. The EU has recently agreed to classify perc as a category 3
carcinogen, a substance of concern to humans owing to possible carcinogenic effects.
1
These concerns are in agreement with the concerns expressed in the Draft HSE Human Health Risk
Assessment for perc. The final version is however yet to be published and may change significantly
from the version used.
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Case Study 3: Tetrachloroethylene
3.3.3 Classification and Labelling
Box 3.4 shows the current (and proposed) EU classification and labelling for perc.
Box 3.4: Current Classification of Perc
The current EU classification is
Carcinogen category 3; R40; N:R51/53
and the current labelling requirement is
Xn; N; R40-51/53; S23-36/37-61
Xn; R40 - Possible risk of irreversible effects
N - Dangerous for the Environment
R51 - Toxic to aquatic organisms
R53 - May cause long-term adverse effects in the aquatic environment
S23 - Do not breathe dust
S36/37 - Wear suitable protective clothing and gloves
S61 - Avoid release to the environment. Refer to special instructions/safety data sheet
Carcinogen category 3 indicates a substance which causes concern for man owing to possible
carcinogenic effects but in respect of which the available information is not adequate for making a
satisfactory assessment. There is some evidence from appropriate animal studies, but this is
insufficient to place the substance in category 2. This classification applies to both the pure
compound and products containing ≥1% of perc.
Proposed classification
In addition to the above classification and labelling the following is proposed for human health:
Xi: R37/38; S23-37
Xi - Irritant
R37/38 - Irritating to respiratory system and skin
3.3.4 Presence in the Environment
Groundwater
As indicated above, perc is not readily soluble in water and can persist in groundwater
for several months or years following improper disposal or seepage into aquifers.
Chlorinated hydrocarbons are reported to be widely distributed in the groundwaters of
Western European countries today, and this has substantial implications for human
health along with the associated costs to society.
A survey of drinking water in the United Kingdom showed average perc levels of 0.4
µg/litre in municipal waters. While this value is well below the 5 µg/l threshold,
significantly higher levels have been found in other groundwater bodies ranging from
as low as 0.05 – 13 µg/l. Sampling of drinking water sites carried out by Anglian
Water in the UK between 1992 and 1995 showed 13 samples with levels ranging from
>10 µg/l to <100 µg/l, with one sample showing levels of up 146 µg/l (EU RAR,
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RPA & BRE
1999). In Switzerland, perc concentrations as high as 954 µg/litre have been found in
contaminated groundwater (Vyskocil et al, 1990), while a survey of aquifers in
Austria in 1994/95 indicated that one out of four sampling sites had perc presence
above 0.1 µg/l.
In France, isolated cases of perc pollution exist, with widespread low levels now
being detected, particularly in the Nievre region and across the Rhône-MéditerranéeCorse basin. Higher concentrations (several µg/l) can be found at sites around larger
towns with industrial areas, and serious accidents which affected public supply wells
in the east of France have been documented (Strasbourg, 1990 in EEA, 1999). In
Germany, the average concentration found in a drinking-water survey in 100 cities
was 0.6 µg/l while the maximum concentration was 35.3 µg/l (Bauer, 1991).
Other cases of groundwater contamination by perc can be found across Europe in :
•
•
•
•
•
Baden-Württemberg (Germany), where perc and trichloroethylene, have been
detected in groundwater in highly industrialised and urbanised areas;
Hungary, around waste disposal sites, landfills and military sites;
Slovak Republic from the chemical industry and military waste dumps;
Estonia, from military air-fields heavily contaminated with fuel; and
Romania, from oil products around pipelines, refineries and storage areas.
Bearing in mind the prevalence of perc in the aquatic environment as shown above, it
should be noted that the EU maximum accepted concentration of perc and
trichloroethylene in drinking water is 10µg/l, which is basically equivalent to the US
EPA limit of 5 µg/l for perc only.
There are no known natural sources of perc in the environment, and as such its
presence in the environment, in any form or concentration, is a testament to the poor
handling and disposal methods which have been in place over the years.
Air
A survey of city air in the United Kingdom detected perc at levels from 0.7 to 70
µg/m3 (Fast & Van Wijnen, 1994). It has also been estimated that 0.1% of the
population in the Netherlands are exposed to an average ambient air concentration of
40 µg/m3, 0.5% to 20 µg/m3, 3.3% to 6 µg/m3, 12.5% to 2 µg/m3 and the remainder to
1 µg/m3. For people living in the close vicinity of dry cleaning facilities (estimated to
represent 0.08% of the population), the average exposure concentration for every
working day was estimated at 1000 µg/m3 (Bessemer et al, 1984).
Considering that the major route of exposure to perc for humans is via inhalation,
studies have shown that based on a perc concentration in air of 6 µg/m3, estimated
exposure for an adult with an air intake of 20 m3 would be about 120 µg/day. While
the values are much lower than the LOAEL or PNEC values, the lack of an in-depth
understanding of the cancer mechanisms, bioaccumulative properties, and other
effects of perc could give rise to substantial concerns in future.
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Case Study 3: Tetrachloroethylene
Data, however, suggest that the concentrations of perc measured in city air may be
sufficient to cause adverse effects to some terrestrial plants, notably trees as seen in
Germany and Finland. Frank and Frank (1985) observing the effects of perc on fir,
Norway spruce, beech and other trees, reported that there was an increased incidence
of chlorosis (bleaching of needles), necrosis (death of needles) with premature needle
loss over the last 2 decades, resulting from exposure to chloroethylenes under photoactivated conditions. Further laboratory investigations indicated chlorosis and
necrosis effects following exposure to 3 to 6 µg/m3 and 40 µg/m3 perc, respectively,
over a period of 1 to 2 weeks. Damage was observed to be dependent on the duration
of exposure and concentration of perc, as trees died after exposure to 100 to 130
µg/m3 perc for 1 to 2 months. There are no known direct adverse effects from air to
aquatic biota or terrestrial wildlife; although indirect food chain effects may exist.
Food
Although data on concentrations of perc in food are limited, studies from Switzerland
and Germany in the early 1980s, have reported relatively high total intakes of 87-170
µg/day (WHO, 1984). Despite the reduction in use volumes, more recent
measurements have still detected elevated concentrations of perc in fatty food
products in residences and markets from nearby dry cleaning firms. Concentrations
of 110 µg/kg to 436 µg/kg have been found in cheese in a supermarket near a dry
cleaning shop in Germany, while concentrations of up to 180 µg/litre have been found
in food in Finland (Vartiainen et al, 1993). Perc concentrations in seafood in the
United Kingdom have been detected ranging from 0.5 to 30 µg/kg (Fast & Van
Wijnen, 1994). The implications for humans are again very low based on LOAEL
and PNEC values, but considering the paucity of knowledge and understanding of
toxic and carcinogenic mechanisms of perc in general, this option cannot be ruled out.
Ultimately, the presence of perc in food remains an undesirable, albeit unknown,
threat to humans.
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RPA & BRE
4.
THE REACH DOSSIER
4.1
Introduction
4.1.1 Overview
In order to develop the REACH dossiers for perc, we have had to go backwards in
time and make assumptions as to the data available to industry and the approach that
they would take to preparing the dossier. As indicated earlier, this has been done here
by drawing on experience gained through the ESR risk assessment in relation to data
availability and the approach taken by industry to that assessment. More generally,
assumptions have been made on:
•
•
•
•
•
production levels and associated uses for the manufacturer/consortia submitting
the dossier;
the level of information available to the manufacturer at time of dossier creation;
the substance-tailored testing that would be undertaken for completion of the
dossier (in line with Testing Option I as presented in Section 2 of the main report);
the assumptions that would be made concerning exposure and hence the
conclusions that would be reached regarding potential risks; and
the manner in which industry would respond to any conclusions concerning
environmental risks or risks to man via the environment with regard to risk
reduction activities.
The remainder of this section sets out the key assumptions and associated results for
the hypothetical dossier compiled for perc. No details of the underlying studies are
included (see the full ESR risk assessment for further information on the underlying
studies).
4.1.2 Basic Assumptions
Perc is produced and used in quantities of 100,000 – 500,000 tonnes in the EU at the
present time. However, given the relatively small number of producers, it is assumed
that all of the manufacturers will join in a consortium to prepare the REACH dossier.
Thus, the dossier must meet the information requirements as set out under Option I for
Dossier D. All current uses of perc are covered by the dossier.
In developing this hypothetical dossier, the following assumptions have been made:
•
•
•
•
the data that were available in the IUCLID submitted to the European Chemicals
Bureau are assumed to have been available to the manufacturers at the start of
dossier preparation;
any further substance tailored testing must be undertaken in line with the
requirements set out for Dossier D (see Section 2 of the main report);
where site specific release data are not available, default data from the TGD and
within the EUSES and EASE models are applied; and
EUSES provides the basis for reaching conclusions as to whether or not
unacceptable risks result from a particular application or sector.
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Case Study 3: Tetrachloroethylene
4.2
Base Data
4.2.1 Identity of the Substance
Perc is produced as a pure substance, at >99%. The CAS number is 127-18-4.
Relevant spectra for substance identification are available.
Methods of detection are available for water, soil, sediments and biota. Detection
limits vary. For water, the lowest in recent studies is 0.01µg/l; in air, levels as low as
0.1 µg/m3 have been reported.
4.2.2 Physico-chemical Data
The basic physico-chemical data are presented in Table 4.1.
Table 4.1: Physico-chemical Data
Property
Value
Molecular weight
165.85
Melting point
-22.0 to -22.7ºC
Boiling point
121.2ºC
Relative density
1.623 at 20ºC
Vapour pressure
1.9 kPa at 20ºC
Octanol-water partition coefficient (log
Kow)
2.53
Water solubility
~149 mg/l at 20ºC
Solubility in other solvents
Miscible with alcohol, ether, chloroform and benzene
Viscosity
0.891 N.m-2.s at 20ºC
Henry's Law constant
2114 Pa m3/mole at 20ºC
Flammability
Flash point: None under test conditions
Autoflammability
n/a
Explosive properties
Not explosive
Oxidising properties
Not considered as an oxidising agent but can oxidise in
presence of air and light
Vapour density
5.8 (Air=1)
Surface tension of aqueous solution
No information
Saturated vapour concentration
25,000 ppm (169,500 mg/m3) at 20ºC
Odour recognition
~180 mg/m3
Conversion factors
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1 mg/m3 = 0.147 ppm at 25ºC
1 ppm = 6.78 mg/m3 at 25ºC
RPA & BRE
4.2.3 Ecotoxicity
Aquatic Toxicity
Table 4.2 below summarises the lowest valid results for each species amongst the
available data set for aquatic toxicity. There are a number of other test results for fish,
invertebrates and algae which support these values, showing effects at higher
concentrations, along with other studies not considered to be valid.
Table 4.2: Aquatic Toxicity Data
Species
Parameter
Concentration (mg/l)
Fish
Jordanella floridae
Oncorhynchus mykiss
96-hour LC50
8.4
10-day NOEC (Survival, Larvae)
1.99
28-day NOEC (Survival, Fry)
2.34
96-hour LC50
5
48-hour EC50
8.5
28-day NOEC (Reproduction)
0.51
72-hour EC50 (Cell multiplication
inhibition test)
3.64
72-hour EC10 (Cell multiplication
inhibition test)
1.77
Invertebrates
Daphnia magna
Algae
Chlamydomonas
reinhardii
Based on these data, a PNEC of 51 µg/l is derived, using a factor of 10 on the long
term NOEC value for Daphnia.
Terrestrial Toxicity
Results are available for terrestrial plants and terrestrial invertebrates, and are
summarised in Table 4.3. Using these data gives a PNEC of 0.01 mg/kg, applying a
factor of 10 to the NOEC for nitrification.
Note that the requirements for Dossier C (and thus Dossier D) include a test on a
higher plant. This is included in the test set above as exposure through the soil, which
is what would usually be expected. There are no indications of a requirement for
testing plants through exposure via the air. The IUCLID has a short note to the effect
that under exposure to UV, chlorinated solvents may lead to bleaching of chlorophyll
in plants, but that this is a hypothesis. There is not a great deal of experience of what
types of chemical do affect plants in this way, with only isolated examples having
been noted (i.e. for specific herbicides). Based on the information available here, no
test through air has been conducted.
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Case Study 3: Tetrachloroethylene
Table 4.3: Toxicity to Terrestrial Organisms
Species
Parameter
Concentration
Reference
100-320 mg/kg
Vonk et al (1986)
Terrestrial Invertebrates
Earthworm
14-day LC50
Eisenia foetida
28-day NOEC
≤ 18 mg/kg
(Production of cocoons)
28-day NOEC
18-32 mg/kg
(Appearance of worms)
14-day NOEC
577 mg/kg
(Mortality, weight and behaviour)
Carabid beetle
14-day LC50
945 mg/kg
14-day NOEC
5.0 mg/kg
Römbke et al
(1991)
1-day LC50
113 mg/kg
Heimann and
Härle (1993)
Poecilus cupreus
Springtail
Römbke et al
(1991)
(549 mg/kg)
Folsomia candida
Terrestrial Plants
Lettuce
16-day NOEC (Growth)
Avena sativa
100 mg/kg
(148 mg/kg)
16-day NOEC (Sublethal effects)
1 mg/kg (1.48 mg/kg)
16-day EC50 (Growth)
580 mg/kg (861 mg/kg)
Bauer and Dietze
(1992)
Soil dwelling bacteria
Pseudomonas
putida
16-hour EC10
Other bacteria
Soil respiration: NOEC
> 45 mg/l
Knie et al (1983)
< 2,000 mg/kg (wet)
Vonk et al (1986)
Nitrification with humic sand: NOEC
< 40 mg/kg (wet)
Nitrification with loam: NOEC
≤ 0.1 mg/kg (wet)
Source: References taken from Environment Agency, 2002
Sediment Toxicity
There are no sediment toxicity data for perc. The properties of the substance mean
that the equilibrium partitioning method should be appropriate to estimate the effect
concentrations for sediment, so no testing for this endpoint is proposed. As the
sediment PEC values are also estimated from water by the equilibrium partitioning
method, the risk characterisation for sediment will be the same as that for water. The
sediment compartment is therefore not discussed separately in the rest of this dossier.
Avian Toxicity
No data are available on avian toxicity. Perc shows a low potential for accumulation,
and so an assessment of secondary poisoning is not considered necessary. Therefore,
no testing to fill this gap is proposed.
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RPA & BRE
4.2.4 Environmental Fate
Biodegradation
Standard ready biodegradability tests show that perc is not readily biodegradable.
Other tests also confirm this, with removal in most cases being due to volatilisation.
It is assumed that the substance is not biodegradable.
In relation to microbial inhibition, an EC50 of 112 mg/l was determined in a test on
nitrifying bacteria. This is considered suitable for assessing the risks to waste water
treatment plants, and so no further testing is proposed.
The requirements of Dossier D do not specifically mention abiotic degradation, but
this is relevant for this substance. Perc reacts in air with OH radicals, and possibly
with Cl radicals. The half life of the reaction with OH radicals is estimated as 3.2
months, based on a measured reaction rate constant and the default concentration of
radicals. The reaction with chlorine radicals is a minor pathway under normal
atmospheric conditions, and will not lead to significant amounts of carbon
tetrachloride (of concern as an ozone depleter). Perc itself is not considered to be a
major contributor to ozone depletion, and is not reactive enough to be a significant
contributor to low level ozone formation.
Adsorption/Desorption
Several experiments have been carried out to determine Koc values and a value of 2.4
(log value) has been selected as most relevant.
Accumulation
The bioconcentration factor in fish has been determined in a number of experiments,
with values up to 50 on a whole fish basis. The value predicted by EUSES giving a
BCF of 28 from log Kow has been used in the calculations.
4.3
Environmental Exposure
The manufacturers have developed predicted environment concentration (PEC) values
for air, water and soil for their own sites, using direct measurements in these
compartments where possible, and using site specific information with some default
assumptions in other cases. These PEC values include the contributions from the use
of perc as an intermediate, as this occurs only on sites where the substance is
manufactured.
The manufacturers have also discussed possible emissions with the dry cleaning
industry in the EU, and have estimated releases from different types of dry cleaning
machine. For the local assessment, emissions have been considered from the worst
case machine (open type). Regional and continental emission estimates take account
of the relative proportions of each type of machine in use.
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Case Study 3: Tetrachloroethylene
Similarly, the metal cleaning industry has been consulted to provide information on
releases from this area of use. This information has been used to make estimates of
local, regional and continental emissions from this area of application.
The estimated emissions have been used in the EUSES program, together with the
information on the properties of the substance, to estimate PEC values (along with
those measured directly at some production sites). The resulting values are reported
in Table 4.4.
Table 4.4: PEC Values for Perc
Use step
Production/intermediate
Soil (µg/kg)
Air (µg/m3)
Site
Water (µg/l)
A
0.02
36
B
0.011
7.3
C
5
1.2
D
0.85
E
9.1
3.6
F
4.2
1.2
3.9*
0.88
Dry cleaning
0.02
0.06
4.4
Metal cleaning
1.6
2.5
7.7
Regional
0.011
0.005
0.88
Note: * - soil concentration calculated for combination of largest emissions to air and to water, which
are not from same site, so the value is a worst case for production
In general, measured levels in water are below 1 µg/l, with occasional higher values
near to sources. There are few measurements in soil. Levels in urban air are
generally below 10 µg/m3, and most are below 1 µg/m3. The values seem to be in
reasonable agreement with the calculated values.
Concentrations in foodstuffs have been estimated using the methods in the TGD and
the concentrations in the environment given above. The resulting estimated daily
intakes for humans are:
•
•
•
•
4.4
Production/intermediate use
Dry cleaning
Metal cleaning
Regional
0.01 mg/kg bw/day
1.3x10-3 mg/kg bw/day
1.7x10-3 mg/kg bw/day
2.5x10-4 mg/kg bw/day
Environmental Risk Assessment
Given the above findings, the following risk assessment conclusions are drawn:
•
Page 3- 22
All PECs for the aquatic compartment are below the predicted no effect
concentration (PNEC) of 51 µg/l, hence no risk to this compartment exists.
RPA & BRE
•
The conclusion for the aquatic environment also applies to the sediment
compartment, as explained earlier. There is also no risk for micro-organisms in
waste water treatment plants (data not shown).
•
All soil PEC values are below the PNEC of 10 µg/kg, hence, no risks to soil exist.
•
As the substance is not accumulative, there are not expected to be risks for
secondary poisoning.
On the basis of the above findings, no further testing is proposed.
4.5
Human Health Exposure and Risk Assessment
Occupational exposure to perc will already be controlled through occupational
exposure limits within the workplace. These exist in all of the main workplace
environments. The current occupational exposure standard (OES), set by the Health
and Safety Executive (HSE) in the UK, for long-term exposure to perc is 50 ppm as
an 8-hour time weighted average (TWA).
With regard to carcinogenicity and reproductive toxicity, it is not believed that the
findings on toxicity and carcinogenicity of perc in rodents applies to humans given
both the evidence from human studies and the species differences which exist.
No classification as a CMR is required.
4.6
Risk Management
Note that it is assumed that proper use and disposal of perc takes place. Owing to the
controls already in place, this should protect against any further issues with regard to
man via the environment arising from ‘normal’ use and disposal in relation to
contamination of ground or surface waters.
Because perc is not a PBT chemical, it does not need to go to Authorisation. On the
basis of the above findings, it should also not be subject to any form of Accelerated
Risk Management.
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Case Study 3: Tetrachloroethylene
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RPA & BRE
5.
THE REACH DOSSIER CONSIDERED
5.1
The Evaluation Approach
The aim of developing the hypothetical dossiers is to provide a basis for comparing
what might have happened had REACH been introduced earlier with what happened
under the existing regime. In order to do this, we discuss below whether REACH
would:
•
•
•
•
•
5.2
require the same level of test data as required under ESR or other regulatory
regimes;
raise any concerns for the example substance and, if so, for which endpoints and
risk compartments;
identify the same endpoints and risk compartments as those identified
(historically) and controlled by the existing legislative arrangements;
recommend through this retrospective application, similar risk reduction measures
to those implemented at present; and
lead to action being taken sooner than under the current system and hence reduce
levels of environmental damage and risk to man via the environment.
The ESR Risk Assessment
5.2.1 Conclusions of the ESR Environmental Risk Assessment Compared to the
REACH Dossier
The Draft Risk Assessment for perc was issued in August 2001 by the UK, as
rapporteur, on behalf of the EU. The conclusions of the assessment, in terms of the
environmental risk characterisation, are presented by sector in Table 5.1
The ESR risk assessment found no risks from perc production or use for surface
water, sediment, waste water treatment plants or soils. Questions were raised during
the discussions on the assessment about the possible effects of perc on plants exposed
through the air, and about possible effects of breakdown products produced through
the degradation of perc in air. As a result, studies were instigated in both of these
areas. Tests on plants in open-topped chambers were carried out and a PNEC for
effects on plants exposed through the air was established. This indicated a risk at one
production site, based on measured emissions, and a conclusion (iii) was reached for
this.
The situation as regards the breakdown product is still under investigation. Some
member states believed that the available evidence was sufficient to reach a
conclusion of risk for this endpoint; other member states thought that further study
was required. These studies are still in progress and effectively a conclusion (i)
currently applies to this endpoint.
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Case Study 3: Tetrachloroethylene
Table 5.1: Risk Characterisation for Perc
(i) need for further
Activity/sector
information and/or
testing
Dry cleaning
Terrestrial
X
Aquatic
Air
Metal cleaning
Terrestrial
X
Aquatic
Air
Chemical synthesis
Terrestrial
X
Aquatic
Air
Source: Environment Agency, 2002
(ii) no need for
information or risk
reduction
(iii) there is a need
for limiting risks
X
X
X
X
X
X
The REACH assessment comes to the same conclusions for surface water, sediment,
waste water treatment plants or soils. The other two issues do not emerge directly
from the data requirements. The discussion on degradation in air in the dossier does
include some comments on the formation of the specific breakdown product,
trichloroacetic acid (TCA), which is the subject of the investigation under ESR in
relation to the terrestrial environment (TCA levels in soil have been identified as
posing a risk in some local scenarios).
However, there is an issue in relation to the lack of information on carbon
tetrachloride, which was of more concern at the time as a possible ozone depleting
chemical. The remarks on TCA indicate that its relatively quick removal from the
atmosphere is probably a benefit. The information requirements indicated so far for
the BIR Dossiers (see Section 2 of the main report) do not appear to require a detailed
consideration of potential breakdown products. Such a consideration may have been
included if there was evidence that biodegradation led to the production of a stable
product in high yield, but in this case the yield of TCA is relatively low (a few
percent).
In relation to effects on plants through the air, the BIR Dossier requirements do not
include any mention of testing by this route. At the time of the ESR assessment, there
were a small number of references in the literature to possible effects, but these were
not well reported or convincing, and the industry position was that they were not
scientifically valid. Under such circumstances, it is unlikely that the submitter would
have pursued this aspect, or even considered it. Since this assessment (and that for
dibutylphthalate) under ESR, there has been more consideration of whether
substances might affect plants through air exposure. Although no strategy for this has
yet been devised, this is seen as an issue particularly for volatile substances which
may be released to air in quantity. This should perhaps be considered for REACH at
higher tonnages with appropriate use or release patterns.
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RPA & BRE
5.2.2 ESR Human Health Risk Assessment Compared to the REACH Dossier
The draft ESR risk assessment for human health is not currently publicly available, so
no comparison can be made here between it and the conclusions of the hypothetical
industry dossier. However, it has been agreed that it be classified as a category 3
carcinogen and is a candidate for the 29th ATP. There is also some concern for
reproductive toxicity. Taken together, these may affect its use both in an occupational
setting and within consumer products.
It is understood that the last draft of the human health assessment did include
estimates of the uptake of perc from drinking water based on measured levels in
drinking water. This assumed that groundwater is the source of drinking water, and a
mixture of measured concentrations and calculated concentrations provided the basis
for estimating concentrations in drinking water. The assessment assumes though that
seriously contaminated water would not be used as a drinking water source. Hence,
contamination results in the loss of the resource, rather than increasing the exposure
of man to perc.
This issue was not addressed in the environmental risks assessment in terms of ‘man
via the environment’ in the environmental assessment. The reason for this is because
the assessments assume that regulations are in place to prevent any further
contamination. Thus, any further contamination would result from mis-use or misdisposal of perc.
5.2.3 ESR Risk Reduction Strategy Compared to the REACH Dossier
No risk reduction strategy has yet been prepared for perc under ESR, given the draft
nature of the risk assessments. The preparation of a strategy is not likely to take place
until the further information identified by the environmental risk assessment is
available, allowing for firmer conclusions to be reached with regard to trichloroacetic
acid and risks to the terrestrial environment and possible effects of perc on plants
exposed through the air. It may also await finalisation of the human health risk
assessment.
Given the conclusions reached to date, risk reduction in relation to the environment
will only address emissions of perc to air from its use in chemical synthesis. (If TCA
formed from the breakdown of perc in the atmosphere were shown to be a risk, then
this would apply to background levels and so all emissions to air would need to be
considered.) As noted above, a similar need for risk reduction does not arise in the
hypothetical REACH dossier prepared for this case study.
Risk reduction in relation to human health will have to address its new classification
as a category 3 carcinogen.
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Case Study 3: Tetrachloroethylene
5.2.4 General Conclusions
Given that the dossier has been produced to represent a production volume of greater
than 1,000 t/y, it would be subject to evaluation by a Competent Authority under
REACH. This would provide an opportunity for the Competent Authority to raise
questions concerning TCA, effects on plants exposed through the air, and
carcinogenicity and reproductive toxicity. REACH may fail to identify such issues in
cases such as this unless they were raised by the Competent Authority during
evaluation. It suggests that a forum such as the current Technical Meetings under
ESR could be important as part of the overall evaluation process within REACH.
Assuming that such issues are raised, it is likely that the Competent Authority may
request further information be provided or propose perc for accelerated risk
management and a Community risk assessment.
5.3
Historic Damage Costs Avoided
5.3.1 Overview
From the above, it is more difficult to draw conclusions for perc than for the other
case studies concerning damage costs avoided. Taking into account the discussion
provided in the preceding sections, two main kinds of damage costs emerge:
•
Damage costs in relation to the lack of test data: These are costs which have or
may be incurred simply due to the delays in further testing to validate early
research findings on the potential effects of perc. Such delays mean that the
potential costs of certain damages cannot be quantified, until the uncertainties are
resolved. Examples include the cost of the atmospheric releases of perc on plants,
or the carcinogenic and reproductive toxicity effects of perc on man; and
•
Damage costs in relation to the historic regulation on use and disposal of perc these are costs which have been incurred despite the different regulatory
initiatives which have been introduced since 1976, aimed at reducing releases of
perc to the environment. The effects of perc releases into groundwater despite the
76% decline in perc use in the EU since 1974 is an example of such costs.
The first point is discussed further in Section 5.3.2 below, with the second discussed
in Sections 5.3.3 and 5.3.4.
5.3.2 Damage Costs in Relation to a Lack of Test Data
The fact that further testing has been required under ESR highlights the fact that the
current regime places no duty on manufacturers of a substance to undertake new
testing and to prepare risk assessments concerning the different uses of their
chemicals. Under REACH, manufacturers would be required to undertake the further
testing now being sought through the ESR process to confirm whether or not perc
presents risks across all endpoints of concern. REACH thus effectively places a duty
Page 3- 28
RPA & BRE
of care on manufacturers with regard to any potential risks arising from release of
chemical to the environment.
In assessing the ability of REACH to lead to action being taken sooner than under the
current system, the chronology of research on perc highlights four landmark years at
which definitive action would have been taken under REACH and considerable
damage costs could have been avoided. These are:
•
in 1976, when perc was established as a List II substance;
•
in 1982, when perc was classed as a potential List I substance;
•
in 1986, when the US toxic program moved to classify perc as a possible human
carcinogen, but the proposal was defeated for lack of concrete evidence; and
•
in 1990, when perc was identified as a List I substance.
Should the testing for plants now being required under ESR highlight unacceptable
risks, then examination of the chronology of research would suggest that damages
first identified as being of potential concern have been on-going for almost 20 years.
Similar conclusions might also be drawn with regard to the potential risks posed by
TCA breakdown products on the terrestrial environment.
The same applies for the human health impacts. Had there been further testing
requirements placed on manufacturers, when epidemiological studies concluded in the
late 1980s and early 1990s reported that exposure presented an increased risk of
developing a range of different types of cancer, there may be less uncertainty
remaining today as to the carcinogenic and reproductive toxicity effects of
occupational exposure to perc. If perc is found to be a category 1 or 2 carcinogen in
the future as further testing is undertaken, then the lack of data may have resulted in
an increased number of cancers within the EU worker population.
5.3.3 Resource Damage Costs in Relation to Historic Regulation and Use
Groundwater
Although the discharge of perc to the water environment was regulated as early as
1976 through Directive 76/464/EEC, poor handling and disposal practices continued
well past this date. Indeed, as perc was only a List II substance under Directive
80/68/EEC, it could still be discharged direct to groundwater under a consent system.
It was not until 1990, when perc became a List I substance, that guidance on the use
and disposal of perc (along with trichloroethylene and other similar solvents) was
formally introduced in the EU under Directive 90/415/EEC (as a Daughter Directive
to 76/464).
It is obviously debatable whether having additional test data in relation to
carcinogenicity and a risk assessment could have changed this situation and led to a
reduction in the number of groundwater resources which have been contaminated.
Given the present knowledge relating to persistence of perc in groundwater, it is likely
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Case Study 3: Tetrachloroethylene
that more conclusive action would have been taken. Overall, the case for less perc
contaminated groundwaters is less debatable given the associated enormous
remediation costs involved, as shown in Section 5.3.4.
5.3.4 Remediation Costs of Perc Contaminated Groundwater
The remediation of perc contaminated groundwater is an expensive environmental
clean-up exercise, with costs of over €30 million incurred for a particular site in the
US (see Box 5.2). In order to appreciate the magnitude of the damage costs incurred
in these incidents, an overview of the technical issues arising is required.
Perc has peculiar physico-chemical properties which pose technical challenges to
conventional groundwater treatment technologies. One of the key issues is that perc
is denser than water. It thus does not mix well with water and once it reaches
groundwater, tends to sink to the bottom of the aquifer. In the process, it gets trapped
between soil particles in the aquifer. At the bottom of the aquifer, perc accumulates
in pools, from where it dissolves very slowly into groundwater - a process which
could take many years. Common treatment methods are basically ineffective for
dealing with these pools, and the cooler temperatures found in aquifers also make the
natural removal processes very slow.
There are two fundamental methods applicable to the remediation of perc
contaminated groundwater:
•
the Pump and Treat method uses wells to pull the contaminated water from the
aquifer, treating it above ground and then discharging it to a sewage treatment
plant or other approved location. Pump and Treat is expensive, however, with an
average cost of $0.25 per 1,000 gallons (€0.07/m3) of treated contaminated
groundwater (ESTCP, not dated). This is exclusive of the costs of well
installation and construction of overhead treatment systems. It is considered the
most versatile groundwater restoration technique and is highly effective when
used with other techniques, such as vapour extraction or physical barriers (Holden
et al, 1998); and
•
Bioremediation uses micro-organisms to remove perc from groundwater and soil.
The micro-organisms take in perc as they absorb water and nutrients, which can
then be stored in the bodies and/or changed into less harmful chemicals within the
organism. While the merits of bioremediation for groundwater remediation are
currently being investigated, particularly with regard to the capture of breakdown
products and the relatively lower costs, the case for being just as versatile and
effective as the pump and treat method is less clear cut.
The following case studies in Box 5.2 present the costs associated with perc
remediation in four sites in the US, Japan and UK.
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RPA & BRE
Box 5.2: Perc Remediation in US, Japan and UK
Case Study 1: San Francisco, USA
In 2002, the US Environmental Protection Agency announced a $37.25 million (€38 million)
Superfund settlement to pay for drinking water aquifer restoration costs in the San Fernando Valley in
California. The San Fernando basin was a primary source of drinking water for more than 800,000
residents in Los Angeles, Burbank, Glendale, and the La Crescenta Water District prior to
contamination by perc and other chlorinated solvents. The agreement included $13.25 million (€14
million) to recover the EPA's costs for feasibility studies, groundwater sampling, monitoring, and
oversight of treatment system design and construction and another $24 million (€25 million), kept
aside for the maintenance and operation of the system for 12 years over which the clean up is
expected to last. Cost estimates were, however, exclusive of $20 million (€21 million) incurred by the
responsible parties for the design and construction of a treatment plant and eight extraction wells.
Source: US EPA (2002)
Case Study 2: Connecticut, USA
In 1999, it was estimated that 312,000 residents were drinking water from 54 wells tainted with
contamination by organic chemical contaminants. In subsequent investigations into the clean up of
these contaminants, it was estimated that an average cost of $2.5 million (€2.7 million) would be
required to investigate, remediate, supply alternative water and/or provide treatment for contaminated
drinking water over a ten year period, approximately €5,000 per week. Based on this figure, it was
further estimated that the total costs to remediate and treat the 54 contaminated wells over a ten year
period would be approximately $135 million (€148 million), with the annual cost of treating only perc
estimated at about $500,000 (€550, 000) per year.
Source: CFE (2000)
Case Study 3: Japan
In May 1998, a large European chemical manufacturer in Japan reported soil contamination at their
factory, after the closure of the plant in 1992. The company announced a plan to clean up the factory
site over three years at an estimated cost of 7 billion yen (€55 million).
Source: US Department of Commerce (1999)
Case Study 4: Cambridge, UK
In 1983, perc contamination was discovered in a borehole in Cambridge, England. Between 1997 and
1999, the UK Environment Agency spent £150,000 (€210,000) investigating and modelling the
available remediation options, and had committed another £1.5 million (€2.1 million) to a pump and
treat scheme for aquifer remediation. A further £3 million (€4.2 million) would be needed to cover
operating expenses over the next 15 years, giving an estimated project cost of £4.65 million (€6.5
million). These costs were exclusive of €1.5 million) incurred by the water company which owned
the borehole for the development of alternative water sources.
Source: ENDS (1999) confirmed by Personal Communication, 2003
As can be appreciated from the figures presented in Box 5.2, the cost of cleaning up
perc leakages could, in some cases, exceed the value of the contaminated aquifer.
More often than not, the costs of the preliminary damage control measures - site
investigation, connection of alternative water supplies, and costs of operation and
maintenance of a water treatment system - prove to be quite significant. For instance,
the costs of providing emergency water supply for some residences near Santa Rosa,
California for perc contaminated groundwater affecting 30 domestic drinking water
wells, site characterisation, wellhead treatment systems and public outreach
associated with human health concerns amounted to well over $600,000 (€660,000) in
2000 (SWRCB, 2002). This is close to the UK figures from Holden et al (1998)
which range from £180,000 (€280,000) to £330,000 (€520,000) for initial response,
site investigations and consultants’ fees.
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Case Study 3: Tetrachloroethylene
Even in cases where a decision to remediate is agreed, the required groundwater
standard and the time scale of restoration work impose further. Table 5.5 shows the
impact of the required groundwater standards and time on costs of pump and treat
systems.
Table 5.5: Effect of Groundwater Standards on Costs of Pump and Treat
Target Removal
Calculated Years to Achieve
Present Worth ($ million)
80%
15
2.8
90%
21
3.25
99%
42
4.75
99.9%
63
5.6
Source: US National Research Council, 1994 cited in Holden et al, 1998
Calculations assume presence of chlorinated solvent of concentration of 1 mg/l
Considering the enormous amounts of resources, time and skilled manpower involved
in perc contamination, responsible authorities find that wells have to be abandoned
permanently or temporarily (usually years), and alternative sources of drinking water
supply have to be developed; this is illustrated by the following cases in Finland and
the UK (Box 5.3).
Box 5.3: Perc in UK and Finnish Groundwaters
Perc in Finnish Groundwaters
In Finland, there were 10 documented cases related to perc in groundwater between 1950 and 1975.
Nine of these were caused by drycleaners and the last by a metal surface plating plant. Overall
concentrations ranged from 2.5 – 100 mg/l, and in three cases were between 100 and 700 mg/l. In five
of the cases, the contamination had a more than strictly local effect on groundwater status, with the
water supply permanently closed down in four cases, and temporarily closed down in one case.
Source: Personal Communication with Chemicals Division, Finnish Environment Institute
Perc in UK Groundwater.
In 1992, the UK National Rivers Authority (NRA) began an investigation into a woollens manufacturer
after the discovery of chlorinated solvents in an aquifer near Berwick-upon-Tweed, England.
Investigations began after swans, whose plumage had been washed by the natural oils of the pollutant,
began to sink on the river Tweed. The investigation of these swans led to the discovery of solvent
contamination in the local Fell sandstone aquifer. The water, which was being abstracted by a
company from its own borehole, contained up to 35µg/l of perc, well over the statutory limit of <10µg/l
in drinking water. This aquifer acted as a drinking water supply for the area while also serving the firm
and measures had to be taken to protect other nearby sources.
Source: ENDS (1992)
Inevitably, for every drinking water source closed down, another source has to be
found for the region affected. Calculations have shown that the cost for the
development of a new groundwater source for each source lost to pollution is between
£1.8 - £2.2 million per Ml/day in the UK (RPA, 2002). Thus, for every water source
which is saved from perc pollution, the potential benefits/savings could amount to £2
million/Ml/day (€3.2 million/) multiplied by the number of sources polluted by perc.
In an assessment of the extent of groundwater pollution in the UK by the Environment
Agency, 250 confirmed occurrences of solvent contamination were discovered. Perc
was one of the top three contaminating solvents, particularly in the Midlands area
Page 3- 32
RPA & BRE
where overall groundwater contamination was specifically attributed to the
widespread industrial use of chlorinated solvents (Environment Agency, 1996).
Assuming that perc is detected in just 10% of these 250 abstraction sources polluted
in the UK, the savings could come to a minimum of £50 million. If this is considered
in terms of Europe, then the savings could become extremely significant. Although
there are no definite figures on the number of contaminated sites in Europe, a
significant number are known to exist, particularly in industrial areas and cities2. This
is highlighted by the results of monitoring activities reported in Section 3.3.3, for
France (in the Nievre region and across the Rhône-Méditerranée-Corse basin),
Germany (Baden-Württemberg) and a range of other countries.
In addition to the direct costs of groundwater remediation, health studies have shown
that if drinking water contains 0.5 µg/l of perc per litre, the average daily exposure
would be 1 µg for an adult consuming 2 litres of water per day. The US EPA has
estimated that one part per billion (1 µg/litre) perc in drinking water could lead to one
or two additional cases of cancer in a population of one million people who drink
water containing perc for a 70-year lifetime (University of Wisconsin, not dated).
In assessing the historic damage costs avoided from perc use, it can basically be said
that, for every case of perc contaminated groundwater in the EU:
•
•
•
•
the costs of preliminary damage control are likely to range from €280,000 to
€520,000;
the direct replacement costs, could be between €2.8 - €3.5 million per Ml/day;
the costs of remediation could be between €4 - €30 million; and
the costs associated with restoring to drinking water standard (99% purity) could
be around €2 million over 20 years.
It should be noted that in many cases of perc contaminated groundwater, only very
small quantities of perc are enough to contaminate large volumes of water. It is
estimated that 7 litres of chlorinated solvent can impact on 108 litres of groundwater
with an average concentration of 100 parts per billion (100µg/l) (Feenstra and Cherry,
1987).
According to Article 4 of the Water Framework Directive (WFD) (2000/60/EC),
member states would be required to protect, enhance and restore all bodies of
groundwater in their areas. One of the key concerns to be addressed by this
legislation is the abandonment of polluted water sources due to contamination. Water
is a scarce resource in the world, and the WFD aims at encouraging member state
authorities to restore contaminated waters to acceptable levels. This could impose
serious costs in cases where perc contamination is discovered as explained above.
Considering that in most cases, the authorities, taxpayers or water rate payers incur
the costs of remediation, (with the exception of when a responsible party is directly
2
The intensity and history of industrial activity has been found to influence the levels and type of
contamination observed in an area. For instance, solvent contamination in the UK was found to be
highly linked with industrial presence (Environment Agency, 1996); with perc contamination reported
in 78% of boreholes sampled in Birmingham, a historically industrial area of the UK (Rivett et al,
1990).
Page 3-33
Case Study 3: Tetrachloroethylene
implicated and can be held liable), it is evident that the prevention of contamination is
preferable to remediation.
It could be argued that only a short acceleration in perc becoming a List I substance,
given the potential persistence of perc in groundwaters and the magnitude of the
potential damage costs associated with the loss of groundwater resources as drinking
water supply sources across the EU, could have resulted in significant savings in
resource costs, as guidelines on use and disposal may have been developed earlier. In
addition, Member States may have taken more care to ensure a more robust
implementation of the Directives in relation to perc.
Such an accelerated listing of perc as a List I substance, could have equally
accelerated the decline in the use of perc in certain applications, and would most
likely have speeded up the voluntary movement of sectors such as dry cleaning and
metal degreasing/finishing away from the use of perc to substitutes; although this will
have been constrained to some degree by the fact that perc has itself been adopted as a
substitute for CFCs and 1,1,1-trichloroethylene.
Again, should the further testing now underway, highlight risks to both plants and to
the terrestrial environment, these are damages that could have been minimised by
earlier voluntary action on the part of downstream users, to adopt less damaging
substitutes.
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RPA & BRE
6.
REFERENCES
Alexander HC et al (1978): Toxicity of Perchloroethylene, Trichloroethylene, 1,1,1-trichloroethane, and Methylene Chloride to Fathead Minnows, Bull. Environ. Contam.
Toxicol., Vol 20, pp344-352.
Anttila et al (1995): Cancer Incidence Among Finnish Workers Exposed to Halogenated
Hydrocarbons, Journal of Occupational and Environmental Medicine, Vol 37, No 7,
pp797-806.
Bauer U (1991): Occurrence of Tetrachloroethylene in the Federal Republic of Germany,
Chemosphere, Vol 23, No11-12, pp1777-1781.
Bessemer AC et al (1984): Criteria Document over Tetrachloroetheen, The Hague,
Ministerie van Volkshuisvesting, Ruimtelijke Ordening en Milieubeheer, 1984
(Publikatiereeks Lucht, No. 32).
Broderius S & Kahl M (1985): Acute Toxicity of Organic Chemical Mixtures to the Fathead
Minnow, Aqua. Toxicol., Vol 6, pp307–322.
Bruckmann et al (1988): Immissionsmessugen halgenierter organischer Verbindungen in
Hamburg, VDI-Berichte, Vol 745, pp209-234.
Buben JA & O’Flaherty EJ (1985): Delineation of the Role of Metabolism in the
Hepatotoxicity of Trichloroethylene and Perchloroethylene: a Dose-effect Study,
Toxicol. Appl. Pharmacol., Vol 78, pp105–122.
CFE (2000): Testimony of the Connecticut Fund for the Environment on Proposed
Aquifer Protection Areas-Land Use Controls May 31, 2000, downloaded from the
Connecticut
Fund
for
The
Environment
Internet
site
www.cfenv.org/aquifer%20testimony.pdf.
CINET (not dated): European Market of Chlorinated Solvents 1974-2000, Figure 16,
article
downloaded
from
the
CINET
Internet
site
www.cinetonline.net/figures.figure16.htm.
Class TH & Ballschmiter K (1987): Chemistry of Organic Traces in Air. VI: Distribution of
Chlorinated Cl-C4 Hydrocarbons in Air over the Northern and Southern Atlantic
Ocean, Chemosphere, Vol 15, pp413-427.
Deipser A & Stegmann R (1994): The Origin and Fate of Volatile Trace Components in
Municipal Solid Waste Landfills, Waste Management and Research, Vol 12, pp129139.
DETR (1999):
Regulatory and Environmental Impact Assessment for the
Implementation of the EC Solvent Emissions Directive, Appendix 2.1: Sector
Profile – Dry Cleaning, report prepared for the Department of the Environment,
Transport and the Regions.
Page 3-35
Case Study 3: Tetrachloroethylene
ECSA (1999): Retail Dry Cleaning: Challenges on all Fronts, Solvents Digest Special
Edition on Dry Cleaning, EuroChlor, July 1999.
ECSA (1996): Dry Cleaning: Your Options Considered, article downloaded from the
EuroChlor Internet Site www.eurochlor.org/chlorsolvents/ publications.htm.
ECSA (not dated): Tetrachloroethylene, article downloaded from the EuroChlor Internet
site www.eurochlor.org/chlorsolvents/science/science1.htm.
ENDS (1999): Sixteen Years on, Agency Discovers New Risks at Eastern Counties Leather,
ENDS Report, Vol 293, June 1999.
ENDS (1992): £1 Million Award in Historic Aquifer Pollution case, ENDS Report, Vol 214,
November 1992.
EEA (1999): Groundwater Quality and Quantity in Europe - Data and Basic
Information, Technical Report No.22, European Environment Agency,
Copenhagen, July 1999.
Environment Agency (2002): Draft European Union Risk Assessment Report –
Tetrachloroethylene, Final Environment Report, March 2002 (R021_0203_env).
Fast T & Van Winjen JH (1994): Exposure to Perchloroethylene in Homes Nearby
Drycleaners Using Closed Systems, Fourth European Meeting of Environmental
Hygiene, Wageningen, Netherlands, June 9-11, 1993, Zentralblatt für Hygiene und
Umweltmedizin, Vol 195, pp147-148.
Feenstra S & Cherry JA (1987): Dense Organic Solvents in Groundwater: An
Introduction, Ontario, Institute for Groundwater Research, University of Waterloo.
Fielding M et al (1981):
Levels of Trichloroethylene, Tetrachloroethylene, a pdichlorobenzene in Groundwaters, Environ. Tech. Lett., Vol 2, pp545-550.
Frank H & Frank W (1985): Chlorophyll-bleaching by atmospheric pollutants and
sunlight, Naturwissenschaften Vol 72, pp139–141.
Gay BW Jr et al (1976): Atmospheric Oxidation of Chlorinated Ethylenes, Environ. Sci.
Technol., Vol 10, pp58-67.
Guth et al (1997):
Categorical Regression Analysis of Acute Exposure to
Tetrachloroethylene, Risk Analysis, Vol 17, No 3, pp321-332.
Holden JM et al (1998): Hydraulic Measures for the Control and Treatment of
Groundwater Pollution London, CIRIA Report 186.
HSIA (1999): Perchloroethylene White Paper Halogenated Solvents Industry Alliance,
article downloaded from HSIA internet site www.hsia.org/white_papers/perc.htm
Page 3- 36
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IARC (1995): Tetrachloroethylene, IARC Monographs of the Evaluation of Carcinogenic
Risks to Humans, Vol 63, pp159-221.
Kenaga EE (1980): Predicted Bioconcentration Factors and Soil Sorption Coefficients of
Pesticides and Other Chemicals, Ecotoxicol.Environ.Safety, Vol 4, pp26-38.
Keynote (2002): Dry Cleaning and Laundry Services Market Report, Executive
Summary, downloaded from the Keynote Internet site www.keynote.co.uk
Konemann H (1981): Quantitative Structure-activity Relationships in Fish Toxicity Studies.
Part 1: Relationships for 50 Industrial Pollutants, Toxicology, Vol 19, pp209-221.
Lawrence AR et al (1990): A Method for Determining Volatile Organic Solvents in Chalk
Pore Waters (Southern and Eastern England) and its Relevance to the Evaluation of
Groundwater Contamination, Journal of Contaminant Hydrology, Vol 6, pp377-386.
McConnell G et al (1975): Chlorinated Hydrocarbons and the Environment, Endeavour, Vol
34, pp13-18.
Neely WB et al (1974): Partition Coefficient to Measure Bioconcentration Potential of
Organic Chemicals in Fish, Environ. Sci. Technol., Vol 8, pp1113–1115.
NTP (National Toxicology Program) (1986): Toxicology and Carcinogenesis Studies of
Tetrachloroethylene (Perchloroethylene) (CAS No. 127-18-4) in F344/N rats and
B6C3F1 Mice (Inhalation Studies), United States Department of Health and Human
Services, Technical report series no. 311.
Organisation & Environment (1991): Reduction of Volatile Organic Compounds from
Dry Cleaning Facilities, Final Report to the European Commission DGXI
(Environment).
OSPAR (1998): OSPAR Strategy with Regard to Hazardous Substances, Sintra: 22-23
July 1998
Pearson CR & McConnell G (1975): Chlorinated C1 and C2 Hydrocarbons in the Marine
Environment, Proc. R. Soc. Lond. B., Vol 189 pp305-332.
Rivett MO et al (1990): Chlorinated Solvents in UK Aquifers, J. Inst. Water and Env.
Mang., Vol 4 No3, pp242-250.
RPA (2002): Charges Scheme Review: Costs of Compensation, Risk & Policy Analysts
Limited, report for the Environment Agency, June 2002.
Schwarzenbach RP et al (1983): Behaviour of Organic Compounds During Infiltration of
River Water to Ground water. Field Studies, Environmental Science and Technology,
Vol 17, pp472-479.
Singh HB et al (1975):
Atmospheric Formation of Carbon Tetrachloride from
Tetrachloroethylene, Environ Lett, Vol 10, pp253-256.
Page 3-37
Case Study 3: Tetrachloroethylene
Smith AM (1995): Chlorinated Solvents: A Sustainable Future, article downloaded from
the
EuroChlor
Internet
Site
www.eurochlor.org/chlorsolvents/
publications/chsol.htm.
SWRCB (2002): State Water Resources Control Board Meeting - Division of Clean
Water Programs June 20, 2002, downloaded from the SWRCB of California,
USA Internet site .
University of Wisconsin (not dated): Drinking Water Contamination: Understanding the
Risks, Extension publication G3339.
US Department of Commerce (1999): Japan - Environmental Remediation - Market
Assessment - ISA990301, downloaded from the US Department of Commerce
Internet site www.strategis.ic.gc.ca/SSG/dd72412e.html.
US EPA (2002): The U.S. EPA Consumer News Release: Tetrachloroethylene. Region 9
News Releases: 51 parties Ante up for Glendale Groundwater Clean Up,
downloaded from the US EPA Internet site www.epa.gov/region09.htm
Vartiainen T et al (1993): Population Exposure to Tri- and Tetrachloroethene and Cancer
Risk: Two Cases of Drinking Water Pollution, Chemosphere, Vol 27, pp1171-1181.
Veith et al (1980): An Evaluation using Partition Coefficients and Water Solubility to
Estimate Bioconcentration Factors for Organic Chemicals in Fish, Aquatic Toxicol.,
pp116-129.
Vonk et al (1986): Comparison of the Effects of Several Chemicals on Micro-organisms,
Higher Plants and Earthworms, paper presented at the Contaminated Soil, 1st
International Conference, Dontrecht, Netherlands.
Vyskocil A et al (1990): Study on Kidney Function in Female Workers Exposed to
Perchloroethylene, Human and Experimental Toxicology, Vol 9, pp377-380.
Walbridge CT et al (1983): Acute Toxicity of Ten Chlorinated Aliphatic Hydrocarbons to the
Fathead Minnow (Pimephales promelas), Arch.Environ.Contam.Toxicol., Vol 12,
pp661–666.
Walker et al (1997): Cancer Mortality among Laundry and Dry Cleaning Workers,
American Journal of Industrial Medicine, Vol 32, No 6.
WHO (1984): Tetrachloroethylene, Environmental Health Criteria, No. 31, World Health
Organisation, Geneva.
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CASE STUDY 4:
TRIBUTYLTIN (TBT)
Case Study 4: Tributyltin
RPA & BRE
1.
INTRODUCTION
1.1
Background to the Case Study
Organotin compounds are man made substances that were first developed in the 1920s
as moth proofing agents, later being used as bactericides and fungicides (Moore et al,
1991 in Santillo et al, 2001). Of these organotin compounds, tributyltin (TBT) is
considered to be the most hazardous and it is this that gives TBT its biocidal
properties.
There are two major uses of TBT:
•
•
antifouling paints; and
wood preservation.
Minor uses include as a biocide in decorative paints, in-can preservatives and
preservation of film. It is also used as a biocide in carpets, through a substance called
‘Ultrafresh’. TBT has also been found where PVC has been used, and has been
identified as a contaminant from the use of butyltin stabilisers. It typically occurs in
concentrations of up to 1% by weight (as tin) in butyltin products (Kemi, 2000).
TBTs were chosen as a case study chemical for a number of different reasons
including:
•
•
•
1.2
their use first began about 70 years ago, during which, severe ecological effects
came to light;
the case study highlights the types of damages that could be avoided in relation to
chemicals that, although highlighted as a priority, are not linked to pollution
incidents and, thus, relate to less obvious environmental impacts caused by normal
(approved) use; and
although international, European and national measures limiting the use of TBT
based on its known impacts, have been in place, in some countries for almost 20
years, discussions and debates are still on-going into the environmental impacts of
TBT, and the continued use as antifouling paints on ocean going vessels.
Format of Case Study
A profile of the market for TBTs used in antifouling coatings and wood preservatives
is provided first (Section 2), with this including a brief description of how they have
been used in different applications. This is followed in Section 3 by the historical
review of how TBTs became an issue of concern, and when regulatory and other
action in response to concerns was initiated.
The hypothetical REACH dossier is then presented in Section 4. This includes a
summary of what we assume for this dossier in terms of production volumes, uses,
test data, exposure, risk assessment conclusions, further testing and risk management
recommendations. The dossier is then considered further in Section 5 and risk
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Case Study 4: Tributyltin
management measures from the hypothetical REACH dossier are compared with the
measures in place today. Further hypotheses are then made as to the damages that
could have been avoided had REACH been in place earlier.
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2.
MARKET PROFILE
2.1
TBT in General
Production data for TBT (mainly in the forms of TBTO and TBTCl) are provided in
Table 2.1.
Table 2.1: Production of TBT
Year
Location of TBT Production (tonnes per annum, tpa)
Worldwide
EU
Germany
US
1992
no data
8,000 tpa
8,000 tpa
no data
1995
35,000 tpa
no data
no data
no data
1996
4,000 tpa
3,000 tpa
3,000 tpa
800 tpa
2000
5,000 tpa
no data
no data
no data
Sources: DoE (1992); Pesticide Outlook (1995); WS Atkins (1998); GuT (2000)
The total market for TBT in the UK in 1992 was 1,000 tonnes per annum (tpa). Of
this, almost 90% was used in the production of antifouling paints, 10% was used in
producing wood preservatives and the remainder (<0.5%) was used in plastics
applications (DoE, 1992).
2.2
TBT in Antifouling Paints
The first antifouling marine paint was invented in 1915 by a Danish manufacturer, JC
Hempel, although copper plating has been used on ships for many hundreds of years.
The use of TBT in marine antifouling paints dates from the 1960s, initially as a
booster biocide in copper-based systems. Use of TBT-based formulations grew
rapidly in the 1970s, as their greater effectiveness was realised, to the point where
they had captured a major proportion of the antifouling market (Evans, 2000 in
Santillo et al, 2001).
The initial TBT-based formulations were ‘free association’ paints. Such paints
release the biocide rapidly and demand frequent application. The invention of ‘selfpolishing copolymer’(SPC) paints in the late 1960s offered a more constant release of
biocide, reduced resistance and allowed repainting intervals of 60 months to be
achieved (Drescher, 2000; Santillo et al, 2001). These SPC paints became widely
available from 1974 (DoE, 1992).
An estimated 4,000 organisms are believed to foul ships, ranging from sea grass and
algae to barnacles (Crisp, 1972 in WS Atkins, 1998). Fouling of a ship’s hull
increases resistance, requiring greater fuel consumption. The impacts can be such that
the level of fouling after six months without protection can increase fuel consumption
by 50% (Haak, 1996 in WS Atkins, 1998). Handling of the ship also becomes much
more difficult.
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Case Study 4: Tributyltin
Champ (2001) estimates that 70% to 80% of the approximately 28,000 commercial
ships used worldwide use TBT antifouling paints. This is very similar to the level of
use of TBT paints in 1996, when CEFIC estimated that TBT paints make up 70% of
all TBT paints used (CEFIC in WS Atkins, 1998).
The EU used an estimated 600 to 1,100 tpa of TBT in antifouling paints in 1992.
Average sales in the UK were 197,400 litres per annum of SPC paints and 37,000
litres per annum of free association paints. This level of sales relates to use of TBT
paints on 50% of all vessels greater than 25m in length, or 20% of all vessels,
including pleasure craft (DoE, 1992).
2.3
TBT in Wood Preservatives
TBT in the form of TBTO was first introduced as a broad spectrum biocide in wood
preservative formulations to UK markets in the early 1960s (SCL, 1995). By 1992,
wood preservatives containing TBT made up 90% to 95% of industrially treated
timber and joints used in above ground applications with an estimated market of 100
tonnes per year (t/y) (DoE, 1992).
The use of TBT in wood preservatives is believed to have been much lower in other
EU countries (DoE, 1992). For example, only four to five tonnes of organotin
compounds were used for surface-treated wood in Denmark in 1998, making up
between 3% and 8% of biocides used for surface treating of wood (Danish
Environmental Protection Agency, not dated). In Sweden, the use of tin-based
preservatives was always minor compared to other types (e.g. copper or chromiumbased or creosote) at around 40 t/y (Jermer, 2000).
Preservatives containing TBTO and TBTN have the advantage of offering long-term
proven protection which allows them to meet the 60 year lifetimes required by British
Standards. They are also non-swelling and non-grain raising organic solvents which
makes them ideal for joinery applications (DoE, 1992).
Each cubic metre of wood that is treated requires between 20 to 40 litres of wood
preservative, depending on the type of wood and its final use. Average usage is
around 25 litre per m3. The cost of treating a cubic metre of wood with TBT
preservatives was estimated at £13.75, or about 9% of the total cost of producing
treated wood (DoE, 1992).
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3.
ENVIRONMENTAL AND HUMAN HEALTH IMPACTS
3.1
Introduction
The release of TBT into the environment through the use antifouling paints has
resulted in measured environmental and human health effects. A historical overview,
highlighting the links between the science that has identified (mainly) environmental
impacts and the regulatory response in relation to TBT is given below.
This
historical overview is neither meant to provide a comprehensive summary of
scientific and other research concerning TBTs nor to question or validate research
conclusions.
The aim of this section is to:
1) review the scientific and academic literature to identify when research on different
hazardous properties began and when concern started to arise;
2) make chronological links between the scientific research and the introduction of
either voluntary or regulatory measures aimed at reducing risks to the
environment or to public health;
3) present monitoring data (where available) to illustrate the possible scale of
environmental damages that have occurred as a result of TBT use; and
4) to analyse the history of testing and risk management activities in relation to
properties of concern (persistence, bioaccumulation and toxicity, carcinogenicity,
etc.) and developing conclusions on the avoidable damages.
Figure 3.1, overleaf, provides a summary of the key points in the identification of
environmental impacts and the regulatory and voluntary responses to this. The Figure
and the supporting text provided below are arranged chronologically, beginning with
the initial indications of environmental problems in the late 1970s.
3.2
Development of Environmental and Health Concerns
Imposex was first described by Smith (1971 in Santillo et al, 2001) from studies on
the American mud snail. Work undertaken by Blaber (1970 in Santillo et al, 2001)
recorded the appearance of a penis in female dog whelks (Nucella lapillus) in
Plymouth Sound in the UK. The causal agent was, however, unknown until analytical
capabilities improved in the late 1970s and early 1980s (Santillo et al, 2001).
Simple correlations of the presence of deformed oysters in areas where there were
high numbers of boats painted with TBT-based antifouling paints provided the first
indications of problems. Levels of TBT in the aquatic environment were not
measured until this correlation was reported at the International Council for
Exploration of the Seas (ICES) and, subsequently, published. There were several
reasons behind this, including (Champ, 2000):
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Case Study 4: Tributyltin
difficulty in analysing for TBT which was (then) at the limit of detection; and
lack of standardised laboratory analytical procedures and standard reference
materials.
•
•
This meant that the first observations of impacts were made by biologists studying
affected oysters. It was such observations that brought about the ‘critical’ evidence
from France. Box 3.1 summarises the findings in the Bay of Arcachon, on the
Atlantic coast of France.
Box 3.1: Evidence of Environmental Effects of TBT from Arcachon Bay, France
Arcachon Bay was an important oyster area, producing 10,000 to 15,000 tonnes of Crassostrea giga
per year (Evans, 2000). The Bay was also an important area for boating, with numbers increasing
from 7,500 in the mid 1970s to 15,000 by the 1980s.
Evidence of imposex was first observed in the oyster drill (Ocenebra erinacea), quickly followed by
impacts in the oysters themselves. Fishermen noticed that few of the oyster larvae from a spawning
event in 1976 had survived. By 1981, the production of oysters had fallen to 3,000 tonnes per year
(Ruiz et al, 1996). Reproductive failure and shell thickening and deformation of adult oysters were
largely to blame.
The first reliable survey of organotins in the water was undertaken by Alzieu in the mid 1980s when
analytical techniques had improved sufficiently to allow TBT concentrations to be measured in
detail. Concentrations in sediment were not available until the 1990s.
Source: Based on Santillo et al (2001).
Financial losses for the oyster fishermen and the severity of the impacts on marine
species prompted action by the French Government. In January 1982, a temporary
ban on TBT paint containing more than 3% by weight of organotin was announced by
the French Ministry of the Environment. The ban applied to boats of less than 25
tonnes on the Atlantic or English Channel coasts. A decree in September 1982
extended the ban to the whole coastal area and to all organotin paints. The only boats
permitted to use organotin paints were those with hulls exceeding 25m in length or
hulls made of aluminium (Santillo et al, 2001).
The process towards legislation in the UK took longer than in France. Box 3.2,
overleaf, describes the steps taken by the UK Government to reach its decision to ban
the use of TBT-based paints on vessels less than 25m in length. The legislation also
meant that all antifoulants had to be registered as pesticides with the Advisory
Committee to approve sales and use (various sources including DoE, 1992; SCL,
1995; Champ, 2000; Santillo et al, 2001).
Concerns over the impact of TBT began to spread more widely with bans on the use
of TBT-based paints on boats of less than 25m length being introduced in the United
States, Australia, Canada, New Zealand, Norway, Sweden, Germany, the Netherlands,
and Japan between 1988 and 1990. Switzerland and Austria also banned all use of
TBT-based paints in freshwater bodies, while in Sweden, the use of TBT paints was
prohibited in inland waters and the Baltic and North Sea. Many of these countries
also require all antifoulants to be registered and set maximum leaching rates of 4 or 5
µg/cm2/day (ORTEPA, nd). The leaching rate limits were set by the registrants of
antifouling paints (Champ, 2000).
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RPA & BRE
Measured Effects and Concentrations
Regulation and Legislation
Voluntary Actions
French oyster industry reports lack of
spatfall and shell thickening in young and
adult oysters in late 1970s
Smith (1981) linked imposition of male
characters with TBT contamination
Production of oysters in Arcachon Bay
fallen to 3,000 tonnes per year (1981)
from 10,000 to 15,000 tonnes per year
(start of the 1980s)
Alzieu et al (1981) reported levels of
7,030 to 17,370 ng/g total tin in oysters
showing TBT accumulation
Bryan et al found TBT concentrations in
UK yacht marinas in mid 1980s of 1,000
ng/l
Alzieu et al (1986) first reliable survey of
organotins in water
Concentrations of TBT at Arcachon Bay
in water reduced from 900 ng/l tin in 1983
to <100 ng/l tin in 1985; not confirmed
below 10 ng/l tin until late 1980s
Hall et al (1987) identified maximum
concentrations of 1,800 ng/l in
Chesapeake Bay
France (1982) : Prohibited the use of
TBT-based paints on vessels <25m on
length, except for aluminium-hulled
vessels
UK (1985): Controls on sale of TBT
paints for use on small vessels.
UK (1987): Total ban on retail sale of
TBT paint in May 1987 for use on vessels
under 25m and on fish cages
PARCOM (Recommendation 88/1) :
Harmonised ban on retail sales for
application to pleasure boats and fish
cages
Bacci et al (1989) found concentrations
of up to 3,930 ng TBT/l in Italian
harbours
1989: Similar bans in:
Norway, Sweden, US, Canada, Australia,
New Zealand
Concentrations in UK waters close to
pleasure boat activity in 1989 were half
those of 1987 (but EQS of 2 ng/l only
achieved at one of 12 monitoring sites)
1990: Similar bans in:
Germany, Netherlands, Japan
No clear trend in UK sediment
concentrations from 1986 to 1989
(probably due to paint particles in
sediment)
Imposex found in over 100 species of
sea snails, bioaccumulation of TBT in
squid livers around Japan
Improved monitoring led to discovery of
more widespread TBT contamination,
indicating:
- relationship of imposex to density of
shipping traffic
- poor recovery of some areas
- accumulation of butyltin in marine
animals
US research found a lack of observed
relationships between bulk sediment TBT
and adverse ecological effects (Puget
Sound showing no relationship between
sediment levels and bioaccumulation
Surface water with high contaminants
found in North Sea - implication that the
contamination is from ocean-going ships
Movement of application of TBT-based
paints to non-regulated countries.
Deformities in oysters have been found
similar to those seen in Europe
1984: Attempt made to persuade the
paint industry in the UK to withdraw TBTbased paints.
European Union (1991) : Prohibited the
use of TBT-based paints on vessels
<25m on length; TBT antifoulants only
available in 20 litre containers
1991: Similar bans in:
non-EU European countries, South Africa
Japan (1992): TBT banned for all
vessels
PARCOM (1995) : Ministerial declaration
of fourth North Sea Conference (Esbjerg)
commits to working for global phase-out
of TBT paint within IMO
International (1997): Concept of global
phase-out of organotin containing paints
agreed at MEPC's 40th Session
Some operators stop using TBT paints
altogether or on some vessels (e.g. P&O,
Shell, British Petroleum, Stena,
Wallenius Lines)
Marine Coatings Group of the
Paintmakers Association of Great Britain
make voluntary move to stop production
of antifouling paints containing free TBT
European Union (1999) : Free
association paints with organotins as a
biocide banned from 1 September 2000
(Directive 99/51/EEC)
1998: Marine Environment Protection
Committee (MEPC) of International
Maritime Organisation (IMO) urged
member states to encourage the use of
alternatives to organotin antifouling
systems (pending the entry into force of a
mandatory instrument)
International (1999) : Resolution
adopted agreeing that the application of
all antifoulings containing TBT should be
banned worldwide by 1 January 2003 and
that a complete ban on the presence of
TBT antifoulings on ship's hulls be in
place by 1 January 2008
Cunard announces its entire fleet would
use TBT-free products from 2001
European Union (2002): Amendment to
the EU Marketing and Use Directive
(Directive 2002/62/EC) formally bans the
application of TBT antifouling paints to all
ships in the EU from 1 January 2003
Jotun Paints (2002) : no sales or supply
of its TBT antifoulings will be made after
31 December 2002
Figure 3.1: Environment Effects and Regulatory Responses
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Case Study 4: Tributyltin
Box 3.2: The Steps Towards Legislation in the UK
Work by Bryan & Gibbs in the mid 1980s in the UK found concentrations of TBT in yacht marinas of
1,000 ng/l during the summer. Even higher concentrations were measured where boats were cleaned
prior to retreatment with TBT paints. It was also recognised that concentrations varied seasonally,
with highest concentrations being in the spring when newly painted boats were launched. Other peaks
occurred due to cleaning of boats before old paint layers were removed (DoE, 1992).
In response, the UK government attempted to persuade the UK paint industry to withdraw its TBTbased paints. The paint industry considered that the new copolymer paints would result in a 3 to 5
fold reduction in emissions compared with the free association paints that were being replaced.
Government scientists, however, estimated leach rates to be 5 to 10 times higher than that stated by
the paint manufacturers. It was proposed, therefore, to introduce regulations that would make it an
offence to sell paints containing more than 0.4 grams of tin per 100 ml of paint for use on boats with a
hull length of 12m or less (SCL, 1995).
Such proposals prompted a strong response from the paint industry and yacht owners because it would
have prevented boat owners from buying any of the copolymer paints or TBT-boosted copper paint.
There was also a strong lobby in Parliament against the proposals (SCL, 1995).
A package of measures was announced in July 1985 by the UK Environment Minister that was
designed to reduce environmental concentrations whilst giving the paint industry and boat owners
time to adjust (SCL, 1995). These included:
•
•
•
•
•
regulations to control the retail sale of organotin-based paints;
a voluntary notification scheme for new antifouling agents;
the preparation of guidelines for cleaning and painting small boats coated with antifoulants;
the setting of an environmental quality target (EQT) for the concentration of TBT in water; and
monitoring the effectiveness of the action taken through enhanced and co-ordinated research
effort.
However, work being undertaken by Plymouth Marine Laboratory at the same time was
demonstrating that the degree of imposex in populations of the dog whelk (Nucella lapillus) was
related to small boat and shipping activity, increasing with greater tin concentrations, but showing no
correlation with any other elements. Significantly, the increase in occurrence of imposex coincided
with the increase in use of TBT-based paints (SCL, 1995).
The UK Department of the Environment chaired a meeting of experts in September 1985 to discuss
the use of organotin in antifouling paint. This included discussion on:
•
•
the establishment of an ambient water quality target concentration for TBT at 20 ng/l. This was
approximately 3 to 5 times lower than the concentration at which harmful effects had been
recorded; and
research and monitoring needs, including the setting up of a Government monitoring programme
to determine the effectives of the proposed controls.
By 1986, the additional toxicity data collected made it clear that the Environmental Quality Standard
would have to be set substantially lower than the 20 ng/l provisionally determined. This was
eventually set at 2 ng/l TBT for marine waters and 20 ng/l for freshwaters. High concentrations of
TBT in freshwater and from the use of TBT-based net dips in fish farming were also of concern.
By February 1987, it was clear that all antifoulants would be brought under statutory pesticides
approval, while Parliament indicated that a ban on the use of organotin compounds on small boats was
likely as the existing controls had not been effective. The use of TBT for fish cage/net treatment was
banned at the same time.
Sources: based on DoE, 1992; SCL, 1995; Champ, 2000; Santillo et al, 2001.
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3.3
The Need for Harmonised Controls
In response to growing concerns across the Member States of the European Union, the
Commission of the European Communities proposed an amendment to Directive
76/769/EEC (the marketing and use of certain dangerous substances and
preparations). This resulted in ‘organostannic compounds’ being listed such that they
could only be ‘sold to professional users in packaging of a capacity not less than 20
litres for use as substances or constituents of preparations intended for use to prevent
the fouling by organisms, plants or animals of:
(a) the hulls of boats of an overall length, as defined by ISO 8666, of less than 25m;
and
(b) cages, floats or nets and any other appliances or equipment used for fish and
shellfish farming’.
The important distinction between Directive 76/769/EEC and much of the national
legislation is that it relates to ‘organostannic’ compounds, which is much wider than
the banning of TBT-based paints. The EC banned the use of free association paints
with organotins as the biocide from 1 September 2000 through Directive
1999/51/EEC.
The Paris Commission identified that the use of TBT paints was causing ‘serious
pollution in the inshore waters of the north east Atlantic’ and recommended that
contracting parties should take effective action to eliminate pollution by TBT within
the Convention. PARCOM Recommendation 87/1 called for a harmonised ban on
retail sales of TBT-based paints for use on pleasure boats and fish cages. A further,
key, recommendation was to consider the use of organotins on seagoing vessels.
However, a debate in 1988 showed that a ban on this use was not achievable.
PARCOM instead focused on docking activities and hull maintenance activities. This
led to contracting parties agreeing to ‘develop procedures and technologies aimed at a
reduction of the amount of organotins released from boatyards and drydocks due to
sand blasting, dust, paint chips, overspray, etc. and to implement them in the near
future’ (based on Champ, 2000; Santillo et al, 2001).
3.4
On-Going Regulation
The introduction of regulations has driven maritime companies to look for cheap
labour in countries with little or no environmental controls. Environmental scientists
in some of these countries are beginning to find deformed oysters. This suggests that
regulations designed to protect local waters have resulted in shipping companies
taking their business abroad, leading to environmental degradation and human health
hazards in non-regulated countries (Champ, 2000).
The International Maritime Organisation (IMO) agreed at the 42nd Session of the
Marine Environment Protection Committee (MEPC) to a draft Assembly resolution to
ban organotins acting as biocides in antifouling systems on ships. Two dates were
provisionally set:
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Case Study 4: Tributyltin
•
•
1 January 2003: global ban on the application of organotins acting as biocides in
antifouling systems on ships; and
1 January 2008: complete ban on the presence of organotin compounds acting as
biocides in antifouling systems on ships.
Subsequent sessions of the Marine Environment Protection Committee have involved
discussions on the development of a legal instrument. A conference to adopt the
proposed instrument was held in London on 1st to 5th October 2001. Several key
issues were discussed and resolved such that the final text for the treaty was prepared.
However, the time taken to agree the text has meant that entry into force will not take
place until after 1 January 2003 (IMO, 2002). As of January 2003, only nine
countries, representing almost 5% of the World’s fleet had signed or ratified the
Convention; the Convention is due to enter into force twelve months after 25
countries representing 25% of the gross tonnage of the world’s merchant shipping
have ratified it (IMO, 2003).
Following the recognition that the entry into force of the IMO Antifouling Systems
(IMO-AFS) Convention ban on antifouling paints containing organotins as a biocide
would be delayed, the European Commission developed a proposal for a regulation on
the prohibition of organotin compounds on ships. This regulation involved the
amendment of the Marketing and Use Directive (76/769/EEC) (European
Commission, 2002). This has resulted in a ban of all applications of TBT antifoulings
to all vessels flying the flag of an EU Member State from 1 January 2003, with the
presence of TBT on all ships irrespective of flag banned from 1 January 2008 (IMO,
2002).
In response to the future IMO-AFS Convention, the US EPA has asked registration
holders of TBT antifouling products to voluntarily agree to cancel their registrations
by 1 January 2003. Procedures have also been undertaken to implement the
Convention in domestic US legislation (IMO, 2002). Paint companies, such as Jotun,
responded to the proposed IMO-AFS Convention by stating that they would withdraw
their TBT antifoulings from the market after 31st December 2002 (Arnold, 2002).
3.5
Key Properties and Presence in the Environment
3.5.1 Presence of TBT in the Environment
TBT may be released to the environment from its use in antifouling paints through:
•
•
•
application of antifouling paints to ships’ hulls;
washdown or removal of antifouling paints from the hull surface; and
leaching of TBT from the paints while the vessel is in service (DoE, 1992).
Leaching from ships’ hulls can be up to 8 µg TBT/day/cm2 from SPC paints when the
ship is at sea. Concentrations from washdown water from drydocks can contain up to
14,000 ng/l to 10 mg/l TBT and is often the major point source of TBT to marine
waters (DoE, 1992).
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RPA & BRE
The introduction of regulations controlling the use of TBT-based paints on small
boats was followed by a general decline in TBT concentrations in water, sediment and
biota. Declines have been demonstrated in surface marine waters in France, the UK,
the US, the Gulf of Mexico and Australia (Champ, 2000). Studies in the UK showed
TBT concentrations in 1989 close to pleasure craft activity had fallen to half those
measured in 1987 (Waite et al, 1991 in DoE, 1992). Declines in three specific
locations in France and the UK are shown in Figure 3.2. While concentrations of
TBT in Arcachon Bay fell by 800 ng/l between 1983 and 1985, it took until the late
1980s before concentrations fell below 10 ng/l in most parts of the Bay (Alzieu et al,
1986 in Santillo et al, 2001).
Blackwater Estuary
2,500
2,000
1 ,500
1 ,000
500
0
1 986
1 990
Crouch Estuary
1 ,8 0 0
1 ,6 0 0
1 ,4 0 0
1 ,2 0 0
1 ,0 0 0
80 0
60 0
40 0
20 0
0
19 86
19 90
Arcachon Bay
1 000
900
800
700
600
500
400
300
200
1 00
0
1 982
1 985
Figure 3.2: Change in Concentration of TBT (ng/l) in Water (Arcachon Bay) and Oysters
(Crouch and Blackwater Estuaries) (based on various sources including Alzieu et al, 1986 in
Santillo et al, 2001; Waite et al, 1991 in DoE, 1992).
Page 4-11
Case Study 4: Tributyltin
Shellfish farming improved in France, southern England, Ireland and Australia. The
flame shell (Lima hians) recovered quickly in Mulroy Bay, Ireland, having been
decimated by the use of TBT-based paints on salmon cages between 1981 and 1985
(Santillo et al, 2001). There has also been a reported decrease in imposex and
population recovery in dog whelks (Nucella lapillus) in England, Scotland, Ireland,
Norway and Canada (Champ, 2000).
However, not all measurements have shown an improvement. Waite et al, 1991 (in
DoE, 1992) found no clear trend in sediment concentrations, probably due to paint
particles within the sediment. Hot spots associated with ship channels, ports,
harbours and marinas are also exceptions to the general decline. Examples of such
hot spots have been found in the Netherlands, Iceland, Israel, Hong Kong and
Galveston Bay in the US (Champ, 2000).
Improvements in the level of monitoring have also found concentrations of TBT in
the open ocean. For example, a zone of water 100 to 200 km offshore (German Bight
to the North Sea) has been found to have concentrations of TBT exceeding 20 ng/l (or
ten times higher than that needed to induce imposex in dog whelks) (Stebbing &
Dethlefsen, 1992; Hardy & Cleary, 1992 both in Champ, 2000). Research in Japan,
where TBT use on all vessels has been banned since 1992, has correlated marine ship
traffic to TBT levels in water and sediments, and determined that ocean-going vessels
were the source of TBT.
3.5.2 Properties
The key properties of TBT are as follows.
•
Persistence: TBT has been found to show slight to moderate persistence in water.
It is, however, significantly more persistent in sediments, with studies indicating a
half-life of up to 15 years in sediment. In terms of the PBT criteria, TBT meets
the persistence criteria.
•
Bioaccumulation: TBT shows significant bioaccumulation in the aquatic food
chain. The Japanese were the first to assess the impact of TBT in the deep sea,
finding that squid livers accumulated TBT 48,000 times ambient concentrations.
Substantial bioaccumulation has also been found in studies with algae, aquatic
invertebrates and fish, with whole body bioconcentration factors ranging from up
to 50,000 in fish to 500,000 in clams. Although TBT does not appear to
significantly biomagnify up the food chain, it has been found in the tissues of
marine mammals. Iwata et al (1995 in Santillo et al, 2001) found levels of up to
10 ppm (10,000,000 ng/l) in porpoise liver. Kannan et al (1996 in Champ, 2000)
noted butyltin accumulation in dolphins, tuna and sharks from the Mediterranean.
TBT meets the bioaccumulative criterion of the PBT assessment.
•
Toxicity: TBT has been found to be toxic to many aquatic organisms. Toxic
effects have been noticed in polychaetes, amphipods, marine benthic organisms
and fish, with endocrine disruption effects observed in dogwhelks. Imposex was
found in more than 100 species of sea snails (Japanese submission to the NIEPC
conference in Champ, 2000), and has now been documented in 150 species
Page 4-12
RPA & BRE
worldwide (Vos et al, 2000 in Santillo et al, 2001). Acute toxicity to some fish
occurs at a few milligrams per litre, while chronic toxicity, particularly in oysters
and clams have been discovered at fractional micrograms per litre concentrations.
Concern is now expressed that accumulation of butyltins may affect the immune
system. This follows the discovery of TBT in the tissues of stranded dolphins in
Florida and the suggestion that TBT may have been the cause of numerous
dolphin deaths off the US coast in 1987-88, the Gulf of Mexico in 1990 and the
Mediterranean Sea in 1990-91. However, there is no evidence to link TBT with
these deaths (ORTEPA, nd-a; Champ, 2000). TBT meets the toxicity criterion for
the PBT assessment on this basis.
•
PBT: overall TBT is classified as a PBT substance.
3.5.3 Human Health Concerns
There is limited human epidemiological data on the toxic effects of TBT in man.
Most research has focussed on workers and ship painters, which indicate that TBT
may have primary irritant effects on contact with skin or nasal mucosa.
These correlate with animal studies which have also shown skin and eye irritant
effects, and possible corrosive effects at higher concentrations. Immune system, liver
and haematological system effects have also been noticed, as have teratogenic effects
in the rat, mouse and rabbit (DoE, 1992).
Box 3.3 below presents the current classification of TBT.
Box 3.3: Current classification of TBT
TBT has been assigned the following risk phrases under Directive 67/548/EEC:
T; R25-48/23/25 Xn; R21 Xi; R36/38 N; R50-53
R25-48/23/25 – Toxic if swallowed; danger of serious damage to health by prolonged exposure
through inhalation and if swallowed.
R21 – Harmful in contact with skin
R36/38 – Irritating to eyes and skin.
R50/53 – Very toxic to aquatic organisms, may cause long-term adverse effects in the aquatic
environment.
3.6
Substitutes
A review of the available information on alternatives to TBT use in antifoulants is
given below. This highlights the importance of an integrated database system, such as
proposed under REACH, in which the information required to make informed and
proper decisions on substitutes to hazardous substances is readily available.
Much of the work on finding alternatives to tin-based antifoulings has been
undertaken in Japan, where a ban has been placed on the use of TBT in antifoulants
since 1992.
Page 4-13
Case Study 4: Tributyltin
The available alternatives can be divided into two major types (Chapman, 2002):
tin-free self-polishing technology; or
biocide free (also called foul release).
•
•
Table 3.1 shows a summary of the main technologies and some of their properties.
Table 3.1: Summary of Tin-Free Technologies
Technology
Basis
Self-Polishing Technology
Based on
Based on
Based on
gum-rosin
metallic
organo-silyl
ion-sensitive
acrylate resin
acrylate resin
polymers
Based on hard
silicone
polymers
Based on
softer silicone
polymers
-
-
Restricted to
ships faster
than 15 knots
60% to 80%
depending on
amount of
biocide
90%
95%
(equivalent to
TBTantifouling)
90%
Hydration
Hydrolysis
Hydrolysis
None – relies
on ‘non-stick’
principle
Restricted to
ships faster
than 5 knots
95%
(equivalent to
TBTantifouling)
None – relies
on ‘non-stick’
principle
3 years
5 years
5 years
5 years
5 years
Restrictions
-
Performance
rating1
Polishing
mechanisms
Claimed
service life
Biocides (tinfree)
Biocide Free
cuprous oxide, cupric oxide, zinc pyrithione,
None
None
Zineb, Irgarol, Diuron
1.5 times the
2.5 times the
2.5 times the
1 to 6 to 5 to 7 times the cost of
Cost (relative
cost of TBT
cost of TBT
cost of TBT
TBT antifoulings (average of
to TBT)
antifoulings
antifoulings
antifoulings
around 6 times)2
Based on: Chapman (2002); Danish EPA (not dated); Head & Klijnstra (2002); Nygren (2002)
Notes: 1 performance level is measured in terms of the percentage of ships drydocking with less than
10% weed or animal fouling (slime fouling is not taken into account).
2
cost estimates vary according to author
The main environmental concerns are associated with the biocides used in the selfpolishing technology paints. A range of metal and organic biocides are available,
with the most common systems using a copper-based biocide with an organic booster.
The use of copper-based biocides makes up almost all of the market. In Denmark, for
example, 4.5 to 9 tonnes per year of copper are used, compared with less than one
tonne of other compounds (Danish EPA, not dated). Copper is most effective against
hard fouling organisms such as barnacles, with the organic biocides being more
effective against soft organisms such as grasses and algae. It is generally recognised
that cupric (Cu2+) oxide is the most toxic to aquatic life (Drescher, 2000). The US
Navy found, however, it is 7 to 40 times less toxic than TBT in short-term tests (DoE,
1992). Copper has a lower toxicity than TBT which means that larger quantities have
to be incorporated into and released from antifouling coatings to provide sufficient
protection to a ship’s hull (Ranke & Jastorff, 2000).
Overall, the toxicity of the biocides currently being used is often not well known, with
copper compounds being the most researched of the metal or organic biocides.
Page 4-14
RPA & BRE
Potential biocides from natural sources, such as seaweed and sponges are also being
researched (Head & Klijnstra, 2002).
Madsen et al (1999) have undertaken an assessment of the environmental risk from
three alternative active ingredients to TBT-based antifouling paints and two biocide
free (foul release) coatings. These alternatives are:
•
•
active ingredients:
- copper;
- DCOI; and
- zinc pyrithione.
foul release coatings:
- epoxy-based paint; and
- silicone-based paint.
The main conclusions were (Madsen et al, 1999):
•
copper: binds to anoxic substances (particularly sulphides) in harbours such that
the bioaccessibility is low in the aquatic environment. However, if the sediment
is disturbed, the bioaccessibility may increase such that sensitive organisms near
harbours or waters used for the disposal of dredged material could be affected;
•
DCOI (4,5-dichloro-2-n-octyl-4-isothiazolin-3-on): breaks down rapidly into
substances that are at least 10,000 times less toxic than the original substances1.
Chronic effects from DCOI on algae, crustaceans and fish may occur, however, in
marinas (based on risk coefficient). Bioaccumulation of DCOI is rated as high by
Ranke & Jastorff (2000), with BCFs greater than 100 due to its association with
body tissues;
•
zinc pyrithione: also found to break down rapidly to substances at least 2,500
times less toxic than zinc pyrithione2, but may pose a risk to aquatic life from
chronic effects in marinas (based on risk coefficient). Synergistic effects between
zinc pyrithione and copper have not been investigated, although a stable copper
complex is formed when the two substances are released together from
antifouling coating (Ranke & Jastorff, 2000);
•
epoxy-based paint: tests showed it to be 1,000 times less toxic than TBT-based
paints. Effects were only found on crustaceans exposed to undiluted water
samples from tests where plastic sheets coated with the paints were submerged in
aquariums (NOEC chronic = 100 ml/l); and
•
silicone-based paint: showed effects on both algae and crustaceans exposed to
diluted water samples from tests on submerged plastic sheets coated with the
1
DCOI - acute toxicity: 2.7 to 14 µg/l EC50/LC50; breakdown products – 90,000 to 160,000 µg/l
EC50/LC50. DCOI – NOEC of 0.63 µg/l; breakdown products – NOEC of 16,000 µg/l (Madsen et al,
1999).
2
Zinc pyrithione – acute toxicity: 2.6 to 28 µg/l EC50/LC50; breakdown products – 29,000 to 72,000 µg/l
EC50/LC50. Zinc pyrithione – NOEC of 1.2 µg/l; breakdown products - no data (Madsen et al, 1999).
Page 4-15
Case Study 4: Tributyltin
paint (NOEC acute <100 ml/l; NOEC chronic <10 ml/l). This paint showed
chronic NOECs that were 100 times higher than TBT-based paints.
Madsen et al (1999) notes that risks from DCOI and zinc pyrithione have not been
calculated for harbour environments or sailing routes and that the biocide free paints
tested had not been fully developed by the manufacturers at the time of testing.
Variations in production and application of the biocide free paints may also affect
leaching.
Ranke & Jastorff (2000) report bioconcentration factors (BCF) for copper of greater
than 1,000 for some algae, macroalgae and bivalves (including Crassostrea virginica
which has a BCF of 28,000), with BCFs greater than 10,000 for crustaceans. The
BCFs for fish are lower, however, at 150 to 700 suggesting that higher organisms can
regulate copper by excretion. This means that bioaccumulation along the food chain
is not found.
Irgarol is a symmetric triazine which inhibits photosynthesis. Bioaccumulation is
considered to be high, with BCFs of at least 1,000 for macrophytes (Ranke & Jastorff,
2000). Recent work has indicated that Irgarol, an organic biocide, may be killing
small algal organisms on which shellfish and other marine creatures feed (SCL,
1995). It is now banned in Denmark, as is Diuron, for use on boats of less than 25m
length (Danish EPA, not dated). Use of DCOI in Sweden is subject to the same
restrictions and regulatory status as Diuron (Ranke & Jastorff, 2000).
Table 3.2 provides the results of a comparative risk assessment of different antifoulant
coatings based on work undertaken by Ranke & Jastorff (2000). The Table provides
an indication of relative risks, but does not give an indication of the difference in
risks. For example, TBT has much higher bioaccumulation and biological activity
than the other biocides. It is not possible, therefore, to place a definite ranking or
scoring of the substances (Ranke & Jastorff, 2000).
Table 3.2: Comparative Risk Assessment of Antifouling Coatings
Release
Spatiotemporal
Biocide
Bioaccumulation
Rate
Range
TBT acrylate
Lowest
Lowest
Highest
Other TBT
Mid
Lowest
Highest
compounds
Cu acrylate
Mid
Highest
Mid
Other Cu
Highest
Highest
Mid
compounds
Irgarol
Lowest
Lowest
Mid
DCOI
Lowest
Highest
Zinc
Lowest
Highest
pyrithione
Based on: Ranke & Jastorff, 2000.
Page 4-16
Biological
Activity
Highest
Remaining
Uncertainty
Lowest
Highest
Lowest
Mid
Mid
Mid
Mid
Lowest
Mid
Mid
Mid
Mid
Lowest
Mid
Highest
RPA & BRE
4.
THE REACH DOSSIER
4.1
Basic Assumptions
This dossier has been produced by a single producer acting independently. The
producer manufactures around 500 tonnes of the substance per year, and so has
addressed the requirements of Dossier C. Data are presented as a single dossier,
rather than by the requirements of Dossier B first and then the additional items for
Dossier C.
The data have been taken from the IUCLID submission on this substance,
supplemented with a small amount of data from other sources. An exception to this is
the information on measured levels in the environment. For this exercise, measured
levels relate to the time before restrictions on the use of TBTO in anti-fouling paints
were put in place.
The use associated with the submission is as an anti-fouling agent in paints for boats,
ships, etc. More specific use is considered in the conclusions to the dossier.
4.2
Base Data
4.2.1 Identity of substance
The substance is identified by the CAS Number 56-35-9. The EINECS name is
bis(tributyltin) oxide. In this document, it is referred to as TBTO. Tributyltin species
in general are referred to as TBT.
Methods of detection are available for water. The speciation of the substance may
change between oxide, chloride, carbonate, etc., in water, so concentrations are often
reported as TBT. In addition, concentrations may be reported as total tin, which will
also include any dibutyl and monobutyl substances present.
4.2.2 Physico-chemical data
The basic physico-chemical data are presented in Table 4.1.
Table 4.1: Physico-chemical Data
Physical state (at ntp):
Liquid
Melting point:
circa -45°C
Boiling point
220-230°C at 13 hPa
Relative density:
1.17 at 20°C
Vapour pressure:
circa 0.3 Pa at 25°C
Octanol-water partition coefficient:
3.8 (Log Kow)
Water solubility:
71.2 mg/l at 20°C
Flash point
~190°C
Page 4-17
Case Study 4: Tributyltin
4.2.3 Ecotoxicity data
Aquatic toxicity
A wide range of species have been tested for the effects of TBTO. The function of
the substance in use is to prevent the growth of organisms on the hulls of boats, and
this effect is achieved through toxicity to certain types of aquatic organisms. Hence,
the substance has a high toxicity to aquatic organisms. The data are summarised in
Tables 4.2 and 4.3.
Table 4.2: Toxicity of TBTO to freshwater organisms
Trophic level
Fish
Invertebrate
Algae
Species
Effect
Concentration (µg/l)
Lepomis macrochirus
96 hour LC50
3.4
Oncorhynchus mykiss
96 hour LC50
2.9
48 hour EC50
4.6
21 day NOEC
0.08
Anabaena flos-aquae
4 hour EC50
13
Ankistrodesmus falcatus
4 hour EC50
20
Scenedesmus quadricauda
4 hour EC50
16
Effect
Concentration (µg/l)
Cyprinodon variegatus
96 hour LC50
5.05
Fundulus heteroclitus
96 hour LC50
24
Alburnus alburnus
96 hour LC50
6-8
Mysidopsis bahia
96 hour LC50
1-2
Acanthomysis sculpta
96 hour LC50
0.41
Crangon crangon
96 hour LC50
1.5
Acartia tonsa
6 day NOEC
0.011
Neanthes arenaceodentata
70 day NOEC (reprod)
0.05
Nucella lapillus
365 day NOEC (reprod)
0.008
Skeletonema costatum
72 hour EC50
0.33
Thalassiosira pseudonana
72 hour EC50
1.03
Daphnia magna
Table 4.3: Toxicity of TBTO to saltwater organisms
Trophic level
Fish
Invertebrate
Algae
Species
There are no toxicity data for micro-organisms relevant to the assessment of the risks
to a waste water treatment plant.
Terrestrial toxicity
No test results for higher plants or soil dwelling organisms are available. The pattern
of use in anti-fouling paints means that release to soil is unlikely, and, hence, it is
considered that there is no need for testing on terrestrial organisms.
Page 4-18
RPA & BRE
Avian toxicity
Although not required at this level of dossier, a study on toxicity to Japanese quail
(Coturnix coturnix japonica) has been conducted. The NOEC for the reproduction
rate from a 56 day study was 24 mg/kg food.
4.2.4 Environmental fate
Biodegradation
A standard test indicates that the substance is inherently biodegradable. Metabolism
by bacteria and micro-organisms in the environment has been noted; the rate varies
considerably with conditions, and estimated half lives range from 3 to 60 days in the
aquatic environment. Studies in sediments from Toronto Harbour showed a half life
of around 4 months.
Abiotic degradation
Studies have shown that TBTO can undergo photolytic degradation in water, and that
the process can be sensitised by the presence of fulvic acids. However, the
significance of this route for degradation in the environment is not clear; attenuation
of UV light in water by particles and turbidity means it is probably limited.
Sorption
There are no specific test data on sorption available. Strong sorption to sediments has
been observed. A Koc value of 90,800 has been measured in sediments from Toronto
Harbour.
Accumulation
A bioconcentration factor of 2,600 in whole fish has been measured (the species was
Cyprinodon variegatus). In oysters, Crassostrea gigas, values of 2,200 – 11,400 have
been measured.
4.2.5 Exposure
Regular monitoring of the effluent from the production pant is carried out for
organotin compounds. Levels are always within those permitted by the local
authority.
The nature of the use of the substance means that the relevant concentrations are those
in water. The circumstances of application of the paint, and the degree of leaching
from vessels in use vary considerably, so calculation of exposure concentrations is not
considered appropriate in this case. Instead, use will be made of measured levels in
water relating to the use of paints containing the substance.
Levels of 14 µg/l to 10 mg/l have been measured in wash down water from dry docks
(as total tin in this case).
Page 4-19
Case Study 4: Tributyltin
The concentrations in water in areas with a high concentration of boats can reach up
to 1 µg/l, with values more usually ~0.1 µg/l. Concentrations in open waters are
lower.
4.2.6 Risk Assessment
The purpose of this substance is to prevent the growth of organisms on the hulls of
vessels. Hence, it is intended to have a toxic effect. In theory at least this effect is
desired against any species that could attach to the surface of the vessel and promote
the growth of other organisms. However, the data base on toxic effects includes a
number of organisms which do not fit into this category, such as fish, sediment
worms, and molluscs such as the dog whelk (Nucella lapillus). The lowest no effect
level for these species is 0.008 µg/l for Nucella. This concentration is clearly
exceeded in areas where large numbers of boats are moored, as well as in wash waters
from dry docks. It is, therefore, concluded that the current use pattern can lead to a
risk to aquatic organisms. In the absence of specific data on sediment organisms, the
equilibrium partition method would be used, and so the conclusions for sediment
would be the same.
Considering the possible risk through secondary poisoning, the BCF for fish is 2,600.
At the higher measured concentrations, this would lead to a concentration in fish of
2.6 mg/kg. A NOEC for birds of 24 mg/kg leads to a PNEC of 0.8 mg/kg (assessment
factor of 30). This indicates that there may be a risk in areas with higher
concentrations. There would not be a risk at concentrations of ~0.1 µg/l.
There is no route to terrestrial exposure through the normal use pattern, and so no
assessment for this endpoint has been carried out. The conclusions of the risk
assessment are provided in Table 4.4, where conclusion (iii) means risk reduction
would be required.
Table 4.4: Conclusions of the Risk Assessment for TBTO Coatings
Processes and Uses of TBTO Coatings
Risks to the Aquatic Environment
Vessels operating on inland waterways
(iii)
Vessels frequently operating in inshore
(iii)
waters/harbours (e.g. servicing/dredging vessels,
tugs, pilot vessels, ferries)
Deep sea vessels
(iii)
Dry dock operations
(iii)
Private vessels
(iii)
4.2.7 Risk Management
Owing to the fact that TBT is classified as a Persistent, Bioaccumulative and Toxic
substance according to the standard PBT criteria, the Dossier for TBTO (although less
clear whether this would be classified as PBT) would qualify for Authorisation and
subsequent controls on the use of the substance in the full range of applications.
Page 4-20
RPA & BRE
5.
THE REACH DOSSIER CONSIDERED
5.1
Control of Risks through REACH versus Current Measures
Unlike the other case studies, there is no single risk assessment document or
mechanism for developing the risk reduction strategy for the substance with which to
compare the results.
While it is unlikely that Authorisation under REACH would propose any controls that
are less stringent than the existing ones, it is possible that Authorisation would
propose wider controls than the existing controls.
The significant difference between the hypothetical situation of REACH being in
place versus the ‘real life’ situation is likely to be both the speed at which a suite of
controls would be proposed, initiated and in place compared with the fairly piecemeal
approach that has actually occurred. One of the possible reasons for this piecemeal
approach is the international dimension of the problem and shipping in general, which
complicates risk management and enforcement issues. In this respect, reducing the
risks of shipping originating/registered outside European waters would still require
international agreement and action from the IMO.
However, REACH could have facilitated a unified European mechanism and position
and, at the very least, reduced the problem in continental European inland waters. It
would probably also have had a significant effect on the occurrence of TBT and TBT
related damages in European inshore and offshore waters. By virtue of action in
Europe, which is clearly an important origin of shipping (as well as a destination), it
follows that there could have been reduced inputs of TBT to international waters and
other non-EU waters.
In addition to the probability that, had REACH been in place, risks and risk reduction
measures would have been put in place faster, there is the issue of substitutes. The
consideration of what comprises a suitable and safe alternative to TBT has also
caused both a slowdown in regulatory controls on TBT itself and a shift, in some
cases, to poor substitutes from a toxicological perspective. Because the objective (or
the result of the operation) of REACH is the consideration of all existing chemicals,
this means that full dossiers for the chemical-based substitutes (as opposed to nonstick methods and possibly polymeric materials) and associated toxicological and fate
data would be available. This may have increased the speed at which a final decision
and substitution was made.
Thus, in assessing the ability of a system such as REACH to deliver environmental
and health benefits, the following advantages over the existing system are highlighted:
•
•
•
the faster rate at which a suite of controls probably would have been initiated and
put in place when TBT effects became known;
the facilitation of a unified European mechanism and position which, at the very
least, could have reduced the problem in continental European inland waters; and
the availability of full dossiers for the chemical-based substitutes, resulting in
faster and more appropriate decisions on substitution.
Page 4-21
Case Study 4: Tributyltin
5.2
Historical Damage Costs Avoided
The case study highlights historical damage costs which have resulted from the use of
TBT including:
•
•
•
•
the reduction in shellfish stocks on a widespread geographic basis around the
world;
the documented discovery of imposex in as many as 150 species of marine snails,
with the exact number of organisms affected unknown;
shell deformity effects and larval mortality in aquatic organisms; and
corresponding financial losses suffered by the aquaculture industry and costs
imposed on harbour authorities.
While the exact numbers of organisms affected worldwide could never be known, it is
possible to make a number of assumptions to allow an estimate of the damages at one
site, Arcachon Bay, and consequently get an idea of the possible scale and extent of
damages in the EU and globally.
Calculated Damages at Arcachon Bay based on TBT Use
In calculating damage costs, it is assumed that the reduction in oyster landings seen at
Arcachon Bay where production fell from 10,000 to 15,000 tonnes per year in the
1970s, to just 3,000 tonnes per year in the 1980s, was solely as a result of TBT (Ruiz
et al, 1996 and Evans, 2000 as quoted in Santillo et al, 2001). Given the causal links,
this seems a reasonable assumption, and is very well accepted in scientific circles. It
is, however, referred to as an assumption because of a similar reduction of oyster
stocks sometime in the 1920s, due to a combination of disease and environmental
factors. Many of these fisheries recovered in the 1940s but some, such as the Solent
fishery in the UK, did not become economically exploitable until the early 1970s
(Guillotreau & Cunningham, 1994).
The price of oysters harvested in the 1970s/1980s is not available, therefore, the
estimates presented here are based on the average production and value of
‘crustaceans, molluscs, etc. prepared or preserved’3 which is €4,650 per tonne.
Taking an average reduction in oysters over a ten year period (i.e. from the mid-1970s
to the mid-1980s when the oyster beds were most affected) of 25% - 45%, at a value
of €4,650 per tonne, the ‘lost’ income to oyster fishermen is in the range of €14
million - €26 million per year4, equating to a minimum of €140 million over the 10
year period of total losses to oyster fishermen.
The above calculations only includes losses to oyster (and other crustacean) fisheries,
whereas there are an estimated 150 species affected by imposex worldwide, not to
mention new research showing bioaccumulation of TBT in marine mammals and
numerous contaminated water courses equally affected. The total environmental costs
are, therefore, under-estimated by the above losses, in addition to the fact that these
3
4
Based on data from Eurostat for 1993 to 1998 and product code 15201600.
12,500 oysters x 0.25/0.45 x €4,650 per tonne
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RPA & BRE
figures are just for one estuary, whereas population-level effects have been widely
documented throughout the EU and elsewhere (e.g. the US and Japan).
Given the above, it is probable that, had REACH been in place sooner, risks to the
environment and potentially to man via the environment from TBTs could have been
reduced sooner, and the damages could be much lower, with a quicker recovery time
for affected environments.
In analysing the damage costs avoided, TBT could be regarded as a peculiar case
study, in that:
•
•
•
concerns arose early in its use and led to actual restrictions at the regional/national
level;
a speedy and conclusive linkage of the chemical to its impacts based on evidence
of imposex along shipping lanes and in proportion to the density of shipping
traffic was possible; and
its impacts on molluscs reflect a highly sensitive, chemical specific phenomenon
which shows the potential implications that continued widespread use of PBT or
vPvB substances could have on the environment (and man via the environment).
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Case Study 4: Tributyltin
Page 4-24
RPA & BRE
6.
REFERENCES
Arnold DEJ (2002): The Challenge of Converting from Tin to Tin-Free Antifoulings, paper
presented at the WWF Conference TBT-Freie Antifoulinganstriche für die
Seeschifffahrt, Hamburg, 3 June 2002.
Champ MA (2001): Potential Economic and Environmental Consequences of the
Convention to Ban Harmful Marine Antifouling Systems, Proceedings of the Pollution
Prevention from Ships and Shipyards Symposium, Oceanology International 2001.
Champ MA (2000): A Review of Organotin Regulatory Strategies, Pending Actions, Related
Costs and Benefits, The Science of the Total Environment, Vol 258, pp21-71.
Chapman R (2002): Selecting a Suitable Tin-Free Antifouling – How Ship Operators can
Learn from the Japanese Experience, paper presented at the WWF Conference TBTFreie Antifoulinganstriche für die Seeschifffahrt, Hamburg, 3 June 2002.
Danish Environmental Protection Agency (not dated): Inventory of Biocides used in
Denmark, Chapter 4: Main Group 2: Preservatives (Full reference details not
available)
DoE (1992): Risk-Benefit Analysis of Hazardous Chemicals, report prepared by RPA and
ACER Environmental, December 1992.
Dresher WH (2000): Copper in Third-Generation Antifoulants for Marine Coatings,
Innovations, September 2000.
European Commission (2002): Proposal for a Regulation of the European Parliament and of
the Council on the Prohibition of Organotin Compounds on Ships, Official Journal of
the European Communities C262 E/492, 29 October 2002.
Guillotreau P & Cunningham S (1994): An Economic Appraisal of the Solent Oyster
Fishery: Historical and Institutional Aspects, Centre for the Economics and
Management of Aquatic Resources (CEMARE), Research Paper 65: University of
Portsmouth.
GuT (2000): Tributyltin (TBT), GuT Fact Sheet, 4/2000.
Head R & Klijnstra J (2002): New Fouling Control Coatings for Shipping: What is Realistic
in the Next Ten Years?, paper presented at the WWF Conference TBT-Freie
Antifoulinganstriche für die Seeschifffahrt, Hamburg, 3 June 2002.
IMO (2003): International Convention on the Control of Harmful Antifouling Systems
on Ships, Treaty Status, article downloaded from www.antifoulingpaint.com
IMO (2002): IMO TBT Antifouling Ban – Account of Discussions at IMO-MEPC47 (4-8
March 2002) and Summary of Reaction to IMO Antifouling Systems Convention
(IMO-AFS), article downloaded from the IMO Internet site.
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Case Study 4: Tributyltin
Jermer J (2000): Wood Preservation in Sweden, paper presented at the 96th Annual Meeting
of the American Wood Preservers’ Association, 7th-10th May 2000: San Francisco.
KEMI (2000): Organotin Stabilisers in PVC: Assessment of Risks and Proposals for
Risk Reduction Measures, No 6/00, Kemi; Solna, Sweden.
Madsen et al (1999): Økotoksikologisk Vurdering af Begroningshindrende Biocider og
Biocidfrie Bundmalinger, Project Number 11324, English Summary, Miljøstyrelsen:
Copenhagen.
Nygren C (2002): Wallenius Lines Experiences from Using Tin-Free Antifouling, paper
presented at the WWF Conference TBT-Freie Antifoulinganstriche für die
Seeschifffahrt, Hamburg, 3 June 2002.
ORTEPA (nd): Current Worldwide TBT Regulations by Country, article downloaded from
the ORTEPA Internet Site (http://www.ortepa.org).
ORTEPA (nd-a): Questions Remain on the Cause of Marine Mammal Deaths, article
downloaded from the ORTEPA Internet site (http://www.ortepa.org).
Pesticide Outlook (1995): Tributyl Tin in the Marine Environment: A Review of Past and
Present Risks, Pesticide Outlook, Vol 6(6), pp18-24.
Ranke J & Jastorff B (2000): Multidimensional Risk Analysis of Antifouling Biocides,
Environmental Science & Pollution Research (ESPR), Vol 7(2), pp105-114.
Santillo D et al (2001): Tributyltin (TBT) Antifoulants: a Tale of Ships, Snails and Imposex,
in European Environment Agency (2001): Late Lessons from Early Warnings:
The Precautionary Principle 1896-2000, Environmental Issue Report, No 22, EEA,
Copenhagen.
SCL (1995): Department of the Environment Report on an Evaluation of the TBT
Monitoring and Research Programme, May 1995.
WS Atkins (1998): Assessment of the Risks to Health and to the Environment of Tin
Organic Compounds and of Arsenic on Certain Biocidal Products and of the
Effects of Further Restrictions on their Marketing and Use, Volume A. Report for
DG-Enterprise, European Commission.
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