The environmental aspects of the use of PVC in building products

Transcription

The environmental aspects of the
use of PVC in building products
Second Edition
A study
carried out for the
Plastics and Chemicals
Industries Association Inc.
CSIRO Molecular Science
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Second Edition
A study
carried out for the
Author: Dr Russell Smith
June 1998
Bag 10
Clayton South 3169
AUSTRALIA
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Published by
CSIRO PUBLISHING
PO Box 1139
(150 Oxford Street)
Collingwood VIC 3066
Australia
Tel: (03) 9662 7666
Fax: (03) 9662 7555
email: [email protected]
ISBN 0 643 06383 8
© CSIRO Australia 1998
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The environmental aspects of the use of PVC in building products – Second Edition
REFERENCE
The Environmental aspects of the use of PVC in building
products, Smith, R., CSIRO Division of Chemicals and Polymers,
September 1996.
STUDY BRIEF
Review any pertinent scientific information relating to the
environmental aspects of PVC in building products which has
become available since the preparation of the above report. As in
the above report, please adopt a cradle-to-grave approach and,
where possible, review the indicated environmental impacts in
comparison with the impacts of comparable or alternative
materials. Please take into account comments received in response
to the initial report, where appropriate.
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CONTENTS
Executive Summary
7
1.
Introduction
9
2.
Scope and methodology
9
3.
PVC manufacture in Australia
9
4.
Raw materials and intermediate products
10
5.
Polymerisation
11
6.
Environmental effects
11
7.
Compounding and manufacture of PVC products
13
8.
Use of PVC products in buildings
18
9.
Waste management of used PVC building products
21
10. Environmental oestrogens
26
11. Alternatives to PVC in building products
27
12. Conclusions
29
13. Acknowledgments
30
14. References
31
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EXECUTIVE SUMMARY
This updated study reports on the environmental effects of the use of polyvinyl chloride (PVC) in its major building product applications, and compares
these effects with those of the available commercially viable alternatives.
Having noted the many comments received on the first report, and having
reviewed in detail the literature which has become available since then, the
author has concluded that there is still no basis on which to believe that the use
of PVC in its major building product applications causes significantly more
overall harm to the environment than the commercially available alternatives.
The major uses of PVC in building applications are as waste water pipes, as
cable insulation and as floor covering.
The possible environmental hazards arising from the use of PVC in the above
applications result mainly from the presence of additives which may find their
way out into the surroundings, from the emissions which may result from
accidental fires, and from the disposal of waste PVC building products.
The quantitatively important additives to PVC are heat stabilisers, plasticisers
and flame retardants. Lead compounds are the most commonly used heat
stabilisers - in view of the well known toxicity of lead, its use in building
products is potentially of concern. However, because the stabiliser is held
within the PVC matrix, only limited and temporary losses from the surface will
occur. Extraction of lead from waste water pipes by running water is the most
likely contribution in this area. It has been shown that the level of lead
extracted from pipes by water rapidly declines to a low value after a short
period of use. It is concluded that the possibility of lead losses to the
environment is limited, and will make a relatively small contribution to the
environment compared to other sources.
While there are alternatives to PVC in pipes and fittings, and in floor
coverings, there are essentially no commercially viable and cost-effective
substitutes for PVC cable insulation in building applications. The only
comparative life cycle analyses carried out so far, on pipe materials, fail to show
any greater environmental harm due to PVC.
The effects on health of phthalates, by far the most commonly used plasticisers
for PVC, has been the subject of a great deal of controversy. However, the
evidence that Di-ethylhexyl phthalate (DEHP), the most common plasticiser
in PVC, does not present any threat to humans has been strengthened by one
recent study. At the present time there is little evidence, if any, to indicate that
phthalates have a significant effect on the environment..
The use of flame retardants in plasticised PVC is sometimes necessary because
of the flammability of the plasticiser, and the lower overall chlorine content of
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plasticised PVC. While these are inherently toxic, the same limitations to
spread in the environment apply as with lead stabilisers.
Available information concerning the performance of PVC in accidental
building fires is reviewed in this study. Evidence for the formation of toxic
emissions in excess of those produced by other plastics is so contradictory that
no firm conclusion can be reached. Likewise the contribution of the
production, use, and disposal of PVC products to global dioxin production is
the subject of continuing debate. It can, however, be concluded that other
industrial processes such as municipal solid waste incineration (little used in
Australia) are far more significant sources of dioxins.
Waste PVC building products can be disposed of by recycling, landfill or
incineration. Recycling is by far the preferable alternative, provided that PVC
products are identified, as the plastic is easily recycled. In the event that a
proportion of waste PVC is taken to landfill or incinerated, the environmental
effects are unlikely to be significant. In the case of landfill, degradation is
extremely slow and the dangerous compounds are immobilised, while
incineration under strictly controlled conditions in municipal solid waste will
not lead to significant increases in dangerous pollutants.
From the evidence cited in this study, it can be concluded that the adverse
environmental effects of using PVC in building products do not appear to be
greater than those for other materials. However there are several aspects which
require further study because the available evidence is either inconclusive or
contradictory. These environmental questions include the use of phthalates as
plasticisers in flexible PVC, and their health effects in the environment; the
ultimate fate of heavy metals used as stabilisers; and the toxicity of emissions
from accidental fires.
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1. INTRODUCTION
The original report (referenced above) became the subject of some criticism:
Greenpeace, (1996); Rae (1997); Andrew (1997); Selinger (1997); Ruchel
(1997). Additionally, more technical and scientific information has become
available over the last year. This edition attempts to address those criticisms,
and incorporate the more recent relevant information.
Useful general information about PVC has been published in a lay-person’s
guide by Emsley (1996), who concluded that the human health and
environmental risks from the manufacture, use and disposal of PVC products
“are so miniscule that they can be ignored”.
Polyvinyl chloride (PVC), with a world-wide market second only to low-density
polyethylene (LDPE), is the plastic most widely used in building products
(about half of all the plastics used for this purpose (Dunlap and Desch, 1983)).
The 1993 figure (ECVM, cited by Møller et al., 1995), was 56%.
The major building product uses in Australia are for (a) pipes, fittings, and
conduit (rigid or unplasticised PVC ), (b) low voltage electrical insulation
(plasticised PVC), and (c) flooring and interior trim.
In Europe and the USA, PVC is used for exterior cladding as well as door and
window frames and sashes. These applications are minor in Australia, but
increasing (Faulkner, 1996). Of the world-wide PVC production in 1993,
10% was used in rigid profiles (ECVM, cited by Møller et al., 1995).
2. SCOPE AND METHODOLOGY
This report surveys the accessible information on the environmental
consequences of using PVC in building products in Australia. Extensive
recourse has been made to detailed overseas studies on the environmental
problems associated with all uses of PVC (SFT, 1993; Tukker, A. et al, 1995;
Møller, S. et al., 1995). Information has also been derived from other reputable
sources such as peer-reviewed papers in the scientific literature, and reports by
government agencies and universities. These sources cover PVC in its many
and various applications.
3 . P V C M A N U F A C T U R E I N AU S T R A L I A
In the last year, the two Australian PVC manufacturers merged to form the
Australian Vinyls Corporation Ltd (Australian Financial Review, 6/6/97), and
consolidated their operations to two plants, one in Laverton, Victoria, and the
other in nearby Altona. Both plants are based on imported vinyl chloride
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monomer (VCM), and the total production capacity is reported to be 240,00
t.p.a. The plant at Botany, N.S.W. has ceased operation.
In addition to Australian production, much smaller amounts (15,000 t.p.a.) of
PVC have been imported when demand exceeded production capacity at the
peak of the economic cycle.
Information about Australian PVC manufacture was provided by industry
sources. (Faulkner, R., 1996; Winstanley, M., 1996).
The world-wide production of PVC was of the order of 20 million tonnes in
1995, and is expanding, with further plants planned in South East Asia and
Qatar (CMR, 1996).
4 . R AW M AT E R I A L S A N D I N T E R M E D I AT E
PRODUCTS
Because PVC has a chlorine content of 57 wt.%, it is more conservative of nonrenewable resources than other commodity plastics, and as such, can be seen as
an environmental benefit (Barton et al., 1997).
The starting raw materials for PVC production are sodium chloride and
hydrocarbons from crude oil. The hydrocarbons are converted to ethylene
which is reacted with chlorine, obtained from the electrolysis of sodium
chloride, to produce 1,2-dichloroethane (ethylene dichloride, EDC). The
production of chlorine is claimed (Greenpeace, 1996) to be environmentally
unacceptable because of mercury and asbestos releases from the older
production processes (mercury cells and diaphragm cells). In Australia, and
other developed countries, these older cells are steadily being replaced with the
more energy-efficient membrane cells, which do not result in mercury or
asbestos releases. There is, however, some concern that the older plants are
being relocated to developing countries (Brotherton, 1998).
To put things into perspective, the Eurochlor 1993 Mercury Phase-out Task
Force estimated mercury emissions as follows:
all industry world-wide
natural sources
European chlorine industry
14,700 t.p.a.
6,000 t.p.a.
19 t.p.a.
The EDC is purified before undergoing pyrolysis to VCM, which is purified
by distillation to achieve the quality necessary for polymerisation.
Detailed technical descriptions of the VCM production process are widely
available (see, for example, Davidson and Gardner, 1983; Newman, 1989).
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5 . P O LY M E R I S AT I O N
PVC is manufactured by the free radical initiated polymerization of VCM.
Suspension polymerisation is the most common of the three processes used,
providing 80% of the world production, and all of the Australian production.
VCM is suspended in water, and emulsifier (concentration up to 0.5%) and
buffer are added. An initiator, typically a peroxide, at concentrations of 0.03 to
0.1% relative to the monomer, triggers the polymerisation of VCM to PVC.
The polymer particles produced are typically 120 - 150 µm in diameter, the
size being controlled by the type of emulsifier used, the stirring rate and the
polymerisation temperature. The suspension is then stripped to recover
unreacted VCM for recycle, and the PVC particles separated and dried.
Detailed technical descriptions of VCM polymerisation are available in the
technical literature (for example, Bunten, 1989).
6 . E N V I R O N M E N TA L E F F E C T S
In the PVC production process, substances with detrimental effects to the
environment and health are used or formed as byproducts. There is, therefore,
the possibility of emissions to the environment and exposure of the workforce.
Of most concern are EDC, VCM, and dioxin-like substances formed during
their manufacture. For example, Greenpeace (1996) state that the groundwater
in Altona is seriously contaminated with EDC from earlier production on that
site, and disposal of EDC-containing wastewaters by deep-well injection.
Papp (1996) gives an extensive review of waste management in the EDC, VCM
and PVC industries.
6.1 Ethylene dichloride
Ethylene dichloride is toxic by ingestion and inhalation, affecting the liver,
kidneys, the central nervous system, and is a suspected carcinogen (Sax and
Lewis, 1989a). It is slightly soluble in water (< 1%), and toxicity data for fish
(Christiansen et al., 1990) and crustaceans (Nikunen, 1990) have been
reported. The major producers of EDC/VCM in Europe have agreed through
the European Council of Vinyl Manufacturers (ECVM) to voluntarily limit, by
1998, EDC emissions to air to less than 5 mg/Nm3, and emissions to water
to less than 5 g/tonne EDC purification capacity (ECVM, 1995).
6.2 Vinyl chloride monomer
VCM is a gas with low water solubility, and any emissions would mainly be to
the atmosphere, where it undergoes photochemical oxidation to hydrogen
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chloride and carbon monoxide with a half-life of about 20 hr. (Møller et al.,
1995). In 1971, a rare cancer of the liver, angiosarcoma, was traced to VCM
exposure amongst workers in the PVC industry, and strict workplace exposure
limits were established by the United States Occupational Health and Safety
Administration. It is a human carcinogen which can also cause blood tumours
(Sax and Lewis, 1989b), and accordingly has to be carefully handled. The level
of unreacted monomer in PVC product is tightly controlled. The ECVM
charter, for example, limits this to < 1 g/tonne (1 ppm).
The OECD (1982) ranks VCM as having low aquatic toxicity, high human
risk, and no persistence. Accordingly, there are strict standards covering
emissions, and as emission control technology improves, the levels are
becoming smaller. Data from 1984 show an annual average concentration of
about 8 µg/m3 in the atmosphere within a radius of 1 to 3 kilometres from
three industrial point sources in the Netherlands. The USOHSA permissible
exposure level as a time-weighted average is 2700 µg/m3 (Sax and Lewis,
1989b). The background concentration in the Netherlands in that period was
0.02 µg/m3. Over the last ten years, VCM emissions from VCM and PVC
production in the Netherlands have been reduced ten-fold (Janus et al., 1994).
The ECVM charter limits VCM emissions to air to less than 5 mg/Nm3.
The ECVM charter also specifies standards for emissions from PVC plants:
VCM to air less than 100 g/tonne of PVC; VCM in aqueous effluents to less than
1 g/m3; and VCM in final product to 5 g/tonne (5 ppm) of PVC for general use
and to 1 g/tonne (1 ppm) for food and medical applications (ECVM, 1995).
The Australian PVC manufacturer operates under discharge licence regulations
for VCM; these correspond to maximum allowable emission limits for VCM
to air of approx. 70 g/tonne of PVC, and VCM in aqueous effluents of 0.5
g/m3. These limits are well within the limits set out in the ECVM charter.
Malin and Wilson (1995) cite the USEPA’s toxic release inventory, which
stated that in 1991, fugitive and stack emissions of VCM in the USA totalled
20g/tonne of PVC resin produced.
Selinger (1997) mentions the alleged repeated release of VCM from
production facilities. As stated by Greenpeace (1996), “Because Australia still
has no publicly accessible system for reporting annual emissions or waste from
particular facilities, it is difficult to quantify the pollution and waste
contribution of the chlorine/EDC/VCM/VCM industry in Australia”. More
recently, the National Environment Protection Council has set up the National
Pollutant Inventory, to be implemented in July, 1998, and VCM is one of the
pollutants listed for future inclusion in this inventory.
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6.3 Dioxin-like compounds
“Dioxin” is frequently used as the generic name for the group of some 210
compounds more strictly termed the polychlorinated dibenzo-p-dioxins
(PCDD) and the polychlorinated dibenzofurans (PCDF). They were first
identified as trace organics in incinerator emissions in 1977 (see para. 9.3).
Some dioxins are very toxic and some are not. Also, some animals are more
sensitive to dioxins than others, and humans appear to be among the least
sensitive (Baldwin, 1997). Dioxins can be formed when chlorinated organic
compounds or mixtures of organic material and inorganic chlorides are
exposed to high temperature conditions, as in fires, incinerators, and
crematoria. For example, HMIP (1995) estimated that the annual emissions of
dioxins to air in the UK from halogenated chemicals manufacture, including
VCM, to be 0.02 g TEq, and the emissions from crematoria to exceed 1 g TEq.
(The toxicity of a mixture of dioxins/furans is expressed in terms of the toxicity
of the most toxic compound in the family, which is 2,3,7,8-tetrachlorodibenzo-p-dioxin.) It is generally recognised that dioxins can be formed during
the production of EDC (Evers, 1989). Baldwin (1997) states, for example, that
the Norsk Hydro VCM plant at Rafnes (production 450,000 t.p.a.) produces
about 6g TEq/year, of which 0.5 g TEq/year is emitted to air and water. The
ECVM charter limits dioxin emissions to less than 0.1 ng TEq/Nm3 for vent
gas, and for emissions to water to less than 1 µg TEq/tonne of oxychlorination
capacity.
Environmental controls covering EDC, VCM, and dioxins are becoming
steadily more stringent (Møller et al., 1995).
The Australian situation with regard to dioxins in the environment is less clear.
Greenpeace (1996) stated that “Australia still has no legally enforceable
national standards for emission of dioxin to air”, and also describe analyses of
high dioxin levels in the sludge produced by overseas VCM manufacturers.
Abad et al., (1997) state that, at present, emissions from municipal, hazardous
and hospital waste incinerators can be considered as the major contributor of
dioxin emissions to air.
7. COMPOUNDING AND MANUFACTURE OF
PVC PRODUCTS
Virgin PVC is thermally and photochemically unstable, and various additives
are used to reduce these problems before the fabrication of PVC into useful
products.
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The thermal instability of the material can result in dehydrochlorination and
the formation of polyene sequences, at the temperatures required for melt
processing and fabrication. The major substance evolved is hydrogen chloride,
while carbon monoxide, benzene and other aromatic hydrocarbons may also be
formed.
There is the potential for photochemical degradation in outdoor applications
(see, for example, Decker, 1984). In the absence of UV stabilisers,
dehydrochlorination can occur, leading to discolouration, chain scission and
cross-linking resulting in embrittlement and mechanical failure.
A wide range of additives are used to counter these and other problems,
including the addition of plasticisers to give flexible formulations for specific
uses. The quantitatively important classes of additives are inert fillers, heat
stabilisers, plasticisers and flame retardants. Other classes of additives for
specific applications include pigments, impact modifiers, lubricants, fillers, UV
stabilisers, biocides (to prevent fungal growth on flexible PVC) and antistatic
agents.
In PVC building products, only a few additives are necessary. Naturvårdsverket
(1996) state that a typical rigid PVC contains a few per cent of lead sulphate,
a maximum of 10% chalk filler, and less than 1% of a processing aid such as
stearic acid, paraffin wax or calcium stearate. For flexible PVC cable insulation,
a typical additive composition is di-2-ethylhexyl phthalate (DEHP) 23%, lead
sulphate 2%, lead stearate 0.5%, processing aid (as above) 0.1%, and chalk
filler 17%.
7.1 Heat stabilisers
The dominant stabilisers are lead compounds (e.g. basic lead sulphate and lead
stearate), and to a much lesser extent organotin compounds and barium/zinc
and calcium/zinc salt systems (see para. 8.1). These stabilisers are held within
the PVC matrix (Naturvårdsverket, 1996). More significant sources of lead in
the aquatic environment derive from leaded petrol, and dumped lead/acid
batteries.
In Sweden, shotgun pellets and bullets and fishing line sinkers consume nearly
three times the amount of lead that is used in PVC (KEMI, 1994).
There is also concern over the use of cadmium compounds in PVC
formulations, now prohibited by the EU, and the ultimate fate of the heavy
metal compounds used as heat stabilisers and pigments, particularly if PVC
waste is incinerated.
Detailed calculations on the environmental impact of lead leaching from PVC
sewage waste and vent pipes have been reported by Burn and Schafer (1997),
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who concluded that the use of lead stabilisers in PVC pipe was not a significant
source of lead in sewage effluent nor in the resultant biosolids which are used
for fertiliser applications
Burn and Schafer (1997) comment that Ca/Zn stabiliser systems are 80% more
expensive than lead stabilisers, which would increase the production cost of
PVC pipes by approximately 3.5 %. This is partly due to co-stabilisers such as
diketones, epoxies, organophosphites, hindered phenols, and polyols making
the formulations more complex. At the present time, the amount of Ca/Zn
stabilisers used is quite small, however Lilja (1996) states that this is increasing.
PACIA (1997) states that 30% of the PVC heat stabilisers used in Australia are
Ca/Zn based. Organotin stabilisers are even more expensive, and some
organotins have proved to be ecotoxic (The Nordic Council of Ministers,
1994). Substitutes with a more environmental profile than organotin and lead
additives are becoming increasingly available (VROM, 1998).
The contribution of lead stabilisers leached from PVC products to the level of lead
in the environment is relatively small compared to contributions from other sources,
and could be eliminated by using less-toxic Ca/Zn stabiliser systems.
7.2 Plasticisers
In the manufacture of flexible PVC, plasticisers are used to give the desired
mechanical properties. The three main types are dicarboxylic esters, phosphate
triesters and trimellitates, with the dominant group (on cost grounds) being
the dicarboxylic acid esters of long chain alcohols, particularly phthalates such
as DEHP, diisononyl phthalate (DINP) and diisodecyl phthalate (DIDP)
(Cadogan, 1995). Phthalates have good compatibility with other
compounding ingredients, good low temperature properties and processing
characteristics, and low volatility.
The plasticisers may be used alone or in combinations to give the required
properties. As stated by Davidson and Gardner (1983), plasticisers are added at
15-20 parts per hundred of resin (phr) for semi-rigid compounds, and at > 100
phr for soft flexible compounds such as gasketing material.
There is the possibility that there may be plasticiser emissions to the external
environment during the processing of flexible PVC into finished products.
Cadogan (1994) estimated that these constitute 0.02 to 0.07 % of the
plasticiser used, and most of these emissions are to air (Christiansen et al.,
1990).
Cadogan (1996) reviewed the occurrence of phthalates in the environment, as
well as their degradation, bioaccumulation and aquatic toxicity. He concluded
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that the levels of phthalates in the environment are decreasing, and that
phthalates pose no significant health risk to humans.
Møller et al., (1995) discuss the environmental effects of phthalate plasticisers
in detail, and point out that since DEHP has a global distribution pattern,
evaporation is most likely to be the major route to the environment. The
phthalate esters are strongly adsorbed by particulate matter, and only very
slightly soluble in water. Accordingly, more will be found in sediments and soil
than in water. The atmospheric photodegradation of DEHP appears to be
rapid, but hydrolysis at neutral pH is very slow. DEHP biodegrades under
aerobic conditions, but only very slowly in an anaerobic environment. A very
detailed review of the literature on the environmental fate of phthalate esters
was published in 1997 by Staples et al. Photodegradation via free radical attack
is expected to be the dominant degradation pathway in the atmosphere with
predicted half-lives of ca. 1 day for most of the phthalates investigated. For
phthalates released to surface waters, soils and sediments, biodegradation is
expected to be the dominant degradation mechanism, with half-lives ranging
from 1 to 14 days in waters, and from 1 week to several months in soils. Longer
half-lives are expected in anaerobic, nutrient-poor, or cold environments.
Certainly the level of phthalates in sediments is of concern to some
environmental authorities (VROM, 1998).
DEHP appears to bioaccumulate in aquatic organisms, but has little effect on
terrestrial organisms. As stated by Møller et al., (1995), the results from
ecotoxicology tests vary over several orders of magnitude, probably because of
the experimental difficulties associated with the very low water solubilities of
phthalate esters. Further studies by the US EPA were due to be completed late
in 1996.
Nielsen et al. (1985) found no dose related health effects following phthalate
exposure during PVC processing. Toxicological studies on rats and mice
showed some potential effects on the liver; testicular toxicity was observed in
rats (at dietary levels of 69 mg/kg/day for 60 days), and carcinogenic effects
were observed after long term feeding trials (674 mg/kg/day for two years) ( see
Nielsen, 1994). Siddiqui and Srivastava (1992) found DEHP (500 mg/kg/day
for 15 days) reduced sperm counts in rats by 17%, and that the effect was doserelated. The effect could be correlated with changes in the activity of several
enzymes involved in spermatogenesis. Fredricsson et al. (1993) found that
human sperm motility was modestly affected by phthalates in a dose-related
fashion.
DEHP was once suspected as being carcinogenic to humans, but was
subsequently evaluated for carcinogenic effects and not classified ( EU 1990).
The potential for phthalates to have other health effects has been studied for
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some time, but the long-term effects on living organisms are not well
understood (VROM, 1998).
Øie et al. (1997) suggest that inhalation exposure to DEHP as aerosols
adsorbed on particulate matter is as important, or more important than vapour
phase exposure, and propose that DEHP in indoor air may be associated with
inducing inflammation of the airways, which is a characteristic of asthma.
Ganning et al. (1985) found that DEHP had a pronounced but selective effect
of cellular structure and metabolism in rat liver. More recently, Huber et al.
(1996) reviewed the information available on the hepatocarcinogenic potential
of DEHP in rodents and its implications on human risk, and concluded that
an actual threat to humans by DEHP seems rather unlikely. However, Hardell
et al. (1997) put forward the hypothesis that there may be an association
between occupational exposure to PVC (and by implication phthalates) and
testicular cancer, and suggested that further studies are warranted in view of the
weak association found.
Barry (1988) identified mono- ethylhexyl phthalate, a DEHP metabolite, as
having cardiotoxic effects. This finding was confirmed by Rock (1990), who
was concerned with DEHP in PVC blood bags affecting multiply transfused
patients. It is also known that DEHP has some beneficial effects when used in
blood bag formulations, such as reducing spontaneous storage haemolysis, and
improving red cell viability upon prolonged storage by stabilising the red cell
membrane. Rock (1990) stated that there was no satisfactory alternative to
DEHP in this PVC application. This view was later contradicted by Snyder
(1992 and 1993) who showed that a proprietary plasticiser, n-butyryl trihexyl
citrate, had equivalent performance to DEHP. It is perhaps relevant to note
that citrate esters, such as diethyl citrate, have been used as PVC plasticisers in
food contact applications (Davidson and Gardner, 1983). No toxicological
data could be located for these citrate plasticisers, which would be alternative
plasticisers for PVC at presumably higher cost. Sundmark (1995a) reported on
the environmental risks from phthalates, and concluded that phthalate
emissions from PVC processing or use were not an environmental problem in
Norway, where some 10,000 t.p.a. of phthalates are used.
At the present time, there is little evidence, if any, to indicate that phthalates have
a significant effect on the environment. There is the additional difficulty of
extrapolating the results from animal experimentation to human health risks. It
should be noted that the Swedish National Chemicals Inspectorate (KEMI, 1996)
and the Swedish EPA (Naturvårdsverket, 1996) have recommended that emissions
of plasticisers such as phthalates be reduced, but that a final decision on their use
should await the current EU risk assessments on these materials, which should be
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completed during 1998. Likewise, the Dutch government will draw up a policy
statement at that time (VROM, 1998).
7.3 Flame retardants
Compared to most other plastics, and materials such as wood, PVC itself is
inherently flame retardant because of its high chlorine content, and will not
burn without an external heat source. For some flexible (and thus containing
plasticiser which lowers the overall chlorine content) PVC products used for
building applications, it may be necessary to incorporate flame retardants, such
as antimony trioxide, or hydrated alumina/magnesium hydroxide, which are
held within the PVC matrix.
8. USE OF PVC PRODUCTS IN BUILDINGS
The above comments apply to the use of PVC in all its applications.
The dominating application for PVC, however, is its use in building and
construction materials where the product life span can be viewed as long-term.
Møller et al. (1995) cite several reports which estimated that the proportion of
PVC production used in this application was 69% in Denmark in 1992, 60%
in Germany in 1992, and 56% on a world-wide basis. At the moment, there is
more PVC being used in building applications than the amount of PVC in
building waste (CML, 1993). There is thus a net accumulation of PVC in
buildings.
The problems associated with the use of PVC in building products arise from
plasticiser evaporation to the atmosphere (cable insulation) and extraction to
waste water (from cleaning vinyl flooring), extraction of stabiliser from the
surface of rigid PVC pipes to running water, and the formation of hydrogen
chloride, and possibly dioxins and other substances in fires in buildings where
large amounts of PVC are present.
8.1 Stabiliser extraction
Lead stabilisers are primarily used in rigid PVC for pipes, gutters, etc. in
concentrations of about 0.6 weight % (Christiansen et al., 1990), and in cable
insulation at a lower concentration.
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There is the possibility that lead may enter the environment by slow leaching
from PVC piping into waste water. Møller et al. (1995) cite a study in Holland
which showed that the lead concentration in the water was 0.1 mg/l after 5
days of flushing, and dropped to 0.1 µg/l after 20 days. (The current Australian
water quality guideline for lead is 10 µg/L.) The concentrations seen in waste
water are typically 70 - 100 µg/l (Henze, 1982, cited in Møller et al., 1995),
but the contribution of lead stabilisers to this is unknown. Wilkie et al. (1996)
for example, reported that the mean lead concentration in Melbourne domestic
sewage is 13 µg/L, and that the mean lead concentration in water treatment
plant influents is 68 µg/L. It has been shown that there can be some extraction
of stabiliser by running water. As the stabiliser is trapped in the PVC matrix,
the extraction rapidly declines after the pipe is put into service. (See Burn and
Schafer, 1997, and Burn and Sullivan, 1993.) Lead leached from buried PVC
piping is almost invariably retained in the soil organic matter, and is essentially
immobilised as complexes and precipitates.
Sundmark (1995b) concluded that the use of lead-based stabilisers posed a very
small risk to humans and the environment. The most promising alternatives
include calcium/zinc systems, at 2 to 3 times the cost. It should be noted,
however, that these systems have not been technically tested to the same extent
as lead-based stabilisers (KEMI, 1996), and it seems established that this type
of system is not optimal in terms of stabilising action, and has limited
applicability where good clarity is required in the product.
The Australian PVC industry recently commenced a review with the aim of
developing an Industry Code of Practice for the use of lead stabilisers in PVC
products manufactured in Australia (PACIA, 1996).
The Swedish National Chemicals Inspectorate (KEMI, 1996) has proposed
that the use of lead compounds in PVC products should be sharply reduced (by
90% compared to 1994 levels) by 2005, in line with the Swedish goal of
completely discontinuing the use of lead and lead compounds. Similarly, the
Danish EPA (1991) hold the view that the use of lead-containing stabilisers
and pigments should be reduced to the extent that is technically and
economically feasible.
The Norwegian and Swedish governments are urging manufacturers to phase
out lead stabilisers, not as a result of safety concerns during the use of the PVC
product, but rather safety in disposal.
Because the stabiliser is held within the PVC matrix, only limited losses from the
surface of PVC building products will occur. Provided that these products are reused or recycled, the possibility of lead losses to the environment appears to be
limited. Further, there are strong indications that lead stabilisers will be reduced or
phased out, at least in some countries.
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8.2 Plasticiser losses
Plasticiser evaporation has been discussed earlier in this report. World-wide,
the amount of flexible PVC products used in the building sector was less than
a third of the total amount of PVC used in building products (ECVM, cited
by Møller et al., 1995).
Møller et al. (1995) cite a laboratory study on the concentration of phthalates in
the waste water from washing vinyl flooring which showed that concentrations
ranged from 1.8 - 2.5 µg/l. It would be reasonable to expect these concentrations
would decrease with prolonged washing, through surface depletion.
8.3 PVC building products in relation to fires
Rigid PVC is difficult to ignite and burns only with the continuous addition
of heat from another source. Flexible PVC can be easier to ignite because of the
plasticiser content, which lowers the overall chlorine content of the material,
and may continue burning after the external heat source has been removed.
When burning, PVC yields a number of combustion products of which carbon
dioxide, carbon monoxide, water and hydrogen chloride make up the largest
part. The fire performance of PVC has been the subject of considerable
investigation both in Australia (see Ramsay, 1995) and overseas.
Ramsay (1995) differentiated between the ‘narcotic’ gases, carbon dioxide,
carbon monoxide and hydrogen cyanide, and the irritant gas, hydrogen
chloride. Hydrogen chloride will form hydrochloric acid fumes, which are very
corrosive to, for example, computing equipment, and destructive to mucous
membranes. Alajbeg (1987) also found aromatic compounds, largely benzene,
in the products of the non-flaming combustion of PVC. The severe practical
difficulties associated with measuring, for example, hydrochloric acid levels, in
actual fires has led to the development of several test methods which may or
may not be relevant to the real situation. Some of these test methods have been
reviewed by Kaplan et al. (1988).
Burning PVC also yields large amounts of soot-containing smoke, which may
contain polychlorinated dioxins and furans formed during the fire, the amount
depending on the fire conditions such as oxygen availability, temperature, and
the amount of PVC involved in the fire. Sack (1988a) confirmed this finding,
and found that in the presence of air a low concentration of chlorinated
aromatic compounds was formed which may be implicated in dioxin
formation. Odhuis et al. (1990) found that the dioxin production from
pyrolysis in an inert atmosphere to be very low (parts per trillion level) but
several orders of magnitude higher in air.
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Earlier studies by the U.S. National Bureau of Standards (Levin, 1986) used a
variety of bioassay toxicity protocols to assess the toxicity of the gaseous
atmospheres generated by the thermal decomposition of seven plastics,
including PVC. Their conclusion was that these plastics did not evolve
unusually or extremely toxic pyrolysis or combustion products when
compared to other synthetic or natural materials. Hirschler (1994) supported
this view, and argued that the toxic potency of carbon monoxide from the
combustion of other materials was more significant in a fire situation.
On the other hand, Detwiler-Okabayashi et al. (1995) found that the thermal
decomposition products from PVC were more toxic to guinea pigs than those
from three other common plastic materials (ABS, PP-PE copolymer and PP
homopolymer), and Lee (1995) studied the mutagenicity of acetone extracts of
smog particulates using the Ames test. The extracts from PVC were more
mutagenic than from three other commonly used plastics (PS, PET and PE),
and this could be correlated with the nitropyrene content of the particulates.
Møller et al. (1995) reviewed several European studies on the dioxin and furan
content of soot from actual fires, and came to the conclusion that the dioxin
formation in fires played a minor role in the total dioxin formation and
emission to the environment.
Carroll (1996a) estimated the annual generation of dioxins and furans in the
United States as a result of PVC burning in house fires to be between 0.47 and
23 g TEq, and pointed out that this source constitutes a negligible contribution
to the dioxin burden, given that the USEPA estimates annual air emissions to
be 9300 g TEq.
While there are obvious difficulties in reproducing actual fire conditions in the
laboratory, and the results consequently inconclusive, the indications are that PVC
is no worse in most respects than other materials when burnt, apart from releasing
hydrogen chloride.
9 . WA S T E M A N A G E M E N T O F U S E D P V C
BUILDING PRODUCTS
There are three options for handling waste PVC building products, namely
recycling, landfill, and incineration.
9.1 Recycling
PVC is a thermoplastic material and is as such recyclable. As stated by the
Danish EPA (1991), there is a strong case for recycling compared to disposal
in waste dumps. In 1991, it was the Danish EPA’s aim to recycle 41% of PVC
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building products by 1995, increasing to 77% by 2000. To what extent this
was/will be achieved is not known.
The Australian PVC recycling situation was discussed by Greenpeace (1996),
who reported that PVC in all its uses was the least recycled plastic in this
country in 1992, at 0.23 % of consumption (BIE, 1994).
Rigid PVC used in building products can be easily recycled, after pulverisation
and mixing with virgin PVC, to manufacture the same product. Production
scrap can be similarly utilised. As pointed out by O’Connell et al. (1996), PVC
is mainly used in Australia in long-life products such as plumbing pipes and
electrical conduit, and because these products have not reached the end of their
useful life time, little recycling of these is occurring at this stage other than inhouse production waste. Flexible PVC used for electrical cable insulation can
be recycled to recover copper metal, and the PVC recovered. However, it is
important to note that the PVC and copper must be separated before recycling.
For example, van Wijnen et al. (1992) were able to demonstrate significant
dioxin contamination of soil samples taken in the vicinity of small open-air
(illegal) scrap wire incineration operations in Amsterdam.
Menges (1996) and Baldwin (1997) have given extensive reviews of PVC
recycling management in Europe. The present level of PVC recycling is small,
but is estimated to treble by 2000 compared to 1994 levels. For example, 1995
data for PVC waste disposal in the Netherlands show that only 3,000
tonnes/year was recycled from the 46,000 tonnes/year of waste PVC produced
(VROM, 1998). Australian data is available for recycling rates for the major
plastics, including PVC, but there is no dissection according to prior use.
PVC building products offer an ideal opportunity for recycling, given that they are
easily identified and that increasing quantities will become available. It is noted
that the Swedish National Environment Protection Board (Naturvårdsverket,
1996) advocate that the greatest potential for the reuse or recycling of PVC is to be
found in the building and construction industries. This should be facilitated worldwide.
9.2 Landfill
Møller et al. (1995) reviewed the limited number of studies that have been
carried out on this aspect of PVC waste management, and pointed out that the
leaching of heavy metals from PVC is small compared to the amount of heavy
metals present as a result of the corrosion of the metal contained in ordinary
municipal waste. PVC is a very resistant material both to water and to most
chemicals, and its degradation in landfills will be very slow. One study found
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that no degradation had occurred in PVC that had been buried for 25 years.
(Hjertberg and Gevert, 1995.)
Hjertberg (1995), in a theoretical study on the degradation of PVC, concluded
that rigid PVC will not degrade at a practically important rate, and plasticised
PVC will degrade slowly with the PVC chain remaining intact. Ejlertson and
Svensson (1995) commented on the paucity of information on the subject of
PVC degradation in landfills, and could not find any report of organochlorine
compound formation from that source. There is the possibility that landfill
leachate may contain traces of plasticisers. However, Lundberg et al. (1992)
point out that “because of their high affinity for organic soil particles and their
low water solubility this is not likely to be a major route into the environment.”
The Swedish EPA (Naturvårdsverket, 1996) states that current knowledge
indicates that emissions from land-filled PVC sources do not constitute a major
environmental problem.
There is little information on the environmental consequences of dumping PVC
building products in landfills, but given the small quantities of heavy metals and
plasticisers involved, and the degree of immobilisation, it seems unlikely that any
significant effects will eventuate.
9.3 Incineration
The contribution from PVC in the formation of dioxins in incinerators remains a
topic of scientific debate, in part due to improvements in incineration technology.
Møller et al. (1995) give an extensive review of the incineration of PVC waste in
Europe where a considerable fraction of municipal waste is handled this way, as it
is in the U.S.A. In Australia, the amount so treated is very small. The problems
associated with the incineration of PVC include particulate and acid gas emissions,
which can be restricted with filters and flue gas neutralisation systems respectively.
There is also the potential for the emissions of dioxins (to the atmosphere and by
adsorption onto fly ash) and heavy metals (in the incinerator ash, the disposal of
which should be tightly regulated). Lead additives in PVC account for some 10%
of the undesirably high lead levels in municipal solid waste (MSW) incineration
residues. The USEPA has recently introduced regulations to reduce air pollution
from medical waste incinerators so that dioxin emissions will be reduced by 95%
(Anon., 1997).
At sufficiently high temperatures, dioxins will break down to non-toxic
compounds, so the incineration conditions influence the dioxin emission.
As inorganic chloride and organic material will always be present in municipal
waste, there is the potential for dioxin formation even in the absence of PVC
in the waste stream. Recent trials on laboratory, pilot and full-scale plant have
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tended to confirm the lack of a beneficial effect on dioxin emissions when
chlorine-containing components of waste are withdrawn from the feedstock to
an incinerator, and have indicated the importance of the combustion and postcombustion processes (Eduljee and Cains, 1996).
Wikström et al. (1996) studied the combustion of a synthetic fuel in a pilot
reactor and found that there was no correlation between the quantities of
dioxins formed and the chlorine content of the fuel when the chlorine content
was below 1%; above this level, a positive correlation was found. Lenoir et al.
(1990) found that the addition of 2% sodium chloride to polyethylene
produced no effect on the dioxin formation during combustion, whereas the
addition of 3% PVC gave a moderate increase. Seys (1997) cited ten studies
that have demonstrated that the same amount of dioxin is emitted from
incinerators burning MSW, with or without PVC present.
Greenpeace (1996) cites several studies in support of there being a link between
the PVC/chlorine content of the incinerator feed and the resulting dioxin
emissions. The Swedish EPA (Naturvårdsverket, 1996) state that incineration
at MSW plants is not an environmental problem, and that the chlorine content
of the PVC waste fraction does not affect the formation of dioxins as long as
the chlorine load does not exceed 1%. As pointed out by the Dutch
government (VROM, 1998), improvements in combustion technology and
flue gas treatment have led to reductions in both the formation and emission
of dioxins, and that eliminating PVC from incinerators would have little or no
effect on their dioxin emissions.
Useful reviews of this area have been provided by Carroll (1996b) and
Royston et al. (1993).
Møller et al. (1995) note that there are conflicting reports concerning dioxin
formation and the PVC content of the waste, and ascribe this to the difficulties
involved in maintaining the combustion parameters constant while varying the
chlorine content of the waste.
Baldwin (1997) states that minimisation of dioxin emissions is not dependent
on the presence or absence of PVC in the incinerator feed, but rather on the
incinerator operating conditions, and also on the flue-gas clean-up techniques.
In a report published by the American Society of Mechanical Engineers
(ASME, 1995), it was concluded from existing data that the dioxin
concentrations in flue gas from MSW incinerators could not be correlated with
fuel chlorine content. Any effect that chlorine had on the dioxin
concentrations from commercial scale systems was masked by the effect of the
air pollution control system temperature, ash chemistry, combustion
conditions, measurement imprecision, and localised flow stratification.
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Sack (1988b) reported that PVC produces chlorobenzenes upon combustion,
and that these may be potential precursors of larger, more highly chlorinated
molecules that are of current environmental interest. Using regular and
modified samples of MSW Kanters et al. (1996) studied the chlorophenols
produced in incineration. They showed that about 45% of the chlorine content
in MSW is due to PVC with the remainder due to inorganic chloride. They
demonstrated that the chlorophenol emission levels slightly increased with an
increase (up to 8-fold) in the PVC load in the MSW.
The contribution of the production, use and disposal of PVC products to
global dioxin production is the subject of continuing debate. Although PVC
production continues to rise, Alcock and Jones (1996) point out that dioxin
concentrations in a wide range of environmental media are on the decline.
Pearson et al. (1997) studied the chronology of dioxin accumulation in
sediment cores from the Great Lakes, and showed that the accumulation rates
began increasing about 1940, peaked at 1970 ± 10 years, and then declined to
the present rates that are 30 - 70% of the maxima. A similar chronology is
given by Alcock and Jones (1997). Toshiro et al. (1990), for example, have
measured dioxin in air samples over forest fires and concluded that large forest
fires could produce significant amounts of dioxins.
Harrad and Jones (1992) have reported that only 12% of the dioxins in the
environment can be accounted for by present source estimates.
HMIP (1995) reported on the annual dioxin emissions to air in the UK from
23 industrial processes and from 5 non-industrial sources. The industrial
sources accounted for approximately 90% of the total inventory, with MSW
incineration being the dominant source. HMIP estimated that the total
dioxin emissions to air from VCM manufacture in the UK to be of the order
of 0.025 g TEq. (compared to, for example, emissions from MSW
incineration of 460-580 g TEq.).
UKEA (1997) reviewed dioxin releases to land and water in the UK, and
found that more dioxins are released to land than to any other media, and the
majority of releases to land involve disposal at landfill sites. Because dioxins are
not soluble in water, they are very unlikely to leach from landfills, making this
the safest means of disposal, but this needs to be verified by field testing. Total
releases to land are estimated to be between 1,500 and 12,000 g TEq/year,
which is significantly higher than estimated releases to air (upper bound estimate
1,000 g TEq/year). Dioxin releases to land from PVC/EDC production were
estimated to be 27 - 82 g TEq/year.
Barton et al. (1997) stated that the release of dioxins from PVC sources is a
relatively small proportion of the total dioxin emissions to the environment,
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confirming the statement by Smit et al. (1994) that the amount of PVC in
waste has little or no effect on dioxin formation in MSW incineration.
The other aspect resulting from the incineration of PVC is the production of
hydrochloric acid, although PVC is not responsible for the total hydrogen
chloride content of the flue gas. Organic household waste is another source.
Where this is collected and treated separately on a large scale, the proportion
of chlorine processed by municipal incinerators accounted for by PVC has
grown significantly (VROM, 1998).
Flue gas cleaning to remove hydrogen chloride (and other acid gases such as
sulphur and nitrogen oxides) results in extra operational costs.
As with fires involving PVC, the evidence relating to incineration is inconclusive.
It seems clear that the PVC component in present-day MSW is small, and that
provided the incinerator is operated efficiently, any consequent public health and
environmental damage will be insignificant. However, the goal to keep PVC away
from ordinary waste incineration where technically and economically justifiable
(Danish EPA, 1991) seems reasonable.
1 0 . E N V I R O N M E N TA L O E S T R O G E N S
In recent years, there has been a vigorous debate about phthalates and/or
dioxins and many other substances acting as “hormone disrupters” compounds which can act as hormones, or affect the mode of action of natural
hormones. Part of the problem derives from the lack of a generally accepted,
validated method to screen chemicals for possible hormone disruption.
Ginsberg (1996) argues that there is a need to develop reliable and practical
tests for defining hormonal influences of chemicals released by industrial
processes.
A detailed discussion of oestrogen receptor activation was recently published
by Wiese and Kelce (1997).
The oestrogenic pollutants presumed responsible for reproductive changes
range from the organochlorine pesticides to polychlorinated biphenyls, dioxins
and furans, alkylphenol polyethoxylates, and phthalates. DEHP has recently
been tested for xenoestrogenic effects, and showed only weakly positive results
(Toppari et al., 1995). Lee (1996) discusses the inconclusive experimental and
epidemiological results that have been published in this area, and also
comments that while the subject is very complex, the hypothesis is plausible.
On the other hand, Barton et al. (1997) concluded that the oestrogenic
activity of phthalates is “unlikely to be significant enough to be of major
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concern”, and state that the more likely candidates are persistent pesticide
residues such as the DDT metabolite DDE, which has been shown to be a
potent androgen receptor agonist (Kelce et al., 1995).
Arnold (1996) claimed to have demonstrated synergistic effects with mixtures
of environmental oestrogens. However, several groups of researchers were not
able to reproduce these synergistic effects (Ashby, 1997; Ramamoothy, 1997)
nor could the authors of the original paper, who have formally withdrawn their
original paper (McLachlan, 1997).
A number of phthalate plasticisers (and their metabolites) were screened for
oestrogenic activity by Harris et al. (1997), using a recombintant yeast screen.
The relative potencies descended in the order BBP>DBP>DIBP>DEP>DINP,
and were a million to 50 million times less than 17ß-oestradiol. DEHP showed
no oestrogenic activity.
It seems clear that the most toxic dioxin (2,3,7,8-tetrachloro-dibenzo-p-dioxin)
harms reproductive capability in male rats, and that some phthalates reduce
testosterone levels in male rats and mice.
However, it should be recognised that many animal studies involve doses far
above environmental levels. The USEPA and the UK Medical Research
Council’s Institute for Environment and Health continue to be involved in the
development of policy and regulations addressing the area.
While the whole area of environmental oestrogens is currently the subject of major
research and debate, there does not appear yet to be the evidence required to justify
a ban on the use of the most widely used PVC plasticiser, DEHP.
1 1 . A LT E R N AT I V E S T O P V C I N B U I L D I N G
PRODUCTS
There seems to be no commercially viable alternative to the use of plasticised
PVC in electrical cable insulation. As pointed out by Hogan (1983), PVC
provides a low cost material for low voltage insulation. This results from its low
material cost and its fast extrusion speed. It also has excellent mechanical
properties and abrasion resistance at normal operating temperatures, and
excellent moisture, chemical, flame and weathering resistance.
KEMI (1996) suggest that polyethylene and polypropylene can constitute
alternatives to PVC for cable insulation, albeit at 3-5 times the cost. However,
these materials are not self-extinguishing, and increase the cable rigidity. In so
doing, the likelihood of cracking during installation and use is increased.
The alternatives to rigid PVC in pipe and fittings in the building sector are
concrete, polyethylene, polypropylene, ductile iron, vitreous clay, and fibreglass
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reinforced plastics. The choice of pipe is determined by price, design pressure,
and the effects of chemical and biological activity. In acid soils, for example,
cement and iron-based pipes can corrode. Polyolefins have lower melting
temperatures, and heavy pipe materials are worse from a work-environment
point of view. Detailed life cycle analyses (LCA) for vitreous clay and concrete
products could not be found in the literature. An extensive review of LCA
studies on PVC packaging has been reported by Smit (1994).
An LCA by van den Berg et al. (1996) compared iron, polyethylene and PVC
piping for low pressure gas distribution systems in Holland. The following
environmental themes were considered
●
●
●
●
●
●
●
Photochemical oxidant formation potential: reactions of NOx with
volatile organic substances leads, under the influence of UV light, to
photochemical oxidant creation, which causes smog.
Nutrification potential: the addition of nutrients to water or soil will
increase production of biomass and may threaten biodiversity.
Global warming potential: an increasing amount of carbon dioxide, for
example, in the earth’s atmosphere leads to an increasing absorption of
heat radiation and consequently to an increase in temperature.
Human toxicity potential: exposure of humans to toxic substances causes
health problems. Exposure can take place through air, water or soil and
especially through the food chain.
Abiotic depletion potential: abiotic depletion refers to the extraction of
raw materials such as ores and energy sources faster than they are created.
Odour threshold limit: exposure to odorous compounds is measured as
the volume of air polluted to the odour threshold.
Ozone depletion potential: depletion of the ozone layer leads to an
increase in the UV light reaching the earth’s surface. This may lead to
human diseases and may influence ecosystems.
Using a cradle-to-grave approach, their conclusions were that in terms of all of these
environmental themes, PVC was better than or equal to the other materials
considered.
The other LCA located in the course of the present study was that by
Finnveden et al. (1996), which dealt with concrete, cast iron, polyolefin and
PVC pipes. Due to a lack of reliability in the basic information and a lack of
comparable data for all pipe materials, the authors were unable to draw any
sound conclusions. In the majority of assessment methods, PVC pipe ended up
between the other alternatives (KEMI, 1996).
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Although little detailed comparative, conclusive LCA information is available, it
does not seem that PVC in its major building product applications has significantly
more effect on the environment than alternative materials.
12. CONCLUSIONS
In the production of PVC, there is the possibility of accidental emissions of
such compounds as EDC, VCM, and dioxins. The concern of regulatory
authorities, environmental groups, and industry itself has led to stringent
environmental measures in most developed countries.
There are, however, several environmental aspects of the general use of PVC
products which require further study, as the available evidence is either
inconclusive or contradictory.
These include
●
the use of phthalates in flexible PVC, and the consequent possibility of
loss to the environment
At the present time, the scientific evidence concerning the effects of phthalate
plasticisers on the environment is inconclusive. There is the additional difficulty of
extrapolating the results from animal experimentation to human health risks. It
should be noted that the Swedish National Chemicals Inspectorate (KEMI, 1996)
has recommended that emissions of plasticisers such as phthalates be reduced, but
that a final decision on their use should await the current EU risk assessments on
these materials, which should be completed during 1998.
●
the ultimate fate of heavy metals used in stabilisers
The contribution of lead stabilisers used in PVC products to the level of lead in the
environment is relatively small compared to contributions from other sources, and
seems destined to diminish. The environmental effects of heavy metals (from all
sources) is an area of continuing research.
and
●
the toxicity of emissions from accidental fires
While there are obvious difficulties in reproducing actual fire conditions in the
laboratory, and the results consequently inconclusive, the indications are that PVC
is no worse in most respects than other materials when burnt, apart from releasing
hydrogen chloride.
However, the use of PVC in buildings has minor environmental consequences,
for the following reasons
●
the rigid PVC products used in buildings can be easily recycled when the
building reaches the end of its life cycle
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PVC building products offer an ideal opportunity for recycling, given that they are
easily identified and that increasing quantities will become available. It is noted
that the Swedish National Environment Protection Board (Naturvårdsverket,
1996) advocate that the greatest potential for the reuse or recycling of PVC is to be
found in the building and construction industries.
●
the inherent non-flammability of PVC is a positive attribute in a building
fire situation
Although little detailed comparative, conclusive LCA information is available, it
does not seem that PVC in its major building product applications has significantly
more effect on the environment than alternative materials.
1 3 . A C K N OW L E D G M E N T S
The author would like to thank Ms. Jenny O’Connell (ANU) for access to a
file of papers on the subject.
During the course of a study tour by the author in Europe in December 1997,
helpful discussions were held with the following:
Dr. Ian Boustead, Boustead Consulting Ltd., London; Dr. John Svalander,
European Council of Vinyl Manufacturers, Brussels; Dr, David Cadogan,
European Chemical Industry Council, Brussels; Dr. Lea Hansen, Danish EPA,
Copenhagen; Inger Klöfver, Swedish EPA, and Eva Ljung, Swedish National
Chemicals Inspectorate, Stockholm; Rolf Bühl, European Vinyls Corporation,
Brussels; Dr. Stellan Marklund and Professor Christoffer Rappe, Institute of
Environmental Chemistry, University of Umeå, Umeå, Sweden.
The author also acknowledges the helpful comments made on a draft of this
report by six referees.
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Andrew, B., (1997): Chemistry in Australia, 64(1), 2.
Anon. (1997): Waste Treatment Technology News, September 1997.. p.5
Arnold, S.F. et al., (1996): Science, 272, 1489.
Ashby, J. et al., (1997):Nature, 385, 494.
ASME, (1995): “The relationship between chlorine in waste streams and dioxin emissions from waste
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Baldwin, J., (1997): in “Chemical Aspects of Plastics Recycling”, Royal Society of Chemistry
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Barry, Y.A. et al., (1988): Blood, 72, 1438.
Barton, P. et al., (1997): “PVCs in packaging and construction materials:- an assessment of their
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15. GLOSSARY
36
ABS
ANU
ASME
BBP
CML
CMR
DBP
DEP
DEHP
DIBP
DIDP
DINP
ECVM
EDC
EU
HMIP
KEMI
LCA
LDPE
MSW
Nm3
NOx
OECD
PACIA
PCDD
PCDF
PE
PET
PP
PS
PVC
SFT
TEq
t.p.a.
Acrylonitrile-butadiene-styrene copolymer
Australian National University
American Society of Mechanical Engineers
Butyl benzyl phthalate
Centre for Environmental Science, University of Leiden
Chemical Marketing Reporter
Di-butyl phthalate
Di-ethyl phthalate
Di-2-ethylhexyl phthalate
Di-isobutyl phthalate
Diisodecyl phthalate
Diisononyl phthalate
European Council of Vinyl Manufacturers
1,2-dichloroethane or ethylene dichloride
European Union
Her Majesty’s Inspectorate of Pollution
The Swedish National Chemicals Inspectorate
Life cycle analysis
Low-density polyethylene
Municipal solid waste
Cubic metre at normal temperature and pressure
Nitrogen oxides
Organisation for Economic Co-operation and Development
Plastics and Chemicals Industries Association
Polychlorinated dibenzo-p-dioxins
Polychlorinated dibenzofurans
Polyethylene
Polyethylene terephthalate
Polypropylene
Polystyrene
Poly(vinyl chloride)
Norwegian State Pollution Control Agency
Unit of toxicity for a dioxin mixture
tonnes per annum
US EPA
The United States Environment Protection Agency
USOSHA
The United States Occupational Safety and Health Agency
UV
Ultra-violet radiation
VCM
Vinyl chloride monomer
Second Edition
A study
carried out for the

The environmental aspects of the use of PVC in building products

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