Publications archive - Waste and recycling
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Much of the material listed on these archived web pages has been superseded, or served a particular purpose at a particular time. It may contain references to activities or policies that have no current application. Many archived documents may link to web pages that have moved or no longer exist, or may refer to other documents that are no longer available.
Prepared by Dr. John Scheirs,
ExcelPlas Polymer Technology (EPT) for
Environment Australia, June 2003
PVC resin is never used neat (i.e. as virgin resin). It is always mixed with additives to create a broad range of PVC compounds with various properties. Actual PVC resin often constitutes only about 70wt.% of PVC end-products and can sometimes be as low as 35-40wt.% (Bühl, 1996). The most common formulants added to PVC are heat stabilizers (to prevent degradation) and plasticizers (to impart flexibility and softness). These additives are dispersed in PVC by compounding. The table below shows some typical PVC formulations for common PVC products.
|Flooring||Calcium/zinc (Ca/Zn) stabilizers, butyltin compound, epoxidized soy bean oil||Di(2-ethylhexyl) phthalate (DEHP)n-butylbenzyl phthalate (BBP)|
|Cable||Basic lead (Pb) sulphateCa/Zn stabilizers||DEHPDi-isononyl adipate (DINA)|
|Pipe||Lead (Pb) salts of fatty acidsCa/Zn stabilizers||--|
|Rigid Sheet||Methyltin compound||--|
|Construction Sheet||Butyltin compound||--|
|Blister Pack||Octyltin compound||--|
|Source: adapted from Mersiowsky, (1999).|
The additives employed in PVC that are most often cited as creating environmental impacts are:
A variety of lead-based stabilizers are used in PVC as shown in the following table:
|tribasic lead sulphate||very effective heat stabilizer|
|dibasic lead stearate||heat stabiliser and a lubricant|
|dibasic lead phthalate||particularly used in cable insulation|
|dibasic lead phosphite||both a heat and light stabilizer|
|Plastics Additive Handbook (Hanser, 2000)|
Lead stabilizers scavenge and react with the hydrochloric acid (HCl) liberated by PVC during its thermal degradation. In this way lead compounds are very effective heat stabilizers for PVC. A major advantage of lead stabilizers is that the lead chloride formed during the stabilizing process has no destabilizing effect on the PVC. In contrast, cadmium and zinc carboxylate stabilizers can have a destabilizing effect on PVC. Furthermore, certain lead stabilizers such as lead sulphates do not increase the conductivity of PVC and therefore are stabilizers of choice for PVC cable insulation (Bacalogulu, 2000). The percentage of lead concentration in lead-stabilized PVC products is typically about 0.1-1.8 wt% (Argus, 2000).
Lead stabilizers such as lead stearates are classified as inorganic lead compounds since the lead is not directly attached to a carbon atom, but rather attached to an oxygen as a carboxylate (the salt of a carboxylic acid). This is an important distinction since organic lead compounds in contrast are readily assimilated by living organisms and are many times more toxic than inorganic lead salts.
Table 5.3. Australian Market Data for Lead-based Stabilizers
|Supplier||Proportion of Market|
|Source: Industry interviews (2002)|
The total usage of lead stabilizers for PVC in Australia is in excess of 2000 tpa. The main suppliers of lead stabilizers in Australia are Sun Ace (Dandenong South, VIC) and Chemson Pacific (Rydalmere, NSW). Sun Ace now supply just over half of the lead stabilizers consumed by the PVC industry (Note: Ferro one of the main suppliers of lead stabilizers until very recently was acquired by Sun Ace). Some 70% of all lead stabilizers are used in pipe with the remainder in cable insulation. An estimated 70% of lead stabilizers are made domestically and the balance are imported. The domestic manufacturing process involves a direct reaction between litharge (lead oxide) and stearic acid (Stewart, 2002).
Numerous studies have confirmed that lead stabilizers are locked into the polymer matrix and that the lead contained in PVC does not migrate from the plastic. Also, while leaching is possible under specific conditions the leachability of lead from PVC is extremely low (Argus, 2000). The inertness of lead-stabilized PVC has been validated by extraction studies that have demonstrated that the amount of lead extracted from lead-stabilized water pipes is negligible (Burn, 1997). Essentially the encapsulation effect of the PVC immobilizes the lead stabilizer and prevents it from entering the environment. The situation is less clear in the case of finely divided PVC waste products such as cable fluff or granulated profiles where the increased surface area of the PVC may facilitate extraction in the presence of acidic landfill leachates (Argus, 2000; Hjertberg, 1995; Griebenow, 1992). A study has shown a 10% release of lead stabilizer from one type of flexible PVC cable containing a mixture of plasticizers (Mersiowski, 1999). It should be noted that the percentage of lead stabilizer used in cable formulations is normally quite low. Therefore if ~10% of the total is leached then the leached amount is very small.
Lead stabilizer replacement has gradually occurred with the advent of improved Ca/Zn (calcium/zinc) stabilizers. It is of interest to note that Australia has led the way in the replacement of lead stabilizers in pressure pipe. As early as 1988 lead stabilizers were phased out in PVC pressure pipe applications. Some of the challenging technical issues with the alternative stabilizers, such as their effectiveness in suppressing 'pinking' in PVC window frames have now been overcome (Yernaux, 2002).
Across Australian industry as a whole the most important use of cadmium in terms of volume is in battery applications. Next to this, pigments for plastics and PVC stabilizers represent the main applications of cadmium. Cadmium usage is shown in the table below:
|Application||Usage Level (%)|
|Alloys and other||2%|
|Source: Cadmium Organization Data, 2001.|
Most cadmium compounds are regarded as harmful and dangerous for the environment (ecotoxic). Some compounds are also classified as carcinogens. Cadmium is persistent in the environment and certain cadmium compounds accumulate in living organisms (EU Green Paper, 2000).
The use of cadmium stabilizers in plastics was banned in Sweden, followed by bans in Denmark and Holland many years ago. Subsequently the EU issued a Directive in 1988 which effectively banned the use of cadmium in most PVC products by limiting its use to those applications where satisfactory technical alternatives did not exist (ECVM, 2002).
Cadmium stearate, cadmium adipate, cadmium oxalate and cadmium laurate represent the main cadmium-based heat stabilizers. Cadmium stabilizers are rarely used on their own and are generally available in combination with barium and/or zinc in mixtures denoted as Ba/Cd or Ba/Cd/Zn. The latter is the most common. Due to health and environmental concerns cadmium stabilizers are being phased out and substituted by Ba/Zn and Ca/Zn stabilizers (Ede, 2002).
It is important to note that cadmium stabilizers contribute only a small amount of cadmium to any PVC product. In a typical Ba/Cd stabilizer, the cadmium concentration is only around 4% of which only approximately 2.5 phr (or ~1.5wt.%) Ba/Cd is added to the PVC. This gives a final cadmium concentration in the polymer of only 0.06% (Ede, 2002).
Cadmium stabilizers such as cadmium adipate and cadmium oxalate are relatively stable within the temperature range 150-200°C. Above 200°C however they can start to decompose (Ikhuoria, 2002).
Australia's largest merchant compounder of PVC (Welvic) ceased usage of cadmium-based stabilizers in September 2000. However, the Australian PVC industry still uses cadmium stabilizers in many flexible PVC applications (mainly in the calendering area). Also PVC compounds and products that are imported can still contain cadmium-based stabilization.
The total usage of cadmium stabilizers for PVC in Australia at present is approx. 120 tpa with the majority of this manufactured locally (Ede, 2002). The main suppliers of cadmium stabilizers in Australia are Sun Ace and APS Chemicals. Market data for cadmium-based stabilizers used in Australia for PVC is summarized in the table below.
Table 5.5. Australian Market Data for Cadmium-based Stabilizers
|Supplier||Proportion of Market (in 2002)|
|Total market||120 tpa|
|* (by difference)|
Zinc stabilizers are a common replacement for cadmium stabilizers. Although zinc is sometimes referred to as a heavy metal like cadmium, the two behave totally differently in biological systems. Zinc is an essential element in almost all biological systems and plays an important role in metalloenzyme catalysis, metabolism, and the replication of genetic material. Cadmium, on the other hand is toxic, particularly in soluble and respirable forms. Cadmium can adversely affect the renal and respiratory systems, depending upon exposure time and concentration, and is not readily excreted.
Waste incinerators and landfills are sources of diffuse spreading of lead and cadmium. During the incineration of PVC essentially all lead and cadmium ends up in the bottom and fly ashes of the incinerators. Since the fly ash and bottom ash are already rich in heavy metals these must be disposed in prescribed landfills in any case. If the bottom ash is reused or incorrectly landfilled, then a dispersion of heavy metals may occur which could leach into the environment.
Cadmium pigments produce intense colourings such as yellow, orange and red. They are widely used in plastics, artists' colours, glasses, ceramics and enamels. The main pigment compositions are as follows:
Table 5.6. Colour of common cadmium-based pigments
|Pigment Colour||Cadmium compound|
|Plastics Additive Handbook (Hanser, 2000)|
The majority of cadmium pigments (~90%) are used for thermoplastic pigmentation. The proportion of cadmium pigments used in various thermoplastics is shown in the table below.
Table 5.7. Usage data for cadmium-based pigments in thermoplastics
|Source: Industry interviews (2002)|
Cadmium pigments are generally the preferred pigments in polymers that are processed at relatively high melt temperatures. At temperatures greater than 250°C the cheaper lead-chrome pigments begin to breakdown. Cadmium pigments have good high temperature stability but also carry a higher price (e.g. $30-40/kg versus $10-20/kg for lead-chrome pigments). Thus since PVC is processed at temperatures lower than 250°C the more economical Pb-Cr pigments are generally used.
While organic substitutes exist for most cadmium pigments these however do not perform adequately in most other polymers that are processed at higher temperatures due to their limited thermal stability. Organic substitutes for cadmium pigments also often lack the brightness and chemical stability of those containing cadmium.
All cadmium pigments used in Australia are imported from either the USA, UK or France. Johnson Matthey (UK) is the largest importer of cadmium pigments into Australia (Gaunt, 2002).
Cadmium pigments are very stable, inert and non-migratory compounds. Since the cadmium is in a non-soluble, calcined form, they have low toxicity and low environmental impact. The excellent dispersion, non-leachable and non-bleeding properties make cadmium pigments useful in plastic applications where uniform colouring is required.
Cadmium pigments are unlike almost all other cadmium compounds because of their extremely insoluble nature. This results from their chemically bound form and high temperature calcination. Even in the acidic conditions representative of stomach acid, they impart negligible acute toxic effects.
It has been shown in an independent risk assessment that emissions of cadmium to the environment and to man associated with the entire lifecycle of cadmium pigments are negligible (<0.5%) compared to total cadmium emissions from all known sources. It was concluded that cadmium pigments do not pose any significant risk to man or the environment (Atkins, 1997).
When considering human and environmental toxicity of cadmium compounds the question of speciation becomes relevant. The form or species of the metal is critical in determining how accessible the metal is for biotic systems. For example, cadmium in the form of cadmium pigments is complexed and calcined and not available to enter into biochemical pathways. Cadmium in cadmium stabilizers on the other hand is more labile and therefore very toxic.
The European Union Directive (91/338/EEC) recognized that cadmium pigments impart thermal and chemical stability that cannot be equalled by other pigments of a similar hue. The directive does however state that they should not be used to colour low-temperature plastics including PVC (European Union Directive (91/338/EEC, 'the Cadmium Directive').
In summary, cadmium pigments are not widely used in PVC but even when used they have low human and environmental toxicity due to their stable, non-leachable and inert (calcined) nature.
In the case of lead and cadmium stabilizers the metal atom is not attached directly to carbon. In contrast, for tin stabilizers the bond is formed between tin and carbon and thus these heat stabilizers are called organotins. There are three major types of organotin stabilizers distinguished by their respective alkyl groups - octyl, butyl and methyl:
Organotin stabilizers are used extensively in PVC packaging applications such as bottles (e.g. vegetable oil, fruit juices and wine), non-food containers (e.g. shampoo and cosmetics), as well as blister packs. Tin stabilizers are the preferred choice of heat stabilizer for the PVC packaging market because of their high clarity and transparency.
The main suppliers of organotin stabilizers in Australia are Atofina (ThermoliteTM) and a number of smaller traders - Plastral Fidene, Ciba and Barlocher.
Methyl-, dibutyl- and dioctyltin stabilizers have had widespread approvals for food-contact applications in Australia. The majority of the studies in vitro and in vivo do not indicate a mutagenic potential for mono- and di-alkyltin compounds, although some contradictory studies exist in the case of dibutyltins and di-octyltins.
Octyltin stabilizer based on a mixture of 80% dioctyltin diisooctylthioglycolate (DOTTG) and 20% of monooctyltin triisooctylthioglycolate (MOTTG) is used as a heat stabilizer for rigid PVC materials such as packaging of foodstuffs. Exposure to humans could occur via migration of DOTTG/MOTTG from PVC materials. The developmental toxicity of DOTTG/MOTTG in mice was investigated by Faqi (2001) who found that the octyltin stabilizers are embryo-fetotoxic and can induce developmental defects and fetal anomalies such as bent forelimbs, cleft palate, and exencephaly (a condition in which the brain is located outside of the skull). Such defects were reported in the group of mice treated with 100 mg/kg/day DOTTG/MOTTG. Doses of 20, 30, and 45 mg/kg/day elicited a significant increase in supernumerary lumbar ribs (Faqi, 2001). Note that it is not known if this study has been reproduced by other researchers.
The organotin stabilizer industry maintains that the available results from long-term studies in rodents do not provide evidence of a tumor risk for humans at the established levels of exposure to organotin stabilizers.
Even if leached from PVC into the environment in the case of say, landfills, organotin stabilizers are not persistent in the environment on account of microbial activity. All organotin stabilizers eventually degrade into inorganic tin. Due to low aqueous solubilities, a high affinity to soil and organic sediments as well as a rapid conversion to inorganic tin in water, the potential of mono and dialkyl tin compounds for ecotoxic effects is low. The purported environmental toxicity of mono- and dialkyl tin compounds however is sometimes confused with the known toxicity of trialkyl (including tributyl) tin compounds to aquatic life (Mersiowsky, 2001c).
Phthalates are a class of diesters that are used widely in industry. The higher molecular weight phthalates are used predominantly as plasticizers for PVC to impart flexibility and softness. The following are common phthalates:
Table 5.8 Common Phthalate Plasticizers Used in PVC
|Plasticizer Chemical Name||Abbreviation|
|benzyl butyl phthalate||BBP|
|dioctyl tere-phthalate||DOTP or DEHT|
Some 12,000 tonnes of phthalates are used annually in Australia as plasticizers in the manufacture of flexible PVC products. The majority of this is used in PVC cable applications. The main suppliers of phthalates in Australia are Exxon Chemicals (mainly DEHP and DIOP) and Orica (mainly DINP).
The most extensively used phthalate in the world is DEHP. The majority of DEHP is manufactured for use as a plasticizer for the PVC industry. The reason for this is that it offers good all-round plasticizer properties. The widespread sales of DEHP plasticizer are indicative of its all-round plasticising performance and its provision of adequate properties for a great many cost-effective, general purpose PVC products. It possesses reasonable plasticising efficiency, fusion rate and viscosity (of great importance for slush moulding applications) and is used predominantly in the manufacture of soft PVC articles ranging from flooring to cable insulation. For example, cable insulation usually contains around 30-35wt.% DEHP.
DEHP migrates at a constant rate from plastics to various environments - it has been detected in water, soil and food. It is therefore considered a widespread environmental contaminant (Latini, 2000). While biodegradation of DEHP occurs in aerobic, nutrient-rich environments over weeks, its degradation can be extremely slow (many years) under anaerobic conditions.
DEHP can enter the environment via a number of different routes:
Under landfill conditions, phthalate plasticizers can be leached from soft PVC. This has now been widely reported in a number of published studies (Argus, 2000; Mersiowsky, 2002). The recent EU Green Paper recognized this as a potential issue but concluded that the quantities and the associated risks need to be assessed further.
Phthalates are now being detected routinely in a range of environmental water samples including drinking water. This is in part due to their ubiquitous presence and in part because of the increased sensitivity of modern analytical methods. Recent advances in detection methods have further decreased the detection threshold of phthalates in water (George, 2002; Luks-Betlej, 2001; McDowell, 2001; Saito, 2002).
Phthalates are widely dispersed in the environment. They are ubiquitous environmental contaminants on account of their use as plasticizers and as constituents in many other commercial products since about the 1940's. Although the acute toxicity of phthalates is low, concerns over the last several years regarding their potential to disrupt the gonadal development in laboratory rodents has prompted interest in determining the distribution of these compounds in the environment (McDowell, 2001).
Due to the poor solubility of phthalates, sediments are the ultimate sink for phthalates released into the aquatic environment (Legler, 2002). Fromme et al. (2002) collected 116 surface-water samples; 35 sediments from rivers, lakes and channels; 39 sewage effluents, and 38 sewage sludges. These were tested for phthalates in order to obtain a better understanding of the levels of these compounds in different environments. It was found that DEHP dominated the phthalate concentrations, which ranged from 0.33 to 97.8 µg/L (for surface water), 1.74 to 182 µg/L (for sewage effluents), 27.9 to 154 mg/kg (for sewage sludge) and 0.21 to 8.44 mg/kg (for sediment). Dibutyl phthalate was found only in minor concentrations and BBP only in a few samples in low amounts. Very high concentrations of phthalates were confirmed in landfill leachate and compost water samples as well as in liquid manure samples (Fromme, 2002).
Fatoki and Noma (2001) report on the use of solid phase extraction and capillary GLC which enables the selective determination of phthalate ester plasticizers in rivers and marine water samples. The rivers and marine water samples were found to be grossly polluted by DMP, DEP, DBP and DEHP which were found to be present at levels up to 2306 µg/L (i.e. grossly polluted).
The majority of exposure of the general population to phthalates is through food (NTP-CERHR, 2000). A recent paper by Koo et al (2002) estimated steady state intake of parent phthalates from the fractional urinary concentrations of phthalate metabolites measured in adults by Blount (2000). They assumed a log normal distribution for the data, steady state intake of phthalates and used statistical methods to overcome censored data. The paper is primarily concerned with associations of overall phthalate intake with socioeconomic variables. It was found that DEHP intake was higher in males and/or in urban populations and/or in people who had family income less than US$1,500 per month. The authors suggest that there may be significant demographic variations in exposure and/or metabolism of phthalates, and also that health-risk assessments for phthalate exposure in humans should consider different potential risk groups (Koo, 2002).
An important piece of information in the Koo et al (2002) publication is that although there is some variation in exposure according to socioeconomic status in the American population, the overall intake of phthalates is very low: mean estimated intake in µg/kg/d is 10.1 for DEP, 1.66 for DBP, 0.84 for BBP, 0.41 for DEHP and 8.99 x 10-7 for DINP. This intake of DEHP from all general sources is approximately 2,500 - 10,000 times less than the various no observed effect levels for sensitive effects observed in experimental animals and is somewhat less than the estimated intake for DEHP of 3 - 30 µg/kg/d calculated by CERHR (2000) using a variety of exposure and consumption assumptions. Because the estimations by Koo et. al. (20002) are based on measurements in humans they are more likely to be reflective of the actual intakes that may be occurring in the American population than estimates not founded on empirical data.
Most phthalates are readily biodegradable by a wide range of microorganisms in aerobic environments. Biodegradative attack occurs at the ester linkage leading to hydrolytic scission of one of the ester bonds (common phthalates have two ester bonds and are chemically known as diesters). The intermediate biodegradation product is the monoester phthalate (mono (2-ethylhexyl) phthalate in the case of DEHP) which then undergoes biodegradative attack on the remaining ester bond to yield phthalic acid and the corresponding alkyl alcohol (2-ethyl hexanol in the case of DEHP). These compounds breakdown further to low toxicity species such as pyruvate and oxaloacetate (Staples 1997a).
Angelidaki et al. (2000) used screening tests to assess the capability of various microbes to degrade DBP and DEHP under aerobic and anaerobic conditions. It was observed that aerobic degradation of DEHP was possible with landfill leachate as the inoculum. Anaerobic degradation of some of the compounds was also detected. Leachate showed good capability of degrading these phthalates. The results indicate a great potential for biological degradation of phthalates in landfills (Angelidaki 2000).
Phthalates can however accumulate in environments that are highly anaerobic and poorly colonized by microbes. Furthermore, cold environmental conditions slow their biodegradation.
Phthalates are well known to accumulate in river sediment (Petrovic, 2001). The occurrence and the potential adverse effects of phthalates in stream-bed sediment were assessed at 536 sites in 20 major river basins across the United States, from 1992 to 1995. It was found that concentrations of phthalates were about 10 times higher at sites influenced by urban activities than at sites in other land-use areas (Lopes, 2001).
Call et al. (2001) studied the acute toxicity of six phthalate esters, including DMP, DEP, DBP, BBP, DHP, and DEHP on sediment-dwelling animals (collectively called the benthos). No significant survival reductions were observed when three species were exposed to either DHP or DEHP at concentrations approximating their water solubilities (Call, 2001).
This is supported by numerous earlier studies on the aquatic toxicity of DEHP which collectively found that this plasticizer does not induce adverse effects in aquatic organisms at (or below) its solubility limit in water (Drew, 2001; Staples 1997b).
McDowell et al. (2001) studied phthalates from sediment samples collected near the outflow of a sewage treatment plant at the western end of Lake Ontario. Analysis of sediment samples indicated that DEHP was present at very high concentrations; ranging from a mean of 29.7 µg/g dry weight at a site near the sewage treatment plant outflow, to a mean of 6.5 µg/g dry weight at a site 300 m away. Di-n-butyl phthalate, and benzylbutyl phthalate were judged to be present in some sediment samples but at concentrations below the method detection limits (< 0.3 µg/g) (McDowell, 2001). The study finds that sewage treatment plant effluents are a major source of phthalates (in particular, DEHP) in the aquatic environment.
Long-chain phthalates like DEHP are only partly degraded in landfills and sewage treatment plants. Therefore, accumulation of phthalates to landfill leachates and the aquatic environment cannot be excluded. DEHP in particular is considered to be persistent and to accumulate in anaerobic sediments (Argus, 2000).
Di-n-butyl phthalate and DEHP were detected in various water samples such as river water, well water and tap water in Japan (Hashizume, 2002). The phthalates had degraded to varying extents in the Tempaku River water - almost 100% of DEP, di-isobutyl phthalate and DBP; and approximately 70% of DEHP (Hashizume, 2002).
In drinking water samples from Leipzig (Germany) and Katowice (Poland), four common phthalates (DEP, DBP, BBP and DEHP) were found to be present in concentrations between 0.02 and 0.6 µg/l (Luks-Betlej, 2001).
Exposure of aquatic organisms to concentrations of DEHP above the limits of its water solubility can lead to distortion of experimental results due to artefacts such as boundary layers of phthalates in test chambers. In experiments the test concentrations of phthalates in excess of their aqueous solubilities are not relevant in assessing the aquatic impact of these chemicals (Drew, 2001; Staples 1997a).
Insufficient studies have been conducted to date on the terrestrial toxicity of phthalates. There is scattered data in the literature on the toxicity of phthalates on flies and birds that suggest a low risk.
In conclusion, phthalates are widely dispersed thoroughout the global environment. The risk from a toxicity perspective however is limited since their poor solubility in water means they are generally present at lower concentrations than those required to induce any known toxic, reproductive or developmental effects on living organisms. Furthermore, most phthalates readily biodegrade in aerobic environments.
The majority of leaching studies of DEHP from PVC has been reported for the medical area. A study in the non-medical field for instance reports that DEHP was found to leach out to a considerable extent from automotive shredder residue containing plasticized PVC. The level of DEHP leached was high given that only distilled water was used as the leaching medium (Sakai, 1998).
Exposure of laboratory rodents to phthalates is associated with developmental and reproductive anomalies, and there is concern that these compounds may also potentially cause adverse effects on human reproductive health (NTP-CERHR, 2000). It has been known for a long time that high doses of phthalates can cause testicular toxicity and pursuant fertility problems in laboratory rodents (IPCS 1992). More recent studies in rodents have associated DEHP exposure with morphological defects in development of the male reproductive system, and subtle biochemical or histological effects when exposure has been at the critical times of testicular development and maturation.
An October 2000 report from the National Toxicology Program Center for the Evaluation of Risks to Human Reproduction cited concern that high levels of DEHP exposure in human infants and toddlers could possibly adversely affect male reproductive-tract development (NTP-CERHR Expert Panel Report on Di(2-ethylhexyl)phthalate, October, 2000, Final, NTP-CERHR-DEHP-00).
DEHP was found to produce morphological changes in the testes, including apoptosis (programmed cell death), necrosis, and loss of spermatogenic cells, which resulted in testicular atrophy in rats (Park, 2002). Body weight gain and testicular weight (absolute and relative) were significantly lower in DEHP-treated rats (Park, 2002).
Despite many earlier published reports, the potency of phthalates in comparison with natural hormones is very low. Although there is some evidence for activity in animals, high doses are usually necessary, and there is controversy and uncertainty in the interpretation of some positive findings.
Legler (2002) measured the estrogenic potency of phthalates that can be found in sediments with an estrogen receptor-mediated luciferase reporter gene assay. Of the phthalates tested BBP was the most estrogenic, though with a potency approximately 100,000 times less than 17beta-estradiol. DEHP showed no activity. BBP is not widely used in PVC formulations.
Based on significantly higher levels of phthalates (DMP, DEP, DBP, and DEHP) in the serum of very young Puerto Rican girls with premature breast development compared to those without signs of premature sexual development, Colon et al (2000) hypothesise that the condition may be the result of phthalate oestrogenic activity. Of 41 cases studied, 25(60%) had measurable DEHP serum levels which ranged from 187-2098 µg/l whereas only 5 of 35 controls (14%) had DEHP in the range 276 - 719 µg/l. Surprisingly only 5(12%) of the cases with measurable DEHP had detectable levels of MEHP and this was a very small fraction of the DEHP level (1 - 2%). Of the control samples analysed, only one showed levels of di-isooctyl phthalate. The authors consider the study is suggestive of a possible association between plasticizers and estrogenic and antiandrogenic activity causing premature breast development in a human female population.
In contrast, a number of studies such as Paganetto (2000) have concluded that the common phthalates studied do not exhibit any estrogenic activity. Similarly Blair et. al. (2000) studied eight of the most common phthalates in a oestrogen receptor (rat) competitive binding assay and found none competed strongly for the receptor. Only BBP and DEHP showed slight competition for the receptor. Others conclude that the in vitro studies demonstrating estrogenic activity of phthalates do not extrapolate to in vivo studies and therefore are not relevant to humans (Moore, 2000).
The weight of scientific research suggests that phthalates are not carcinogenic to humans.
A review study has assessed the role that high levels of phthalate esters have in the induction of rodent liver tumors. Rodent studies with DEHP and DINP demonstrated that levels of these phthalates correlated with induction of both liver tumors and markers for peroxisomal proliferation (McKee, 2000). However in hamsters and primate species, phthalate treatment did not induce peroxisomal proliferation or other preneoplastic changes seen in rodents. Thus this data along with those from other studies on peroxisomal proliferation support the view that the carcinogenic effects of DEHP and DINP in rodents is a species specific phenomenon not relevant to humans (McKee, 2000).
Two sets of independent risk assessments have recently been undertaken on phthalates in the US and Europe.
In 2000, The Centre for Evaluation of Risks to Human Reproduction (CERHR) conducted risk assessments on DBP, DEHP, DINP, DIDP and BBP. The CERHR risk assessments have found that there is little risk to the general population with exposure to DIDP, DINP and DBP. However due to limited data on the anti-androgenic effects of DEHP and BBP, these two phthalates are still under review. The CERHR principal recommendation for DEHP was to conduct further studies to determine the effect of DEHP in medical treatment of children and to conduct in-depth animal studies concentrating on late gestation and neonatal periods.
Five phthalates (DBP, DEHP, DINP, DIDP and BBP) have also undergone the EU risk assessment process in line with the requirements of EU Council Regulation 793/93. The assessments involved a rigorous procedure during which scientific data and research from a wide range of sources was evaluated by national scientific institutes and qualified bodies. The EU risk assessments are presently at draft stage and under scientific review by an expert panel. The risk assessments have concluded that there is an issue with DEHP (8 carbons), but not with DIDP (10 carbons) or DINP (9 carbons). The evidence is quite clear that DEHP poses certain environmental issues and is of concern for neonates (babies).
After the risk assessments for DINP, DIDP and DBP go through a final approval process by the EU Commission and EU Parliament they will be published in the EU Office Journal. Finalization of the DEHP risk assessment has been delayed in order for important new data to be taken into consideration from studies conducted in Germany and the US. Final publication of the DEHP risk assessment is likely around the end of 2002/early 2003 (Vinyl2010 Progress Report, 2002).
A two generation toxicity study for DEHP sponsored by ECPI (European Council for Plasticisers) which indicated that the likelihood of adverse health effects with DEHP are much less than previously thought. These findings are now considered crucial to the EU risk assessment on DEHP. Also part of the DEHP risk assessment is a fish multi-generation study to be completed by the end of 2002.
Substitution of DEHP is now occurring in Europe (ECVM private communication, 2002). DEHP is largely being replaced by DIDP and DINP which have been given a low risk assessment rating. Since DIDP and DINP have longer chains than DEHP their volatility is lower and their plasticizing action in PVC is also somewhat lower and thus somewhat higher concentrations are required. In Europe the usage of DEHP has dropped from 55% in plasticized PVC to 35% in just two years (ECVM private communication). The main substitute plasticizer is DINP. EU risk assessment procedures have not identified any negative environmental impacts of DINP.
Phthalate plasticizer substitution is more problematic than cadmium and lead stabilizer replacement because phthalate alternatives can incur some 50% higher cost. Since phthalates can comprise up to 50% of a particular PVC formulation this adds considerably to the overall cost. Three main types of plasticizers have been proposed for replacing phthalates:
The last type of plasticizer is known as DINCH (di-isononyl cyclohexane). Though presently they are twice the cost of phthalates, a 20,000 tpa manufacturing plant is coming on-stream shortly (in Europe) that will lower their price. This plasticizer appears to be a very effective substitute for phthalate plasticizers, since they are essentially hydrogenated phthalates that have similar properties to traditional phthalates but do not share the androgen disrupting behaviour exhibited by some phthalates (ECVM private communication). The first application that these plasticizers are being used in is for sensitive applications such as PVC medical plastic products and toys.
There is a lack of toxicological data on the substitute phthalate plasticizers (namely adipates, citrates, benzoates and mellitates). Presently the scientific body of the EU, the CSTEE (the EU Scientific Committee for Toxicity, Ecotoxicity and the Environment), is planning risk assessment on these plasticizers (ECVM private communication).
It is emphasised that any replacement compound for phthalates in consumer products or medical devices should be properly tested and scientifically evaluated. The EU risk assessments are more likely to identify toxicological data gaps than provide carte blanche clearance of possible substitutes.
Bis-phenol A (BPA) is used in plasticizers for PVC. Its function is to stabilize the plasticizers and is more often encountered in plasticizers used in outdoor applications where degradation of the plasticizer is more likely. Plasticizers for the production of heat resistant PVC cable insulation are generally pre-stabilized with 0.5 wt% BPA and offer excellent electrical properties and long-term stability. The release of bis-phenol A from PVC products has been reported (Yamamoto, 1999). Some findings of weak estrogenic` activity is reported (Howdeshell, 1999) but whether it has effects in humans is still a matter of debate.
Overall the amount of bisphenol A used in PVC products is not great, however more attention could be paid to the use of bisphenol A as a co-stabiliser for PVC.