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End-of-life Environmental Issues with PVC in Australia

Prepared by Dr. John Scheirs,
ExcelPlas Polymer Technology (EPT) for
Environment Australia, June 2003

6. Current End of Life Management of PVC

6.1 In Situ Disposal

The majority of end-of-life PVC pipes in Australia are left in situ rather then being excavated and collected. In the CBD of major cities however space limitation dictates the need for removal of end-of-life PVC pipes.

From practical experience gained from PVC water pipes and cable insulation it can be concluded that the lifetime of PVC is very long (Hjertberg, 1995). It is interesting to note that the analysis of PVC disposed of in a landfill more than 20 years ago still showed considerable amounts of plasticisers and metal-based stabilizers (Griebenow, 1992). One can therefore conclude that PVC is a very stable matrix in the absence of heat and light.

The degradation of PVC in soil burial may be accelerated by catalytic species (e.g. iron compounds) in the surrounding medium. Such catalytic species accelerate dehydrochlorination and cause embrittlement (Hjertberg, 1995). There is a limited amount of data on this topic.

Rigid PVC without plasticizer will not degrade appreciably under burial conditions. Estimates of lifetimes of up to 1000 years have been made (Hjertberg, 1995). If left exposed then it will embrittle due to UV degradation over a period of years.

6.2 Landfilling

Landfilling is the predominant form of waste disposal for PVC waste in Australia and the rest of the world. The issues to consider surrounding PVC in landfills include:

The general view expressed in the literature with regard to the potential release of stabilizers from PVC products under landfill conditions maintains that because stabilizers are encapsulated in the PVC matrix, their migration rate is expected to be extremely low and would only involve the surface of the PVC product as distinct from the bulk of the material (Argus, 1999). No decomposition of rigid PVC articles in landfills has been detected to date. Thus rigid PVC effectively encapsulates heavy metal stabilizers and prevents their extraction in bulk items. The situation is less clear however for ground-up and pulverized rigid PVC.

The early acidogenic stage of the landfill development seems to constitute the critical stage regarding emissions of heavy metals from the PVC (Argus, 2000). Also the release of heavy metal stabilizers may be accelerated by high temperatures that can occur in large landfill sites under aerobic thermophilic conditions. Compared to the total load of lead and zinc in MSW however the contribution of PVC is considered to be negligible (Argus, 2000).

Plasticizer loss from PVC can occur in the case of insufficient compatibility with the PVC compound coupled with microbial consumption at the PVC product surface. Although losses of stabilizers are generally restricted to transient leaching from the product surface, plasticized products show a higher inclination to release their stabilizers (Mersiowsky, 2002).

Thin, plasticized PVC packaging film has been found to have undergone degradation under specific landfill conditions (Argus, 1999). This was under the extreme conditions found in average landfills - that is, under aerobic conditions and at a temperature of 80C. Hot spots of up to 80C are present in some landfills but these are transient.

It has been established that as plasticized PVC ages it can loose its plasticizers causing it to harden and in most cases to crack (Monney, 2001). This cracking can accelerate the loss of plasticizers. Further, as the aged PVC becomes brittle it fragments and disperses in the landfill environment.

The recent EU Green Paper study identified some potential problems with landfilling of PVC. However the European PVC Industry has challenged the conclusions of the EU study on the basis that the extreme temperature used to accelerate aging of materials in the study undoubtedly affected the results. Other independent studies closer to real landfill conditions have concluded that PVC in landfill, including plasticized applications, is environmentally safe (Mersiowsky, 2002).

Landfill simulation experiments have demonstrated the long-term behavior of PVC products and the fate of additives under landfill conditions. The data generated thus far indicates that the PVC polymer matrix is stable (Mersiowsky, 1999; Mersiowsky, 2001b; Mersiowsky, 2002).

A comprehensive review has recently been undertaken on the long-term behaviour of PVC products and their additives under landfill conditions (Mersiowsky, 2001c). This five-year study was undertaken by universities and institutes in Germany and Sweden. The behaviour of PVC under landfill conditions was assessed in laboratory-scale landfill simulation assays. In order to make a long-term assessment, characteristic landfill processes (such as leaching and biodegradation) were enhanced with leachate and gas being monitored. The study yielded the following conclusions:

The overall conclusion of this study was that PVC products do not constitute a substantial impact on toxicity of landfill leachate and gas (Mersiowsky, 2001).

In a related study, the long-term performance of PVC products buried under soil and landfill conditions has been investigated in laboratory scale landfill simulation and no leachates or gases were detected (Mersiowsky, 1999). While the assessment of the environmental impact indicates that there is no significant contribution by PVC waste to concentrations of heavy metals in landfills, these as such were only simulations and the PVC samples were bulk items - importantly the effect of particle size and surface area were not considered. In this respect DEHP has been shown to leach from shredder residues containing PVC (Sakai, 1998). Presumably the smaller particle size of shredded PVC enhances its leaching potential due to the greater surface area over which extraction can occur.

Griebenow et. al. (1992) investigated the leaching behaviour of ground-up PVC window and shutter profiles containing lead phosphite and barium/cadmium as stabilizers in various aqueous solutions. It was found that in an acidic test medium (pH of 3) considerable amounts of stabilisers did dissolve. It was concluded however that under landfill conditions PVC moulded parts in building debris exhibit almost no bleeding of the heavy metal stabilisers and that they have very little effect on landfill leachate.

Another study investigated eight different landfill sites of various ages for organotin compounds. The levels of all target compounds ranged between less than the limit of detection of 0.1 g/L and a maximum level of 4 g/L. Only octyltin compounds can be directly attributed to PVC products with any certainty. In the case of methyltin and butyltin compounds alternative and less distinct sources of origin exist. Organotin compounds were found to be subject to microbial transformations, such as dealkylation and methylation processes (Mersiowsky, 2001a).

The long-term behaviour of plasticised PVC products (cable material and a flooring with different combinations of plasticisers) was investigated by Mersiowsky (2001b) in landfill simulation reactors. The behaviour of the various plasticisers was found to differ significantly. Losses of DEHP and BBP from the flooring were too low for analytical quantification. No loss of DIDP from the cable was detectable whereas DINA in the same product showed considerable losses of up to 70% compared to the original contents. The loss of DINA was attributed to biodegradation rather than leaching (Mersiowsky, 2001b).

6.2.1 Landfilling of Automotive Shredder Residue

PVC has been widely used to date in interior automotive applications. Large automotive parts such as instrument panels and door trim panels are generally based on an injection moulded plastic substrate (e.g. ABS, PP) which is sprayed with adhesive, vacuum wrapped with soft-touch PVC foam and then PVC sheet. It has been reported that the average US family sedan contains a total plastic content of 9.3% by weight, of which PVC constitutes 14% by weight (Sullivan, 1998).

Metal recycling processes for end-of-life vehicles are well established in Australia and are conducted by companies such as Sims Metal and Metacor. The remaining ~30% of the vehicle is converted to automotive shredder residue (ASR) or 'flock' (a mixture of plastics, glass, foam, wire, textiles). Flock by weight is typically made up of 90 % plastic. The composition of ASR is shown in the table below.

Table 6.1 Composition of automotive shredder residue in Australia.

Material % by weight
Polyurethane foam 22.6
Reinforced Polyesters * 21.9
Polypropylene 19.2
Polyvinyl Chloride 5.0
Acrylonitrile Butadiene Styrene 7.3
Nylons 3.7
Acrylic 2.5
Phenolic 2.1
Other 5.2
Source: Roger Sweeney (Australian Composite Technology) (2002).
* Includes bulk moulding compound and sheet moulding compound

The use of PVC in the automotive industry has been diminishing in recent years. However due to the vintage of the motor vehicles currently being shredded (from 1970's and 1980's) PVC still represents around 5% of the ASR of end-of-life vehicles (ELVs) (private correspondence Roger Sweeny, ACT, 2002). The Australian Bureau of Statistics estimates that more than 5% of the national fleet equating to more than 500,000 vehicles are scrapped annually.

Based on the non-metal portion of ELVs that is converted to ASR (approximately 30%) there are approximately 195,000 tonnes of ASR generated from ELVs annually with the majority going to landfill. Given that PVC represents 5 wt.% of the ASR stream this equates to about 9,750 t of PVC that is landfilled per year. Potential environmental impacts of landfilling this material include plasticizer leaching and resource loss.

6.2.2 Cable Insulation

Another area that warrants further attention is PVC cable waste that is lead stabilized. Often PVC cables are shredded to liberate the metal wire leaving the PVC material known as fluff, which may be landfilled or recycled. The high surface area of the fluff may constitute a leaching hazard from the lead stabilizer aspect (Sell, 1993).

6.3 Mechanical Recycling of PVC

Plastic recycling can be divided into a four-tier hierarchy: primary (ie. in-house use of scrap plastic), secondary (i.e. mechanical recycling of post-consumer plastic, tertiary (i.e. pyrolysis and feedstock processes) and quaternary (i.e. incineration with energy recovery).

PVC can be remelted and reprocessed several times due to its thermoplastic nature. Mechanical recycling makes ecological and economic sense wherever sufficient quantities of homogeneous, separated and sorted PVC waste streams are available. In these cases the quality of the recyclate permits manufacture of the same or similar products. PVC can typically be recycled into waste water pipes, hoses, floor tiles and shoe soles.

In mechanical recycling processes the chemical composition remains largely unchanged. Conventional mechanical recycling processes generally involve separation, shredding, grinding, pulverization and extrusion with melt filtration. PVC products that lend themselves favourably to mechanical recycling include profile, pipes, cable insulation, bottles, rigid sheeting and window profiles (Bhl, 1996).

The Dutch Industrial Applied Research Institute (TNO) has recently completed an eco-efficiency study on the mechanical recycling of plastics. A key finding was that the mechanical recycling of polymers taken across all plastic types was only eco-efficient up to a level of about 15% (Eggels, 2001). Above this 15% threshold the collection of plastics waste (including PVC) becomes more expensive and one begins to encounter a more contaminated, mixed waste stream that requires more intensive separation techniques and more cleaning/washing. For example mechanical recycling is not an ecologically sound option for post-consumer PVC film and sheeting, which is generally too contaminated, too thin and too dispersed.

In Europe an emphasis on recycling as a recovery option for plastics packaging is beginning to drive Europe to ever increasing waste management costs with limited environmental gain (Eggels, 2001). An increase in the recycling rates from 15% to 50% increases the end-of-life management costs by a factor of 3 while environmental impacts remain broadly similar (Eggels, 2001).

There are few environmental impacts associated with mechanical recycling of PVC waste. One of the alleged impacts is the dispersion of heavy metal stabilizers into the environment via the various PVC recyclate streams. Lead and cadmium stabilized PVC come from applications with a long service life; e.g. underground pipes (> 50 years), cables (> 25 years) and profiles (> 25 years). Any restrictions on the presence of lead or cadmium in PVC recyclates may jeopardize the mechanical recycling industries. With controlled-stream recycling there is limited environmental dispersion of these heavy metals since the PVC recyclates are usually used again in long-life products. For instance PVC cable insulation is largely recycled into industrial hose (personal communication Basil Siganakis, Cryogrind, 2002). Likewise PVC pipe can be recycled into multilayer PVC pipe or injection-moulded pipe fittings.

No significant environmental impacts occur during the actual mechanical recycling process other than the generation of lead-laden dust that should be trapped using dust cyclones and dust filters hose (personal communication Graham Johnstone, Nylex SRM, 2002). Workers in such PVC recycling plants are blood-tested for lead on a regular basis and no problems have been highlighted (private communication Graham Johnstone, Nylex SRM, 2002). The recent EU Green Paper supports the contention that there are no significant environmental impacts during mechanical recycling (EU Green Paper, 2000).

There are however certain barriers that limit the more widespread collection and mechanical recycling of PVC. These indirectly contribute to more PVC waste being sent towards landfill.

6.3.1 Impediments to PVC Recovery and Recycling in Australia

The primary impediments include the following:

Factors such as the presence of sustainable end markets for PVC recyclate, the existence of sorting processes and availability of recycling technologies for PVC waste streams are not considered significant barriers that limit its greater recycling.

6.3.2 Specific Technical Issues and Barriers for Mechanical Recycling of PVC


6.3.2.1 Heat Stability

It is important to note from the outset that PVC is a heat sensitive material that degrades by dehydrochlorination - that is, by releasing hydrochloric acid. In chemical terms, formation of conjugated double bonds causes colour change and loss of other properties. PVC compounds experience heat history in processing operations such as extrusion/moulding, embossing, thermoforming, laminating and scrap rework. The role of the heat stabilizer is to delay heat degradation so that the compound can be formed into a product before it degrades. The stabilizer does this job by absorption of hydrogen chloride, displacement of active chloride atoms, free radical scavenging, disruption of double bond formation and deactivation of degradation by products (Blh, 1996).

Due to PVC's limited heat stability it is always formulated with heat stabilizers that effectively scavenge the hydrochloric acid evolved when PVC begins to degrade. It is noteworthy that some 40% of the heat stabilizer inventory is consumed in the first processing step. Therefore upon recycling, the heat stabilizer levels begin to get rather low. Restabilization of PVC is often required during reprocessing and is especially necessary after extended periods of outdoor weathering. Supplementary heat stabilizer usage adds extra cost to a product with already slim margins.

Calcium-zinc heat stabilizers are used in applications such as blow moulded PVC for food contact, potable water pipe, packaging film and sheet. They can provide compounds with crystal clarity, low odour properties and can meet statutory requirements for food packaging. However, the stabilization power of calcium-zinc in PVC is relatively weak and therefore mechanical recycling capability is poor (personal communication, Graham Johnstone, Nylex SRM, 2002).

6.3.2.2 Cross-staining reactions

Recycling post-consumer PVC products is difficult because of the wide range of additives and adjuvants used in PVC compounds. An issue that exists due to the large diversity of PVC formulations is incompatibility between different heat stabilizer systems used in PVC. This incompatibility results from a reaction between the various metal salts to produce dark stains, objectionable odours and lower overall stability. The most well-known of these arises from the antagonistic reaction of certain tin stabilizers with lead stabilizers and is known as lead cross-staining (Bhl, 1996). These antagonistic reactions can impact negatively on recycling since it is difficult to readily ascertain which stabilization systems exist in different PVC products. Some of these reactions are summarized in the table below.

Table 6.2. Possible antagonistic interactions of different stabilizer groups

Stabilizer combination Colour of the reaction product
Ba/Cd + SnS Yellow cadmium sulphide
Pb + SnS Brown to black lead sulphide
Sb + SnS Orange antimony sulphide
Ca/Zn + SnS White zinc sulphide
Ba/Zn + Sn White zinc sulphide
Ba/Cd + Pb No discoloration
Ba/Cd + Ca/Zn No discoloration
Ca/Zn or Ba/Zn + Pb No discoloration
Source: Blh (1996).

Therefore melting a mixture of recycled PVC products is problematic since by-products from antagonistic reactions can detract from the stability of the recyclate. For instance lead cross-staining cannot be tolerated in rigid pipe fittings. However in foam core pipe technology tin-lead admixtures can be tolerated since it is sandwiched between two layers of virgin PVC (Scheirs, 1998, p. 255-56).

6.3.2.3 Suitability of Rigid PVC for Recycling

Rigid PVC is far more sensitive to degradation during reprocessing than plasticized PVC. In addition, the higher melt viscosity of rigid PVC leads to shear-induced heating during extrusion (Johnstone, 2002).

6.3.2.4 Cross-contamination of rigid and plasticized PVC

Flexible and rigid PVC generally cannot be mixed together and recycled. Therefore another impediment to the mechanical recycling of PVC is that it is necessary to separately recycle rigid and plasticized PVC products. Plasticized PVC cannot be added to a rigid PVC stream because of the loss of rigidity that results. Conversely only a limited amount of rigid PVC can be tolerated in a plasticized PVC stream, otherwise embrittlement may occur (private communication Graham Johnstone, Nylex SRM, 2002).

6.3.2.5 Cross Contamination with Other Plastics

A major problem with PVC collection bins for post-industrial PVC scrap is that in many cases the plastics industry does not conduct proper source separation (personal communication, Trevor Walton, Vision Plastics, 2002). Consequently the level of cross contamination with other polymers and even non-polymers becomes a major issue.

PET is a common contaminant in the recycled PVC stream. PET contamination occurs a great deal in PVC bottle recycling since these two polymers types are virtually indistinguishable to the untrained eye. Likewise PET sheet is a major contamination problem in PVC sheet recycling. Since the two polymers share the same specific gravity they cannot be separated by conventional float-sink techniques used in the plastic recycling industry. PET comes out as a lump in finished PVC products and causes drag marks.

It is interesting to note that just 400 ppm of PET contamination can initiate pipe failure (Watson, 2002). This is because PET fragments remain as hard, unfused lumps even after PVC melt extrusion.

6.3.2.6 Particulate Contamination

Particulate contamination cannot be tolerated in pressure pipe applications since it can initiate brittle failure by acting as a stress concentration (personal communication, Peter Chapman, Vinidex, 2002). To effectively restabilize rigid PVC it needs to be micronized to a fine powder - a step that adds cost. This is in contrast to other thermoplastics such as polyethylenes which can be mechanically recycled simply by granulation to flake followed by melt compounding. PVC fragments when extruded have a marked tendency to degrade if they contain insufficient thermal stabilization. Without adequate heat stabilizer PVC quickly forms hydrochloric acid which exerts an autocatalytic effect on PVC degradation. If the HCl is not quickly mopped up by a heat stabilizer that scavenges the HCl it will cause a solid burn of the PVC resulting in decomposition particles (Chapman, 2002). Such decomposition particles are harder than the surrounding PVC matrix and act as stress concentrations. Such stress concentrating particles can cause premature failure in high stress applications such as PVC pressure pipe.

For this reason in-house recycled PVC from pipe production is generally used lower down the specification chain into less demanding and more tolerant applications. The lowest specification pipe product is 90 mm stormwater pipe. Other low specification PVC products include drain-waste-vent (DWV) pipe and conduit (Chapman, 2002).

In addition during any in-house regrind operation the opportunity exists for the introduction of contamination from dust, cross-contamination by different resins and by metal fragments from grinding blades. Because of contamination concerns recycled PVC pipe is generally used in the production of injection moulded pipe fittings which are less demanding from a hoop stress perspective (personal communication, Peter Chapman, Vinidex, 2002). It is also recycled into the foamed core middle layer of conduit.

6.3.2.7 PVC textile composites

Fabric-backed PVC contains far too much fibre to be effectively recycled with conventional mechanical recycling technologies. Furthermore the fibre and textile reinforcement is generally adhesively bonded to the PVC and therefore difficult to liberate even after size reduction steps (personal communication, Charles Hrubos, Nylex Mentone, 2002).

Considerable PVC scrap is generated in Australia during the production of vinyl-coated cloth and fabrics used for automotive trim. For example, 15% PVC scrap rate occurs with the production of PVC trim and this material is landfilled because the fibre reinforcement and scrim are too difficult to separate. The largest manufacturer of this type of product in Australia currently landfills some 500 tonne of fibre-reinforced PVC and foam-backed PVC (personal communication, Kevin Thomson, Nylex Mentone, 2001).

6.3.3 PVC Recycling in Australia

The major PVC recyclers and the details of their recycling operations are described in Appendix A. The table below summarizes the status of PVC recycling in Australia at present.

Table 6.3. Major PVC recyclers in Australia and their approximate production tonnages
Company Input Material Type
T/m
Nylex SRM Flexible PVC
250
Cryogrind Flexible PVC
180
Cryogrind Bottles
10
Key Plastics Profile, Sheet, Pipe
100
Vision Plastics Clear, rigid sheet
50
Pacific Chemicals Skeletal sheet
30
Pacific Chemicals Cable insulation
20
Australian Plastics Recycling Bottles/Sheet
20
Repeat Plastics Cable insulation
20
Silverfox Sheet, Profile
20
Aust. Synthetic Fibres Sheet
20
APN Compounding Profile
<10
Visy (export) Bottles
30
Total (minimum)  
760
Source: multiple industry interviews (2002).

The total quantity of PVC recycled in Australia is approximately 9,000 tpa. Note that exported PVC bottles and sheet are included in the above recycling figure - the rationale being that if this material were not exported it would be recycled domestically given the present demand for such clear, rigid PVC. Aside from this exported PVC waste still represents a diversion of PVC waste from domestic landfills.

It is important to emphasize that the majority of the PVC recycled in the above table is of the pre-consumer type (that is, post-industrial).

There is very high demand for clear, rigid PVC, which is achieving very high levels of recovery and recycling. Furthermore approximately 70% of clear, rigid (post-consumer) PVC is being exported (personal communication, Trevor Walton, Vision Plastics, 2002).

Filled, coloured and textile-reinforced PVC however are presently being recycled only at low rates. Thus significant opportunities exist for the recycling of filled and coloured PVC. This is also the case for textile scrim-reinforced PVC sheet (e.g. Nylex trimmings) (personal communication, Graham Johnstone, Nylex SRM, 2002).

6.3.4 Economics of PVC Recycling

Recycled PVC feedstock pricing varies widely depending on purity, from ~$100/t to ~$650/t for the cleanest material. Baled PVC bottles are $300/t on average. From the table below it is evident that the margins between the cost of the recycling and the final price of PVC recyclate are slim.

Table 6.4. Indicative Costs of Various PVC Recycling Steps
Recycling Step Approx. Cost
Granulation 30 cents/kg
Pulverization 35 cents/kg
Extrusion and pelletization 50 cents/kg
Addition of formulants (e.g. pigments) 20 cents/kg
Recyclate price 90 cents/kg - $1.60/kg
Source: Shaw (2002) and Johnston (2002).

6.3.5 Comparison of PVC Recycling in Australia and Europe

In Australia, PVC is predominantly being landfilled with the remainder recycled. A very small fraction of PVC is incinerated as part of the medical waste stream.

Presently in Europe some 4-5% of PVC waste (total available) is mechanically recycled with ~32% being incinerated with energy recovery and ~64% of PVC waste is still being landfilled. The percentage being landfilled varies widely between individual member countries (from 10% in Switzerland to 90% in the UK).

In order to obtain a precise measurement of the level of PVC in the waste stream it is useful to draw upon European experience in this area. In Europe it was found that it was neither practical nor accurate to attempt to measure the PVC content of the waste stream directly through waste audits. Instead a theoretical model was found to be superior. Such a model uses a number of inputs that include:

Of course, many assumptions have to be made and comprehensive data bases need to be compiled. This type of model has been developed by the Association of Plastics Manufacturers of Europe (APME) and is denoted as the Eckstein Model (Source: J. Eckstein, EuPC, Sept., 1999). This model allows one to predict the level of PVC entering the waste stream, in say the next ten years. The forecasts indicate increasing levels of PVC waste, as many building and construction applications reach their end of life.

6.3.6 Calculating the Level of PVC Recycling

To calculate an accurate figure for the level of PVC that is being recycled, it is necessary to clearly define the waste streams into the following categories:

  1. Processing waste (also known as in-house regrind) which is now being recycled at levels of upto 85-90%.
  2. Production waste streams (e.g. off-cuts, skeletal sheet, trimmings, purgings) which are generally collected by an independent recycler.
  3. Installation waste (e.g. off-cuts at point of installation such as pipe off-cuts, vinyl flooring off-cuts, vinyl siding off-cuts) which is generally landfilled by the installer.
  4. Post-consumer waste (e.g. bottles, containers and cable strippings) which at present achieves fairly low recycling rates.

An important distinction needs to be made between the first three categories above, which relate to pre-consumer waste and the last category, which is post-consumer waste. Any presentation of recycling rates should delineate these two groups.

In Australia, no model exists as yet to calculate the available quantities of PVC in the waste stream. Until such estimates are made it is not possible to accurately calculate the real percentage of PVC that is recycled. In Europe today the values in the following table apply.

Table 6.5. Volumes of pre- and post-consumer PVC recycled in Europe

PVC waste
Available (kt/a)
Recycled (kt/a)
% recycled
Pre-consumer
500
420
84
Post-consumer
2670
100
4
Source: Prognos Consultants, Basel and ECVM industry data.

The installation waste stream in Europe has a high potential for recycling since it is generally of high quality and comprises a readily identifiable single product. Opportunities exist to achieve higher recycling rates of this material by initiatives such as take-back schemes to recover this type of installation waste and via back-loading operations.

6.3.7 Main streams of PVC waste being recycled

The following table shows the main streams of PVC pre-consumer and post-consumer waste that is being recycled in Australia today.

Table 6.6. Volumes of PVC recycled in Australia today.
PVC applications
Approx. Volume (kt per year)
Cable
1.5-2.0
Sheet
2.2
Pipe and Profile
0.1
Flexibles (mainly post-industrial)
4.8
Bottles
0.4
Source: Multiple industry interviews (2002).

When comparing Australian figures with European data consideration should be given of course to prevailing differences in population density, consumption patterns, infrastructure etc.; all of which play an important role in determining recycling volumes. The situation in Europe is given in the following table.

Table 6.7 Volumes of PVC recycled in Europe
PVC applications
Approx. Volume (kt per year)
Cables
35
Window frames
5
Bottles
10
Pipe
6
PVC packaging in mixed plastics
15
Source: ECVM, 2000.

It is also important to note that an additional 12 kt of PVC (as part of a mixed plastic waste stream) is feedstock recycled by incineration with energy recovery. The comparatively high level of pipe recycling is due mainly to initiatives (for example in the Netherlands) for collection of PVC pipes aided by the high population density and the strong social conscience/attitude towards recycling. Note also that recycling of PVC window frames is prevalent in Europe whereas it is non-existent in Australia since PVC for window frames is not a preferred material locally. From the European figures it is clear that recycling of PVC cable insulation is occurring at a high level. This is because the copper content of the cables pays for its collection. Much of the PVC recyclate from recycled cable insulation goes into traffic management products such as the sleeping policeman and also industrial flooring such as compression moulded tiles.

Taken collectively the overall level of post-consumer PVC recycling in Europe today is estimated at approximately 100 kt/a. The voluntary commitment of the European PVC industry (first issued March 2000 and subsequently updated March 2001) has committed to recycle an additional 200 kt/a of post-consumer PVC waste by 2010 (in addition to the 1999/2000 volumes).

6.3.8 Voluntary commitment of the European PVC industry

The voluntary commitment in Europe proposes the phasing out of lead additives by 2015. The ECVM maintain that lead is perfectly safe in PVC products but nevertheless acknowledges that there is public concern. Though the time-line seems protracted, technical products such as window frames require long test times and actual sun exposure in order to evaluate the effectiveness of new stabilizer systems in preventing discolouration and embrittlement.

The European PVC industry has also given the following commitments for the mechanical recycling of PVC:

Table 6.8 European timing for achieving PVC recovery targets
Sector
25% CAW
50% CAW
Pipe and Fittings
in 2003
in 2005
Window profiles
in 2003
in 2005
Flooring
in 2006
in 2008
Roofing membranes
in 2003
in 2005
Where CAW = Collectable Available post-consumer PVC Waste
Source: ECVM, 2000.

Pipes and fittings are being recycled largely by Wavin (PVC pipe manufacturers) in the Netherlands, while window profiles are being recycled chiefly by Veka (PVC window profile manufacturers) in Germany. Both flooring and roofing membranes are presently being processed by a plant in Cologne, Germany, which has a capacity of 1000 tpa. Therefore in order to meet the 2005 and 2008 targets new technology such as solvent-based separation is required (see table below).

Table 6.9. How the 200 kt target by 2010 will be achieved
Product Contribution Technology
Pipe, Windows ~ 65 kt Conventional mechanical recycling
Cables, Flooring, Coated fabrics, Membranes ~ 135 kt Novel processes i.e. Vinyloop and Feedstock recycling
Source: ECVM, 2000.

An issue with the mechanical recycling of PVC products such as profile and pipe which generally contain cadmium and lead stabilizers is that cascade recycling of these products would serve to spread the heavy metals into other products. This is presently of concern in Europe where even though many new PVC products are heavy metal free, PVC recyclate still has a significant heavy metal content and will do so for many years. The PVC industry in Europe has addressed this concern by proposing that PVC products be recycled within controlled chains (or controlled-loops). While controlled chains is not the same as closed-loop recycling it essentially means that PVC products will be recycled into the same application sector. For instance window frames could go into building profiles. Similarly pipes could go into pipe fittings.

Australia has now introduced its own Product Stewardship Commitment in 2003. The commitment has been signed by al major players in the Australian PVC industry. See www.vinyl.org.au for further details.

6.3.9 Directives affecting PVC recovery and recycling in Europe

Quite apart from the voluntary commitment of the PVC industry there are three other directives involving PVC that enforce targets for recycling and recovery:

It is noteworthy that the commitment to recycle 200 kt of post-consumer PVC waste is in addition to the 1999 volumes and in addition to the above directives. This makes the actual total target closer to 380 kt of post-consumer PVC waste (ECVM private communication).

6.3.10 Solvent-Based Mechanical Recycling

The Vinyloop technology developed by Solvay (Europe) allows the recycling of PVC together with most of its additives into a compound that can easily be used for production of high quality products. The Vinyloop process can recover pure PVC compound from PVC composite waste. PVC compounds recovered in this process can be reused in closed loop recycling schemes (e.g. cable insulation) or processed into a large variety of high value applications by calendaring, extrusion and injection moulding. Details of the technology are included in Appendix B.

Since this process precipitates PVC compound out of solution into free-flowing grains (that have a bulk density similar to that of virgin compound), it avoids the cost of compounding and repelletizing. Thus the fundamental advantage of the Vinyloop process is that it is profitable because there is no need to repelletize, which typically is an energy intensive and costly step in alternative recycling processes. Vinyloop can handle all sorts of PVC waste such as wire and cable or flooring. It yields a compound that potentially can go right back into the same application. The reclaimed PVC is typically grey because of the mixture of colours in the original feedstock and it contains all the various additives and stabilizers of the original compounds as well. PVC molecular weight and hence properties are not diminished because no heat history has been added. It cannot however process cable scrap in combination with PVC flooring waste due to the incompatibility between lead and tin stabilizers.

The process is ideally suited for recycling of PVC composites such as:

6.4 Thermal Processes for End-0f-Life PVC


6.4.1 Introduction

Thermal processes such as pyrolysis and gasification fall under the classificaton of tertiary recycling while incineration with energy recovery is considered quaternary recycling. The table below summarizes the main waste-to-energy processes for plastic waste and their defining characteristic.

Table 6.10. Main waste-to-energy processes for waste plastics
Process Characteristic feature
Combustion (incineration) performed at stoichiometric or excess air
Gasification partial combustion (less than stoichiometric amount of air)
Pyrolysis absence of air

Although the above definitions are quite well accepted there are numerous new processes that are hybrids of the above classifications since they incorporate two of these processes often with some variance. For example some modern incinerators work on a two-stage pyrolysis-combustion process. Some other systems use a combination of pyrolysis-gasification. Also while gasification relies on low levels of oxygen, steam gasification processes rely on the injection of steam and not air.

Waste-to-energy processes for PVC utilize some form of thermal treatment step to extract heat content, combustible gases and/or hydrochloric acid from the PVC waste. PVC is not an ideal polymer for energy recovery owing to its high chlorine content due to the fact it is 57% derived from salt and only 43% derived from petroleum products, unlike other polymers. Rigid PVC (with 46 wt.% chlorine) has an energy content of 16 MJ/kg while flexible PVC (with 29 wt.% chlorine) has an energy content of 20 MJ/kg due to the presence of chlorine-free, high calorific plasticizer (Rijpkema, 1999). For comparative purposes polyethylene, polypropylene and coal have calorific values of 46, 44 and 29 MJ/kg respectively.

End of life PVC may be incinerated via various routes including:

The major issue that the incineration of PVC raises is the potential for the generation of dioxins. Dioxins are considered to be a toxic by-product of incineration. Since this is such important topic a discussion on the factors contributing to dioxin formation is provided below.

6.4.2 Dioxin and PAH Formation

Chlorinated dioxins and related compounds (chlorinated furans) are extremely potent toxic substances, producing toxicological effects in humans and animals at extremely low levels. These compounds are persistent in the environment and accumulate in the food chain. They are now dispersed globally and every member of the human population is exposed to them, primarily through the food supply and via mothers' milk. An emerging body of information suggests that dioxin contamination has reached a level that may pose a large-scale, long-term public health risk. Of particular concern are the potential effects of dioxins on reproduction, development, immune system function and carcinogenesis (Thornton, 1996). It is important to note however that dioxin emissions peaked in the 1960s-70s and have fallen dramatically since especially in the US and Japan.

The incineration of PVC (or other chlorine-containing hydrocarbons) can lead to releases of polychlorinated dibenzo-p-dioxins (PCDD) and polychlorinated di-benzo-p-furans (PCDF) (Hoenich, 2002). Collectively these dioxins and furans are denoted as PCDD/F. PCDDs chlorinated in positions 2,3,7,8 have the highest toxicity, the most hazardous being 2,3,7,8 tetrachloride dibenzo-p-dioxin (2,3,7,8 TCDD). Dioxin concentrations are commonly expressed as their toxic equivalents (TEQ) relative to the most toxic form, 2,3,7,8 TCDD. Polycyclic aromatic hydrocarbons (PAHs) are another class of highly toxic incineration by-products. The toxicity of PAHs is thus usually expressed in dioxin equivalents based on the toxicity of 2,3,7,8 TCDD (Blomqvist, 2001).

Regarding the formation of dioxins during combustion processes, a clear distinction should be drawn between the incineration of:

In the first category there appears to be little relationship with the proportion of PVC in the waste and dioxins produced, since chlorine from other sources is already present. In the second category there is clear evidence that the presence of PVC contributes directly to dioxin formation as there are no other major sources of chlorine in end-of-life commodity polymers.

Katami et al. (2002) studied the exhaust gases from the combustion of PVC, polyethylene, polystyrene, PET and their various mixtures, for PCDDs, PCDFs, and coplanar polychlorinated biphenyls (PCBs) by gas chromatography/mass spectrometry in order to investigate the role of PVC in these chlorinated compounds. Total amounts of dioxins (PCDDs + PCDFs) found in the samples are shown in the table below.

Table 6.11 Total dioxins and furans from combustion of plastics
Polymer System
Total Dioxins (PCDDs + PCDFs) (ng/g)
PE alone
11.7
PS alone
1.17
PET alone
25.3
PE with PVC
448
PS with PVC
140
PET with PVC
126
PVC alone under low-CO level*
824
PVC alone under high-CO level*
8920
*CO level in high-CO level condition was 880 ppm which was 20 times greater than that in low-CO level condition.
Source: Katami (2002).

This study found there is a strong correlation between dioxin formation and chlorine content. The formation of coplanar PCBs ranged from 0.095 ng/g (PE alone) to 77 ng/g (PVC alone under high-CO level). The results indicate that PVC contributes significantly to the formation of PCDDs, PCDFs, and coplanar PCBs from mixtures of plastics upon combustion (Katami, 2002).

Other work investigating the effects of PVC and sodium chloride on the formation of dioxins during the combustion of paper (newspapers) has been conducted by Yasuhara (2001). Exhaust gases from the combustion of newspaper alone, wood alone, and from newspapers mixed with sodium chloride or PVC were analyzed for dioxins by gas chromatography/mass spectrometry.

Table 6.12 Total dioxins and furans from various waste materials
Material
Total Dioxins (PCDDs + PCDFs) (ng/g)
newspapers alone
0.186
wood (branches) alone
1.42
newspapers impregnated with sodium chloride (Cl wt % = 3.1)
102
newspapers impregnated with sodium chloride mixed with PVC (Cl wt % = 2.6)
101
newspapers mixed with PVC (Cl wt % = 5.1)
146
Source: Yasuhara, (2001).

Samples with a higher chlorine content produced more dioxins and again a definite correlation was established between dioxin formation and chlorine content (Yasuhara, 2001). NaCl was found to vaporize at the temperature of the flame used for combustion of the samples (760-1080 deg.C). The results indicate that PVC and NaCl contribute significantly to dioxin formation from specific waste materials combusted in incinerators (Yasuhara, 2001).

In another recent study, the influence of the level and chlorine source on the formation of PCDD/F during combustion of MSW has been investigated by Wikstrom (2001). Eight artificial municipal solid waste fuel mixtures containing pure PVC plastic, two PVC products (PVC flooring and PVC cable insulation) and sodium chloride were evaluated. The level of chlorine as well as the chlorine source was varied for the different fuel mixtures. The total chlorine level varied between 0.28% and 1.1%. A correlation between the total chlorine in the fuel and the formation of the PCDD/F was found. However the most important variable for changes in the PCDDs/Fs formation was disturbance in the combustion condition rather than the variation in chlorine content of the fuel. Furthermore no differences in formation of dioxins between the chlorine sources could be observed (Wikstrom, 2001).

6.4.2.1 Formation of Copper Chloride (a catalyst for PCDF formation)

PVC is a strong chlorine source for the formation of copper chloride which in turn is a potent catalyst for the formation of PCDFs. Combustion experiments in a fluidized-bed reactor have demonstrated the effects of copper chloride as a catalyst in polychlorinated dibenzofurans formation in municipal waste incineration containing PVC as a chlorine source (Hatanaka, 2002). The results revealed that copper chloride in the waste increased the amount of PCDFs formed and made the homologue profile shift towards the highly chlorinated species. Hatanaka (2002) has shown that copper chloride contributes to PCDFs formation by promoting chlorination via catalytic reactions of copper compounds and carbon.

The effects of temperature and the amount of copper chloride on the types and yields of the toxic emissions from combustion of polyvinyl chloride have also been investigated by Wang (2000). The results suggest that a large amount of polycyclic aromatic hydrocarbons are released from combustion of pure PVC, accompanied by a small amount of polychlorinated dibenzo - p - dioxins and polychlorinated dibenzofurans. Copper chloride can effectively enhance the formation of PCDD/Fs while possibly inhibiting the formation of PAHs. Metal chlorides are cited as one of the main factors to enhance the PCDD/Fs formation from combustion of PVC (Wang, 2000).

Other recent studies (Liu, 2002) have shown that the presence of PVC in the incinerator feedstock leads to the improved removal of PAHs from the flue gases emitted by the incinerator. PVC also purportedly assisted in the removal efficiency of SO2 from the flue gases by a spray dryer used as the air pollution control device (Liu, 2002).

6.4.2.2 Combustion Conditions

If an incinerator is running above 950C with adequate air flow it actually destroys dioxins in the main chamber. Even under such conditions dioxins tend to form on carbon and fly-ash particles which contain carbon, chlorine and trace metals during the cooling down phase of the off-gases (Altwicker, 1993).

The de novo synthesis is the main mechanism by which dioxins form in the flue gases upon cooling. This mechanism involves the formation of dioxins from organic and inorganic compounds which are chemically unrelated to PCDD/CDF but combine in pollution control equipment that is operated at temperatures between 200-400C (the so-called de novo temperature range) (Milligan, 1993)

The main operating strategies to reduce dioxin formation are:

PCDD/F levels in flue gases can be reduced by quenching of the flue gases (if energy recovery is not an option) or by activated carbon injection (where energy recovery is performed). Quenching reduces the time that particulates entrained in the flue spend in the temperature range associated with PCDD/CDF formation (ie. the de novo temperature range). In activated carbon injection mode, activated carbon is injected into the flue gas stream prior to the particulate collector. The carbon adsorbs the PCDD/F onto its surface and then the particles are removed by the particulate collector (private communication, Paul Wooten, Brightstar Environmental, 2002).

6.4.2.3 PAH Emissions

Essentially PAHs form during incineration due to an incomplete combustion processes. The carbon monoxide concentration in flue gases has been found to correlate well with the total PAH (R2 > 0.89) and thus can be used as a surrogate indicator for PAH emissions (Li, 2001). Excess amounts of air supply in the incineration of plastic wastes could decrease not only the concentration of the PAHs in the bottom ash but also the emission factor of the total PAH in the stack flue gases.

The concentration and composition of individual polycyclic aromatic hydrocarbons in the raw wastes, flue gas (gas and particle phases) and ash from the incineration of PVC, HDPE and PP in a batch-type, controlled-air incinerator has been reported by Li (2001). HDPE was found to contribute to the highest mean emission of the total PAHs (462.3 mg/kg waste) from the flue gas stack. Incinerating of PVC was found to result in a higher emission of PAHs (195.4 mg/kg waste) to the bottom ash. Moreover when PVC plastic wastes were incinerated, higher-ringed PAHs constituted a larger percentage in the bottom ash as compared to those from PP and HDPE (Li, 2001).

In an interesting study by Gotlib (2001) it was found that total concentration of polyaromatic hydrocarbons (in dioxin equivalents) in plasticized PVC is approximately 2 times higher in the case of a DEHP plasticizer as opposed to an EDOS (a mixture of 1,3 dioxane derivatives) plasticizer. Gotlib and colleagues from Kazan State Technological University in Russia incinerated vinyl flooring samples in a specially constructed unit simulating combustion at temperatures of 600-1100C (Gotlib, 2001). Total trapping of the incineration products was ensured by vacuum collection and absorption by a filter set.

The specific PAHs formed from the combustion of DEHP-plasticized PVC relative to those formed from EDOS-plasticized PVC are shown in the table below:

Table 6.13 PAHs formed from the combustion of DEHP-plasticized PVC
Polyaromatic Hydrocarbon
Times higher Concentration in the DEHP samples
Phenanthrene
2 X
Fluoroanthrene
45 X
Benzoanthracene
100 X
Pyrene
200 X
Anthracene
250 X
Source: Gotlib (2001).

The study by Gotlib (2001) concludes that the EDOS plasticizer (which provides good-quality flexible PVC decorative materials) significantly reduces the environmental toxicity upon incineration of the PVC.

The carcinogenic potential of polycyclic aromatic hydrocarbons is widely recognized. The mutagenicity of combustion products of synthetic polymers such as PVC, PET, PS, and PE have been studied (Lee, 1995). The mutagenicity of the combustion particulates was evaluated with strains of Salmonella typhimurium. The greatest mutagenicity was observed from the extracts of particulates from combustion of PVC followed by that of PS, PET, and PE (Lee, 1995).

6.4.3 MSW Incineration

PVC constitutes a very low component of MSW (0.6-0.8%). Estimates of PVC contribution to chlorine content in MSW range from 38% to 66%. The remainder comes mainly from putrescibles, bleached paper and miscellaneous combustibles. These values vary somewhat depending on the source. For instance the Bertin report (2000) concludes that "Chlorine mainly comes from PVC (49 %), putrescibles (20 %), paper (12 %) and miscellaneous (20 %)".

Table. 6.14. Estimated PVC contribution to chlorine content in MSW (from literature sources)

Study
Contribution of PVC to Cl content in MSW (%)
Rijpkema (1993)
38
Mark (1995)
42.4
Rasmussen (1995)
66
Nieuwenhuysen (1996)
41.4
Reimann, 1999 (from Bernard, 2000)
50

For MSW incineration the primary combustion is generally achieved at about 700C to 900C in the presence of an overall excess of air.

6.4.3.1 Impact of PVC Waste on Solid Waste Streams from MSW Incineration

There are three solid waste streams from MSW incinerators, namely:

The most prolific waste stream is the bottom ash which accounts for more than 90% of the total solid waste output. The fly ash and neutralization residues (collectively known as air pollution control residues - APC residues) account for the remaining 10%. PVC incineration contributes to approximately 10% of the APC residues with negligible contribution to the bottom ash ('Solid Residues from PVC in Municipal Solid Waste Incineration' 2001).

Generation of neutralization residues from PVC is between 0.4-0.92 kg/kg of PVC. This is due to the large degree of hydrochloric acid (HCl) evolution from PVC. The generation of residues from the neutralization of acidic gases such as HCl is highly dependant on the type of gas treatment system fitted to the incinerator, as shown in the following table.

Table 6.15. Residues generated by incineration of PVC waste in kg per kg PVC for various gas treatment systems (ie. scrubber types)
  Dry Dry Semi-dry Wet Semi-wet
Neutralization agent Lime Bicarbonate Lime Lime Lime
Quantity of neutralization residue in kg per kg PVC 0.92 0.51 0.83 0.4 0.57
Source: Kirrmann, 2000.

As is evident from the table the presence of PVC in MSW contributes to APC residues, especially in dry and semi-dry gas treatment systems and thus adds to the overall cost of MSW incineration (Musdalslien, 2000).

6.4.3.2 Metal Speciation and Partitioning

PVC and food residue chloride content in MSW are known to affect the formation and partitioning of metal chlorides in incineration by-products (Wang, 1999). Studies indicate that the heavy metal partitioning behaviour is mainly affected by the presence of chloride, alkaline metals (i.e., Na, K) and moisture in the incoming wastes (Wang, 1999).

The effect of organic chloride (derived from PVC and other possible sources in wastes) on the partitioning and speciation of heavy metals and their compounds has been studied in detail by Wang et al. (1997) using simulated waste spiked with organic chloride. The results indicated that an increase in the organic chloride content and combustion temperature increases the volatility of heavy metals and/or their compounds in the wastes, thus increasing metal partitioning in the fly ash or the flue gases (Wang, 1997).

The effect of PVC on speciation of metallic (Cr, Pb, and Cd) incineration by-products during fluidized-bed incineration of waste has been investigated (Wey, 1999). Chemical analysis indicated that Cr2O3, PbCl2, and CdCl2 were the dominant species of Cr, Pb, and Cd when PVC was present, whereas PbO, Pb2O3, CdO, and CdCO3 were the dominant species under non-PVC operating conditions (Wey, 1999). The former metal salts exhibit a higher potential for water leaching. Chromium oxide was not produced in significant amounts when PVC was absent in the waste stream.

It has also been established that the addition of PVC or NaCl during fluidized bed incineration leads to a shift in the particle size of the ash towards a finer distribution. Further the concentration of lead and cadmium was found to be higher in the finer particles (Wey, 1998).

The variation of speciation and partitioning (i.e. the split between the bottom ash and fly ash) of heavy metals in MSW combustion is affected by the chlorine/metal ratio. The effect of PVC-derived chlorine on heavy metal emissions from MSW incineration has been investigated using simulated MSW spiked with PVCs and heavy metals (Cd, Cr, Cu, and Zn) (Chiang, 2001). The ratio of the chlorine content to that of the heavy metal (Cl/M ratio) was varied and it was found that the emission of heavy metals and/or their compounds, with extreme and medium volatility, tended to be enhanced within the lower Cl/M range. Those with a low volatility were likely to be affected within the higher Cl/M range.

6.4.3.3 Bottom Ash

The contribution of PVC to heavy metals in the bottom ash is shown in the table below. It is clear that PVC contributes significant quantities of both cadmium and lead. These originate primarily from rigid PVC products. It is also evident that cadmium and lead concentrate predominantly in the fly-ash.

Table 6.16. Contribution of PVC to bottom ash and fly ash metal content

Heavy metal in bottom-ash (mg/kg of waste)
MSW
PVC waste
Hg
0.2
0.10
Cd
11
368
Cu
8383
3832
Pb
2474
9222
Sb
108
104
Zn
3447
2373
Heavy metal in fly-ash (mg/kg of waste)
MSW
PVC waste
Hg
1.2
0.60
Cd
371
13008
Cu
6325
2891
Pb
9749
36340
Sb
688
661
Zn
17462
12021
Source: Rijpkema (1999).

The presence of PVC in MSW actually has some positive effects on the quality of the bottom ash as it contributes to reducing the content of heavy metals and total organic carbon in the bottom ash. The reduction of heavy metals is due to the chlorine content of the PVC, while the energy content of the PVC contributes to a better burn-out of the waste thereby decreasing the content of total organic carbon in the bottom ash (Rijpkema, 1999).

During incineration of PVC containing organotin stabilizers, all the tin compounds are transformed to tin dioxide which is found in the ashes and in the dust collected by the flue gas purification equipment. A significant amount of this tin however, originates from products other than PVC (Rijpkema, 1999).

Chen (1999) investigated the adsorption behaviors of Cr, Cu, Pb, and Cd on a sorbent system (e.g. silica sand with limestone) during fluidized-bed incineration with various chloride additive conditions:

The results showed that the NaCl additive generally led to a higher heavy metal adsorption capacity than did a PVC additive or no additive at all. Cadmium was extremely difficult to adsorb and sequester due to the high volatility of cadmium chloride. The efficiency of trapping heavy metals with a sorbent in the presence of PVC during incineration was in the order Cr>Pb>Cu>Cd.

6.4.3.4 Fly Ash

Incineration trials in Germany show that the volatile heavy metals, particularly cadmium, show a significant transfer into the fly ash with increasing chlorine load in the feedstream. The findings from the German Tamara incinerator which uses a grate system showed that with increasing PVC content in the feedstream there is a greater transfer of volatile heavy metals from the bottom ash into the fly ash (i.e. filter ash). This transfer of heavy metals significantly increases the toxicity of the fly ash. While the main components of fly ash from MSW are SiO2 and Al2O3, the concentration of Pb and Cd is about 750 times and 2000 times the natural concentration in the earth's crust (Balg, 1995). Since the fly ash stream from MSW incinerators is already classified as a prescribed waste in Europe, a higher loading of heavy metals does not alter its disposal requirements. On the other hand, the enhanced volatilization of heavy metals from the bottom ash is advantageous since the bottom ash can be utilized in civil engineering applications (Rijpkema, 1999).

6.4.3.5 Neutralization residue

PVC is often targeted as the main source of HCl in incinerator emissions. However there are other sources of chlorine in municipal solid waste requiring the use of pollution control equipment (such as HCl scrubbers) and these will also control any additional emissions from PVC. In other words HCl scrubbers will be required anyway even in the absence of PVC in the waste stream. A more pertinent point may be the increase in quantity of neutralization residues that are created by the presence of PVC in the input stream and the cost to dispose of these.

The chlorine content of PVC places a high demand on the use of alkaline reagents in air pollution control systems on incinerators depending on the type of air pollution control technology employed. Each unit of PVC incinerated requires the same amount of these reagents as up to 70 units of MSW, for the average mix of air pollution abatement systems. This thus increases the amount of residue generated and the cost of disposal (Brown, 2000). In this respect it is important to properly allocate the contribution that PVC makes to the neutralization residues. For example when calcium oxide is added to scrub the HCl that is evolved from PVC, a proportion of the lime will react with sulphur oxides to form calcium sulphate (gypsum). Therefore when calculating cost of disposal of the neutralization residues, the relative contribution made by PVC needs to be properly apportioned.

As discussed the amount of neutralization residue varies widely with the type of incinerator scrubber systems employed (e.g. dry, semi-dry, wet, semi-wet). For example the Neutrec Process will reduce the amount of neutralization residue to be landfilled by 95% compared with dry and semi-dry scrubbers that use lime. The reason for this is that the neutralization residue is a commercial product used as feedstock for soda plants (Bernard, 2000).

With tightening emission limits, wet scrubbing of combustion gases is favoured over dry scrubbing (Hyzik, 2001). Wet scrubbers for the absorption of HCl produce a neutralization residue after evaporation of the water component. The HCl-based residues can either be upgraded to a commercial grade hydrochloric acid or upgraded to calcium chloride (which has widespread use in Europe as a de-icing salt) (Hyzik, 2001).

In Germany, a growing number of incineration plants (nine at present) have a radically different way of operating the wet scrubber. These plants absorb the HCl from the flue gases separately and upgrade this solution to a commercial HCl grade. This technology has the potential to fully eliminate HCl-related residues (Anon., 'Solid Residues from PVC in Municipal Solid Waste Incineration' (2001).

Trials with PVC incineration have been conducted at a modern state-of-the-art incineration plant built in 1999 at Mullverwertungsanlage Rugenberger Damm (MVR) in Germany. Under normal operation chlorine is recovered as 30% HCl at levels of approximately 4000 tonnes per annum. Trials were conducted in order to verify if increased HCl could be recovered from additional PVC without any negative impacts on the incinerator operation. The PVC input stream was increased from the normal 0.7% to 5% and maintained at this level for 4 weeks. The results showed that the HCl production increased by approximately 50% without any adverse change in acid quality. Significantly there was no detrimental impact on emissions with respect to cadmium, lead, PAH, dioxins, PCB and also, there was no observed increase in corrosion. Due to the success of these trials, the MVR incinerator will now incinerate an additional 6500 tpa of shredded PVC waste (Anon., 'Solid Residues from PVC in Municipal Solid Waste Incineration' (2001).

6.4.3.6 Landfilling of Incineration Waste Streams

The contribution of PVC to heavy metals in the output streams of incinerators (ie. bottom ash and scrubber residues) needs to be put in perspective however against the contribution of other waste fractions. For example, though PVC might make a relatively large contribution to the amount of residues, the hazardous waste nature of the residues is mainly determined by heavy metals from other fractions in the MSW (Bernard, 2000).

Data by Bernard (2000) however indicates that PVC incineration is responsible for an increase in the amount of leachate generated from the landfilled residues corresponding to:

In general the incineration of PVC (which has a concentration in MSW of 0.6 to 0.8 %) appears to increase the content of leachable salts (primarily chlorides of Ca, Na and K) by a factor of 2 (Bernard, 2000).

Air pollution control residues (with or without PVC) have to be managed appropriately and placed in hazardous landfills (Bernard, 2000). The influence of PVC on the hazardousness of the gas treatment residues are difficult to quantify due to the lack of reliable data related to leachate compositions and quantities corresponding to different storage scenarios (Bernard, 2000).

There is little data showing that the leaching of trace elements/heavy metals will increase significantly as a result of the incineration of PVC. There is a possibility that the leaching of cadmium may increase due to increased chloride complexation caused by PVC incineration, but no data is available to substantiate this (Bernard, 2000).

The risk of lead and cadmium leaching from APC residues is of more concern than lead and cadmium leaching from PVC in landfills. For this reason the Swedish Environment Protection Authority recommends that APC residues should also be pretreated (stabilized) by cement encapsulation. A high chlorine content in the ash, caused in part by incineration of PVC waste may cause problems where APC residue stabilization is performed using cement or vitrification (Swedish EPA report, 1997).

PVC contributes less than proportionally to the heavy metal content of the APC resides but overproportionally to its volume (in the form of calcium chloride) (Rijpkema, 1999). It is thus clearly established that the PVC in the waste stream contributes significantly to the production of the neutralization residue. Since this neutralization residue is typically landfilled, a number of strategies have been developed to reduce its environmental impact through diversion from landfill (see table below).

Table 6.17. Characteristics of technologies for reduction of the amount of neutralization residues to be landfilled.
Technology Commercial products Status of development % reduction of residue to be landfilled1,2
Production of HCl Conc. HCl (15-30 wt%) 5 plants in operation, 3 under construction (all in Germany) 33% (B)
Neutrec process Feedstock for soda plant -Scrubbers (~20) in operation since 1991-Recycling of residue: 1 industrial pilot plant in operation; more capacity planned 94% (C)95% (D)
Recycling of salt residue Calcium chloride Pilot plant experiments executed 31 to 51% (B)
Recycling of salt residue: Special case Calcium chloride Industrial pilot plant just started operation ~50% (B)~65% (C)~72% (D)
Adjustment of flue gas cleaning system Calcium chloride Concept, composed of excisting technology 51 to 52% (B)
1 The percentage reduction is indicative and is specified relative to flue gas cleaning configurations, with the amounts of solid residues produced as specified in above table. A negative value indicates an increase of the amount to be landfilled.
2 The configurations are indicated by letters: A = wet scrubber with effluent discharge, B = wet scrubber with effluent evaporation, C = semi-dry scrubber and D = dry scrubber
Source: Anon., 'Solid Residues from PVC in Municipal Solid Waste Incineration' (2001).

6.4.3.7 The Australian Situation

No MSW incinerators currently operate in Australia although a number of waste-to-energy installations are at the trial stage (e.g. Visy Gasifier, SWERF). Since any such units considered for installation in Australia would represent state-of-the-art technology the PCDD/CDF emissions are expected to be low.

6.4.4 Incineration of PVC in Medical Waste Incinerators

PVC is widely used in medical applications such as for blood bags, drips, IV solution bags and tubing. Clinical waste containing PVC is generally incinerated or chemically treated. It has been shown that PVC influences the chlorine content of clinical waste more than in the case of MSW; Ranges of 50% to 86% of the chlorine in clinical waste is contributed by PVC while ranges of 38% to 66% of the chlorine in MSW is contributed by PVC (Bernard, 2000).

Due to the relatively high level of PVC in medical and clinical waste, medical waste incinerators are often cited as having high levels of dioxin emissions (Thornton, 1996). PVC the dominant source of organically bound chlorine in the medical waste stream is claimed by some studies to be the primary cause of dioxins produced by the incineration of medical wastes (Thornton, 1996).

In the ASME study by Rigo (1995, 1996) however it is concluded that there is no statistically significant relationship between the composition or amount of PCDD/F concentrations in the gases emitted from medical waste incinerators and the amount of chlorine in the waste infeed. Notably however the authors did not actually measure the chlorine content in the incoming waste stream, instead they assume HCl emitted is proportional to the incoming chlorine level. This is clearly flawed since some of the HCl will be tied up in the solid residue. Furthermore not all the dioxins produced in incineration are emitted in the stack emissions.

Huang and Beukens (1995) report that stack emissions account for less than 12 percent of dioxin output. The major share is distributed amongst the fly ash, bottom ash, and other residues.

The Rigo studies (1995, 1996) did not evaluate the relationship between the quantity of chlorine fed into the incinerator over a specific period of time and the quantity of dioxins that is released in stack gases, fly ash, and other residues during the same period.

Currently almost all the medical plastics employed in renal therapy (e.g. dialysis) are PVC-based and their incineration can lead to releases of PCDD/F during incineration. A recent study by Hoenich (2002) calls for the introduction of PVC-free materials whose incineration is less controversial environmentally for use in renal therapy items.

In another study El-Hamouz (2002) examined the type and quantity of the medical waste generated from four overseas hospitals and two medical centres with daily recordings made over one month. It was found that PVC waste represented the largest fraction of waste and that the dioxins/furans emissions from the medical waste incinerator increased with increasing HCl concentration and decreasing combustion temperature. This study confirms the need to control these pollutants from the medical waste incinerators and to control combustion conditions.

It has been shown that the operating conditions of a medical waste incinerator can have more of an effect on dioxin formation than the relative volume of PVC. In particular bed temperature, residence time and air turbulence are important controlling factors. Thus the influence of PVC on dioxin formation is far less significant than the actual incineration operating conditions (Bernard, 2000).

Also it is important to note that medical incinerators rapidly quench the flue gases thus minimizing dioxin formation in this area. This is because medical incinerators do not capture energy from the combustion process. Municipal waste incinerators on the other hand, generally have co-generation capabilities and therefore the flue gases cannot be rapidly quenched after they leave the combustion zone since their heat energy needs to be first captured. For this reason WTE incinerators allow dioxin formation to some extent and subsequently remove these from the flue gas stream with activated carbon filters (private correspondence Tony Truman, 2002).

Alternative disposal technologies exist for the sterilisation of clinical waste which do not involve incineration. These include mechanical/chemical treatment (hammer mills followed by treatment with disinfectant solution), plasma torch, thermal deactivation, electro-thermal deactivation, autoclaving, microwaving and electron beam sterilisation. Technologies such as autoclaves, plasma disintegrators and microwave-based units are claimed to be more environmentally friendly than incineration (Startech, 2002).

6.4.4.1 The Australian Situation

A number of medical waste incinerators operate in Australia today. There are tight controls on flue gas emissions particularly with respect to PCDD/CDF emissions. The rapid cool down rate of flue gases permitted by medical waste incineration (since there is no energy recovery step) means that dioxin formation via the de novo† synthesis is minimized. The de novo synthesis is the main mechanism by which dioxins form in the flue gases upon cooling (Milligan, 1993).

†The de novo synthesis is the main mechanism by which dioxins form in the flue gases upon cooling. This mechanism involves the formation of dioxins from organic and inorganic compounds which are chemically unrelated to PCDD/CDF but combine in pollution control equipment that is operated at temperatures between 200-400C.

The load of dioxins produced by the incineration of clinical waste in Australia is estimated at 0.9-19 g/year but this figure is regarded as small compared with dioxin emissions from other sources (some 149-2080 g/yr) ('Sources of Dioxins and Furans in Australia: Air Emissions, by Pacific Air, Revised Edition, May 2002.)

Some medical incinerators in NSW have closed down since the mid-1990s and others have been drastically reconfigured to meet the dioxin limit. The dioxin limit in Australia for medical incinerators is 0.1 ng/m3.

6.4.5 Incineration in Cement Kilns

Co-incineration in cement kilns is a viable solution for plastic waste treatment disposal with energy recovery. In the production of cement raw materials consisting of limestone, clay, iron oxides and silica are calcined to form clinker. The clinker is then ground to a fine powder to produce cement. The principal components of Portland cement are tricalcium silicate, dicalcium silicate and tricalcium aluminate.

In a cement kiln the temperature reaches around 1450C and the combustion gases stay above ~1200C for five to six seconds. These conditions are necessary to ensure the resulting clinker quality (Lemarchand, 2000). Thus the rotary cement kiln provides the high temperatures and long burning time needed to completely destroy hazardous wastes. Cement kilns easily achieve a Destruction and Removal Efficiency (DRE) of at least 99.99 percent, as required by US EPA regulations for most organic wastes (http://www.temarry.com/Cement%20Kilns/Cement-Kiln.htm). Cement kilns easily reach this destruction level and are many times more efficient than are other thermal processes and rather than simply destroying wastes, the energy is recovered to make Portland cement.

There are many advantages in using a cement kiln to incinerate waste plastic. These include:

Cement kilns can handle low-level PVC waste quite effectively since as a result of the very large quantity of lime present in the clinker process, the kiln may be said to act as a huge scrubber (Lemarchand, 2000). Because of the neutralisation of acids by the lime, no additional scrubber system is required.

The cement process is not well adapted however to treat plastic wastes containing appreciable levels of PVC. Some cement kilns in France limit the incoming chlorine content to 0.3 to 0.5% in order to prevent plugging of the cyclones due to the crystallisation of chlorinated salts. Other types of cement processes (e.g. the wet process) can tolerate a maximum of 6% chlorine in the feed waste, however operators do not exceed 1% (preferably 0.1 to 0.5%) in order to meet the specification for chlorine content in the manufactured cement (Bernard, 2000). The calcium chloride content can alter the setting and strength properties of Portland cement. In addition, all clinker production facilities include a dust collection system (generally an electrostatic precipitator or baghouse) that collects fugitive dust which contains volatile heavy metals such as lead and cadmium. All collected dust (cement kiln dust) is then recycled and reintroduced into the process so there is no stream for landfill disposal and hence no leachate issues (Lemarchand, 2000). During the process of cement kiln incineration, heavy metals (the majority of which come from waste PVC) are fixed by the clinker into a very stable silicate chemical combination (Lemarchand, 2000).

Moreover one hundred per cent of the calorific value contained in plastic waste can be recovered in cement kilns, whereas waste incineration in conventional combustors is less efficient owing to the use of boilers and turbines (Lemarchand, 2000).

In the Australian context Blue Circle Southern Cement (Geelong, Vic.) do not currently use plastics as an alternative fuel but do use whole tyres as a fuel source. Waste plastic feedstock on the whole are deemed to be to variable and inconsistent.

Queensland Cement (QCL) in Gladstone, Qld. has the capacity to use a range of alternative fuels such as solvents, resins, paint, tyres as well as plastics at a level of 20% substitution of the normal coal feed as a toll service.

6.4.6 Iron Ore Smelting

The effect of chlorine-based additives (ie., PVC and NaCl) on the dioxins emissions from an iron ore smelting process has been studied by Kasai (2001) in Japan. Kawaguchi (2002) examined the presence of trace chlorine in sintering bed and its effect on dioxins concentration in exhaust gas of iron ore sintering in Japan. Various sources of chlorine were investigated such as PVC, KCl, NaCl and CaCl2. It was found that volatile Cl content and organic content are important factors governing dioxins formation. Dioxins concentration in exhaust gas is proportional to Cl content added to iron ore mix (Kawaguchi, 2002). No PVC or other waste plastics are used in iron ore smelting in Australia.

6.4.7 Incineration of Automotive Shredder Residue

As discussed previously automotive shredder residue can contain between 5-14% of PVC compounds. Trouve et. al. (1998) studied the behaviour of some heavy metals during the incineration of ASR and found that the chlorine content of the ASR is the most important parameter influencing the volatility of the studied metals. It was shown that in the absence of chlorine in a system containing several metals, copper, lead and zinc remain in their metallic form or in the form of oxides. However in the presence of a PVC mastic (sealing compound) the lead in the ASR was completely transported into the gas phase (Trouve, 1998). Thus it is apparent that PVC has the potential to increase the volatility of heavy metals during incineration which is undesirable. Furthermore the heavy metals form their corresponding metal chlorides which are generally water soluble and thus leachable.

In Australia the majority of ASR is landfilled. Currently there is no incineration of ASR occurring in Australia.

6.4.8 Gasification

Gasification is the conversion of organic waste into combustible gas (synthesis gas) via controlled heating in an oxygen-starved, reducing atmosphere (Scheirs, 1998). Gasification can be regarded as a highly controlled and disjoint form of incineration.

Many new gasifiers are forms of two-stage combustors (or incinerators). A discussion is warranted to distinguish between these technologies.

The general design for a controlled air combustor is as follows: a close-coupled gasifier is essentially a two-stage incinerator with a fixed bed (fixed grate) burner. In the primary (or lower) chamber a controlled amount of air is used of about 30-40% of full combustion requirements (a sub-stoichiometric level of air). Here the waste is transformed into volatile compounds - the moisture content of the waste is reduced and the volatile components of the waste are volatilized. Due to the low rates of controlled air addition and the correspondingly low velocities of flue gases, the level of solids entrained in the gases exiting the primary chamber is minimized. Low air velocity is critical for maintaining low particulate carry over. The hot gases then flow to the secondary (oxidizer) chamber where they undergo full burn-out with excess combustion air. Subsequently, energy is captured from the hot flue gases (Cordell, 2000). A low air velocity is critical for maintaining low particulate carry-over. The characterizing feature of this system is that there is no cooling or treatment of the intermediate hot gases (ie. those gases exiting the first chamber and entering the second chamber). Thus there is no provision for scrubbing the HCl from the gases before the second combustion chamber. This places severe chlorine limitations on incoming waste stream.

In contrast, a steam gasifier uses no additional oxygen or air injection at any stage of the process. In the first stage the waste is essentially pyrolyzed in the absence of oxygen to give hot pyrolytic gases. These gases are then passed to the steam gasifier where high pressure steam is injected to convert the carbon-based waste to syngas (C + H2O --> H2 + CO). The syngas is then passed through a water scrubber (typically a high-velocity venturi scrubber) to remove HCl. The gases are then pressurized in an accumulator and combusted in a reciprocating engine to produce electricity. The reciprocating engine is an efficient thermal oxidizer and allows complete combustion of the syngas.

An exergy efficiency study of PVC waste gasification and PVC waste incineration has been conducted (van Schijndel, 1998). Exergy is a measure of the quality of mass and energy streams. It was found that the exergetic efficiency of PVC waste gasification is 60% higher than for PVC burning (see table below). The main reasons for the high efficiency of the gasification process are the use of a high temperature gas-turbine and the controlled gasification process compared to the chaotic PVC burning process. Some small optimisation calculations showed that the gasification process has potential for further optimisation.

Table 6.18. Burning versus gasification efficiencies
Process Efficiency (%) (with 10% heat loss) Efficiency (%) (with zero heat loss)
PVC incineration 29 32
PVC gasification 49 50
Data from van Schijndel (1998).
Both processes produce electricity and a 30% HCl-stream

Here in Australia a close-coupled gasifier has been installed by Visy at Coolaroo (VIC) for paper and plastic wastes. This installation is still being optimized. The process has a chlorine limit of just 60 ppm.

A steam gasifier has been installed as part of the Solid Waste Energy Recovery Facility(SWERF) in Woolongong (NSW) by Energy Developments Ltd. (Brightstar). Since the SWERF process has a water scrubber between the gasifier and the energy recovery step it has the capability to handle large quantities of PVC in the incoming feed stream without adverse consequences. Nevertheless Energy Developments Ltd. are still to conduct a PVC spiking trial under independent supervision in accordance with EPA recommendations. In this trial up to 5% PVC in the infeed (1 tonne of PVC in 20 tonnes of MSW) will be evaluated (private communication, Rick Ralph, Brightstar Environmental, 2002).

6.4.9 Pyrolysis

Pyrolysis of waste plastics involves heating them under vacuum or inert atmosphere to temperatures in the range 350-450C. Under these conditions the plastic molecules crack to give shorter chain hydrocarbons resembling oil, diesel, kerosene and paraffins. These products can be used as fuels directly or as chemical feedstocks in the petrochemical industry. The presence of PVC in the feedstock stream for pyrolysis processes is undesirable as HCl is evolved and chlorinated by-products can be formed.

The formation of PCDDs/PCDFs during PVC pyrolysis has been studied by McNeill et al. (1998). The tars produced by PVC pyrolysis under different conditions of temperature and oxygen concentration were analysed for PCDDs and PCDFs content. The tar collected after low temperature pyrolysis (500C) and in reduced oxygen atmosphere (11.6% O2) had the highest PCDDs/PCDFs content. The toxicity of the tars is mainly due to the high PCDF concentration. The most toxic compound, 2,3,7,8-TCDD, had a contribution of only 0.5-4.4% of the total toxicity of the tar (McNeill, 1998). Thus pyrolysis and combustion conditions play a critical role in the extent of dioxin formation.

Plastic pyrolysis installations are presently being considered in Australia. For example, Ozmotech (Dandenong, Vic.) have developed a low temperature pyrolysis process to convert waste plastics into diesel fuel. This process excludes all oxygen and no measurable dioxins are produced.

6.4.10 Impact of PVC on Waste to Energy (WTE) Processes


6.4.10.1 Chlorine Limitations

An important distinction between thermal treatment processes for waste is their chlorine limitation (see table below).

Table 6.19. Chlorine Limitations of various Waste-to-Energy Processes
Process Upper Limit of PVC
MSW incinerator Limited by APC system*
Clinical waste incinerator Limited by APC system*
Pyrolysis 1-2%
Visy 'Gasification' 0.006%
SWERF steam gasification No limit
Cement kilns <1%**
* Specific chlorine limitation determined by design of flue gas cleaning system
**1% chlorine (preferably 0.1 to 0.5%) not to be exceeded in order to meet the specification for Cl content in some manufactured cement.
6.4.10.2 Low Energy Recovery

Most waste-to-energy processes for waste plastics are limited by the extent of PVC that they can tolerate since PVC has a chlorine content of 57% and is thus effectively an inorganic polymer. Furthermore since rigid PVC only has a hydrocarbon content of 43% it is less well suited to feedstock recycling or other waste-to-energy forms of recycling as it contains less recoverable energy - some 56% less than say polyolefins (20 MJ versus 46 MJ). Fillers and additives may reduce the hydrocarbon content of the PVC even further. It can nevertheless still be used in feedstock recycling or waste-to-energy processes providing appropriate counter measures are implemented.

6.4.10.3 Corrosion and Low Steam Parameters

Problems with PVC in waste-to-energy systems are related mainly to corrosion. In order to minimize corrosion low steam parameters must be used and this in turn results in thermal efficiencies of the order of 20% (Zevenhoven, 2000). The presence of HCl leads to corrosion and pitting of construction metals. HCl attack is exacerbated by increasing temperature and pressures. However the hotter the steam temperature of a waste-to-energy system, the higher the efficiency of the process. Hence the presence of PVC necessitates the need to strike a balance between minimizing pitting and maximizing steam pressure and efficiency (private correspondence, Paul Wooten, Brightstar Environmental, 2002).

6.4.10.4 APC Residues

As discussed previously the presence of PVC during WTE conversion leads to the increased production of air pollution control residues. There are potential environmental impacts associated with leaching of these metal chlorides.

6.4.11 Approaches to mitigate the environmental impact of PVC in WTE Processes


6.4.11.1 Dehydrochlorination

PVC behaves differently to most other commodity plastics in that it decomposes quantitatively to hydrochloric acid at temperatures in the range 250-400C. What remains is a hydrocarbon coke-like residue that has good fuel value. The carbonaceous residue can be combusted as any other chlorine-free solid waste-derived fuel (Zevenhoven, 2000). Given that the chlorine content of PVC can be thermally removed, many waste-to-energy processes for waste plastics incorporate a pretreatment step whereby the PVC fraction is dehydrochlorinated by heat treatment above 300C. The other advantage that pretreatment offers is that the HCl that is evolved can be recovered and used commercially.

Dehydrochlorination of plastic mixtures containing significant levels of PVC is an essential reaction step in waste incineration, pyrolysis and chemical recycling of polymers (Bockhorn, 1999). The difficulty with dehydrochlorination of PVC in mixed plastic waste is that the temperature must be such that dehydrochlorination of PVC is maximized and pyrolysis of the other polymers (e.g. polyolefins) as well as the dehydrochlorinated PVC itself is suppressed. In other words an optimum dehydrochlorination temperature must be used that removes the HCl but at the same time preserves the majority of the hydrocarbon material for the subsequent thermal process.

In the vacuum pyrolysis of mixed-plastics waste, PVC is generally problematic for the pyrolysis process (de Marco, 2002). Therefore efforts are largely made to mitigate the chlorine effect either by a dehydrochlorination step or by neutralization with an alkali earth hydroxide or carbonate (e.g. slaked lime).

Zevenhoven et al. (2002) studied the pyrolysis of PVC as well as mixtures of PVC with wood (Finnish pine) and low density polyethylene in nitrogen at 250-400C. The study was intended to establish the optimum temperature range for producing low-chlorine or chlorine-free fuel in a dehydrochlorination reactor without pyrolysing any of the other combustible fractions. It was found that the PVC can be effectively dehydrochlorinated at approximately 350C and that secondary pyrolysis is suppressed when LDPE is present.

Some systems rely on stepwise pyrolysis where the PVC fraction is first dehydrochlorinated and then the hydrocarbon polymers are pyrolyzed (Bockhorn, 1999). The advantage of these systems is that other plastic components in the solid waste stream do not decompose during the PVC dehydrochlorination phase. In this way high levels of PVC can be tolerated in waste plastics pyrolysis. Stepwise pyrolysis involves slowly ramping up the pyrolysis temperature and holding the pyrolysis reactor isothermally at various temperatures to separate different chemical reactions. In this way PVC in a mixture of waste plastics can be dehydrochlorinated at moderate temperatures and prior to the thermal degradation of the polymer backbone.

In stepwise low temperature pyrolysis, mixtures of for example PVC, polystyrene and polyethylene can be separated into HCl, styrene monomer and aliphatic compounds from polyethylene decomposition respectively. The degree of conversion of chlorine from PVC into hydrogen chloride in the low temperature (330C) first step is about 99.6% (Bockhorn, 1999).

Dehydrochlorination of mixed plastic from MSW is currently being undertaken in Europe (the Redop project). This project is investigating the use of mixed plastics with limited chlorine content (0.5 to 5.0 wt.% chlorine) as a reducing agent in blast furnaces. The process begins with dehydrochlorination and granulation of the waste. The granules are then injected into a steel-making blast furnace as a substitute for coke (Vinyl 2010, The Voluntary Commitment of the PVC Industry Progress Report 2002, ECVM, Brussels.). The technical feasibility of this process which is being managed by DSM Research was demonstrated in 2001.

The generation of HCl during the incineration of PVC can also be controlled at low levels by means of mixing calcium carbonate (CaCO3) and lithium carbonate (Li2CO3) with PVC (Nakazawa, 2001). It has been shown that the capturing effect of HCl with CaCO3 and Li2CO3 combined was superior to the single compound. The sample with CaCO3 : Li2CO3 = 3 : 1 (mole ratio) resulted in the highest capturing effect of HCl in this experiment. Furthermore the addition of a small amount of FeO(OH) was found to be synergistic in increasing the capturing effect of HCl by catalyzing HCl formation (Nakazawa, 2001).

PVC can also be dehydrochlorinated to well below 0.1% chlorine by heating to 250C in the presence of a strong alkali such as sodium hydroxide. This is the basis of the Stigsnaes hydrolysis process currently operating in Denmark. This process involves two distinct and separate steps. First hydrolysis at 250C of PVC waste products occurs in the presence of caustic soda, yielding sodium chloride and a dechlorinated fraction. The sodium chloride is purified so that it can be discharged directly to the sea without any risk to the environment. Secondly pyrolysis of the dechlorinated fraction is performed to produce a liquid organic phase and a solid residue containing the inorganic components of the waste. The organic phase can be used as feedstock for petrochemical processes or for energy recovery (Vinyl 2010, The Voluntary Commitment of the PVC Industry Progress Report 2002, ECVM, Brussels.).

Another strategy for dealing with HCl is not to neutralize it but rather recover it for industrial use. HCl removed during the incineration process can be captured for recovery using a two-stage process. A study has been undertaken comparing a two-stage combustion process for high-PVC solid waste with HCl recovery at a conventional solid waste incineration plant (Saeed, 2002). The two-stage process has a theoretical thermal efficiency of approximately 36.8% depending on the pyrolysis temperature, PVC content in the solid waste, and moisture content. HCl recovery achieved was above 90% at pyrolysis temperatures above 310C combined with low HCl flue emissions. In comparison the conventional waste incineration plant was calculated to have a thermal efficiency of 33.4% with no HCl recovery and requiring significant flue gas neutralization (Saeed, 2002). In a commercial example the Mullverwertungsanlage Rugenberger Damm incineration plant in Germany recovers the chlorine content of the waste as 30% HCl at levels of approximately 4000 tpa (Anon., 'Solid Residues from PVC in Municipal Solid Waste Incineration' (2001).

6.4.11.2 Slaked Lime Addition

An alternative strategy to a separate dehydrochlorination step is a
HCl removal process based on calcium hydroxide (slaked lime) injection. This is essentially a form of in situ scrubbing in the pyrolysis reactor itself. This method is used by a number of pyrolysis processes since the fuel oil end products need to be low in chlorine to minimize engine corrosion problems.

It has been demonstrated that oil recovered from PVC can be used as a fuel. Cost is the main obstacle since it requires a large amount of slaked lime to neutralize the HCl gas that is produced by the thermal cracking. The table below shows the effect that slaked lime has on reducing the chlorine content of fuel oil derived from the pyrolysis of PE (55wt%), PP (28wt%) & PS (17wt%) at a decomposition temperature of 420C.

Table 6.20. Effect of slaked lime addition on the pyrolysis of plastic containing 1% of PVC

Slacked lime added (%)
Total chlorine in oil (mg/L)
Total hydrochloric acid in oil (mg/L)
0.0
1139.3
490
1.2
72.5
26
2.0
54.2
20
Source: Masataka Tsukada (Environment Systems Co. Ltd., Japan)

As shown in the table above, adding 1.2% of slaked lime enables the desired level of chlorine in recovered oil to be achieved thereby allowing the oil to be suitable as a fuel. However some chlorine still remains. There is always a small percentage of moisture excisting in the waste plastic. It is later separated out and condensed as water. This water reacts with chlorine species in the fuel and produces hydrochloric acid.

The dehydrochlorination of refuse-derived fuels (RDF) containing PVC as the chlorine source has been evaluated by Naruse (2001). The dehydrochlorination efficiency was calculated by analysis of Cl in the residue after combustion. HCl was found to be emitted mainly during the volatile matter combustion. For RDF with PVC and calcium hydroxide addition, the dehydrochlorination reaction depends on the furnace temperature with the optimum temperature being about 650 deg.C. At lower temperatures the rate of HCl evolution is faster than the dehydrochlorination reaction with the calcium hydroxide, whereas at higher temperature the calcium chloride which is produced by the dehydrochlorination reaction is decomposed during the char combustion (Naruse, 2001).

The NKT-Watech pyrolysis process in Europe involves a two-step pyrolysis of PVC wastes in a stirred vessel. The HCl liberated at 220C, reacts with filler and added lime/calcium carbonate to form calcium chloride. Then as the temperature is raised above 350C the polymer chains break down. Light organic material escapes leaving a solid coke residue. The residual calcium chloride is treated to make it suitable for selling (Vinyl 2010 document).

The dehydrochlorination capacity of calcium hydroxide for HCl formed from refuse-derived fuels containing PVC and NaCl in a reactor with two reaction zones (pyrolysis and combustion) was studied by Wang (2002). High temperatures did not favour the dehydrochlorination capacity of calcium hydroxide since it was shown to have poor thermal stability at high temperature. The results by Wang (2002) demonstrate that calcium hydroxide is not optimal for scavenging HCl during combustion processes.

6.4.12 Emerging Technologies

A new waste-to-energy process suited for high-PVC solid wastes has been developed in which the incoming waste is heated to between 200-400C in a fluidized bed reactor. The PVC fraction decomposes into HCl and a coke-like residue that can be combusted as any other chlorine-free solid waste-derived fuel (Zevenhoven, 2000). The process is composed of two fluidised bed reactors and a heat recovery system. The output products of the process are electricity and recovered HCl. A thermal efficiency analysis using PROSIM software showed that efficiencies of ~36% (for electricity generation) can be reached, depending on pyrolysis temperature and the PVC content in the solid waste. HCl recovery can be in excess of 90% for pyrolysis temperatures above 310C, combined with low HCl emissions from the combustor that may be below legislative emission limits even without gas cleaning (Zevenhoven, 2000).

Slapak et. al. (2000) describe a recycling process for PVC waste using steam gasification in a bubbling fluidized bed reactor. The main products of the process are syngas (a mixture of hydrogen and carbon monoxide) utilizable for energy recovery as well as HCl that can be reused for PVC production in an oxychlorination plant. The authors demonstrate the technical and economical feasibility of this process. The design capacity is 50 kton waste/year. The syngas that is produced proves to have a heating value of 8.6 MJ/Nm3. Half of the chlorine present in PVC waste is converted to pure HCl gas. The rest leaves the process as a CaCl2 waste stream due to the presence of calcium-based additives (added as filler material to many PVC compounds).

Slag bath gasification of PVC (the Tavaux Project) is a new technology currently under development. The core aspect of this process is the reactor, where decomposition of the PVC waste takes place in a molten slag bath at 1400 - 1600C. HCl and syngas are the products targeted for recovery. A pilot plant based on this technology has been built in Tavaux (France) by Linde in Germany (Vinyl 2010 document).

Jaksland et. al. (2000) present a new environmentally sustainable technology for chemical recycling of PVC waste whereby the PVC is subjected to combined thermal and chemical degradation in a reactor. The HCl evolved from the PVC reacts with fillers producing calcium chloride. The metal stabilizers (lead, cadmium, zinc and/or barium) are converted to the corresponding metal chlorides. Metals and calcium chloride are sequentially extracted from the reaction medium by exploiting the influence of pH, temperature, liquid to solid ratio and particle size on the metal solubility. Separation occurs in a downstream multi-stage extraction-filtration procedure. The products from the process are: (1) calcium chloride which satisfies the specifications as 'thaw-salt' (de-icing salt), (2) lead product which may be further purified and re-used, (3) coke and (4) an organic condensate (fuel oil), which may be used as an energy source for the process (Jaksland, 2000).

In Japan, a dehydrochlorination process to utilize PVC in ironmaking has been developed in a 1,000 tpa semi-commercial facility with a newly-designed indirect heating rotary kiln (Asanuma, 2001). In this process coarse coke is added to the PVC waste in order to prevent agglomeration of PVC particles and adhesion to the inner wall of the kiln during PVC dehydrochlorination. A dehydrochlorination efficiency of over 95% was claimed at a processing temperature of 325C and a residence time of 30 minutes. In addition to HCl, C1-C4 hydrocarbons, benzene and tar were generated by PVC pyrolysis. The supply of coarse coke with PVC was also found to be effective in promoting heat transfer to the PVC layer in the kiln. From results obtained in the 1,000 tpy semi-commercial plant, it was verified that this dehydrochlorination process could be continuously operated due to the process advantages afforded by the coke particles. It was concluded that this process is a viable means for the feedstock recycling of significant quantities of waste PVC (Asanuma, 2001).

In summary the precise conditions employed in waste-to-energy processes involving feedstock containing PVC have a large bearing on the types of emissions and residues. It must be stressed that operating conditions during incineration are very important in determining the extent of dioxin emissions. Numerous technological developments are taking place in the WTE area and PVC can increasingly be accepted and dealt with by these types of emerging processes.

6.5 Fires


6.5.1 PCDD/F Emissions from Uncontrolled, Domestic Waste Burning

The formation of PCDD/Fs from the uncontrolled burning of dosmetic waste has been studied by Gullet (2000) and Lemieux (2000). Lemieux et. al. (2000) varied the quantity of PVC in domestic waste and then burned it in barrels. As the table below shows significant levels of dioxin were emitted.

Table 6.21 Dioxin and furans levels from the uncontrolled burning of PVC
%PVC
TEQ (ng/kg)
Standard deviation(ng/kg)
Total PCDD/F(ng/kg)
Standard deviation(ng/kg)
0
14
18
1549
1758
1
201
43
11518
817
7.5
4916
2146
336642
125306
Source: Lemieux (2000).

These results suggest that backyard burning of domestic waste could be a major source of PCDD/F emissions. Tests with 0.2 % PVC (considered base runs) illustrating the normal content of PVC in domestic waste gave 80 TEQ/kg.

It has been argued however that PVC has only a secondary effect in that it lowers the burning temperature and/or augments the residence time, which results in higher dioxin releases (Engelbeen, 2002). Part of the confusion surrounding the precise role that PVC plays in dioxin formation during uncontrolled burning is due to the influence of other controlling factors that have an impact on dioxin formation.

It has already been established that combustion conditions have a major influence on the generation of dioxins thus uncontrolled burning of waste can be a major source of dioxins regardless of the source of chlorine. Factors which increase dioxin formation in the presence of PVC include:

Thus it is apparent that burning of waste which is wet, compacted and contains PVC or alternative chlorine sources, will lead to considerable dioxin formation. These are generally the conditions experienced in backyard barrel burning and small scale inadvertent fires. To complicate matters, the presence of PVC or alternative chlorine sources can affect some of these factors to the extent of depressing the average temperature and extending residence times.

Uncontrolled burning of domestic waste is not widely practised in Australia and should continue to be discouraged.

In the case of accidental fires of houses and warehouses, incineration conditions in general are better (that is higher temperatures, less compaction, less moisture) and therefore dioxin formation in the presence of PVC is expected to be less (Engelbeen, 2002). In uncontrolled fires consideration should also be given to the emissions of polyaromatic hydrocarbons the formation of which may be suppressed in fires containing high levels of PVC. Moreover the carcinogenic potential of PAHs is higher than that of TCDD in some strains of rats (Engelbeen, 2002).

6.5.2 Uncontrolled Large Scale Fires (Accidental fires)

PVC in uncontrolled fires can lead to high emissions of polychlorinated dioxins and polychlorinated dibenzofurans into the surrounding environment due to the generaly large scale of these fires. The presence of chlorine is necessary for the forming of dioxins, with PVC being a significant source of chlorine. The chlorine content of wood is just 0.028% compared to 57% for PVC, so wood used in construction can be viewed as a relatively insignificant source of dioxins on a comparative weight basis. However when a building burns the ratio of dioxin contributed from PVC and wood respectively is close to 1:1 due to the preponderance of wood in most houses and dwellings. Uncontrolled fires such as those that occur in accidental building fires have high emissions comparable to the emissions of controlled incineration of combustibles. In uncontrolled fires there exist "poor" combustion conditions, leading to incomplete combustion (Engelbeen, 2002).

It should be noted that there are other important products of incomplete combustion produced by accidental fires besides dioxins. In all large-scale accidental fires and conflagrations the local conditions are such that large quantities of soot are also produced, which in turn contains high levels of polyaromatic hydrocarbons. In general the main carcinogenic component of soot from accidental fires is benzo(a)pyrene (BaP), which can be found at 10,000 times higher levels than TEQ-dioxin (Engelbeen, 1998). Even in fires where large quantities of PVC are involved, the BaP to I-TEQ dioxin ratio is still above 1,000 (Engelbeen, 1998).

Toxic chlorinated hydrocarbons (polychlorinated biphenyls, benzenes and dioxins and furans) and polyaromatic hydrocarbons were examined in combustion gases and deposited soot wipe samples from simulated house fires. The concentrations of these substances were high during the fires. The amounts of polychlorinated dioxins and furans (PCDD/Fs) in the combustion gas varied from 1.0 to >7.2 ng/m3 in International Toxic Equivalent (I-TEQ) values and those of polyaromatic hydrocarbons from 6.4 to 470 mg/m3. Thus it is established that large amounts of toxic organic compounds are released in house fires (Ruokojarvi, 2000).

The formation of dibenzo-p-dioxins during the uncontrolled combustion of PVC construction materials in house fires has been reviewed by Carroll (2001). The dioxin emissions were estimated on the basis of standardized amounts of wood (21,000 kg) and PVC (180 kg) in a new house of standard construction and using previously published emission factors for wood (up to 173 g/t) and PVC (up to 6554 g/t). The overall emissions from combustion of these two construction materials in house fires on an as-used basis is estimated to be similar (of the order of 1 g/yr respectively) and a small contributor to the PCDD/F sources inventory in the US (Carroll, 2001).