Section 6. Case Studies
6 CASE STUDIES
(a) Platinum Group Metals in Urban Environment
(b) Sustainable Urban Drainage
(c) Artisanal activities in Vicenza, Northern Italy
(d) Pharmaceuticals in the Urban Environment
(e) Personal Care Products, Fragrances in Urban Waste Water and Sewage
Sludge
(f) Surfactants in Urban Wastewaters and Sewage Sludge
(g) Use of Polyelectrolytes; The Acrylamide Monomer in Water Treatment
(h) Landfill leachate
(i) Potentially Toxic Elements (PTE) transfers to Sewage Sludge
(j) Effect of Chemical Phosphate Removal on PTE Content in Sludge
113 Section 6. Case Studies
(a) Platinum Group Metals in the Urban Environment
Introduction
The platinum group of metals (PGMs), sometimes referred to as the platinum group elements (PGEs), comprise the rare metals platinum (Pt), palladium (Pd), rhodium (Rh), ruthenium (Ru), iridium (Ir) and osmium (Os) and are naturally present in a few parts per billion (mg/kg) in the earth’s crust. The elements are noble chemically unreactive metals, and are found in nature as native alloys, consisting mainly of platinum.
Recently these metals have gained importance as industrial catalysts including vehicle exhaust catalysts (VECs). This use and possible implications for human health were the subject of an earlier review undertaken by Imperial College, London for the UK Department of the Environment (Farago et al, 1995; 1996).
Increasing understanding of the environmental damage of vehicle emissions has led to the introduction of stringent emission control standards throughout the western world. Since 1974 all new cars imported or produced in the United States have had catalytic convertors fitted, cutting down hydrocarbon and carbon monoxide emissions. In 1977 they were fitted to a substantial proportion of all cars sold in America, where at the time, this application accounted for 32% of the total Pt usage (Herbert, et al., 1980).
Vehicle exhaust catalysts have also been used in Japan since 1974. Vehicle exhaust catalysts were also introduced in Germany in 1985, in Australia in 1986, and into the UK at the beginning of 1993 in response to the emission standards equivalent to the US standards which were introduced in the EC at that time. Other uses of PGMs are noted in later sections.
Sources
The PGMs are found in nickel, copper and iron sulphide seams (Bradford, 1988). They are currently mined in South Africa, Siberia and Sudbury, Ontario. World mine production of the PGMs, of which 40-50% is platinum, has steadily increased since 1970. This reflects the increasing world-wide use of PGM vehicle catalysts (IPCS, 1991). From 1988-1992 world mine production was essentially constant at around 255 tonnes per year (WMS, 1994). The amount of PGMs present in the earth’s crust down to a depth of 5km, and hence technologically attainable, are still enormous when compared with present requirements, but only a fraction of the pertinent ores is sufficiently rich for commercial exploitation. Of the total of 3x1011 tonnes of PGMs in the earth’s crust, 3x103 tonnes have been mined, and 7x1010 tonnes are minable (Renner and Schmuckler, 1991).
The total worldwide supply of Pt for 1999 and 2000 was 138 tonnes and 153 tonnes respectively for Pd 230 tonnes and 224 tonnes respectively, and for Rh 14.2 tonnes and 20.9 tonnes respectively (Johnson Matthey, 2000)
Uses of platinum group metals.
By far the greatest use of PGMs both in Europe and worldwide is in vehicle catalysts, with additional major uses in the chemical industry, electrical and electronics industries, petroleum industry, the manufacture of jewellery, as a cancer treating drug in medicine, as alloys in dentistry and in the glass industry.
Demands by application for 1999 and 2000 for PGMs are shown in Table a.1. (Johnson Matthey, 2000)
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TABLE a.1 Platinum Group Metals Demand by Application (Worldwide)
Application kg 1999 (kg) 2000 (kg) PLATINUM Autocatalysts: gross 45600 51000 Autocatalysts: recovery -12000 -13000 Jewellery 79400 83300 Industrial 38400 41400 Investment 5100 -1420 Total Demand (Pt) 159000 146000 PALLADIUM Autocatalysts: gross 166700 146000 Autocatalysts: recovery -5530 -6520 Dental 31500 24700 Electronics 56100 58600 Other 16600 15000 Total Demand (Pd) 265000 238000 RHODIUM Autocatalysts: gross 14400 16000 Autocatalysts: recovery -1870 -2240 Chemical 964 992 Electronics 170 170 Glass 851 1050 Other 312 312 Total Demand (Rh) 14900 16200 RUTHENIUM Chemical 2440 1930 Electrochemical 2040 2270 Electronics 5560 6580 Other 1160 1360 Total Demand (Ru) 11200 12100 IRIDIUM Automotive 964 397 Chemical 198 170 Electrochemical 794 680 Other 936 1450 Total Demand (Ir) 2890 2690
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Trends over time in platinum and palladium uses by application for Europe are shown in Table a.2 and a.3.
TABLE a.2 Platinum demand by application in Europe (kg) Platinum demand 1992 (kg) 1994 (kg) 1996 (kg) 1998 (kg) 2000 (kg) Autocatalyst: gross 16300 17200 14600 15500 17900 Autocatalyst:recovery -142 -284 -567 -851 -1130 Chemical 1420 1420 1700 1700 2410 Electrical 851 709 709 1280 2270 Glass 425 851 1130 709 709 Investment:small 992 1276 142 142 0 Jewellery 2410 2840 3540 4540 5670 Petroleum 567 709 425 425 284 Other 1560 1840 2130 2410 2840 Totals (Pt) 24400 26500 23800 25800 30900
TABLE a.3 Palladium demand by application in Europe (kg) Palladium demand 1992 (kg) 1994 (kg) 1996 (kg) 1998 (kg) 2000 (kg) Autocatalyst: gross 1130 7370 24400 38800 51600 Autocatalyst:recovery 0 0 142 -142 -425 Chemical 2100 1700 1840 1840 2690 Dental 8500 7230 7230 5950 3120 Electronics 5950 7230 8500 7660 7370 Jewellery 992 851 851 1420 1280 Other 425 709 567 709 567 Totals (Pd) 19100 25100 43200 56300 66200
Of particular interest is the increased demand for palladium in Europe, largely in response to the introduction of Euro Stage III legislation from January 2000; palladium – rich catalysts will meet stricter emission limits for petrol models, resulting in a further move away from platinum technology (Johnson Matthey 2000).
Catalytic convertors
A catalytic converter is a unit about the size of a small silencer that fits into the exhaust system of a car. The metal catalyst is supported on a ceramic honeycomb monolith and housed in a stainless steel box similar in shape to that of a conventional silencer. About 1-3g of PGM is contained in some vehicle exhaust catalysts, approximately 50g of PGM per cubic foot of catalyst (Steger, 1994). Due to the commercial sensitivity of these products it is difficult to obtain data on the exact amounts in each of the many different formulations of catalyst. The honeycomb made of cordierite contains 300 to 400 square channels per square inch (6.45cm2), and is coated with an activated high surface area alumina layer called the washcoat (Farruato, 1992) containing small amounts of the precious metals, platinum, palladium and rhodium in varying proportions. The conventional three-way catalysts typically contain 0.08% platinum, 0.04% palladium and 0.005-007% rhodium (Hoffman, 1989).
These metals convert over 90 percent of carbon monoxide (CO), hydrocarbons (HC) and nitrous oxides (NOx) into carbon dioxide (CO2), water (H 2 O) and nitrogen (N2). Platinum is an effective oxidation catalyst for carbon monoxide and hydrocarbons, but it is more sensitive to poisoning than palladium and so can only be used in cars which use unleaded petrol. Palladium is becoming increasingly used instead of platinum due to the higher costs of the latter. The rhodium oxidises the hydrocarbons and reduces the NOx emissions. Base
116 Section 6. Case Studies metals are also incorporated, cerium being the most frequently used; others include calcium, strontium, barium and iron.
Chemical fingerprinting of ground autocatalyst materials has been undertaken by laser ablation and analysis by ICP-MS for 31 elements (Rauch et al, 2000). Variations in composition were found to occur in agreement with the known fact that variations occur from one manufacturer to another and from one year to another. An association between PGMs and Ce in road sediments was ascribed to the emission of PGMs as abraded washcoat particles onto which PGMs are bound and of which Ce is a major component.
Recycling
Of the total platinum consumption in the United States, approximately thirty per cent is accounted for by vehicle catalysts (IPCS, 1991). The recovery of spent autocatalysts from vehicles at the end of their lives is regarded as important and substantial secondary sources of platinum as well as palladium and rhodium (Torma and Gundiler, 1989). The quantity of spent autocatalysts greatly increased in the United States from 1984 to 1988. These scrapped autocatalysts present an important secondary source of the platinum group metals.
On current projections it is expected that 3.5 million catalysts will be available for recycling in the UK by 2000.
Platinum group metals in the environment.
The average concentration of platinum group metals in the lithosphere is estimated to be in the region of 0.001-0.005 mg.kg-1 for Pt, 0.015 mg.kg-1 for Pd, 0.0001 mg.kg-1 for Rh, 0.0001 mg.kg-1 for Ru, 0.005 mg.kg-1 for Os and 0.001 mg.kg-1 for Ir (Greenwood and Earnshaw, 1984).
Although a rapid increase in Pd in sediments from the Palace Moat, Tokyo, Japan was reported by Lee (1983) between 1948 and 1973, it seems unlikely that this was connected with car catalytic convertors since there were few in use in Tokyo by 1973.
Concentrations of Pt and Pd in Boston Harbour have been investigated to evaluate Pt and Pd accumulation and behaviour in urban coastal sediments (Tuit et al, 2000). Increased levels of both metal of approximately 5 times above background concentrations were ascribed to anthropogenic activity with catalytic convertors a major source. It was concluded that anthropogenic enrichments can significantly influence coastal marine inventories of PGMs. The study also indicated that Pt associated with catalytic convertors is much more soluble than expected or alternatively that there is an additional source of dissolved Pt to the harbour. Further study of the biogeochemical behaviour of Pt and Pd was recommended.
Urban pollution with PGMs from catalytic convertors
Emissions of PGMs arise as a result of deterioration of the catalytic convertors, mainly due to thermal or mechanical strain and acid fume components, and are intensified by unfavourable operational conditions (misfiring, excessive heating) which may even destroy the converter (Schäfer and Puchelt, 1998).
Emission rates range between several ng and mg of Pt per km driven depending on whether they were measured in motor experiments or calculated on the basis of environmental concentrations (König et al., 1992). Platinum is mainly emitted as a metal or an oxide with
117 Section 6. Case Studies particle sizes in the nm range, bound to small articles of washcoat material (Schlögl et al., 1987).
Several workers have reported accumulation of Pt, Rh and Pd in road dusts and soils (Zereini et al, 1993; Schäfer et al 1995; Farago et al 1996; Heinrich et al, 1996; Schäfer et al, 1996).
Mostly inert under atmospheric conditions, the reactivity of Pt increases significantly if these nanoparticles are brought into contact with soil components. Lustig et al., (1996) demonstrated that humic substances considerably enhance the reactivity of Pt clusters in the nm-range under atmospheric conditions. Using road-dust from a tunnel, it was demonstrated that within hours Pt can be fixed to several humic acids with different molecular weights. The low solubility of Pt in deionized water increases significantly even under reducing conditions when certain anions or complexing agents are used (Nachtigall et al., 1996).
A detailed study has been undertaken in several sites in southwest Germany, selected on the basis of traffic density and morphology, including roads in Stuttgart with 120,000 vehicles per day and near Heidelberg with 100,000 vehicles per day (Schäfer and Puchelt, 1998). At these two locations, Pt concentrations in the 0-2 cm surface soil adjacent to the road ranged from several hundred mg/kg to local background values (£ 1 mg.kg-1) at less than 20 m from the road. Maximum Pd and Rh values were 10 and 35mg.kg-1 respectively. The PGM concentration decreased significantly with depth.
However the maximum PGM concentrations in soils at Heidelberg were only 25 percent those at Stuttgart, even though the traffic density was only 20% lower. The authors suggested that this could be due to frequent traffic jams at the Stuttgart site “leading to excessive emissions due to unfavourable working conditions of the engines”.
Urban road dusts collected in Stuttgart at the same time showed concentrations of Pt ranging up to 1000 mg.kg-1, 110 mg.kg-1 Rh and 100 mg.kg-1 Pd; these reflect short-term inputs of PGMs. A ratio of around 6 Pt: 1 Rh in traffic influenced soils and dusts has been reported by Schäfer et al (1996).
Schäfer et al (1999) measured time-dependent changing PGM depositions and contents of dusts and soils at a typical urban location at Karlsruhe in Germany. Daily deposition rates at 2 m distance from the traffic lane were within the range 6-27 ng m2 Pt, 0.8-4 ng m2 Pd. Concentrations of PGMs in the dusts sampled over an 8 monthly period illustrated the steady inputs. Using data for Pt concentration in soil at a site near Pfarzheim and a daily passage of around 15,000 Pt emitting cars per day, the authors calculated for a total number of 11 million converter-equipped vehicles over 2 years, a total emission of at least 3,000 ng of Pt per km along the traffic lane, giving a mean emission rate of 270 ng/km per vehicle. This value significantly exceeds the Pt emission rates of 2-86 ng.kg-1 measured in stationary motor vehicle experiments (König et al, 1992).
A recent estimate of total Pt emission in the vicinity of roads in Germany over the period 1985-2018 was 2,100 kg (using emission factors of 0.65 mg.kg-1 for highways, 0.18 mg.kg-1 for federal and national streets and 0.065 mg.kg-1 for district and city streets) (Helmers and Kummerer, 1999). These different emission rates reflect the increase in Pt load of exhausts with increasing speed of the car.
Accumulation of Pt was clearly shown in road dusts and surface soils adjacent to roads in the UK in 1994 (Farago et al, 1996, 1998). In the heavily trafficking London Borough of
118 Section 6. Case Studies
Richmond, Pt concentrations ranged up to 33 ng.g-1 in road dusts and 8 ng.g-1 in soils. Pt in road dusts was highest at major road intersections (mean 21 mg.g-1) compared with along major roads (13 ng.g-1) and intermediate and minor roads (2ng.g-1). The local background concentration for soils was 1 ng.g-1, similar to that obtained in rural Scotland.
More recently a study in the city of Nottingham, UK, compared Pt and Pd concentrations in garden soils and road dusts taken in 1996 and 1998 and archived samples taken in 1982 (which represented levels before the introduction of catalytic convertors) (Hutchinson, 2001). Significant increases for both Pt and Pd were found in road dusts (see Tables a.4 and a.5 and Figure a.1) with values ranging up to 298 ng.g-1 and 556ng.g-1 respectively for Pt and Pd in 1998 .
Table a.4 Summary results for Pt in garden soils (0-5cm) and road dusts from Nottingham (ng.g-1) aResidential streets with low traffic densities; b Includes major roads with high traffic densities (from Hutchinson, 2001)
Year Sample N Range Mean Geomean Median Nottingham 1982 Soil 42 0.27-1.37 0.61 - 0.59 1996 Soil 42 0.19-1.33 0.80 0.75 0.73
1982 Road dust 10 0.46-1.58 0.90 0.80 0.75 1996 Road dusta 8 0.82-6.59 2.78 2.29 2.06 1998 Road dustb 20 7.3-297.8 96.78 69.55 76.72
Table a.5 Summary data for Pd in garden soils (0-5 cm) and road dusts in Nottingham (ng.g-1) a Residential streets with low traffic densities; b Includes major roads with high traffic densities (from Hutchinson, 2001)
Year Sample N Range Mean Geomean Median Nottingham 1982 Soil 42 0.64-0.99 0.05 - 0.04 1996 Soil 42 0.21-1.11 0.18 - 0.10
1982 Road dust 10 0.69-4.92 1.24 - 0.22 1996 Road dusta 8 0.19-1.43 0.75 0.64 0.60 1998 Road dustb 20 5.6-556.3 92.9 40.95 35.84
An EU-funded study under the Environment and Climate Programme, CEPLACA, involved laboratories in Madrid, Gothenburg, Sheffield, Rome and Neuherberg. Changes in catalyst morphology over time were studied using SEM/EDX and laser induced breakdown spectrometry (LIBS) (Palacious et al, 2000). Catalysts were used up to 30,000 km in a roller dynamoneter following a driving cycle representing urban and non-urban driving conditions. Releases of PGMs were found to decrease with time. For new petrol catalysts mean releases were 100, 250 and 50 ng.km-1 for Pt, Pd and Rh respectively. In diesel catalysts Pt release ranged from 400-800 ng.km-1.
The effect of catalyst ageing was large. At 30,000 km releases were reduced to around 6-8 ng/km Pt, 12-16 ng/km Pd and 3-12 ng/km Rh for petrol catalysts. In diesel catalysts, the Pt release ranged from 108-150 ng/km.
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The difference between diesel and petrol was ascribed to the different composition of the washcoat and the different running conditions of diesel engines.
Soluble forms of PGMs emitted (in dilute HNO3) were significant for the fresh catalyst but less than 5% of the total amount. A previous study had reported 10% of the total Pt emission to be water soluble for fresh petrol catalysts (König et al, 1992).
At 30,000 km the amount of soluble PGMs released was similar or slightly higher than at 0 km. One possible explanation suggested for the relatively high amount of soluble PGMs related to the relatively high chloride concentration in fresh washcoat (i.e. one example quoted of 3.4 wt %). The authors suggested that the formation of soluble PtCl6, PdCl2 , PdCl4 or RhCl3 could be favoured at the high temperature and humidity that can be reached in the catalyst. The chloride concentrations in aged catalysts are normally very much lower. However, further laboratory tests using spiked solutions showed the instability of these chloro-complexes in the final exhaust fumes solution. It was thus thought possibly that the soluble or labile PGM fraction of the exhaust could be higher than those measured (Palacious et al, 2000).
An important conclusion from this study was that “no clear relation could be observed between the labelled amount of PGMs in the different catalysts studied and the measured amount released through car exhaust fumes. Different catalytic converter manufacturers, different car engines, even if running under the same conditions during the sampling period, and the well-characterized non-uniform behaviour of the catalyst could account for the lack of an observed correlation”.
Emissions of Platinum in effluents from hospitals
Effluents from hospitals contain platinum from excreted anti-neoplastic drugs, cisplatin and carboplatin, though workers in Germany have concluded that these are only of minor importance to environmental inputs from other sources and in particular from the use of catalytic convertors (Kümmerer and Helmers, 1997). These drugs were introduced 25 years ago to treat various tumours and are usually administered in the hospital environment. The platinum passes into hospital sewage which is then treated with household sewage in WWTS.
The authors monitored effluent samples from the University Hospital of Freiberg and two communal hospitals and found total inputs of platinum of around 330g/year from the University Hospital and 12 g/year from the Community Hospital. These equated to a consumption per bed per day of 600 mg Pt for the University Hospital and 85 mg Pt the community hospital. Extrapolation on a national basis, this amounted to an upper limit for the input in Germany of 141 kg Pt per year (c. 645,000 hospital beds (German Statistical Federal Agency, 1994) and a lower limit of 20 kg/year, with an average calculated value of 28.6 kg/year. Comparisons with other sources are shown in Table a.6.
Table a.6 Sources and sinks of platinum in Germany (from Kümmerer and Helmers, 1997) Source Amount Pt (kg/year) Reference Catalytic converters 15 König et al., 1992 emissions Zereini, F., personal communication 1996 Hospital effluents 28.6 Kümmerer and Helmers, 1997 Sewage sludge 100.4 - 400.8 Laschka and Nachtwey German Statistical Federal Agency, 1995
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A broader study based on hospitals in Belgium, Italy, Austria, the Netherlands and Germany aimed to provide reliable data with which to quantify sources of platinum in the environment from hospitals with other sources (Kümmerer et al, 1999). This study was supported by the LIFE95/D/A41/EU/24 Project of the European Community.
It was shown that 70% of the Pt administered in carboplatin and cisplatin is excreted and will therefore end up in hospital effluents. Pt concentrations measured in the total effluent of the different hospitals ranged widely from less than 10 ng.l-1 (the detection limit) in the Belgian and Italian hospitals to cca. 3,500 ng.l-1 for the Austrian and German hospitals. In all cases the influent of the WWTS was below 10 ng.l-1 as a result of dilution within the waste water system.
Annual emissions by hospitals and cars in Germany, Austria and the Netherlands are listed in Tables a.7 and a.8 and compared in Table a.9.
Table a.7 Emission of platinum by hospitals (D=Germany, A=Austria, NL=The Netherlands)
D 1994 D 1996 A 1996 NL 1996 Total hospital beds (approx) 645000 645000 77500 60000 Maximum Medical Performance 45000 45000 6500 N/A Pt per bed and year (mg) - maximum medical service spectrum 154.0 130.4 58.7 22.3 - medium medical service spectrum 14.0 14.0 N/A N/A Pt emissions by hospitals - maximum medical service spectrum 6.9 5.8 0.38 1.3 - medium medical service spectrum 8.4 8.4 N/A N/A Total emissions by hospitals (kg) 15.3 14.2 N/A N/A All hospitals as maximum medical service 99.3 84.1 4.6 1.3 spectrum
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Table a.8 Emissions of platinum by cars D 1994 D 1996 A 1996 NL 1996 Number of cars 32 000 000 32 000 000 3 593 588 5 740 489 With catalytic converter 12 800 000 19 200 000 1 607 699 3 307 300
% with catalytic converter 40.0 60.0 44.7 57.6 Kilometres/car 15 000 15 000 14 374 13 538 Total Kilometers (cat only) 1.92 x 1011 2.88 x 1011 2.311 x 1010 4.491 x 1010 Emission ( g km-1) 0.65 0.65 0.5 0.5 Total emission by cars (kg) 124.80 187.20 11.55 22.46
Table a.9 Platinum emissions comparison: hospitals vs. cars D 1994 D 1996 A 1996 NL 1996 Maximum medical service spectrum 5.6 3.1 3.3 6.0 Medium medical service spectrum 6.7 4.5 N/A N/A
Total 12.3 7.6 3.3 6.0 All hospitals calculated as maximum 79.6 44.9 39.4 6.0 medical service spectrum
In this study the highest concentrations of Pt in WWTP influents were found at the beginning of rain periods and at the end of cold periods when snow was melting. It was then concluded that the main inputs of Pt into municipal sewage were from urban and road run-off from traffic and other Pt emitting sources and not from hospital emitted sewage (Kümmerer et al, 1999).
Emissions from other Sources Kümmerer et al (1999) further concluded that “emissions by traffic and hospitals cannot explain the whole amount found in sewage and other sources emitting platinum directly into sewage have to be considered like glass and electronics industries or jewellery manufacturing. For the catalytic ammonia oxidation 92 kg platinum are reported to be lost from the catalyst” (Beck et al., 1995). If all of this is emitted into the atmosphere and washed off from roads and other paved areas in urban areas, which make up 11% of the total area of Germany (Losch, 1997), 10 kg from this source would be the input into sewage. If there are local industries like jewellery and electronic industries (Lottermoser, 1994) which use platinum to a certain extent they might be the most important local contributor to the platinum content of a certain municipal sewage and sewage sludge. Thus, unspecified input directly from industrial processes into sewage must be taken into account. These possible sources include jewellery manufacture, dental laboratories, electronic industries, glass manufacturing, production of platinum-containing drugs and industrial catalysts.
Knowledge on these sources, the species involved and their environmental properties is sparse if not non-existent.
A study of PGMs in sewage sludge incineration ashes from the municipal WWTS at Karlsruhe, Germany, showed Pd concentrations to have increased from 64 to 138 mg.kg-1 from 1993 to 1997, with Rh increasing from 4.8 to 6.3 mg.kg-1; Pd varied from 300 to 450 mg.kg-1 with no significant trend although these concentrations were 10-fold higher than in 1972 (Schäfer et al, 1999). The authors drew attention to the Pt/Rh ratio in the sludge of c. 20:1 which differs greatly from that of 6:1 typically found in environmental samples influenced by traffic emissions. They then estimated that, as more than 90% Rh is used for the production of catalytic convertors, and as Pt and Rh are emitted in a ratio of 6:1, that the contribution of traffic to the Pt concentration in sludge is only c. 30%, a result similar to that found in Munich by Laschka et al (1996). They thus suggested that the greater part of Pt in sewage sludge must come from sources other than catalytic convertors, such as hospital
122 Section 6. Case Studies and medical effluents or industrial emissions. They drew attention to the fact that “in cities with a jewellery industry, noble metal concentrations in sewage sludge far exceeded normal values even before the introduction of catalytic convertors” (Lottermoser, 1994).
Laschka and Nachtwey (1997) analysed Pt on primary and secondary effluents and in primary and digested sludge from two sewage treatment plants in Munich, where platinum pollution from industry, hospitals and traffic is considerable. Samples were taken before and after rainfall in October 1994 and July 1995. In general Pt concentrations in effluents were higher during rainy weather compared to dry weather. The Pt loading in secondary effluents was lowest for the period Monday/Tuesday (i.e. after the weekend), which was considered typical for industrial loads; the authors concluded that during dry weather, the platinum load originated mainly from industry. Comparison of the average Pt load in primary and secondary effluents in Munich, indicated a removal rate of 74% and 70% in the treatment plants. These elimination rates were lower than those typical for other metals such as Pb and Cd, which was attributed by the authors to the stabilizing effect of chloride (>100mg.l-1 in domestic sewage, or to the low Pt content of untreated sewage (<0.1 mg.l-1).
This study confirmed the enrichment of Pt in sewage sludge, which was present in materials from the 2 Munich treatment plants in concentrations ranging from 86 to 266 mg.kg-1. Sludges from other large towns and centres of industry had previously been found to contain 10 to 130 mg.kg-1 Pt and from smaller rural plants <10 to 50 mg.kg-1 Pt (Lottermoser, 1994). In this earlier study, an exceptionally high value of 1070 mg.kg-1 had been found in sludge at Pforzheim, a town with a jewellery industry.
The study of Laschka and Nachtwey (1997) concluded that “in a large industrial centre such as Munich, automobile traffic is not the dominant source of Pt in municipal sewage”.
A recent review article by Helmers and Kümmerer (1999) has attempted to quantify the sources, pathways and sinks of Pt in the environment. The authors noted that there was as yet not enough data to reliably investigate Pd and Rh fluxes, noting the lack of good quality assurance for Pd analysis and the paucicity of environmental data on Rh. An examination of archived sewage sludge ash from Stuttgart, Germany showed a continuous increase in Pt concentrations since 1984. With an estimated 5 x 1010 kg of sewage sludges for Germany in the early 1990’s and c. 250 mg.kg-1 Pt in sewage sludge, this amounts to some 12,500 kg of Pt, some 2 orders of magnitude higher than the Pt flux emitted by traffic. Much smaller Pt concentrations (mean 35 mg kg-1) have been found in smaller German purification plants.
Assuming that 50% Pt emitted by cars is received by sewage systems and taking into account the amounts of Pt in effluents from hospitals being completely received by sewage systems, the authors calculate that for Germany Pt received in the influents of WWTS from both these major sources amounted to 42.9 kg in 1994 and 56.4 kg in 1996. They thus considered an additional input of around 10 kg Pt per year from industrial sources.
If around 70% of the Pt influent is removed within the WWTS into sludge, the remaining 30% is emitted into freshwater (see Table 10). In Germany 30% of the sewage sludge is used on agricultural land and 70% disposed of as sludge or incinerated sludge ash. Losses to the atmosphere from incineration are not yet known.
Table 10. Partition of anthropogenic Pt fluxes (in kg) within German WWTS Year Received by Remaining in Disposal with Deposited Released into the WWTS the sewage sludges or agriculturally freshwater sludges ashes with sludges 1994 42.9 30.9 21.6 9.3 12 1996 56.4 40.6 28.4 12.2 15.8
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Helmers and Kümmerer (1999) consider the possibility of extrapolating these results to other European countries, taking into account traffic densities, catalytic convertor policies etc., with the qualification that “since there is no highway speed limit in Germany, highway Pt emissions of other countries may be halved in comparison with the German situation” (Helmers, 1997).
Solubility and bioavailability of PGMs in the environment
Current scientific opinion would seem to agree that PGMs emitted as autocatalyst particles remain bound to these and have limited mobility in the road and soil environment (Rauch et al, 2000). Experimental studies under laboratory conditions, in which ground catalysts have been added to soils under varying conditions of pH, chloride and sulphur concentrations have indicated that post-deposition processes in soils and waters are of minor importance and that “the risk of a health endangering contamination of the environment, and especially groundwater, at present seems negligible, as the PGM species behave relatively inertly” (Zereini et al, 1997). However transformation of PGMs into more mobile forms has not been ruled out and indeed Rauch et al (2000) suggest that this may occur “in the roadside environment, during transport through the stormwater system or in the urban river”.
In the absence of detailed study, it would seem impossible at this stage to apportion soluble PGM species in the influents and effluents of WWTS’s to specific PGM sources from traffic, hospital or industry, or to transformation/mobilisation of Pt and other PGMs in the environment and/or waste water system, or indeed in the processing plant.
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Conclusions
Platinum group metals are present in the influents of WWTS as a result of
· exhaust emissions from motor vehicles using catalytic convertors (both petrol and diesel) and subsequent runoff from road surfaces and roadside soils; · emissions in effluents from hospitals using the anti-cancer drugs cirplatin and carboplatin, · industrial uses including jewellery manufacture, electronics and glass manufacture.
Several studies over the past decade have shown a steady increase in the use of PGMs. Reliable quantitative information has shown that in general by far the greatest input of Pt and Pd into the environment and into WWTS is from vehicle exhaust catalysts, with hospital effluents accounting for some 6 to 12 per cent Pt. In large industrial centres, such as Munich, inputs from other sources (presumed industrial) may exceed those from catalytic convertors and hospitals. Quantitative knowledge of these sources is not currently available. The solubility of PGMs entering the environment and the influents of WWTS is thought to be low, though reactions within the soil/dust and wastewater environments need further study. Interactions with chloride ions and humic substances may well increase solubility and thus bioavailability.
Around 70 per cent of Pt in the influents to WWTS is removed with treatment into sludge, which may then be applied to agricultural land or incinerated. Where land application is practical, studies into uptake into pasture and foodcrops are recommended. The 30 per cent of Pt emitted into freshwater systems will potentially increase Pt levels in drinking water supplies.
At present there is no evidence of health risks arising from increasing levels of PGMs in the roadside environment, in sewage sludge or in drinking water. However, as levels of use continue to rise, it would seem prudent to focus research into factors influencing their solubility and bioavailability, their uptake and input into food crops and drinking water and into multiple exposure routes into the population.
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(b) Case Study- Sustainable Urban Drainage
Summary
Urban runoff source control practices have been the centre of an ongoing discussion involving maintenance and quality issues. This review will provide a brief overview of available techniques and structures and summarise their design characteristics. A discussion on performance will focus on water quality but comments on maintenance will also be included to allow the reader to form an overall opinion. A number of source control application case studies in Europe will be discussed from the point of view of performance.
Introduction
Urban Runoff has traditionally been treated as a water quantity problem and the usual approach to solving it has been a system of buried pipes designed to convey water downstream as soon as possible (CIRIA, 1999). Several problems in this traditional approach have been identified including possible flooding in downstream areas due to alteration of natural flow patterns, water quality issues that are not dealt with within the pipe system and largely ignored amenity aspects (such as water resources, landscaping potential and provision of varied wild habitat). These considerations have led to an effort of rethinking surface water drainage methods within the following framework:
· Deal with runoff as close to the source as possible · Manage potential pollution at source · Protect water resources from pollution · Increase amenity value
This framework and the practices and drainage systems that were developed from it, are collectively referred to in the UK as “sustainable urban drainage systems” (SUDS) (CIRIA, 1999) or more generally “source control” (Urbonas and Stahre, 1993) or “best management practices” (BMPs)1 (Jefferies et al., 1999). They essentially confirm with the emergence of Agenda 21 as a local action-planning basis for strategic and integrated approaches “to halt and reverse the effects of environmental degradation and to promote sound environmental development” (United Nations, 1992). Source Control includes structures such as:
· Dry Detention Basins · Infiltration Devices · Oil and Grease Trap Devices · Sand Filters · Vegetative Practices · Filter Strips · Grassed Swales · Wetlands, Constructed · Wetlands, Natural and Restored · Wet Retention Ponds
1 The fact that this Report adopts the widely used terms SUDS and BMPs to refer to source control and distributed storage practices does not imply that it necessarily considers them either “sustainable” or “best”. The positive and negative aspects of these practices will be discussed in the following paragraphs.
126 Section 6. Case Studies
Basic design characteristics and principles of use of the most widely used of these systems will be presented in the following paragraph and are summarised in Table 1.
Source Control Systems
The techniques presented will be grouped in four categories according to the CIRIA recommendations: (a) filters and swales, (b) permeable surfaces (c) infiltration devices and (d) ponds (CIRIA, 1996; CIRIA, 1999). The overall structure of an urban catchment with source control can be seen in the schematic in Figure b.1.
Figure b.1. Urban Runoff and Catchment (after CIRIA, 1996)
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Filters and Swales: These are vegetated landscape features with smooth surfaces and downhill gradient. Swales are long shallow channels while filters are gently sloping areas of ground. They mimic natural drainage patterns slowing and filtering the flow and are used for the drainage of small residential areas and roads. The flow depth should be smaller than the height of the grass to ensure filtration. Operational practices include regular mowing and clearing litter. Special care should be taken not to allow the swale to erode after heavy storms. Grass swales have been used extensively in North America, but have only recently appeared in Europe. Information on treatment performance comes mainly from the US (as summarised in Ellis, 1991) and indicates removal potential for solids, potentially toxic elements and hydrocarbons). In the UK however the quality improvement potential of Figure b.2. Grass Swale (after CIRIA 1996) the swales is ignored. A typical swale structure can be seen in Figure b.2.
Permeable surfaces: These include porous pavements, gravelled areas, grass areas and other types of continuous surfaces with an inherent system of voids. The water passes through the surface to the permeable fill, allowing for storage, transport and infiltration of water. The actual amount of water stored is dependent on the voids ratio, the plan area and the structure’s depth. It acts as a trap for sediment thus removing a large number of pollutants from the runoff, but keeps them within the particular site. The principal mechanism for pollutant retention is thought to be adsorption onto materials within the pavement construction (Pratt, 1989). Maintenance should ensure that the voids are not filled by sand and silt and such an operation may prove costly, as the surface structure can deteriorate under external pressure. The US, France, Holland, Austria and Sweden have used porous surfaces for Figure b.3. Porous Pavement (after CIRIA both traffic and pavement areas (Diniz, 1976; 1994) Hogland, 1990). A typical structure can be seen in Figure b.3.
128 Section 6. Case Studies
Infiltration devices: Soakways and infiltration trenches are below ground and are filled with a coarse material. They drain water coming in the infiltration device from a pipe or a swale directly to the surrounding soil. Their operation is based on increasing the natural capacity of the soil for infiltration but effectiveness is ultimately limited by soil permeability. The volume of storage therefore is dependent on soil infiltration potential. Physical filtration can remove solids, while biochemical reactions caused by microorganisms growing on the fill or the soil can degrade hydrocarbons. The level of treatment depends on the size of the media and the length of the flow path (CIRIA, 1999). Extensive use of soakways in Sweden and the US as well as in the UK generally provide positive feedback on maintenance and operation (Pratt, 1989; CIRIA, 1996). Areas that are drained through infiltration structures of Figure b.4. Soakway (after CIRIA, 1996) different types include car parks, roads, roofs pavements and pedestrian sidewalks. Pollution levels in these types of urban runoff can be however significant and there is therefore serious risk of introducing the pollutants to the groundwater. Additionally the introduction of water to the soil may cause geotechnical problems. Figure b.4 describes a typical soakway.
(d) Basins and Ponds: These are areas of storage of surface runoff that are free from water under dry weather conditions. Structures can be mixed with a permanently wet area for wildlife or treatment of runoff and an area that is usually dry to allow for flood attenuation. The ponds are normally situated near the end of the system due to detention and land price constraints (Makropoulos et al., 1999). Flow detention would lead to settlement of the particles and associated pollution loads. Additionally some bacterial die-off and soluble particle removal could be expected (CIRIA, 1994). Annual clearance of the aquatic vegetation and silt-removal every five to ten years should be thought of as an average operational practice. Figures b.5 and b.6 give an idea of on and off stream detention and the pond-wetland principle respectively. Figure b.5. Typical on and off stream storage ponds (after CIRIA, 1994)
129 Section 6. Case Studies
Figure b.6. Typical arrangement of a reed bed treatment pond (after CIRIA, 1994)
Performance
Systematic evaluation of the application of these systems is scarce in literature. Research has been focusing on mathematical modelling of the system’s quality performance and actual data is generally not available in Europe. Scotland is a notable exception. BMPs have been promoted for the past five years in response to the need to combat pollution from diffuse sources in urban areas. To meet this need, a programme of investigations is being undertaken into the performance of BMPs, which have been built in Scotland. An initial awareness survey by Abertay University and SEPA indicated high levels of apparent knowledge of BMPs, but subsequent investigations showed that in many instances knowledge was very superficial and often inadequate. Jefferies et al. (1999) discuss experimental findings and theoretical considerations of that investigation and show that, in most systems, pollutants will form sludge. This in turn must be disposed of, and indeed good household practices may be the only truly sustainable drainage practice. Pollutants removed from runoff in a system such as a pond may accumulate in sediments and biota. Potentially toxic elements and trace organics in rainfall runoff are to some extent associated with soil particulates, as discussed earlier in this report, and will thus tend to be removed by sedimentation. The soluble fraction of pollutants will also to some extent precipitate following changes in pH, oxidation-reduction potential or temperature (Kiely, 1997). The activity of the pollutants however is not ended with their concentration in the sediments. Polluted sediment may be resuspended or pollutants may be released during high stream flows (Pitt, 1995). The quality of groundwater may also be affected by exfiltration of contaminants from BMP systems. Studies in the US have shown that, when disposed in soakways, organophosphates have appeared in watercourses 400 metres away only two hours after disposal (ENDS, 1993). The entire range of toxic pollutants identified as possible input to urban rainfall runoff may leak this way to the groundwater. A new problem has appeared in the form of methyl butyl tertiary ether (additive to unleaded petrol), which is ten times more soluble in water than other constituents in petrol and thus would tend to spread readily in groundwater (Kiely, 1997). When soil is used as a filtration medium in source control systems (as in infiltration trenches and even grass swales), it must be regularly checked as the adsorbed pollutants may be remobilised under various conditions. Furthermore, possible degradation of pollutants inside the systems may give rise to hazardous by-products which may be more soluble or toxic than the original forms (Hallberg, 1989). Biotransformation of TCE for example results in hazardous products such as vinyl chloride, which is a confirmed human carcinogen (Burmaster, 1982). Table b.1 summarises the main functions and water quality attributes of source control.
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Table b.1. Functions and water quality attributes of different source control structures (after CIRIA, 1994)
METHOD PRIMARY SECONDARY WATER QUALITY ATTRIBUTES FUNCTION Infiltration Collection and Sediment and Can remove pollutants associated pavements disposal of pollutant removal with sediments and dissolved surface water pollutants but may be lead to increase in nutrient levels Swales Conveyance of Storage sediment and Can remove suspended and possibly surface waters pollutant removal; dissolved pollutants but may be a risk disposal to groundwater quality if not sealed Infiltration Disposal of Storage; sediment and Can remove suspended and possibly basins surface water pollutant removal dissolved pollutants but may be a risk to groundwater quality Storage Storage of Sediment and Can remove pollutants associated ponds surface water pollutant removal with sediments and provide some biological treatment Wetlands Pollutant removal Storage Can remove and treat various pollutants
France has also had experience in particular aspects of SUDS (for detention basins and ponds). Nascimento et al. (1999) provide an overview of the French experience in detention basins use and performance. In France, detention basin use dates back to the 1960s, together with the construction of the “Villes Nouvelles”. Their use was limited, but these facilities are increasingly popular as indicated in Deutsch et al. (1990) (Reported in Nascimento et al., 1999). Recent research has focused in the quality side of performance of the ponds with the QASTOR database created by CEREVE as a centre point (Saget et al., 1998). The database aims to collect and analyse all French national data related to urban wet weather discharge that have been collected in 19 catchments since 1970. The efficiency of detention basins in reducing pollutants is the result of a large number of variables including physical, chemical and biological characteristics of pollutants, precipitation regime, detention time and quality of maintenance services (Nascimento et al., 1999). Table b.2 identifies the potential annual and short-term efficiency of detention basis recorded by Adler (1993) as reported in Nascimento et al, (1999). The Table draws from studies conducted by the French institution CEMAG-REF. The presented data were for basins installed in separate drainage systems and the results indicate that such storage facilities can have a reasonable performance even over quite short time scales.
Table b.2. Efficiency of detention basins (after Nascimento et al., 1999)
Yearly Inflow Yearly Outflow Reduction Reduction after (kg/ha imp) (kg/ha imp) (%) 2h (%) Pb 0.893 0.054 94.0 65 Zn 5.12 0.66 87.1 77 Cd 0.0310 0.0051 83.7 - Cu - - - 69 Hydrocarbons 65 4 94.2 -
However, despite such evidence, when the issue of integrating detention basins into the urban context is concerned, the outcome may be very different according to the specific case. For the UK the high pollutant loading to urban detention basins, which has been reported, has led to concerns about long-term siltation (and loss of effective storage volume) and water quality (especially in terms of health risks). Sansalone (1999) describes the
131 Section 6. Case Studies results of measurements taken in a field scale infiltration trench. Figure b.7, summarises some of the findings, indicating significant potentially toxic element removal efficiency exceeding 80% after 1 year of runoff loadings.
The Technical University of Denmark (Mikkelsen, et al. 1996a and 1996b) has been involved in a series of tests to examine the effects of stormwater infiltration on soil and groundwater quality. They found that potentially toxic elements and PAHs present little groundwater contamination threat, if surface infiltration systems are used. However, they express concern about pesticides, which are much more mobile.
Recent and ongoing studies in the US have tried to identify the potential hazards from the use of infiltration systems.
Figure b.7 Infiltration Trench removal performance and % of influent exfiltrated to soil for a series of 4 runoff events compared to bench scale results (lab) after 1 year of equivalent loading (after Sansalone, 1999)
In particular, a multi-year research project sponsored by the US EPA addresses the potential problem of groundwater contamination due to stormwater infiltration (Pitt, Clark and Palmer, 1994; Pitt et al., 1997; Pitt, Clark and Field, 1999). In the case of pesticides the research found that heavy repetitive use of mobile pesticides, such as EDB, on site with infiltration devices likely contaminates groundwater. Fungicides and nematocides must be mobile in order to reach the target pest and hence, they generally have the highest contamination potential. Pesticide leaching depends on patterns of use, soil texture, total organic carbon
132 Section 6. Case Studies content of the soil, pesticide persistence, and depth to the water table (Shirmohammadi and Knisel 1989). A pesticide leaches to groundwater when its residence time in the soil is less than the time required to remove it, or transform it to an innocuous form by chemical or biological processes. The residence time is controlled by two factors: water applied and chemical adsorption to stationary solid surfaces. Volatilization losses of soil-applied pesticides can be a significant removal mechanism for compounds having large Henry’s constants (Kh), such as DBCP or EPTC (Jury, et al. 1983). However, for mobile compounds having low Kh values, such as atrazine, metolachlor, or alachlor, it is a negligible loss pathway compared to the leaching mechanism (Alhajjar, et al. 1990).
Restricted pesticide usage in areas with high infiltration potential has been recommended by some U.S. regulatory agencies. The slower moving pesticides were recommended provided they were used in accordance with the approved manufacture’s label instructions. These included the fungicides Iprodione and Triadimefon, the insecticides Isofenphos and Chlorpyrifos and the herbicide Glyphosate. Others were recommended against, even when used in accordance with the label’s instructions. These included the fungicides Anilazine, Benomyl, Chlorothalonil and Maneb and the herbicides Dicamba and Dacthal. No insecticides were on the “banned list” (Horsley et al, 1990).
In the case of potentially toxic elements, problems may appear when infiltrating stormwater using a rapid infiltration system (Crites 1985), such as a dry well. Most metals have very low solubilities at the pHs found in most natural waters and they are readily removed by either sedimentation or sorption removal processes (Hampson 1986). Many are also filtered, or otherwise sorbed, in the surface layers of soils in infiltrating devices when using surface infiltration. Table 3 discusses the pollutants found in stormwater that may cause groundwater contamination problems when allowed to infiltrate through infiltration devices.
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Table b.3. Groundwater Contamination Potential for Stormwater Pollutants (after Pitt et al., 1994)
Compounds Mobility Abundance Fraction Contamination Contamination Contamination (worst case: in storm- filterable potential for potential for potential for sub-surface sandyl-1ow water surface infilt. and surface infilt. with injection organic soils) no pre-treatment sedimentation with minimal pre-treatment Nutrients nitrates mobile low/moderate high low/moderate low/moderate low/moderate Pesticides 2,4-D mobile low likely low low low low g-BHC (lindane) intermediate moderate likely low moderate low moderate malathion mobile low likely low low low low atrazine mobile low likely low low low low chlordane intermediate moderate very low moderate low moderate diazinon mobile low likely low low low low Other VOCs mobile low very high low low low organics 1,3-dichlorobenzene low high high low low high anthracene intermediate low moderate low low low benzo(a) anthracene intermediate moderate very low moderate low moderate bis (2-ethylhexyl) intermediate moderate likely low moderate low? moderate phthalate butyl benzyl phthalate low low/moderate moderate low low low/moderate fluoranthene intermediate high high moderate moderate high fluorene intermediate low likely low low low low naphthalene low/inter. low moderate low low low penta- chlorophenol intermediate moderate likely low moderate low? moderate phenanthrene intermediate moderate very low moderate low moderate pyrene intermediate high high moderate moderate high Potentially nickel low high low low low high toxic cadmium low low moderate low low low elements chromium inter./very low moderate very low low/moderate low moderate lead very low moderate very low low low moderate zinc low/very low high high low low high Salts chloride mobile Seasonally high high high high high
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Conclusions
The control of diverse pollutants requires a varied approach, including source area controls, end-of-pipe controls, and pollution prevention. All dry-weather flows should be diverted from infiltration devices because of their potentially high concentrations of soluble potentially toxic elements, pesticides, and pathogens (Pitt et al., 1999) Similarly, all runoff from manufacturing industrial areas should also be diverted from infiltration devices because of their relatively high concentrations of soluble pollutants. In areas of extensive snow and ice, winter snowmelt and early spring runoff should also be diverted from infiltration devices.
All other runoff should include pre-treatment using sedimentation processes before infiltration, to both minimize groundwater contamination and to prolong the life of the infiltration device (if needed). This pre-treatment can take the form of grass filters, sediment sumps, wet detention ponds, etc., depending on the runoff volume to be treated and other site-specific factors. Pollution prevention can also play an important role in minimizing groundwater contamination problems, including reducing the use of galvanized metals, pesticides, and fertilizers in critical areas. The use of specialized treatment devices can also play an important role in treating runoff from critical source areas before these more contaminated flows commingle with cleaner runoff from other areas (Pitt et al., 1999). Sophisticated treatment schemes, especially the use of chemical processes or disinfection, may not be utilised, provided there is no danger of forming harmful treatment by-products (such as THMs and soluble aluminium).
The use of grass swales and percolation ponds that have a substantial depth of underlying soils above the groundwater is preferable to using dry wells, trenches and especially injection wells, unless the runoff water is known to be relatively free of pollutants. Surface devices are able to take greater advantage of natural soil pollutant removal processes. However, unless all percolation devices are carefully designed and maintained, they may not function properly and may lead to premature hydraulic failure or contamination of the groundwater (Pitt et al., 1999).
It should be clear that although SUDS have great potential in both quantity and quality control in urban runoff, each case should be assessed individually, and an incremental approach containing both high tech and low-tech solutions is the most likely development scenario (Butler and Parkinson, 1997). Direct application of such methods across different regions and countries is not always appropriate and must also include consideration of the local socio-economic and administrative circumstances associated with the operational design, which can be primary inhibitors to the implementation of innovative technology.
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(c) Artisanal Activities: Pollutant Sources and load in Urban Wastewater in Vicenza, Northern Italy; Gold Jewellery – Best Environmental Practice
Pollutant sources and load in urban wastewater in Vicenza, northern Italy
Introduction
In Northern and Central Italy there is a high density of small-scale, artisanal enterprises and activities. For example, in the region of Veneto (North-Eastern Italy), artisanal activities account for 20% of the regions exports. The area of Vicenza, a provincial town in the Veneto region, represents one of the largest agglomerations of artisanal activities in Italy. With a population of 109,000 inhabitants, the municipality of Vicenza has a total of about 1600 small to medium-scale enterprises (SMEs).
The environmental impact of artisanal activities is less clearly understood than the impact from industrial activities. While there are a large number of point sources, each source contributes a very low wastewater flow rate, closer to the wastewater discharge from residential units than from industrial sites. Nevertheless wastewater from artisanal activities may be dramatically different to that from residential units, both in terms of pollutant concentration and the presence of specific pollutants.
Due to the presence of specific pollutants, wastewater discharge is regulated in the same way as industrial wastewater (i.e. in terms of pollutant concentration limits), even though artisanal wastewater flow rates may be orders of magnitude lower than industrial ones [Italian law by decree n. 152, 1999]
EBAV (Ente Bilaterale Artigianato Veneto, a non-profit bilateral organisation representing the interests of both artisanal workers and enterprises) sponsored a study on the origin and contribution of pollutants to urban wastewater. The aim of the study was to assess the load of pollutants from different artisanal activities, in comparison to the total load originating from the urban wastewater system. In addition, the pollutant load from artisanal activities was subdivided into load from discharged wastewater and load from concentrated liquid wastes (which are separated and collected by external firms), to assess if wastewater segregation could significantly affect pollutant load from artisanal activities.
Description of the study The EBAV study was carried on in the period 1994-1995 in the municipality of Vicenza. According to local authorities, the only notable change (in terms of residential population and type of SMEs) since the time of the study, has been the rapid increase in the number of “service” enterprises (for example software companies), which do not contribute specific wastewater. Therefore, this growing number of “service” companies does not affect the conclusions of the study, which still may be considered valid today.
Wastewater for the whole Vicenza area is treated by four WWTS (Table c.1). The total capacity is about 137,900 p.e. (population equivalent) with a total flow rate of about 23.2 million m3 year-1.
136 Section 6. Case Studies
Table c.1 capacity of the four municipal wastewater treatment plant of Vicenza municipality p.e.=population equivalent
WWTS Capacity (p.e.) Treated flow rate m3 year-1 CASALE 71,900 10,202,000 LAGHETTO 3,500 394,000 LONGARA 3,500 826,000 S. AGOSTINO 59,000 11,800,000 Total 137,900 23,222,000 In this area there are about 1,579 artisanal enterprises discharging their wastewater into the UWW collecting system. Table c.2 shows the most common artisanal activities in the area of Vicenza and the number of enterprises involved in each activity. For each activity a representative number of enterprises was selected for further investigation. Typically one wastewater sample was drawn from each enterprise. In some cases an additional sample of concentrated, segregated wastewater was also drawn.
Table c.2 Main artisanal activities in the area of Vicenza Type of activity No. of samples No. of enterprises in the municipality of Vicenza Food workshops 10 53 Car-repairers 20 (14+6*) 175 Ceramic and photoceramic 7(6+1*) 23 Artisanal galvanic shops 8(5+3*) 18 Printing shops 21(14+7*) 140 Wood manufacturing 18(3+15*) 92 Marble manufacturing 5 140 Metallurgists and mechanics 15(8+7*) 155 Dental practices 21 88 Gold manufacturing shops 34 258 Hairdressers 19 310 Laundrettes and dry-cleaners 25 88 Textile shops 2 16 Artisanal glass manufacturing 3 23 TOTAL 208 1579 *concentrated wastewater, segregated and committed to external firms
Total load Along with artisanal wastewater samples, influent and effluent samples from the four municipal WWTS were analysed for a large number of pollutants including: B, Cd, Cr(III), Cr(VI), Mn, Ni, Pb, Cu, Zn, and anionic surfactants. From each of the four WWTS a large number of influent and effluent samples were taken and analysed. Using the specific wastewater flow rate and the influent concentration, the pollutant load was calculated for each plant. Table c.3 reports the total load as sum of the pollutant loads of the WWTS, assuming that urban wastewater is treated by only one hypothetical centralized plant.
Total pollutant load in Table c.3 is reported as average value, based on the average concentration from 20-50 samples. In addition, the maximum load and the upper 95% confidence interval are given. The last column reports the average removal efficiency based on the comparison of influent and effluent pollutant concentrations.
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Table c.3 Pollutant load to the hypothetical centralized WWTS of Vicenza municipality
POLLUTANT TOTAL LOAD REMOVAL (g/day) EFFICIENCY(%) AVERAGE MAX(95%) MAX Cd 64 75 960 40 Cr(III) 636 759 13244 75 Cr(VI) A typical example of the analytical work performed for each category of artisanal activity is reported in Tables c.4 and c.5 for car repair shops. A rough statistical analysis of the results has been performed to obtain the average concentration and the upper limit of the 95% confidence interval (Max 95%). In addition the maximum value (Max) is also reported. Table c.4 Pollutant concentrations (mg l-1) in discharged wastewater of 14 car-repair shops Pollutant Average Max 95% Max CI COD 329 547 1800 Cd 0.01 0.03 0.14 Cr(III) 0.1 0.2 0.8 Cr(VI) 0 0 0 Mn 0 0 0 Ni 0 0 0 Pb 0 0.1 0.4 Cu 0.1 0.1 0.5 Zn 4.5 11.1 55 Table c.5 Pollutant concentrations (mg l-1) in the segregated wastewater of 6 car- repair shops Pollutant Average Max Max 95%CI COD 10293 15346 19120 Cd 0.3 0.7 1.3 Cr(III) 0.2 0.3 0.5 Cr(VI) 0 0 0 Ni 2.5 6.1 12 Pb 22.1 52 102 Cu 33 74.6 144 Zn 31.9 73.8 145 Pollutant load was calculated for the selected enterprises and extrapolated to the total number of enterprises for each category. Even though each enterprise was equipped with its own treatment plant, pollutant loads were calculated on the basis of concentrations in the untreated wastewater, hypothesizing a scenario where no pre-treatment is performed and 138 Section 6. Case Studies the wastewater is discharged directly into the UWW collecting system. Similar calculations were performed using concentrations and volumes of segregated wastewater (spent baths). So, for the enterprises that separate wastewater for treatment by external firms (for recovery and detoxification), two potential pollutant loads were given with reference to the two hypothesized scenarios: · all the wastes (concentrated and diluted wastewater) are discharged into the urban wastewater system without pre-treatment (Table c.6 D and S); · the spent baths are treated externally, whereas wastewater is directly discharged into the urban wastewater system (Table c.6 D only). Car-repairers (175 shops) The most significant pollutants are: suspended material, COD, oils, surfactants, organic solvents, copper, and zinc. Considering both discharged and segregated wastewater the total pollutant load for Zn, Cd, Cu and, above all, Pb is very high. However upon careful segregation of concentrated wastewater (D only), the pollutant load is significantly reduced. The careful segregation of spent baths induces a decrease of about one order of magnitude in the percentage of lead originating from car repair shops. Ceramics and photoceramics (23 shops) The principal pollutants from these artisanal activities are: suspended solids, lead, ammonia, nitric nitrogen, and surfactants. As can be seen in Table c.6 the pollutant load from the ceramic and photoceramic shops is minimal due to the low wastewater flow rate. Only lead seems to represent a significant load. In the case of ceramic shops, segregation of concentrated wastewater is not very significant in terms of reducing pollutant load. Galvanic (18 shops) The principal pollutants from these enterprises are: suspended solids, chromium (VI), nickel, lead, copper, and cyanide. Segregation and external treatment of concentrated wastes does not appear to significantly reduce the lead load because this load originates mainly from the discharged diluted wastewater. Printing shops (140 shops) Printing activities generate several pollutants: suspended material, COD, cadmium, chromium, lead, copper, zinc, sulphites, sulphates, chlorides, ammonia, total phenols, aldehydes, aromatic organic solvents, and surfactants. Due to the low contribution to the overall UWW flow rate (0.15%), only Cd and Cu loads from the printing activities are significant with respect to the total load. Segregation of concentrated wastewater reduces the load of metals to negligible values. Wood processing and furniture making shops (92 shops) The most significant pollutants originating from this activity are: suspended solids, COD, lead, copper, zinc, total phenols, organic solvents, and surfactants. Table c.6 shows that by simply segregating concentrated wastes the pollutant load to the WWTS from wood processing and furniture manufacturing shops is dramatically reduced. Other specific wood processing pollutants such as arsenic were not analysed. 139 Section 6. Case Studies Metallurgists and mechanics (155 shops) Several pollutants originate from metallurgists and mechanic shops: suspended solids, COD, cadmium, chromium, nickel, lead, zinc, copper, sulphates, chlorides, phosphorus, oils, solvents, and surfactants. Pollutant concentrations in wastewater are typically much lower than in segregated wastewaters (see Table c.7) which have high average concentrations of pollutants such as; nickel, lead, copper and zinc. Separation of concentrated wastes reduces the load to the WWTS significantly. Goldsmiths (258 shops) The area of Vicenza represents one of the most important districts for gold manufacturing in Italy, with up to 258 artisanal goldsmith shops. The main pollutants originating from gold manufacture are: COD, boron, cadmium, copper, zinc, and surfactants. However goldsmiths shops are characterized by very low wastewater flow rates, with an average of about 1 cubic meter per day per unit. In terms of contribution to the pollutant load, the 258 goldsmiths shops represent a high contribution of Cd, Cu, and Zn. This is mainly due to the fact that goldsmith shops used to add spent concentrated baths to the diluted wastewater in order to recover precious metals during the wastewater treatment before discharging it into the sewer. Food workshops (confectioners, ice-cream parlours, bakeries) (53 shops) The most significant pollutants originating from this activity are: COD, fat, oils, and surfactants. Cu and Zn are the only metals to be above the limits of detection in wastewater. Dental technicians (88 shops) Principal pollutants originating from this activity are: suspended solids, COD, and surfactants. The contribution of the 88 shops to the total metals pollutant load is generally low, due to the low contribution in terms of flow rate, that is an average of about 0.07% of the total UWW flow rate. Mercury is an important pollutant in wastewater linked to dental practices, which is not considered in this particular study but is referred to in section 2.1.2 of this report. Hairdressers (310 shops) This is the category with the largest number of shops in the Vicenza’s municipality. The main pollutants originating from hairdressers are: suspended solids, COD, and surfactants. Laundrettes and dry-cleaners (88 shops) The main pollutants originating from laundrettes and dry-cleaners are: suspended solids, COD, surfactants, chlorides, and solvents. The most significant load of pollutant from the 88 laundrettes and dry-cleaners is for Cd although overall the levels for this pollutant were very low . 140 Section 6. Case Studies Table c.5 Pollutant load from specific artisanal activities. D=discharged wastewater, S=segregated wastewater, (- = not reported). Flow Average Pollutant Load (g/day) rate m3 per day Cd Cr III Mn Ni Pb Cu Zn Car repair shops 153 4 6.3 Impact of all artisanal activities on urban wastewater treatment plants Pollutant loads from all artisanal activities of Vicenza municipality were calculated, summing the loads of each specific activity as reported in Tables c.7 and c.8. 141 Section 6. Case Studies Table c.7 Total pollutant loads from artisanal activities in the Vicenza municipality, considering both discharged and segregated wastewater POLLUTANT TOTAL POLLUTANT LOAD FROM PERCENTAGE ARTISANAL ACTIVITIES (g/day) OF TOTAL LOAD TO WWTS average max(95%) max B 2909 4774 38832 Cd 30 66 529 46.59 Cr(III) 54 106 243 8.55 Cr(VI) 14 49 43 Mn 14 19 38 0.28 Ni 1180 3153 8114 16.63 Pb 452 1064 2676 71.12 Cu 903 1872 6387 25.48 Zn 2231 3998 10589 19.88 MBAS 53404 93158 353795 25.45 flow rate m3/day 1633 2526 5559 2.57 p.e. 7697 13295 46221 5.58 Table c.8 Total pollutant loads from artisanal activities in the Vicenza municipality, considering discharged wastewater only POLLUTANT TOTAL POLLUTANT LOAD FROM PERCENTAGE ARTISANAL ACTIVITIES (g/day) OF TOTAL LOAD TO WWTS average max(95%) max B 2909 4774 38832 Cd 24 51 498 36.9 Cr(III) 11 27 100 1.8 Cr(VI) In terms of flow rate the contribution of all artisanal activities is typically low (2.6%). It remains relatively low (4%) in the worst scenario case, where the artisanal activities discharge at the upper limit of the confidence interval of their cumulative wastewater flow rate, while the total UWW flow rate remains at the average value. In contrast, the percentage contribution of artisanal activities to the total load is very high for pollutants such as: Pb, Cd, Cu, Zn, Cr III, and surfactants. Figures c.1 and c.2 compare the load of each specific activity to the total load of a hypothetical centralized WWTS, for potentially toxic elements and surfactants, respectively. Figure c.1 clearly shows that only car-repairers, goldsmiths shops, metallurgists and mechanics contribute significant amounts of potentially toxic elements to UWW, with respect 142 Section 6. Case Studies to the total metal load of the all activities. The main activities responsible for the surfactant load in UWW are; hairdressers, goldsmiths, and food workshops (Figure c.2). Figure c.1 Heavy metal load from artisanal activities 9000 8000 7000 6000 5000 4000 3000 average load (g/day) 2000 1000 0 AL AU CE GA GR LG MA ME OD OR PR PU TS VE tot AL: Food workshops AU: Car repairers CE: Ceramics GA: galvanic GR: Printing shops LG: Wood manufacturing MA: Marble manufacturing ME: Metallurg.and mechanics OD: Dental practices OR: Goldsmiths PR: Hairdressers Figure c.2 Surfactant load from artisanal activities 60000 50000 40000 30000 20000 average load (g/day) 10000 0 AL AU CE GA GR LG MA ME OD OR PR PU TS VE tot 143 Section 6. Case Studies With the exception of surfactants and cadmium, the high pollutant loads from artisanal activities are notably reduced when the concentrated wastes are not considered (Table c.23). The poor effect of waste segregation on surfactant load is explained by the fact that the major surfactant contribution derive from activities that do not practice waste segregation: hairdressers, goldsmiths shops, and food workshops (Figure c.2). Impact of artisanal activities on sewage sludge characteristics Assuming that the fate of sewage sludge produced from the hypothetical centralized WWTS is used in agriculture, the maximum admissible pollutant concentration in sludge must be considered. From this value a maximum admissible pollutant concentration in the UWW may be calculated according to the following equation: Ci Qf 100 %H 2O C R% Q where C is the maximum influent concentration permitted for sludge disposal Ci is the maximum pollutant concentration allowed in sludge to land R% is the removal efficiency (%) in the treatment plant 3 -1 Qf is sludge flow rate in m day calculated on the basis of the production of 1.87 l/inhabitant with 95.5% of humidity Q is the influent wastewater flow rate % H2O is the water content of the sludge (95.5%) Table c.9 Pollutant contribution of artisanal activities to the admissible load for sludge disposal in agriculture, considering both discharged and segregated wastewater Pollutant Regulatory limits for Admissible(*) Admissible(*) percentage of total agricultural use concentration load admissible load (mg/kg dry sludge) (%) D.Lgs 99/92 Veneto (mg/l) (g/day) average Cd 20 10 0.01 580 5.11 Cr(III) 500 0.15 9670 0.56 Ni 300 200 0.17 10549 11.18 Pb 750 500 0.21 13599 3.33 Cu 1000 600 0.28 17582 5.14 Zn 2500 2500 0.66 42045 5.34 (*)on the basis of equation (1) 144 Section 6. Case Studies Table c.10 Pollutant contribution of artisanal activities to the admissible load for sludge disposal in agriculture considering only discharged wastewater Pollutant Regulatory limits for Admissible(*) admissible(*) percentage of total agricultural use conc load admissible load (mg/kg dry sludge) (%) D.Lgs 99/92 Veneto (mg.l-1) (g/day) average Cd 20 10 0.01 580 4.04 Cr(III) 500 0.15 9670 0.12 Ni 300 200 0.17 10549 0.31 Pb 750 500 0.21 13599 0.37 Cu 1000 600 0.28 17582 1.92 Zn 2500 2500 0.66 42045 2.35 (*)on the basis of equation (1) In Tables c.9 and c.10 the concentration limits imposed by law for sludge used in agriculture are reported. From these limits the concentration limit in the influent and the consequent admissible load were calculated. In the last column the impact of artisanal activities on the sludge characteristics are reported in terms of percentage load from artisanal activities with respect to the total admissible load for sludge disposal (calculated using average values). Table c.10 considers the highest possible pollutant loads in the hypothesis that waste segregation does not take place. Even in this pessimistic hypothesis, the impact of artisanal activities is typically low. The highest metal contribution is for Ni, representing 11% of the admissible load. In the worst case scenario, considering all the maximum loads, artisanal activities by themselves almost reach the admissible Cd and Ni loads. The average impact on sewage sludges, without considering the concentrated spent baths (Table c.9) is less than 5% for all the metals. Validation of the Case Study with results from a recent study on hairdressers’ shops in a different area of Veneto region To validate the results obtained in the study described above, EBAV undertook an additional study specifically addressed to activities of hairdressers and beauticians. This study, performed during 1999-2000, considered a group of shops representing all the hairdressers and beauticians located in the district of Valdagno (Vicenza), which discharge their wastewater into the UWW collecting system of the municipality of Trissino (Vicenza). In this validation study the number of shops examined was about 22% of those present in the area, compared to only 6% of hairdressers in the Vicenza Case Study. Typically one wastewater sample was drawn from each shop. A rough statistical analysis of the results has been performed to obtain the average concentration; then the average values have been increased by 10% to take into account the effects of highly polluted wastewater (due to typical products such as shampoos or dyes). In Table c.11 the average pollutant concentrations are reported, as well as the average wastewater flow rate per unit. In this case wastewater production was even lower than in the case of Vicenza (600 l.day-1 vs 900 l.day-1). The contribution of hairdressers’ shops to the total wastewater flow rate entering the Trissino WWTS was 0.33%, whereas in Vicenza area this was about 0.45%. From the average concentrations (with a safety increase of 10%) and from wastewater flow rates the pollutant load from the total 135 shops to the Trissino WWTS was calculated. Where data was available (chromium and surfactants), this load was compared with the total pollutant load to the WWTS. Data from Valdagno area confirm the negligible contribution of chromium load, and other potentially toxic elements as well, to the 145 Section 6. Case Studies WWTS. Analysis of surfactants (anionic and non-anionic) in wastewater samples showed that the contribution of beauty shops in Valdagno district to the total surfactants load still remains lower than the estimated contribution in Vicenza municipality, where only anionic surfactants were considered. The data presented above, while confirming the negligible contribution of this artisanal activity to the total load of metallic pollutants, suggest that the extrapolation of the results from Vicenza Case Study may result in an overestimation of the contribution of artisanal activity to the pollutant loads in the WWTS systems. Table c.27 Pollutant concentrations and pollutant loads from hairdressers and beauticians in the district of Valdagno (Vicenza) POLLUTANT AVERAGE POLLUTANT LOAD* PERCENTAGE OF LOAD CONCENTRATION (g/day) TO WWTS (mg/l) (%) Cr(III) <0.1 <5 <0.1 Anionic surfactants 44 3915 8.2 Non ionic surfactants 51 4590 6.2 Total surfactants 99 8775 7.2 Flow rate 0.6 (m3/day x shop) 81 (m3/day) 0.33 *based on average + 10% Conclusions Cases like the Vicenza district, with artisanal activities deeply rooted in residential areas are common in Italy and in other EU regions. As shown in the validation case of Valdagno, in other districts the contribution of artisanal activities to the pollutant load of the UWW system may be lower. The principal pollutants originating from these artisanal shops are potentially toxic elements such as Cd, Ni, Pb, Cu, and Zn, and surfactants. The main conclusion of this study was that, by segregating concentrated liquid wastes, the contribution of artisanal activities to the pollutant load was dramatically reduced, at least for potentially toxic elements. However, only some of the artisanal activities in this Case Study practised wastewater segregation. One issue raised by artisanal representatives was the economic cost of segregation of wastewater. It is felt that the stringent environmental requirements concerning wastewater from Italian artisanal shops, considered industrial wastewater, do not compensate efforts for waste segregation. Artisanal activities that did not practice segregation of concentrated liquid wastes include goldsmiths, hairdressers and food manufacturing shops, which are also the main enterprises responsible for the surfactant load in wastewater. As a consequence, neglecting the contribution of segregated liquid wastes did not significantly affect the total load of surfactants from artisanal shops. This load typically represents one fourth of the total surfactant load to the WWTS. It may be anticipated that, if careful segregation of the concentrated liquid wastes were extended to all the artisanal activities, a dramatic decrease of potentially toxic element and surfactant load from the artisanal activities would be obtained. Even though the wastewater flow rate from artisanal shops would not decrease significantly, the pollutant load could be reduced to negligible values with respect to the total pollutant load. 146 Section 6. Case Studies GOLD JEWELLERY PRODUCTION IN ITALY- BEST ENVIRONMENTAL PRACTICE Introduction In Italy there are about 6,000 small to medium gold and jewellery manufacturing shops, most of which are concentrated in three main production districts: Arezzo (Tuscany), Vicenza (Veneto) and Valenza Po (Piemonte). In the period 1995-1999 the Italian Government and the Association of Artisanal Activities sponsored a large research programme, carried out by the National Research Council, in support of Craft Goldsmiths Production and Trade. The programme tackled problems relating to: · innovation in production cycles, · fast analytical tools for the assay of precious metals and their alloys, · safety and health of artisanal workers, · environmental impact of gold manufacturing shops. The Italian Water Research Institute (IRSA) carried out a survey on management practices for wastewater, produced in small to medium gold manufacturing shops in Arezzo (Marani, 1997). According to the Association of Goldsmiths, the results obtained in the survey of Arezzo district may confidently be extended to draw conclusions about national practices in goldsmiths’ shops. There is no information on the losses of Au, Ag, PGMs from the gold and jewellery shops to the wastewaters in this research programme. More data on PGM is presented in Case Study (a). The gold manufacturing district of Arezzo Gold manufacturing processes The most prevalent processes are those starting from wire or plate to produce rings, chains and medals. Hollow bars may be used to produce lighter objects. Gold or silver goods may also be produced using micro-casting or electrolytic processes (electro-forming). All production cycles have common final steps: object assembly, polishing, finishing and cleaning. Wastewater origin and characteristics Different production cycles and processes generate wastewater in the gold manufacturing shops. Casting operations to prepare wire or plate do not typically require aqueous solutions, with exception of small volumes for washing crucibles. In contrast, the preparation of hollow bars requires nitric acid, hydrochloric acid, caustic soda, and ammonia solutions. The wastewater resulting from micro-casting comes from water used to break the gypsum mould and rinse waters. The “gypsum” waters are segregated from the other wastewater produced in the workshops and recycled after a settling step. Wastewater from electro-forming is derived from specific activities, as well as from operations common to other processes. The former wastes may be highly turbid wastewater, exhaust baths and rinsing waters, mainly derived from the galvanic cycle. Wastewaters characterised by high turbidity are filtered to eliminate the suspended material and then re- circulated several times before collecting them with the other wastewater. Regarding galvanic waters, acid wastes are separated from the wastes containing cyanide. The concentrated cyanide baths are collected by external firms, whereas the diluted rinse waters can be either pre-treated in a separate circuit, then added to the main wastewater stream or 147 Section 6. Case Studies re-circulated after elution through an ion exchange column (anionite). The concentrated cyanide solutions produced by column regeneration are sent to external firms. In the final steps of assembly and finishing, the operations producing liquid wastewater are: · acid pickling, · galvanic treatments, · surface shining, · washing steps. Usually spent pickling baths are treated by external firms. The waters derived from surface shining generally contain metal powder. In addition, the final step generates large amounts of wastewater as well as surfactants present in the spent baths. Finally hand and floor washing waters, do not typically require pre-treatment and could be sent to the sewer. However as they may contain gold and silver powder they are not directly discharged into the sewer. Instead they are sent to the internal wastewater treatment plant. Here the insoluble precious metals are concentrated in the sludge which, is dried and sent to specialised firms for recovery of precious metals. Wastewater may be divided into four classes: Soapy waters contain high concentrations of detergents, along with fatty substances and metal powder. Other pollutants such as phosphates and ammonia typically originate from the detergents used in these workshops. The acids contained in the acid wastewater are: sulphuric acid, nitric acid, hydrochloric acid, hydrofluoric and fluoboric acid. This wastewater may also contain high concentrations of potentially toxic elements such as copper, zinc, iron and nickel. “Gypsum” waters containing suspended gypsum particles, are recycled several times after sedimentation of the suspended material. Then they are committed, together with the sedimented gypsum, to external firms for final disposal. The cyanide waters contain free cyanide and soluble cyano-complexes of gold and silver. For safety reasons these waters are treated separately to oxidize the cyanide before sending them to the wastewater treatment plant. 148 Section 6. Case Studies Wastewater management and fate About 40% of the goldsmiths’ workshops of Arezzo province do not declare any wastewater production. These workshops: · have a particular production step that does not produce wastewater; · accumulate little quantities of wastewater to be treated by external firms specialized in wastewater treatments; · treat their own wastewater with systems that permit the complete recycle of the treated water in the productive cycle; · evaporate the wastewater, obtaining a concentrate to dispose with other solid wastes. Regarding the remaining 60% of workshops with authorisation to discharge their wastewater, Table c.28 reports the number (and relative percentage) of workshops that discharge their wastewater into the UWW collecting system or into surface waters. In terms of flow rate, the 694 workshops discharge about 121600 m3.year-1 of wastewater. Assuming an average number of 6.3 workers per unit, the average specific wastewater flow rate is about 0.5 m3.week-1 per worker. Table c.28 Destination of wastewater of workshops having discharge authorisation Destination No. of workshops % of workshops UWW collecting 567 81.7 system Surface waters 29 4.2 Unknown 98 14.1 TOTAL 694 100 Survey on a representative group of goldsmiths’ shops The workshops were selected with the aim of choosing representative establishments in terms of number of workers and in terms of type of manufacturing processes. Twelve small to medium workshops were selected for the study. The survey included both interviews on production processes, wastewater flow rate, wastewater treatment, direct sampling and analysis of treated and untreated wastewater. Typically, several wastewater samples were collected from each workshop, in March, May and October of 1996. The chemical characterisation of samples was performed analysing a large number of parameters, including: boron, potentially toxic elements like Cd, Cr, Cu, Ni, Pb, and Zn, and surfactants. Conclusions In examining the pollutant concentrations in these wastewaters (considering the results obtained for the 12 shops sampled), the pollutants most often detected with concentrations higher than admissible limits for discharge are: surfactants, copper, zinc, cadmium and boron. Surfactants, copper and zinc are detected in all samples whereas boron and cadmium are present in 75 % and 60% of samples respectively. Surfactants, derived generally from washing processes, are present in wastewater with an average concentration of 34 mg.l-1 MBAS and a maximum value of 118.5 mg.l-1 MBAS. The average boron concentration found in the wastewater of small jewellery shops examined is 13.5 mg.l-1, with a maximum value of 100 mg.l-1. The presence of boron in these wastewaters may be due to processes such as soldering (where boron is used as borace), voiding and washing, or through the use of other materials e.g. hydrofluoboric acid, detergents. For potentially toxic elements, the average concentrations of copper and cadmium detected in wastewater are 14.2 mg.l-1 Cu and 0.4 mg.l-1 Cd, with maximum values of 61 mg.l-1 Cu and 1 mg.l-1 Cd. Zinc concentration is highly variable, which is probably due to the different manufacturing steps of the shops, with an average value of 22 mg.l-1 but a maximum value of 270 mg.l-1. 149 Section 6. Case Studies (D) Pharmaceuticals in the Urban Environment Introduction Pharmaceutical substances are a group of compounds, which until recently, have not been of major concern with regards to their environmental effects. These compounds are developed for their biological effect (primarily in humans), to cure disease, fight infection or reduce symptoms. If these substances enter the environment they may have an effect on aquatic and terrestrial animals, due to these biological properties and the fact that some of them may bioaccumulate. Unlike other organic compounds, such as PCBs whose use has been discontinued over the last 20 years, pharmaceuticals are used widely and they and their metabolites may easily enter the UWW system. Pharmaceuticals use is also expected to increase in Europe with the increasing avearge age of the population. Current research demonstrates that drugs and their metabolites entering water supplies and the food chain may pose a real threat, both to the ecosystem and to human health, and risk assessments are slowly being carried out. However, many problems must be overcome, such as the fact that these compounds are very changeable and are usually present in mixtures and at low concentrations. Furthermore, pharmaceuticals have a wide variety of structures and activities and that they may act synergistically (Alcock, 1999). There are many different pharmaceuticals substances and approximately 3000 pharmaceutical compounds are discharged into UWW collecting systems (ENDS, 2000). Sewage sludge is predominantly disposed of on agricultural land, as is manure from farms and both of these products will contain large amounts of pharmaceutical substances. Unfortunately, very little is known about the fate of these compounds in the environment and the potential long-term impacts. Many pharmaceutical substances though, have the same characteristics as organic compounds; i.e. they are lipophilic, which tends to be a requirement to be able to pass membranes, and some are designed to be persistent so that they are not inactivated before achieving their healing effect (Halling-Sørensen, 1998). Sources and fate in the environment There are two major groups of pharmaceuticals; human and veterinary drugs, and they will enter the environment through different pathways (Figure d.1). A large amount of pharmaceutical products from both categories are prescribed each year. For example in Germany, 100 tonnes of human drugs were prescribed in 1995 (Ternes, 1998c). This probably reflects the amounts prescribed in other countries of Western Europe, relative to population size. Over the counter pharmaceutical sales will also increase this figure. It is likely that a high concentration of drugs may find their way into wastewater, making wastewater and sewage sludge major vectors for the entry of these compounds into the environment. However, this will depend on the chemico-physical behaviour of the pharmaceuticals in question. 150 Section 6. Case Studies Figure d.1: Scheme for the main fates of drugs in the environment after application [after Ternes, 1998c.] The main entry routes of pharmaceutical substances into the environment are through; disposal of wastewater treatment end products: effluent and sewage sludge; and manure spreading onto agricultural land or even from the excreta of grazing animals (Figure d.2). Fish farms also use medical substances as feed additives, but most of the food is not eaten and is deposited straight onto the sea-bed (Halling-Sørensen, 1998). 151 Section 6. Case Studies Figure d.2: Anticipated exposure routes of veterinary and human medicinal substances in the environment [after Halling-Sorensen, 1998]. The compounds For the purpose of risk assessments, pharmaceutical compounds have been divided into four activity groups: antibiotics, antineoplastic drugs, antiparacetic drugs, and hormone disrupters. However, there are many other groups of pharmaceutical compounds, such as lipid regulators, which have been found ubiquitously in the environment and are highly persistent compounds (Daughton, 1999). 152 Section 6. Case Studies Antibiotics: Antibiotics are widely used as medicines for human and animals treatment, also used widely as growth promoters in veterinary use. According to the Swiss environmental research institute, in the EC 54000 tonnes of antibiotics were used in human medicine in 1997. Veterinary use amounted to 3500 tonnes of medicines and 1600 tonnes of growth promoters (ENDS, 2000). Due to the effect of bans the use of growth promoters is expected to decline. According to a study carried out by Halling-Sorensen (1998), most antibiotics are not very persistent in the environment, particularly in soils, and the most widely used growth promoters have been shown to have no effect on invertebrates, even at relatively high concentrations. However soil bacteria may be more sensitive (ENDS, 2000). Veterinary drugs tend to end up in manure and so have the potential to contaminate soils where manure or slurry is spread. Levels of antibiotics in soil around a pig farm studied reached up to 1400mg.kg-1, due to presence of antibiotics in the animals’ feed (ENDS, 2000). The increased use of antibiotics has led to an increase in drug resistant micro flora. This resistance is actually favoured by low concentrations of antibiotics (Jorgensen, 2000), thus, the presence of antibiotics in the environment may be an important problem. Anti-neoplastic drugs These anti-cancer drugs are mainly used in hospitals rather than in households. They are primarily used for chemotherapy and are found sporadically, in a range of concentrations in the environment (Daughton, 1999). Anti-cancer drugs act as non-specific alkylating agents, which means that no receptors are required, hence, they have the potential to act as mutagens, carcinogens, teratogens, and embryotoxins (Daughton, 1999). The most widely used substance is cyclophocamide (CP). In Denmark, approximately 13 to 14 kg of CP is used in hospitals each year, and approximately 6 kg are prescribed by pharmacies (Christensen, 1998). Thus, it is assumed that a total of 20 kg are used per year (Christensen, 1998). Anti-cancer drugs are also referred to in the Case Study on Platinum Group Metals. Analgesic drugs Analgesic drugs are used for pain relief and are probably the most commonly used medicines. Endocrine-disrupting substances There is increasing concern about compounds that interfere with the hormonal system. Endocrine disrupting substances block or trigger oestrogenic effects by binding to receptors. Receptor-specific responses are particularly problematic as they can affect people for which they are not intended (Christensen, 1998). An endocrine-disruptor may have an ‘agonistic effect’ where it binds to the receptor instead of the natural hormone and causes a response, or it may have an ‘antagonistic effect’ where the binding of the compound prevents the natural one from binding and producing the required response (Environment Agency of England and Wales, 1998). Other effects may also occur, showing that the process is very complex and affects many systems in the body. Endocrine-disrupting non pharmaceutical substances include phthalates, some PCBs, and some pharmaceutical compounds, such as oestrogens. Many of these may be persistant in sewage sludge and could enter the food chain as they are potentially taken up by plants and animals. The effects of endocrine-disruptors were discovered about 10 years ago and may occur in concentration ranges of a few nanograms per litre (Jørgensen, 2000). Humans use hormones to cure diseases, as well as in contraception and hormonal replacement therapy. Table d.1 shows some categories of substances with endocrine-disrupting properties: 153 Section 6. Case Studies Table d.1: Categories of substances with endocrine-disrupting activities [Environment Agency, 1998]. Category Examples Uses Modes of action Natural phytoestrogens Isoflavones, lignans Present in plants Oestrogenic and anti- oestrogenic Female sex 17b-oestradiol, Produced in animals Oestrogenic hormones oestrone Man-made Polychlorinated PCBs, dioxins By-products from Anti-oestrogenic organic incineration and chemical compounds processes Organochlorine DDT, dieldrin, lindane Insecticides Oestrogenic and anti- pesticides oestrogenic Alkylphenols Nonylphenol Production of NPE and Oestrogenic polymers Alklphenol NonylPhenol Surfactant Oestrogenic ethoxylates Ethoxylate (NPE) Phthalates Dibutyl phthalate Plasticiser Oestrogenic Bi-phenolic Bisphenol A In polycarbonate plastics Oestrogenic compounds and epoxy resins Synthetic steroids Ethinyl oestradiol Contraceptives Oestogenic Of these, oestrogens are of a major concern, as they are excreted in an inactive form but are found to be reactivated in sewage effluent. Oestrogens are organic molecules derived from cholesterol, which can bind to receptors and cause a physiological response (Montagnani et al., 1996). The purpose of the endocrine system is to regulate metabolic activity, which requires a degree of interaction with the nervous system (Montagnani et al., 1996). Because oestrogen receptors are located in the cell nucleus, oestrogen-like molecules can thus enter the cell and could potentially interact with DNA, causing damage which may lead to tumour formation (Montagnani et al., 1996). Prolonged exposure to these compounds may induce female characteristics in males. There is increasing speculation that these compounds may be linked to reduction in male fertility and reproductive complications (Montagnani et al., 1996). In the UK, research has shown that male fish exposed to the natural hormone: 17b- oestradiol, oestrone, and the synthetic hormone: ethinyloestradiol, from domestic sewage effluent, developed hermaphrodite characteristics (Alcock et al., 1999). It was also found that these hormones were present in the biologically active form, having been transformed and reactivated after excretion and not degraded during wastewater treatment (Alcock et al., 1999). However, research carried out by the Ministry of Agriculture, Food and Fisheries (MAFF) along the river Lea, UK determined that although estrogenic substances are likely to be present in wastewater effluents, the development of female characteristics in male fish present in WWTS lagoons is unlikely to be due to these substances, as the transformation is only possible at a very early stage in their development (Montagnani et al., 1996). Due to the widespread use of estrogenic substances and their entry into the environment via sludge and effluent from WWTS, aquatic environments may be acting as a sink for these substances. Natural and synthetic estrogens are extremely widely present/used. Although the contribution these compounds make to oestrogenic effects is thought to be small, they are are of concern because of their highly persistent and potent nature (ENDS, 2000). 154 Section 6. Case Studies Consumption of pharmaceutical compounds As already stated, it is difficult to obtain information on the quantities of pharmaceuticals used for most countries but in Denmark consumption of the most commonly used drugs is available (Table d.2). Table d.2: Major drugs and drug groups and their consumption in Denmark, in 1997 [ENDS, 2000, Christensen, 1998, and Halling-Sørensen, 1998]. Active ingredient Major use Amount used (kg) Aspirin Analgesic 305250 Paracetamol Analgesic 248250 Ibuprofen Anti-rheumatic 33792 Pencicillin V Antibiotic 19000 Furosemide Diuretic 3744 Terbutaline Anti-asthmatic 475 Enalapril Anti-hypertensive 416 Citalopram Anti-depressant 368 Diazepam Anti-depressant 207 Salbutanol Anti-asthmatic 170 Bendroflumethiazide Diuretic 167 Zopiclone Anti-hypertensive 144 Amlodipidine Anti-hypertensive 132 Oestradiol Hormone replacement 119 Nitrazepam Anti-hypertensive 116 17b-estradiol Oral contraception 45 Budesonide Anti-asthmatic 39 Gestodene Birth control pill 37 Cyclophosphamide Anti-cancer 20 Xylometazoline Nasal decongestant 13 Digoxin Heart drug 4 Desogestrel Birth control pill 3 Medicine groups Antibiotics 37700 Analgesics 28300 Hypotensiva 410 Diuretica 3800 Anti-asthmatics 1700 Psychleptics 7400 It was also found that a total of 110 tonnes of antibiotics were used as growth promoters, feed additives or as medicines, on livestock and fish farms (Halling-Sørensen, 1998). In 1994, the overall production of antibiotics in Germany was 1831 tonnes, of which Penicillin contributed 624 tonnes (Hirsch, 1999). It should be noted that the amount of antibiotics used for human and veterinary purposes, 37.7 tonnes and 110 tonnes respectively, is in the same range as the amounts of certain pesticides used (Hirsch, 1999). A paper published by Goll van (1993) estimates that if the total amount of growth promoters used in the Netherlands was spread on their 2 million hectares of agricultural land, this would give a yearly average of 130mg of antibiotics and metabolites/m2 (Halling-Sørensen, 1998). The pharmaceutical substances predominantly used in hospitals must also take into account compounds such as X-ray contrast media. Iodinated X-ray contrast media are very stable biochemically, so they tend to be excreted unmetabolised. In Germany, 500 tonnes per year of X-ray contrast media are used and iopromide (CAS 73334-07-3) alone accounts for 130 tonnes per year (Ternes, 2000). Other X-ray contrast substances used in the EU are: 155 Section 6. Case Studies diatrizoate (CAS 131-49-7) an ionic X-ray diagnostic drug, iopamidol (CAS 60166-93-0) and iopromide (CAS 73334-07-3) both non-ionic X-ray diagnostic substances, iothalamic acid (CAS 2276-90-6) and ioxithalamic acid (CAS 28179-44-4) both ionic X-ray diagnostic substances. Detection of pharmaceutical compounds Other biologically active compounds are common in wastewater, and there is increasing evidence that such substances are widely present in the environment. Danish research has found that up to 68 different drug residues can be detected in the environment. Compounds such as caffeine, nicotine, aspirin, and paracetamol are all frequently detected (ENDS, 2000). More studies are now looking at the prevalence of these drugs and their metabolites in wastewater and sewage sludge. For example, clofibric acid (2-(4-chlorophenoxy)-2- methylpropionic acid), which is a breakdown product of lipid-regulating drugs, is highly persistent and resistant to wastewater treatment, as usually only 15-51% is removed (ENDS, 2000). Lipid regulating drugs are commonly prescribed and it is thought that the daily load of clofibric acid to UWW collecting systems in Denmark, is around a few kilograms (ENDS, 2000). Concentrations of clofibric acid have also been detected in sewage effluent in micrograms per litres, and in nanograms per litres in water bodies such as rivers and lakes (ENDS, 2000). In the UK, clofibric acid has been detected in the 1 mg.l-1 range in the aquatic environment, and in Germany, it has been detected at concentrations up to 165 ng.l-1 (Ternes, 2000). Antibiotics have been also widely detected. In Germany, concentrations up to 5 mg.l-1 were found in WWTS effluents, which is comparable to the data collected by Richardson and Bowron in 1985 (Hirsch, 1999). Five of the 18 compounds investigated were frequently detected in German WWTS effluent and rivers: erythromycin, roxithromycin, clarithromycin, sulfamethoxazole, and trimethoprim. The highest concentration was detected for erythromycin degradation products, at a median value of 2.5 mg.l-1 in WWTS effluent and a maximum value of 6 mg.l-1 (Hirsch, 1999). The other four antibiotics were only detected at levels below 1 mg.l-1 (Hirsch, 1999). Median values for the concentrations detected in surface waters are one order of magnitude lower than those detected in WWTS effluent (Hirsch, 1999), (see Figure d.3). 156 Section 6. Case Studies Figure d.3: Presence of antibiotics from investigated surface waters in Germany [Hirsch, 1999]. Analyses for tetracycline and penicillins found no detectable amounts in five WWTS effluents or surface waters (Hirsch, 1999). Tetracyclines tend to form stable complexes with calcium and other ions, thus contaminating the sediment rather than the water (Hirsch, 1999). Penicillins tend to be easily eliminated, as they are very susceptible to hydrolysis of the b- lactam ring (Hirsch, 1999). Ground water samples were also slightly contaminated by sulphamethoxazole and sulphamethazine (not used in human medicine), due to infiltration from application of contaminated sewage sludge or manure to agricultural land. The samples collected contained concentrations of sulphonamide residues up to 0.48 mg.l-1 (Hirsch, 1999). X-ray contrast media are also widespread in German wastewater influent and effluent. Loads of the most frequently used compounds, such as iopromide were found to exceed 1mg.l-1 during the working week but decreased at weekends as X-rays did not tend to be performed (Ternes, 2000). Maximum concentrations detected were greater than 3 mg.l-1 (up to 15 mg.l-1 for iopamidol). Median values were around 0.25-0.75 mg.l-1, which indicates their ubiquity in German wastewater treatment effluents (Ternes, 2000). The compounds detected depended on the region and on the practices of particular hospitals in that area. Antiseptics are another major group of pharmaceutical compounds commonly used both in households and in medical practices. It has been found that major antiseptics, such as chlorophene and biphenol, are present at concentrations up to 0.05 mg.l-1 in wastewater (Ternes, 1998b). However, biphenol tends to be eliminated at rates of 98% during treatment and chlorophene at 63% (Ternes, 1998b). These compounds, particularly clorophene, are detected in rivers at similar concentrations. This is probably due to the fact that they are also used in many household detergents and disinfectants, as well as in veterinary medicine on farms, so leading to widespread contamination of the aquatic environment. Analgesics. Salicylic acid, a major metabolite of acetylsalicylic acid, is also detectable in German wastewater at high concentrations (54mg.l-1 over 6 days), but treatment degrades most of it, as the compound is no longer detectable in WWTS effluents (Ternes, 1998b). 157 Section 6. Case Studies Oestrogens. There is increasing concern about the widespread presence of oestrogenic substances in wastewater and other water bodies. The daily production rate of natural oestrogens by humans is in the microgram range, up to 400 mg of 17b-estradiol for women (Ternes, 199b). The maximum daily excretion rate is 64 mg for oestriol (Ternes, 1999b). Oestrogens are mainly excreted as inactive polar conjugates (Ternes, 1999b). Vitellogenin (precursor for production of yolk in all oviparous vertebrates) induction in male or juvenile fish has become a “biomarker” for the presence of estrogenic substances in the aquatic environment (Larsson, 1999). In the UK, caged fish downstream of an WWTS were found to produce vitellogenin. Two possibilities were investigated: the presence of ethinylestradiol and NPE. The effluent from a Swedish WWTS showed high levels of oestrogenic compounds (Larsson, 1999). It was found that exposure to large amounts of ethinyloestradiol caused accumulation in fish, as fish concentrations were found to be 104- 106 times higher than those detected in the water (Larsson, 1999). The estimated use of ethinyloestradiol is 3.5mg/day, which is close to the concentration found in the WWTS effluent (2.9mg/day), showing very low degradation of this compound during treatment (Larsson, 1999). In the UK, analysis of sampled effluents found that natural hormones (17b-oestradiol and oestrone) are present in the range 1.4 to 76 ng.l-1, whereas the synthetic hormone (ethinyloestradiol) was only found in 3 out of the 7 effluents analysed and at comparatively low levels: 0.2 to 7 ng.l-1 (Alcock et al., 1999). The source of these is thought to be mainly from human excretion products. It was also found that the hormones were present in the biologically active form, suggesting that they had been transformed and reactivated after excretion (Alcock et al., 1999). In Germany, raw sewage was found to contain 0.015 mg.l-1 of 17b-estradiol and 0.027 mg.l-1 of oestrone and it was also found that oestrone and 17a -ethinyloestradiol were not efficiently removed during wastewater treatment (Ternes, 1999a). In contrast, 17b-oestradiol and 16a - hydroxy-oestrone were eliminated with a higher efficiency: around 64-68% (Ternes, 1999a) (See Figure d.4). Figure d.4: Elimination percentages and loads of estrogens during passage through a municipal sewage treatment plant located near Frankfurt/Maine over 6 days [Ternes, 1999a]. In discharges from the treatment plants, all compounds could be detected in the ngl-1 range (Ternes, 1999a), however oestrone was predominant with concentrations up to 0.07 mgl-1 158 Section 6. Case Studies and a median value of 0.009 mgl-1. The compounds 17a -ethinyloestradiol and 16a -hydroxy- estrone were found at the detection limit of 0.001 mg l-1 (Ternes, 1999a). Oestrone was the only compound detected in 3 of the 15 rivers sampled at concentrations between 0.7 and 1.6 ng l-1 (Ternes, 1999a). Therefore, it seems that these compounds, particularly natural oestrogens, are not degraded in the treatment system and tend to accumulate in sludge and effluent. However, the loads entering receiving waters are quite low. DEHP is also an endocrine-disrupting compound. In Sweden, it has been found in all sewage sludge samples analysed, with concentrations between 25-660 mg kg-1 dry weight (Alcock et al., 1999). In Italy, drinking water, rivers, and sediments have been analysed to determine the extent of environmental contamination by pharmaceuticals (see Table d.4), (Zuccato, 2000). Table d.4: Concentrations of medicinal drugs in drinking water, river water, and sediments [Zuccato, 2000]. -1 -1 -1 DRUG DRINKING WATER (ngl ) RIVER WATER (ngl ) RIVER SEDIMENTS (ng kg ) Po (Piacenze Po Lambro Adda Lambro Adda Milan Lodi* Varese and, (Piacenze, (Milan)* (Sondrio) (Milan) (Sondrio) Cremona)* Cremona)* Atenolol It can be seen that most drugs were measurable in these media, showing widespread contamination. The concentrations measured could potentially give rise to human exposure in the ng day-1 range, which is 3-4 orders of magnitude lower than the concentrations capable of producing pharmacological effects (Zuccato, 2000). Hence, acute exposure is assumed to be unlikely, but long-term effects must still be studied. In Germany, occurrence of drugs in WWTS and rivers has been studied (Ternes, 1998c). Results showed maximum values in average loads of up to 3kg.day-1 for salicylic acid in the influent and up to 114g.day-1 for carbamazepine in the effluent (Ternes, 1998c). 159 Section 6. Case Studies Figure d.5: Elimination of different drugs during passage through a municipal sewage treatment plant near Frankfurt/Maine over 6 days [Ternes, 1998c]. From Figure d.5, it appears that more than 60% of the compounds in the influent were usually removed during treatment of wastewater: ranging between 7-99% removal (Ternes, 1998c). Only carbamazepine, clofibric acid, and phenanzone were less efficiently eliminated (Figure d.6). However, complete elimination was not usually achieved; thus receiving waters may potentially be contaminated. Subsequently, a screening programme of 49 different German WWTS effluents was carried out, in addition to river sampling. Lipid regulating agents were found in the majority of the WWTS effluents, and in many river samples but in a much lower concentration (Ternes, 1998c). Polar metabolites of the compounds were usually detected. For example, clofibric acid was detected at levels up to 1.6 mg.l-1 in WWTS effluent and in the ng.l-1 range in rivers, which illustrates the importance of metabolites (Ternes, 1998c). Anti-inflammatories , such as, ibuprofen and naproxen, were also detected. Diclofenac was present in the highest concentration at median levels of 0.81 mg.l-1 in treated effluent effluent and 0.15 mg.l-1 in rivers (Ternes, 1998c). In the case of betablockers, the highest median concentration was found for metoprolol at 0.73 mg.l-1 in WWTS effluent and 0.45 mg.l-1 in rivers (Ternes, 1998c). b2-sympathomimetics were also present but in very low concentrations. Anti-cancer agents such as cyclophosphamide and ifosamide were detected at levels of 0.02 mg.l-1 and 0.08 mg.l-1 respectively in WWTS effluent; however, they are associated with presence of hospital effluents, and are not widespread (Ternes, 1998c). Carbamazepine, an anti-epileptic drug was widespread in the aquatic environment, with a high median value of 2.1 mg.l-1 in effluent and 0.25 mg.l-1 in rivers (Ternes, 1998c). Annual prescriptions of carbamazepine amount to approximately 80 tonnes per year in Germany, it then becomes metabolised and glucuronides are excreted. However treatment of wastewater cleaves these metabolites back to the parent compound, increasing the environmental concentrations (Ternes, 1998c). Table d.6 shows specific studies on pharmaceuticals in the environment and the concentrations found for the different substances (Halling-Sørensen, 1998). 160 Section 6. Case Studies Ground water pollution has been detected in some instances, mainly due to leaching from landfill sites containing pharmaceutical wastes (Halling-Sørensen, 1998). In Berlin, clofibric acid has been detected in drinking water at concentrations between 10 ng.l-1 and 165 ng.l-1 and in all surface water samples around Berlin, suggesting extensive contamination (Halling- Sørensen, 1998). River water is often polluted with pharmaceutical compounds. Most groups of compounds, i.e. antibiotics, antineoplastic agents, and ethinyloestradiol, have been detected between 5- 10 ng.l-1. A study conducted by Richardson and Bowron (1985), investigated the exposure of human pharmaceuticals in the river Lea in England and found that over 170 substances are used in excess of 1 tonne per year in the river’s catchment. This allowed them to predict a concentration of at least 0.1mg.l-1 in the river water (Halling-Sørensen, 1998). 161 Section 6. Case Studies Table d.6: Pharmaceutical compounds identified in environmental samples [after Daughton, 1999] Compound Use/Origin Environmental occurrence Acetaminophen Analgesic/anti- Removed efficiently by WWTS, max. effluent 6mgl- inflammatory 1, not detected in surface waters Acetylsalicylic acid Analgesic/anti- Ubiquitous, removal efficiency 81%, max. effluent inflammatory 1.5mgl-1, in surface water 0.34mgl-1. Betaxolol betablocker Max. effluent 0.19mgl-1, in surface water 0.028mgl- 1. Benzafibrate Lipid regulator Removal efficiency 83%, max. effluent 4.6mgl-1, in surface water 3.1mgl-1. Biphenylol Antiseptic, fungicide Extensive removal in WWTS. Bisoprolol betablocker Max. effluent 0.37mgl-1, in surface water 2.9mgl-1. Carazolol Betablocker Max. effluent 0.12mgl-1, in surface water 0.11mgl- 1. Carbamazepine Analgesic, anti- Removal efficiency 7%, max. effluent 6.3mgl-1, in epileptic surface waters 1.1mgl-1. Chloroxylenol Antiseptic In influents and effluents <0.1mgl-1. Chlorophene Antiseptic Influent 0.71mgl-1, removal not very efficient Clenbuterol b2-sympathomometic Max. effluent 0.08mgl-1, in surface waters 0.05mgl- 1. Clofibrate Lipid regulator River water 40ngl-1, not detected in effluent or surface waters. Clofibric acid Metabolite of Removal efficiency 51%, max. effluent 1.6mgl-1, clofibrate surface waters 0.55mgl-1, up to 270ngl-1 in German tap waters Cyclophosphamide antineoplastic Max. effluent 0.02mgl-1, not detected in surface waters, high in hospital sewage: up to 146ngl-1 Diatrizoate X-ray contrast media Resistant to biodegradation, median in German surface waters 0.23mgl-1, locally very high concentrations can occur. Diazepam Psychiatric drug Max. effluent 0.04mgl-1, not detected in surface waters. Diclofenac-Na Analgesic/anti- Removal efficiency 69%, max. effluent 2.1mgl-1, in inflammatory surface waters 1.2mgl-1. Dimethylaminophen Analgesic/anti- Removal efficiency 38%, max. effluent 1mgl-1, in azone inflammatory surface waters 0.34mgl-1. 17a -ethinylestradiol Oral contraceptive Up to 7ng.l in WWTS effluent, not detected in German surface waters above 0.5ngl-1. Etofibrate Lipid regulator Not detected in WWTS effluent and surface waters Fenfluramine Sympathomimetic No studies but is known to be an endocrine- amine disrupting substance Fenofibrate Lipid regulator Efficiently removed, max. effluent 0.03mgl-1, not detected in surface waters. Fenofibric acid Metabolite of Removal efficiency 64%, max. effluent 1.2mgl-1, in fenofibrate surface waters 0.28mgl-1. Fenoprofen Analgesic/anti- Not detected in WWTS effluent or surface waters inflammatory Fenoterol b2-sympathomometic Max. effluent 0.06mgl-1, in surface waters 0.061mgl-1. Flurorquinolone Antibiotics Ubiquitous, led to resistance in pathogenic carboxylic acids bacteria, strongly sorbs to soil. Fluoxetine Antidepressant No studies Fluvoxamine Antidepressant No studies 162 Section 6. Case Studies Gemfibrozil Lipid regulator Removal efficiency 69%, max. effluent 1.5mgl-1, in surface waters 0.51mgl-1. Gentisic acid Metabolite of Efficiently removed by WWTS, max. effluent acetylsalicylic acid 0.59mgl-1, in surface waters 1.2mgl-1. o-hydroxyhippuric Metabolite of Efficiently removed by WWTS, not detected in acid acetylsalicylic acid effluent or surface waters. Ibuprofen Analgesic/anti- Removal efficiency 90%, max. effluent 3.4mgl-1, in inflammatory surface waters 0.53mgl-1. Ifosamide Antineoplastic Max. effluent 2.9mgl-1, not detected in surface waters, hospital sewage 24ngl-1, totally refractory to removal by WWTS. Indomethacine Analgesic/anti- Removal efficiency 75%, max. effluent 0.60mgl-1, inflammatory in surface waters 0.2mgl-1. Iohexol X-ray contrast media Very low aquatic toxicity. Iopamidol X-ray contrast media Max. effluent 15mgl-1, median 0.49mgl-1. Iopromide X-ray contrast media Resistant to biodegradation, yields refractory, unidentified metabolites, max. effluent 11mgl-1. Iotrolan X-ray contrast media Very low aquatic toxicity. Ketoprofen Analgesic/anti- Max. effluent 0.38mgl-1, in surface waters 0.12mgl- inflammatory 1. Meclofenamic acid Analgesic/anti- Not detected in WWTS effluent or surface waters. inflammatory Metoprolol betablocker Removal efficiency 83%, max. effluent 2.2mgl-1, in surface waters 2.2mgl-1. Nadolol Betablocker Max. effluent 0.06mgl-1, not detected in surface waters. Naproxen Analgesic/anti- Removal efficiency 66%, max. effluent 0.52mgl-1, inflammatory in surface waters 0.39mgl-1. Paroxetine antidepressant No studies Phenazone Analgesic Removal efficiency 33%, max. effluent 0.41mgl-1, in surface waters 0.95mgl-1. Propranolol Betablocker Removal efficiency 96%, max. effluent 0.29mgl-1, in surface waters 0.59mgl-1. Propyphenazone Analgesic/anti- Prevalent in Berlin waters. inflammatory Salbutamol b2-sympathomometic Max. WWTS influent 0.17mgl-1, in surface waters albuterol 0.035mgl-1. Salicylic acid Metabolite of Up to 54mgl-1 in WWTS effluent but efficiently acetylsalicylic acid removed in effluent, average in effluent 0.5mgl-1, in surface waters 4.1mgl-1. Sulfonamides Antibiotics Present inn landfill leachates Terbutaline b2-sympathomometic Max. effluent 0.12mgl-1, not detected in surface waters. 3,4,5,6-tetrabromo- Antiseptic, fungicide Found in influents and effluents in Germany o-cresol <0.1mgl-1. Timolol Betablocker Max. effluent 0.07mgl-1, in surface waters 0.01mgl- 1. Tolfenamic acid Analgesic/anti- Not detected in WWTS effluent or surface waters. inflammatory Triclosan Antiseptic 0.05-0.15mgl-1 in water, very widely used. Verapamil Cardiac drug No occurrence data 163 Section 6. Case Studies Table d.6 summarises the occurrence of certain pharmaceuticals in the environment. It can be seen that some pharmaceuticals, such as lipid regulators, X-ray contrast media, antibiotics etc., are ubiquitous and extremely persistent in the environment, some are even present in drinking water: clofibric acid for example has been found at concentrations up to 0.27 mg.l-1 in some German waters. This would breach EC regulations if the compounds were classed as pesticides (ENDS, 2000). However only a fraction of the drugs on the market, have been investigated regarding their occurrence in the environment. The fate of pharmaceutical compounds Pharmaceutical compounds enter the body and are then often metabolised in the liver through oxidation, reduction, or hydrolysis to “phase I” metabolites, which tend to be more toxic than the parent compound. Other reactions, such as conjugation, metabolise the compounds into “phase II” metabolites, which tend to be inactive and more polar and water- soluble. It has also been observed that phase II metabolites are often reactivated into the parent compounds, either during treatment of wastewater and sewage sludge or in the environment. For example, chloramphenicol glucoronide and N-4-acetylated sulphadimidine (phase II metabolites of the antibiotics chloramphenicol and sulphadimidine, respectively), are reactivated in liquid manure (Halling-Sorensen, 1998). This shows the importance of the investigation of metabolites as well as parent compounds. For example, 17b-oestradiol, is administered orally and mainly undergoes first-pass hepatic metabolism, being transformed to oestrone and oestriol, which are less potent (Christensen, 1998). Other metabolites are also formed but to a lesser extent. Experiments using the diluted slurry of activated sludge from a WWTS, were undertaken to investigate the persistence of natural oestrogens and contraceptives under aerobic conditions (Ternes, 1999b). The natural oestrogen 17b-oestradiol, was oxidised to oestrone, which is then linearly removed with time. Rapid elimination also occurred for16a -hydroxy-oestrone. However, the contraceptive 17a -ethinyloestradiol was persistent and highly stable under environmental conditions (Ternes, 1999b). Two glucuronides of 17b-oestradiol were cleaved to their parent compounds and 17b-oestradiol was re-released in an activated form (Ternes, 1999b). This indicates that the microorganisms present have the ability to deconjugate oestrogen glucuronides. It is interesting to note that glucuronide conjugates are the main oestrogen metabolites excreted by humans, so during wastewater treatment, the concentration of free oestrogen increases due to the cleavage of the glucuronide moieties from the compounds. As a result, the predominant presence of oestrone in WWTS effluents and rivers is due to; its high stability during treatment; the cleavage of glucuronide conjugates from oestrone and 17b-oestradiol; and the oxidation of the latter to oestrone (Ternes, 1999b). Penicillin antibiotics are eliminated rapidly and have short half-lives in the body, usually 30- 60 minutes, and very high concentrations are excreted in urine: it has been determined that up to 40% of penicillin V is excreted unchanged (Christensen, 1998). Cyclophosphamide, an anti-cancer drug is administered intravenously or orally. It is not active in itself but undergoes activation in the body when transformed to phosphoramide mustard and acrolein. The parent compound is genotoxic. Some of it is excreted unchanged: 5-20% (Christensen, 1998). Antibiotics are generally believed to leave humans unchanged by the body metabolism (see Table d.7) (Hirsch, 1999) and it has been determined that up to 90% of the parent compounds are excreted unchanged (ENDS, 2000). These active products can be excreted either as unchanged compounds or as conjugates; 30-90% of administered antibiotics are excreted via urine as active substances (Alcock et al., 1999). This introduces the problem at 164 Section 6. Case Studies the WWTS of disruption of biological treatment processes, as pharmaceutical compounds, particularly antibiotics, can potentially affect bacteria. Table d.7: Human prescription amounts and excretion rates of antibiotics [Hirsch, 1999]. Antibiotic Amount prescribed Excretion (%) (t/a) Unchanged Other Metabolites Amoxicillin 25.5-127.5 80-90 10-20 Ampicillin 1.8-3.6 30-60 20-30 Penicillin V 40 40 60 Penicillin G 1.8-3.6 50-70 30-50 Sulphamethoxasole 16.6-76 15 Trimethoprim 3.3-15 60 Erythromycin 3.9-19.8 >60 Roxithromycin 3.1-6.2 >60 Clarithromycin 1.3-2.6 >60 Minocycline 0.8-1.6 60 40 Doxycycline 8-16 >70 A study looking at the amounts of antibiotics in human faeces found trimethoprim and doxycycline at concentrations between 3-40 mg.kg-1, and erythromycin at concentrations around 200-300 mg.kg-1 (Hirsch, 1999). Elimination at treatment plants is usually incomplete, ranging between 60-90% (Ternes, 1998b). Polar antibiotics are probably not removed efficiently because elimination is mainly due to adsorption on activated sludge, which is mediated through hydrophobic interactions (Hirsch, 1999). As a consequence, receiving waters and other environmental media may become contaminated. Furthermore, erythromycin and other drugs such as naproxen and sulphasalazine, have survived in the environment for over a year (Zuccato, 2000). Clofibric acid was also found to survive for 21 years and although its use has been stopped, it is still detected in rivers and lakes in Italy (Zuccato, 2000). Many pharmaceutical compounds have the same physico-chemical characteristics as organic compounds, such as persistence and lipophilicity; much less is known though about their entry into the environment and their subsequent fate (Alcock et al., 1999). Over 30% of all drugs produced between 1992 and 1995 were lipophilic, i.e. solubility less than 100 mg.l-1 (Halling-Sørensen, 1998). The fate of pharmaceutical substances may be divided into three groups: · Mineralisation to CO2 and water, for example aspirin. · Retained in sludge, if the compound is lipophilic and not readily biodegradable. · Emitted to receiving water due to transformation into a more hydrophilic form but still persistent, for example clofibrate. Richardson and Bowron (1985) investigated degradation of pharmaceuticals during wastewater treatment and found that many common compounds are biodegradable, although cortisteroid compounds and ethinyloestradiol, among others, were non- biodegradable (Alcock et al., 1999). In work by Kummerer (2000), two clinically important groups of antibiotics have been studied with regards to their biodegradability. Chinolones and nitromidazoles possess different chemical structures, actvity spectra and modes of action. The study found low rates of biodegradation for ciprofloxacin, ofloxacin, and metronidazole; it also found that the genotoxicity of these compounds remained unaffected during treatment (Kummerer, 2000). 165 Section 6. Case Studies The different groups of antibiotics were active against bacteria present in wastewater (Kummerer, 2000). Only a few compounds have been studied regarding their behaviour in wastewater treatment and the results are varied. Compounds such as the analgesics, ibuprofen and naxoproxen, have been found to have removal efficiencies between 22-90% and 15-78% respectively (ENDS, 2000). It has also been determined that 70-80% of the drugs administered in fish farms, are transferred into the environment (Halling-Sorensen, 1998). Table d.6 gives an overview of the present knowledge on the environmental fate of specific pharmaceuticals. It can be seen that most hormones, such as oestrogen, are persistent in all areas and that most of the antibiotics used for human treatment are not biodegradable. The majority of other compounds used for human treatment are also non-biodegradable, with the exception of the following: paracetamol, nicotinamide, ibuprofen, caffeine, and aspirin. The compounds used in veterinary treatment tend to be more biodegradable than human pharmaceuticals, although the speed of degradation will depend on environmental conditions, such as pH, and temperature. It has also been determined that iodinated X-ray contrast media are not degraded during wastewater treatment, due to their high polarity (log Kow of iopromide = -2.33) (Ternes, 2000). These compounds are designed to be highly stable to give optimum results during X-ray, so are not readily biodegradable. Ninety percent of X-ray contrast media are excreted unmetabolised (Ternes, 2000); hence, receiving waters will also tend to be contaminated. Concentrations up to 0.49 mg.l-1 for iopamidol were detected in receiving rivers (Ternes, 2000). It appears that groundwater may also become contaminated, as concentrations of up to 2.4 mg.l-1 were identified for iopamidol as a result of infiltration by polluted surface water (Ternes, 2000). The concentration of X-ray contrast media in receiving water bodies is lower than that detected in wastewater effluent; however, due to the high persistence of such media in the environment, this reduction seems to be less important than for other pharmaceutical compounds (Ternes, 2000). This shows that pharmaceutical compounds have the ability to infiltrate aquifers and survive for many years. Pentobarbital, clofibric acid, benzafibrate, diclofenac, and carbamezepine, have all been found in aquatic environments, persisting there for up to 20 years (Ternes, 2000). With the exception of Denmark, there is very little data available on the use and quantities of pharmaceuticals, which renders the task of studying the fate of these compounds in wastewater very difficult. The list of pharmaceutical substances could be exhaustive and prioritisation is necessary. Certain physical processes occur that may be used to degrade these contaminants: sorption to solids, volatilisation, chemical degradation, and biodegradation. The effectiveness of sorption and volatilisation can be determined using the octanol water partition coefficient (Kow) and Henry's law constant (Hc) [Rogers, 1996]: · if log Kow is less than 2.5, the compound has a low sorption potential (i.e. it will not adsorb onto soil particles and will not be very lipophilic), · if log Kow is between 2.5 and 4, the compound has a medium sorption potential, · if log Kow is greater than 4, the compound has a high sorption potential and is very lipophilic. -4 -9 · if Hc is greater than 1x10 and Hc/Kow is greater than 1x10 , the compound is thought to have a high volatilisation potential, -4 -9 · if Hc is less than1x10 and Hc/Kow is less than 1x10 the compound is thought to have a low volatilisation potential. 166 Section 6. Case Studies Several models, using Kow, Koc, and Hc, have been developed in order to take all these characteristics into account in order to prioritise pollutants. This has enabled an assessment to be made of the exposure risk to such pollutants, through consuming food derived from sludge-amended soil. Two parameters are important when trying to determine movement of contaminants through the food chain: persistence, and non-polarity. Easily metabolised or polar compounds do not move through the food chain. As bioconcentration increases with lipophilicity, compounds with high log Kow values will tend to accumulate in the food chain (Duarte-Davidson, 1996). Unfortunately, sewage sludge may contain a wide variety of pharmaceutical compounds, and for many, information on their characteristics is not readily available. Therefore, it becomes difficult to eliminate or prioritise pharmaceuticals using this screening process. In addition, without proper information on the physico-chemical properties of these compounds, it is not possible to predict their fate in the environment, or their concentrations in sewage sludge (Alcock et al., 1999). In the absence of detailed knowledge, it can be presumed that most pharmaceutical substances have the same properties as pesticides and other organic pollutants (Halling-Sorensen, 1998). Also, as mentioned already, many of these compounds have lipophilic characteristics, so they are likely to accumulate. Some pharmaceuticals may even be metabolised during treatment or in the soil, to more readily available compounds, increasing their potential for plant and animal uptake (Engwall et al., 2000). The low concentration of individual pharmaceutical compounds, coupled with their metabolic characteristics leads to incomplete removal in WWTS (Daughton, 1999). They tend to be non-volatile, so transport and movement through the environment will occur via aquatic media. In fact, their polarity and non-volatile characteristics will often prevent them from leaving the aquatic environment (Daughton, 1999). As it has been seen earlier, metabolites also tend to be cleaved to the parent compound during wastewater treatment and then released afterwards. Nutraceuticals/Herbal Remedies During the last several years, the popularity of nutritional supplements was codified by the creation of a new term for the subclass of highly bioactive food supplements called nutraceuticals (Daughton and Ternes, 1999) also referred to as nutriceuticals. Nutraceuticals are a rapidly growing commercial class of bioactive compounds, usually botanicals, intended as supplements to the diet. Nutraceuticals and many herbal remedies can have potent physiologic effects. These are a mainstay of alternative medicine and have enjoyed explosive growth in use in the United States and Europe during the last decade. Many are used as food supplements that have either proven or hypothesized biologic activity but are not classified as drugs by the FDA, primarily because a given botanical usually has not one but an array of distinct compounds whose assemblage elicits the putative effect and because these arrays cannot be easily standardized. As such, they are not regulated and are available over the counter (heavily promoted via the Internet). Even in those cases in which the natural product is identical to a prescription pharmaceutical (e.g., the Chinese red- yeast product Cholestin newly introduced to the United States contains lovastatin, an active ingredient in the approved prescription drug Mevacor used to lower cholesterol levels), a recent ruling (Borman , 1998; Zeissel, 1999). prevented the FDA from regulation. The significance of dietary supplements in the United States led to the creation of the Office of Dietary Supplements (ODS) via the DSHEA in 1995 under the National Institutes of Health (NIH) (DSHEA, 1994). The ODS maintains a searchable database (International Bibliographic Information on Dietary Supplements [IBIDS]) of published scientific literature on dietary supplements (NIH Office of Dietary Supplements, 1999). 167 Section 6. Case Studies Although these substances are readily available off the counter, not always in a characterized/standardized forms, an effort is underway to patent various nutraceuticals by standardizing the extracts and thereby making them available only by prescription. The patenting of hundreds of multiple-molecule nutraceuticals for therapeutic purposes could lead to more widespread use of these substances. As an example, a recent addition to this class is a substance called huperzine A, an alkaloid extracted from a Chinese moss, which has been documented to improve memory. It is therefore experiencing strong demand for treating Alzheimer's disease and has captured the attention of those who follow the nutraceutical market because of its true pharmaceutical qualities. The significance of this particular compound is that it possesses acute biologic activity as a cholinesterase inhibitor identical to that of organophosphorus and carbamate insecticides. It is so effective that the medical community is concerned about its abuse/misuse, especially since it is legal. While huperzine A, and alkaloids in general (compounds with heterocyclic nitrogen, proton-accepting group, and strong bioactivity), are naturally occurring compounds, their susceptibility to biodegradation in WWTS or in surface waters is unknown. This is the case for almost all nutraceuticals, therefore more research is needed. Another example is Kava, which is prepared from the root of Piper methysticum, used of its mild narcotic effect among other effects. The active ingredients in Kava are believed to be a suite of lipophilic lactones comprising substituted -pyrones (methysticin, kavain, yangonin, and others) (Shao, et.al.,1998). These compounds display a host of effects in humans, but nothing is known about their effects on other organisms or fate in WWTS. There are many nutraceuticals, both new and ‘traditional’, experiencing increased consumption. These few examples illustrate the unknowns regarding whether these compounds are being excreted, surviving WWT, and then having possible effects on aquatic organisms. Nutraceuticals and herbal remedies would have the same potential fate in the environment as pharmaceuticals, with the added dimension that their usage rates could be much higher, as they are readily available and taken without the controls of prescription medication. However, because these compounds are natural products, they would be expected to biodegrade more easily . Legislation and policy for risk assessment At the beginning of the 1980s, environmental risk assessment was introduced for new chemicals but it took a decade later for drugs to be included in the discussion. In Europe, since the 1990's, there has been a distinction made between compounds for human use and those used in veterinary practice. For several years legislation has been implemented for veterinary medicines. The EU Directive 81/852/EEC, (Amended 1993), introduced the requirement for a tiered environmental risk assessment of new veterinary products, and attempts are being made to implement this for a review of existing substances (ENDS, 2000). Currently, environmental risk assessment consists of examining the likely environmental sectors and if levels of pharmaceutical compounds exceed the trigger values set, such as 100 mg.l-1 in manure, further data is required (ENDS, 2000). The technical directive [Directive 81/852/EEC, amended 1993] concerning veterinary medical products outlines the basic requirements for conducting an environmental risk assessment (Halling- Sorensen, 1998). The technical directive [Directive 75/318/EEC, amended 1993] concerning human medical products does not refer to any ecotoxicology or ecotoxicity tests and no guidance is given on how to carry out an environmental risk assessment for drugs used by humans. However, a draft directive for human pharmaceuticals is currently being devised, proposing that risk assessment should be part of the approved procedure of new medical substances (Halling-Sorensen, 1998). The EC is proposing a similar programme for human 168 Section 6. Case Studies medicines: if drug concentrations in surface waters are predicted to exceed 0.01 mg.l-1, toxicity testing is required to find the no effect level (NOEC) (ENDS, 2000). Although environmental assessment of the potential impacts of newly developed drugs has been expected in the EU since the 1st of January 1995 (Christensen, 1998), it is should be noted that the end point of human exposure is not usually investigated. Also, assessment of individual compounds is usually based on a limited number of tests but pharmaceuticals in the environment may affect a large number of different organisms and species, so this ability should be reflected in the tests carried out (Stuer-Laurisden, 2000). Pharmaceuticals may not affect the standard test species and give rise to false negative results (Stuer-Laurisden, 2000). A risk assessment study was carried out in Germany looking at salicylic acid, paracetamol, clofibrinic acid, and methotrexate (Henschel, 1997). As seen previously in this Case Study, these compounds were present in the environment and had toxic effects in at least one standard ecotoxicological test. The most sensitive reaction however, was to a non-standard test incorporating relevant end points for the pharmaceuticals (Henschel, 1977), so proving the limitations of standard tests. Risk assessment of pharmaceuticals Risk assessment for pharmaceuticals in the environment has not usually been carried out due to the lack of data and the need for more precise and sensitive measures in the environmental sectors. Some high consumption compounds, such as antibiotics and clofibric acid, are being released into the environment and have been found to be widely present in aquatic environments, sediments and soils. Although the liver often metabolises pharmaceutical compounds to more easily hydrolysed compounds, the metabolites can be cleaved back to the more hydrophobic, active parent compounds by bacteria, which can then persist and bioaccumulate. There is a strong opinion that there are more pressing environmental problems than pharmaceuticals and that these compounds do not pose a large risk because they are present in such low concentrations (ng.l-1), with most effects only seen in the mg.l-1 range (ENDS, 2000). As already stated though, disease resistance to pharmaceuticals is favoured by low concentration exposure and compounds such as antineoplastic agents and hormones have effects at very low levels. The effects of active compounds in the low, ng.l-1 range, cannot be excluded, as experience with pesticides shows, impacts can be significant at low levels (Stuer-Laurisden, 2000). At the moment, most toxicity tests performed investigate acute impacts on specific species. However, as most compounds in question are persistent and are discharged continuously into the environment at low levels, it would appear to be more relevant to look at the chronic, long-term impacts of exposure to low concentrations over all trophic levels. Some of the long-term, chronic impacts that may be of concern are genotoxicity and reproduction impairment. It has been found that Daphnia are tolerant to most antibiotics within the mg.l-1 range but that exposure to these levels over several weeks causes death, probably because of toxic effects in the food organisms (ENDS, 2000). The effects of continuous exposure to even low levels of pharmaceuticals in the environment are very complex and affect many different organisms. More studies are necessary with regards to the long-term impacts and the potential synergistic effects of exposure to a mixture of drugs. Exposure route is very important in determining environmental loading, as the dose and duration of exposure are important parameters in risk assessment. Drugs tend to be released in low concentrations, although local discharges, such as those coming from hospitals, may have higher concentrations (Jorgensen, 2000). 169 Section 6. Case Studies A study carried out in Italy (Zuccato, 2000), found that most drugs were present in drinking water, river water and sediments, but that human exposure would only be in the ng day-1 range, which is much lower that the concentration where effects are expected to be observed. The study concluded that risk seems negligible but possible long-term exposures and impacts still need investigation. The same was found with X-ray contrast media, which were ubiquitous in the aquatic environment and highly persistent. No acute toxic effects were observed, in Daphnia magna or in bacteria, algae, fish, and crustaceans (Ternes, 2000). Long-term exposure was not investigated. Risk assessments were carried out on three pharmaceuticals, using the computer program EUSES, which was developed as a support to the technical guidance document for risk assessment on new and existing substances (Christensen, 1998). The three compounds assessed were; the synthetic estrogen, 17a -ethinylestradiol; the antibiotic, penicillin V; and the antineoplastic agent, cyclophosphamide. This program estimated environmental fate and human exposure based on worst-case scenarios, using data on the physico-chemical properties of the compounds and amounts consumed. The results indicated that for all three there was negligible human risk. However, the author stressed the point that many uncertainties are associated with this method and that the drugs, although seemingly insignificant, still contribute to the total toxic load in the environment, and that interactions may have ecotoxicological impacts. Research in Denmark attempted to carry out a risk assessment for the 25 most highly used drugs in the primary health sector in Denmark, including furosemide, paracetamol, ibuprofen, and estradiol (Stuer-Laurisden et al., 2000). Different parameters are used to calculate environmental exposure: biodegradation, bioaccumulation, and bioavailability are three of the most important. Nevertheless, it has been seen that biodegradation of pharmaceuticals does not happen very often. Bioaccumulation in the human body does not happen for drugs, as they are metabolised to more polar compounds that can then easily be excreted. The bioavailability of drugs is different if it is bound to solids, adsorbed, or dissolved. However, as seen above, the octanol-water partition coefficient can be used to determine bioaccumulation, and other parameters can be used to determine bioavailability. However, this information on the properties of the compounds is not readily available. They found that ecotoxicology data was available only for 6 of the 25 compounds, and biodegradation data only for 5 (Tables d.9 and d.10) (Stuer-Laurisden et al., 2000). Predicted environmental concentrations (PEC) should be determined for the system where it is anticipated that the highest values would be found, i.e. aquatic ecosystems and sewage sludge (Jorgensen, 2000). In order to do so, modelling could be a useful tool; however, not many models have yet been developed and validated due to the lack of data in this area (Jorgensen, 2000 and Halling-Sorensøn,1998). The predicted environmental concentrations for the 6 compounds were calculated and all exceeded 0.001mg.l-1, which is the cut off value in EU legislation for carrying out more investigations. The majority of the PECs are between 1-100ng.l-1, it is only for the top 5 compounds that these reach the mg.l-1 range (Stuer- Laurisden, 2000). The predicted no effect concentration (PNEC) is based on ecotoxicological data and the PEC/PNEC ratio was found to exceed one only for ibuprofen, paracetamol, and acetylsalicylic acid and below one for estrogen, diazepam, and digoxin (Stuer-Laurisden, 2000). This showed that data is only partially available, preventing complete risk assessments. Nevertheless, it was concluded, with this data, that ibuprofen, aspirin, and paracetamol may pose a risk; hence, contradicting other studied that had concluded these were efficiently removed by treatment and did nor reach hazardous levels in the environment (ENDS, 2000). The efforts are hampered by the fact that concentrations measured in sludge and effluent vary extensively, and furthermore comparisons of predicted concentrations in sludge based on Kow, sludge-water partition coefficients (Kd), or acid-base constants (pKa) also reveal large variations (Stuer-Laurisden, 2000). 170 Section 6. Case Studies The cost/benefit stage of risk assessment is extremely important: the indirect and direct effects of the drug on the environment and the human body must be known in order to make an informed decision (Jorgensen, 2000 and Halling-Sørenson,1998). This can allow selection of a substitute drug that has the same benefits but fewer environmental impacts. Hence, lack of data for toxicity but also environmental concentrations prevent calculation of risks. Examples of good environmental practice In France, and the UK, there are procedures for returning prescribed but unused pharmaceuticals. In France, these plans are encouraged by the ADEME "RETOUR" initiative (ADEME, 1997a), which also encourages the distributor to include the costs of collection and treatment into the product's selling price. This strategy is useful for reducing the amount of polluted domestic and artisanal (laboratories, photographic shops etc) wastewater through special collections for specific pollutants such as thermometers, medicines, and paint leftovers. Most areas in France have implemented such programmes and they are successful. There are also possibilities of reformulating and substituting certain pharmaceutical compounds with substances incurring fewer impacts. It has been found that cyclophosphamide and ifosfamide, the very widely used antineoplastic drugs, act through an active metabolite, which is highly unstable. German researchers have detected another compound that has the same therapeutic activity but that is much more readily biodegraded (ENDS, 200). However, in order to research more environmentally suitable drugs, the characteristics of the existing ones must be known, and, as yet, there is still very little data available. If the assessment of a drug gives a high risk, the response may not have to be the phasing out of the drug, but maybe just the collection and specific treatment of the faeces and urine containing this compounds (Jorgensen, 2000). Environmental risk assessment should be part of the development of all new drugs and could be used as a marketing tool, as public concern for the environment is increasing (Jorgensen, 2000). Analysis tools and their sensitivity must be improved for the determination of the very low concentrations of drugs. A method obtaining detection limits within the ngl-1 range for various pharmaceutical compounds, particularly neutral basic drugs such as betablockers, has been developed, using advanced solid phase extraction, modified derivatisation procedures and LC-electrospray-MS/MS detection (Ternes, 1998a). This allows detection down to 10ng.l-1, in different aqueous matrices. In another study, determination limits down to 5ng.l-1 were achieved for phenolic compounds and other acidic drugs, such as lipid regulators and acetylsalicylic acid metabolites, using solid phase extraction and methylation or acetylation of the carboxylic and phenolic hydroxyl groups, followed by detection by GC/MS (Ternes, 1998b). In Sweden, a pharmaceutical company, AstraB, was discharging very toxic effluents containing a large amount of persistent organic pollutants and phosphorus, which had caused operational problems at the WWTS (Rosen, 1998). There were large variations in the composition of their effluent over time, as the drugs tend to be produced in discontinuous batches. Hence, a broad and flexible treatment method had to be introduced to treat the wastewater at source. The wastewater was investigated and it was found that the main contribution came from the treatment of packages containing non-approved liquid pharmaceutical preparations (Rosen,et.al. 1998). The washing water had a very high toxicity and could not be treated biologically. This effluent was removed and incinerated and the remaining effluent still too toxic for discharge into the UWWT system is now treated using a 171 Section 6. Case Studies multi-stage biofilm process removing all organic matter and the toxicity is no longer measurable (Rosen, et.al. 1998). Conclusion and Recommendations Pharmaceutical compounds must become priority substances in the same way as persistent organic pollutants are. Judgements on the relative priorities are based on the knowledge at the time and the priority list will obviously change over time as more studies are carried out and more data is gathered. Pharmaceuticals are widely used and mainly disposed of through the sewerage system, allowing their entry into the environment continually, as removal rates can be compensated by replacement rates (Daughton and Ternes, 1999). They are concerning because they are biologically active and are usually lipophilic and potentially bioaccumulating. Many resist biodegradation, within the WWTS and in the environment, and can end up in surface and ground waters, as well as sediment and soils and are found to be highly stable under environmental conditions. Furthermore, metabolites tend to be cleaved and transformed back to the parent compounds once in the environment, increasing the concentrations and justifying the importance of metabolites. Many can have unpredicted and unknown side effects particularly after long-term exposure to low concentrations. Aquatic ecosystems are the most vulnerable, as this is the main environmental compartment where pharmaceuticals are found ubiquitously. Analytical methods must be improved to detect pharmaceuticals at very low levels and sampling procedures must be of very high quality so as not to cause contamination. More information on the physicochemical, ecotoxicity, and ecotoxicological characteristics, using appropriate tests that better accommodate subtle end points, of drugs and their metabolites should be obtained in order to allow environmental risk assessments to be carried out. Furthermore, this may lead to the validation of modelling techniques that could speed up the whole process. Furthermore, use patterns of drugs in all countries is still very limited and should be determined, as it is essential for the elaboration of amounts released into the environment. Screening of high use drugs should be carried out and samples with high potential should then be subject to more analyses (Daughton and Ternes, 1999). Furthermore, a more precautionary view on the potential impacts of the drugs should be adopted and more studies are required to elucidate these effects at the concentrations observed and also investigating additive and synergistic effects of mixtures (Daughton and Ternes, 1999). Risk assessment for pharmaceuticals should include an assessment of their biodegradability and environmental fate and potential impact as occurs for other discharged substances, such as detergent residues. Furthermore, the disposal of unwanted drugs into the wastewater system from domestic sources should be discouraged by encouraging collection of these wastes. 172 Section 6. Case Studies (e) Personal Care Products, Fragrances in Urban Waste Water and Sewage Sludge Personal care products are defined as chemicals marketed for direct use by the consumer (excluding off the counter medication with documented physiologic effects) and having intended end uses, primarily on the human body (products not intended for ingestion) or in the household. In general, these chemicals alter odour, appearance, touch, or taste without displaying significant biochemical activity (Daughton and Ternes, 1999). Most of these chemicals are used as the active ingredients or preservatives in cosmetics, toiletries and fragrances. They are not used for treatment of disease, but some may be intended to prevent diseases (e.g., sunscreen agents). In contrast to drugs, almost no attention has been given to the environmental fate or effects of personal care products, the focus has traditionally been on the effects from intended use on human health. Personal care products differ from pharmaceuticals in that large amounts can be directly introduced to the environment and unlike medicinal compounds, there are rarely recommended doses. These products can be released directly into recreational waters or volatilised into the air. Because of this direct release they can bypass possible degradation in UWWT. Also, in contrast to pharmaceuticals, less is known about the effects of this broad and diverse class of chemicals on non-target organisms, such as aquatic organisms. Data are also limited on the potential adverse effects on humans. For example, common sunscreen ingredients, 2-phenylbenzimidazole-5-sulfonic acid and 2-phenylbenzimidazole, can cause DNA breakage when exposed to UV-B (Stevenson and Davies, 1999). The quantities of personal care products produced commercially can be very large. For example, in Germany alone the annual output was estimated to be 559,000 tonnes for 1993 (Statistisches Bundesamt, 1993). A few examples are given below of common personal care products that are ubiquitous pollutants, which may possess varying degrees of bioactivity. Table e.1 Personal care and fragrances produced in Germany (1993) Product category Tonnes Bath additives 162 300 Shampoos, hair tonic 103 900 Skin care products 75 500 Hair sprays, hair dyes, setting lotions 71 000 Oral hygiene products 69 300 Soaps 62 600 Sun screens 7 900 Perfumes, aftershaves 6 600 TOTAL 559 100 · Preservatives Parabens (alkyl-p-hydroxybenzoates) are one of the most widely and heavily used types of antimicrobial preservatives in cosmetics (skin creams, tanning lotions, etc.), toiletries, pharmaceuticals, and even foodstuffs (up to 0.1% wt/wt). Although the acute toxicity of these compounds is very low, Routledge et al.[1998] report that these compounds (methyl through butyl homologs), display weak oestrogenic activity. Although the risk from dermal application in humans is unknown, the probable continual introduction of these benzoates into wastewater treatment systems and directly to recreational waters from the skin, leads to the question of risk to aquatic organisms. Butylparaben showed the most competitive binding to the rat oestrogen receptor at concentrations one to two orders of magnitude higher than that of nonylphenol and showed oestrogenic activity in a yeast oestrogen screen at 10-6 M . 173 Section 6. Case Studies · Disinfectants/Antiseptics Triclosan, a chlorinated diphenyl ether: 2,4,4´-trichloro-2´-hydroxydiphenyl ether, is an antiseptic agent that has been widely used for almost 30 years in a vast array of consumer products. Its use as a preservative and disinfectant continues to grow; for example, it is incorporated at < 1% in Colgate's “Total” toothpaste, the first toothpaste approved by the FDA to fight gingivitis. While triclosan is registered with the U.S. EPA as a pesticide, it is freely available over the counter. Triclosan's use in commercial products includes footwear (in hosiery and insoles of shoes called Odour-Eaters), hospital hand-soap, acne creams (e.g., Clearasil), and rather recently as a slow-release product called Microban, which is incorporated into a wide variety of plastic products from children's toys to kitchen utensils such as cutting boards. Many of these uses can result in direct discharge of triclosan to UWW collecting systems, and as such this compound can find its way into receiving waters depending on its resistance to microbial degradation. Okumura and Nishikawa [1996] found traces of triclosan ranging from 0.05 to 0.15 µg.l-1 in water. Although triclosan has long been regarded as a biocide (a toxicant having a wide-ranging, nonspecific mechanism(s) of action - in this case gross membrane disruption) McMurry et al. [1998] report that triclosan actually acts as an antibacterial, having particular enzymatic targets (lipid synthesis). As such, bacteria could develop resistance to triclosan. As with all antibiotics in the environment, this could lead to development of resistance and change in microbial community structure (diversity). A wide range of disinfectants are used in large amounts, not just by hospitals but also by households and livestock breeders. These compounds are often substituted phenolics as well as other substances, such as triclosan. Biphenylol, 4-chlorocresol, chlorophene, bromophene, 4-chloroxylenol, and tetrabromo-o-cresol [Ternes et al 1998] are some of the active ingredients, at percentage volumes of < 1-20%. A survey of 49 WWTPs in Germany [Ternes et al 1998] routinely found biphenylol and chlorophene in both influents, up to 2.6 µg/L for biphenylol and up to 0.71 µg.l-1 for chlorophene, and effluents. The removal of chlorophene from the effluent was less extensive than for biphenylol, with surface waters having concentrations similar to that of the effluents. · Sunscreen Agents The occurrence of sunscreen agents (UV filters) in the German lake Meerfelder Maar was investigated by Nagtegaal et al. [1997]. The combined concentrations of six sunscreen agents (SSAs) identified in perch (Perca fluviatilis) in the summer of 1991 were as high as 2.0 mg.kg-1 lipid and in roach (Rutilus rutilus L) in the summer of 1993, as high as 0.5 mg.kg- 1 lipid. Methylbenzylidene camphor (MBC) was detected in roach from three other German lakes. These lipophilic SSAs seem to occur widely in fish from small lakes used for recreational swimming. Both fish species had body burdens of SSA on par with PCBs and DDT. The bioaccumulation factor, calculated as quotient of the MBC concentration in the whole fish (21 µg.kg-1) versus that in the water (0.004 µg l-1), exceeded 5,200, indicating high lipophilicity. The fact that SSAs (e.g., 2-hydroxy-4-methoxybenzophenone [oxybenzone] and 2-ethylhexyl-4-methoxycinnamate) can be detected in human breast milk (16 and 417 ng.g-1 lipid, respectively) [Hany et al 1995] shows the potential for dermal absorption and bioconcentration in aquatic species. No data have been published on newer SSAs such as avobenzene (1-[4-(1,1-dimethylethyl)phenyl]-3(4-methoxyphenyl)-1,3-propanedione). · Perfume Ingredients The raw ingredients in perfumes include essential oils, plant extracts and animal secretions, and synthetic or semi synthetic (natural material that has undergone some chemical modification) compounds. Thousands of these substances can be blended to create perfumes. These can be used directly as perfumes or as scents in other products, for 174 Section 6. Case Studies example in cosmetics, cleaning agents and air fresheners. Perfume ingredients may enter the urban wastewater system directly from domestic sources, such as in the washing agents or from being washed off skin in the case of perfumes and cosmetics. The organic compounds found in perfumes that may be of environmental or health concern include - nitro-musk compound, - polycyclic musk compound, - solvents and fixatives - other fragrances. Details of the physical properties of these compounds are included in Appendix B. Health and environmental effects of the compounds discussed are also briefly introduced within this case study. This case study will focus on musk compounds. Fragrances (musks) are ubiquitous, persistent, bioaccumulative pollutants that are sometimes highly toxic; amino musk transformation products are toxicologically significant. Synthetic musks comprise a series of structurally similar chemicals (which emulate the odour of the more expensive, natural product, from the Asian musk deer), used in a broad spectrum of fragranced consumer items, both as fragrance and as fixative. Included are the older, synthetic nitro musks (e.g., ambrette, musk ketone, musk xylene, and the lesser known musks moskene and tibetene) and a variety of newer, synthetic polycyclic musks that are best known by their individual trade names or acronyms. The major musks used today are, the polycyclic musks (substituted indanes and tetralins), which account for nearly two-thirds of worldwide production and the inexpensive nitro musks (nitrated aromatics), accounting for about one-third of worldwide production. These substances are used in nearly every commercial fragrance formulation (cosmetics, detergents, toiletries) and most other personal care products with fragrance; they are also used as food additives and in cigarettes and fish baits (Gatermann, et.al. 1998) The nitro-musks are under scrutiny in a number of countries because of their persistence and possible adverse environmental impacts and therefore are beginning to be phased out in some countries. Musk xylene has proved carcinogenic in a rodent bioassay and is significantly absorbed through human skin; from exposure to combined sources, a person could absorb 240 µg/day [Bronaugh et al 1998]. The human lipid concentration of various musks parallels that of other bioaccumulative pollutants, such as PCBs [Schmid 1996]. Worldwide production of synthetic musks in 1988 was 7000 tonnes [Gattermann et al 1998] and worldwide production for nitro musks in 1993 was 1,000 tonnes, two-thirds of which were musk xylene [Kfferlein 1998] Synthetic musks first began to be identified in environmental samples almost 20 years ago [Yamagishi 1981 and 1983]. By 1981, Yamagishi et al. had identified musk xylene and musk ketone in gold fish (Carassius auratus langsdorfii) present in Japanese rivers and soon after [Yamagishi et al 1983] in river water, wastewater, marine mussels (Mytilus edulis), and oysters (Crassosterea gigas). This was followed by a number of studies in Europe, some of which are summarised in table e.2. 175 Section 6. Case Studies Table e.2 Concentrations of various musk compounds in environmental samples. Location Tissue/Substance Product Concentration Reference North Germany Freshwater Fish Musk Xylene 10-350mg.kg-1 Geyer et al, 1994 (Fillet) Musk Ketone 10-380mg.kg-1 Geyer et al, 1994 Ruhr River, Bream and Perch Galaxolide, Average Eschke et al 1998 Germany (Fatty tissue) Tonalide and concentrations Celestolide between 2.5 and 4.6 mg.kg-1 (ppm) Berlin, Germany Surface waters Galaxolide, Maximum Herberer et al Tonalide and concentrations 1999 Celestolide above 10m L-1 Elbe River, Particulate matter Musk ketone 4-22 ng/g Winkler et al. 1998 Germany from river samples Galaxolide 148-736 ng/g Tonalide 194-770 ng/g Italy Freshwater Fish Galaxolide, 4 ng.g-1 – Draisci et al 1998 (Fatty tissue) Tonalide 1054 ng.g-1 Musks are refractory to biodegradation (other than reduction of nitro musks to amino derivatives), which explains why they have been detected in water bodies throughout the world [Gattermann et al 1998]. They also are very lipophilic [octanol-water partition coefficients are similar to those for DDT and hexachlorocyclohexane , Winkler et al, 1998] and therefore can bioaccumulate, leading to very high concentrations being measured in some studies. The values for the three most prevalent musks in the Elbe river study (table e.2) were within the same order of magnitude as those for 15 polycyclic aromatic hydrocarbons (PAHs) and exceeded those for 14 common polychlorinated organic pollutants (only hexachlorobiphenyl [HCB] and p,p´-DDT were of similar concentration). Also, all the 31 water samples contained musk ketone (2-10 ng.l -1), Galaxolide (36-152 ng.l -1), and Tonalide (24-88 ng.l-1); Celestolide was found only at 2-8 ng.l-1. These higher values exceeded those for all the polychlorinated organics and the PAHs. The occurrences of individual musks are sometimes correlated as a result of their use as mixtures in commercial products. In Germany, the nitro musks are being replaced by the polycyclic musks, therefore resulting in lower concentrations for musk ketone [Winkler et al, 1998]. Although the significance of the aquatic toxicity of the nitro and polycyclic musks is debatable (genotoxicity from the polycyclics seems not to be a concern) [Kevekordes 1998], the aminobenzene (reduced) versions of the nitro musks can be highly toxic. These reduced derivatives are undoubtedly created under the anaerobic conditions of sewage sludge digestion. Behecti et al. [1998] tested the acute toxicity of four reduced analogs of musk xylene on Daphnia magna. The p-aminodinitro compound exhibited the greatest toxicity of the four, with extremely low median effective concentration (EC50) values averaging 0.25 µg.l-1 (0.25 ppb). Recently, the amino transformation products of nitro musks were identified in wastewater treatment effluent and in the Elbe River, Germany. Gatermann et al. [1998] identified musk xylene and musk ketone together with their amino derivatives 4- and 2-amino musk xylenes and 2-amino musk ketone. In wastewater entering treatment plants, the concentrations of musk xylene and musk ketone were 150 and 550 ng.l-1, respectively. In the effluent, their concentrations dropped to 10 and 6 ng.l-1, respectively. In contrast, although the amino derivatives could not be detected in the influent, their concentrations in the effluents dramatically increased, showing extensive transformation of the parent nitro musks: 2-amino musk xylene (10 ng.l-1), 4-amino musk xylene (34 ng.l-1), and 2-amino musk ketone (250 ng.l-1). It was concluded that the amino derivatives could be expected in wastewater effluent at concentrations more than an order of magnitude higher than the parent nitro musks. In the 176 Section 6. Case Studies Elbe, 4-amino musk xylene was found at higher concentrations (1-9 ng.l-1) than the parent compound. Amino nitro musk transformation products are · more water soluble than the parent musks, · still have significant octanol-water partition coefficients (high bioconcentration potential), · more toxic than the parent nitro musks, therefore more attention should be focused on these compounds. Because synthetic musks are ubiquitous; used in large quantities; introduced into the environment almost exclusively via treated wastewater; and are persistent and bioconcentratable, they are prime candidates for monitoring in both water and biota as indicators for the presence of other personal care chemicals. Their analysis, especially in biota, has been thoroughly discussed by Gatermann et al. [1998] and by Rimkus et al. [1997]. It is thought that musk compounds can bioaccumulate in human tissue [spinnrad website 2000] and act as hormones, because they bind to the hormone receptors of the cells [Gerhard, I umweltmedizin website 2000]. However, there is insufficient data for an adequate toxicologically assessment for both the nitro- and the polycyclic musk scents [Antusch, 1999]. Emission Data At the present time the quota of the polycyclic musk scents amounts to approximately 85 % of total musk production worldwide, and the quota of nitro-musk scents is approximately 12 % [Rebmann et al., 1998]. Musk Compounds in Wastewater: The use of musk compounds in cosmetic and detergent products, which are used primarily in domestic situations or in buildings connected to the UWW collecting system, implies that their presence in surface waters occurs via municipal wastewater treatment plants. The mean concentrations found in biologically clarified wastewater from 25 German municipal WWTP were, o for musk-xylene: 0.12 µg.l-1 (concentration range: 0.03 – 0.31 µg.l-1) and o for musk-ketone: 0.63 µg.l-1 (concentration range: 0.22 – 1.3 µg.l-1). The mean emission levels in Germany were quantified as 20 µg/inhabitant/day for musk- xylene and 90 µg/inhabitant/day for musk-ketone [Eschke et al., 1994]. In Vienna, Austria, extensive testing of wastewater was carried out at the pilot plant of the city’s main WWTP, during 1999. Concentrations of musk compounds in WWTP influent and effluent are shown in Table e.3. 177 Section 6. Case Studies Table e.3: Musk compound concentrations in influent and effluent of the pilot WWTP Simmering, Austria in 1999 [Hohenblum et al., 2000].n: number of samples analysed. LOD: limit of detection Compound Type of Sample Range Mean (µg.l-1) sample number > (µg.l-1) Value LOD (µg.l-1) Musk-xylene Influent 4 0.023 - 0.037 0.031 (n=4) Effluent 0 - - Musk-ambrette Influent 0 - - (n=4) Effluent 0 - - Musk-moskene Influent 0 - - (n=4) Effluent 0 - - Musk-tibetene Influent 0 - - (n=4) Effluent 0 - - Musk-ketone Influent 4 0.049 - 0.069 0.056 (n=4) Effluent 4 0.038 -0.053 0.049 Musk Compounds in Sewage Sludge: The result of analyses into the presence of musk compounds in sewage sludge and the sediment of UWW collecting systems in German commercial and residential areas, are presented in Table e.4. In all samples noticeably high musk scent concentrations were detected. Sediment samples from the UWW collecting system for the residential area had slightly higher concentrations of the three polycyclic compounds ADBI (celestolide), HHCB (galoxolide) and AHTN (tonalide). This is probably due to the more frequent use of perfumes and detergents in domestic areas. Table e.4: Concentration of different musk scents in sewage sludge and UWW collecting system sediment in mg/kg DS, Germany [Antusch, 1999]. N = number of samples analysed. N>LOD number of samples over the limit of detection Compound N> Sediment: Sediment: Sewage sludge LOD industrial area residential area (n=17) (n=2) (n=2) mean range mean range mean range Musk-xylene 6 0.028 <0.005-0.20 0.095 0.066-0.134 <0.005 < 0.005 Musk-ketone 7 0.12 <0.01-1.78 0.25 0.15-0.36 0.03 <0.01-0.06 ADBI 12 0.051 <0.01-0.28 0.35 0.19-0.52 0.20 0.12-0.29 HHCB 17 1.43 0.08-5.2 15.5 9.1-21.8 8.87 4.3-13.4 AHTN 17 2.08 0.13 - 8.9 23.1 9.5 - 36.7 8.30 4.0 - 12.6 178 Section 6. Case Studies Musk Compounds in Watercourses: The pollution of the river Ruhr (Germany) with musk compounds was found to be relatively low, with maximum concentrations of 0.08 µg.l-1 and 0.03 µg.l-1 for musk-ketone and musk- xylene, respectively. Fish from the Ruhr contained residues of musk compounds in their muscle flesh, at concentrations below 10 µg.kg-1 wet weight [Eschke et al., 1994]. The polycyclic musks, HHCB (galaxolide, abbalide) and AHTN (tonalide, fixolide) were found in German receiving waters at concentrations up to the µg.l-1 level. In the Wuhle, a small stream consisting almost totally of wastewater effluent, maximum concentrations were 12.5 µg.l-1 for HHCB and 6.8 µg.l-1 for AHTN. Additionally, the polycyclic musk ADBI (celestolide, crysolide) and musk-ketone were detected at low concentrations in the majority of samples. Two other nitro-musks, moskene and xylene, were only detected in a single surface water sample [Heberer et al., 1999]. The concentration of the three compounds tonalide, celestolide and galaxolide, were measured in different watercourses in Germany. The results of this study are shown in Table e.5. The Elbe concentrations for musk-xylene were approximately 0.2 µg.l-1. Table e.5: Concentration of nitro-musk compounds in different watercourses in the länder Sachsen and Sachsen-Anhalt, Germany [Lagois, 1996]. ADBI (celestolide) HHCB (galaxolide) AHTN (tonalide) [µg.l-1] [µg.l-1] [µg.l-1] Sample date 22.5.95 12.6.95 22.5.95 12.6.95 22.5.95 12.6.95 Elbe at Torgau <0.08 Conclusions There is very limited information on the health and environmental effects of personal care products, such as the musk compounds found in perfumes. Many of these compounds have the potential to bioaccumulate, which is why there is concern about their presence in wastewater. Though these products may be used in large quantities there is insufficient data though to establish whether the presence of these compounds in UWW could cause any detrimental environmental or health effects. 179 Section 6. Case Studies (f) Surfactants in Urban Wastewaters and Sewage Sludge Introduction Surfactants are the largest class of anthropogenic organic compounds present in raw domestic wastewater. They are used in household and commercial laundry and cleaning operations. Surfactants can be classified [Ullman’s Encyclopaedia of Industrial Chemistry, 2000] into: · Anionic surfactants are anion-active, amphiphilic compounds in which the hydrophobic residues carry anionic groups with small-sized counter-ions, such as sodium, potassium or ammonium ions. These counter-ions have only a slight influence on the surface active properties. Examples include soaps, alkylbenzene sulphonates (ABS), alkylsulphates (AS), and alkylphosphates (AP). · Non-ionic surfactants (NIS) – are amphiphilic compounds that are unable to dissociate into ions in aqueous solutions, for example, alkyl- and alkylphenyl polyethylene glycol ethers, alcohol ethoxylates (AE) and alkylphenol ethoxylates (APE), fatty acid alkylolamides, sucrose fatty acid esters, alkylpolyglucosides, trialkylamine oxides. · Cationic surfactants, cation-active amphiphilic compounds in which the hydrophobic groups exist as cations with counter-ions such as chloride, sulphate or acetate. Examples include tetraalkyl ammonium chloride, N-alkylpyridinium chloride and others. · Amphoteric surfactants have zwitterionic* hydrophilic groups (*electrically neutral ions with both positive and negative charges), such as aminocarboxylic acids, betaines and sulphobetaines. Uses and sources of surfactants in the environment The largest proportion of surfactants is used in detergents and cleansing agents for domestic and commercial use [Falbe, 1987]. Surfactants are also used in: · fabric softeners (cationic), · foam cleaning agents (sulphosuccinates, LAS, AE), · general cleansing agents (LAS, alkylbenzenes, fatty alcohol ether sulphates), · domestic washing up liquids (betaines, NPO, alkylpolyglucosides), · industrial cleansing agents (alkylbenzene sulphonates, alkanesulphonates, fatty alcohol ethoxylates, alkylphenol ethoxylates, fatty amine ethoxylates, ethoxylates, propylene oxide adducts, and others), · bodycare products (ethersulphates, ether carboxylates, betaines, sulphosuccinates of fatty alcohol polyglycol ethers, isethionates, amineoxides, alkyl polyglucosides). Textile manufacturers uses surfactants extensively as washing agents, also for cleaning, lubricating, bleaching, de-sizing or shrinking (where the mutual adhesion of fibres is reduced), mercerising (cotton treatment that requires wetting agents), and finishing. Wool washing is done with NIS, while cotton is washed with anionic surfactants. The leather industry uses non-ionic and cationic surfactants as wetting and cleansing agents, and also for leather conditioning. Surfactants are also used as emulsifiers and dispersants, and also as food additives (natural substances only, such as glycerides, fatty acid salts, etc.). Pharmaceuticals manufacturing uses surfactants, and they are also used in agricultural applications such as crop-protection and pest control agents. Metal working and machining, petroleum extraction and processing, ore flotation and dressing, mineral oil industry, road construction and maintenance work, cement industry, plastics production, pulp and paper industry and printing, electroplating, adhesives manufacturing, all use surfactants for their unique properties. 180 Section 6. Case Studies Non-ionic surfactants are detergents which possess specific physicochemical properties, including relative ionic insensitivity and sorptive behaviour [deVoogt et.al, 1997] which makes them particularly suited for use wherever interfacial effects of detergents, foaming- defoaming, de-emulsification, dispersion or solubilisation can enhance product or process performance. The major part of the non ionic surfactants group consists of alcohol ethoxylates (AE) and alkylphenol ethoxylates (APE) of which, nonylphenol ethoxylate (NPE) is the main representative. Because of the formation of persistent metabolites in the environment, OSPARCOM member states have decided to phase out the use of NPE and to replace the APEs with AEs. The production of non-ionic surfactants in the USA and EU amounts to about 750 000 t/a and includes some 300 000 t/a of APE [Holt et.al., 1992]. Ionic surfactant molecules contain both strongly hydrophobic and strongly hydrophilic groups. They thus tend to concentrate at interfaces of the aqueous system including air, oily material and particles. The hydrophobic group is generally a hydrocarbon radical (R) of 10 to 20 carbon atoms. The hydrophilic portion may ionise or it may not. Ionic surfactants may be either anionic or cationic. Ionic surfactants constitute approximately two-thirds of the surfactants used. Cationic surfactants constitute less than 10% of the ionics and are used for fabric softening, disinfection and other specialized applications. The predominant class of anionic surfactants is linear alkylbenzene sulphonates (LAS). The concentrations of linear alkylbenzene sulphonates (LAS) in raw wastewater range from 3 mg.l -1 to 21 mg.l-1 (Brunner et al., 1988, De Henau et al., 1989, Ruiz Bevia et al., 1989). Although LAS and other common surfactants have been reported to be readily biodegradable by aerobic processes, much of the surfactant load into a treatment facility (reportedly 20-50%) is associated with suspended solids and thus escapes aerobic treatment processes, being directed via primary sedimentation into sludge management processes. Because LAS is not biodegraded by anaerobic biological processes usually employed in sludge stabilization (McEvoy and Giger, 1985; Swisher, 1987), it may be found in the gram per kilogram range in anaerobic sludges. Given these concentrations and the major effects of surfactants in particle surface modification, deflocculation, and surface tension reduction, it seems clear that the performance of certain treatment processes as thickening, conditioning, and dewatering may be strongly influenced by these materials. Thus, surfactants may induce significant extra costs in sludge handling. Increased water content in landfilled sludges represents an additional possible impact, adding to the difficulty of proper landfill leachate control. Surfactants may also mobilise otherwise insoluble organic pollutants within the landfill. Similar implications exist for land application of surfactant-laden sludges. Feijtel et al [1995] examined five WWTS across Europe and found the influent of LAS to WWTS in the UK to be higher than in the other plants (see Table d.1). 181 Section 6. Case Studies Table d.1 LAS in influent and effluent in the UK compared with other regions [Feijtel et al 1995] Country Influent mg/l Country Effluent mg/l Mean (+/- 95% CIs) Mean (+/- 95% CIs) UK 15.1 (2.3) UK 0.010 (0.002) Germany 5.4 (6.1) Germany 0.067 (0.076) Italy 4.6 (5.1) Italy 0.043 (0.065) Netherlands 4.0 (4.0) Netherlands 0.009 (0.008) Spain 9.6 (9.6) Spain 0.14 (0.14) As it can be seen from this study, the UK plants had significantly higher influent concentrations than in Germany, Italy and the Netherlands (Spain had a very small data set and therefore there is uncertainty in these results, which is reflected in the large confidence intervals, which overlap with those of the UK). Having a high influent of LAS was unrelated to the concentration in the effluent and the UK samples had a lower concentration of LAS in its effluent than Germany, Italy or Spain. The level of biodegradation of LAS during the wastewater treatment process is high and also varied between the plants, with the UK WWTS having the highest level of biodegradation breaking down greater than 99.9 percent of the LAS. The study above, used UK data from Holt et al [1995] in which the levels of LAS in the influent corresponded to estimates of LAS usage in homes. However a subsequent study [Holt et al 1996], based on usage data on the influent entering six WWTS in the Yorkshire region, found much lower levels of LAS than expected. In no cases did the concentrations of LAS reaching the plant approach the level predicted from consumer usage. The previous study had been carried out in March while the second was carried out in a ‘warm dry summer’. This study suggested there was significant biodegradation of LAS under certain conditions in the UWW collecting systems. Even in the relatively short residence time in the Yorkshire UWW collecting system, for a couple of hours, up to 60% of the LAS was removed prior to the wastewater treatment plant. This was then followed by removal of between 70 to 99 percent of the remaining LAS in the treatment plants. Given the right treatment conditions, LAS are biodegradable and more research is necessary to compare risks associated with alternative chemicals used in detergents. The impact of LAS in wastewater effluent has been studied on several UK and Dutch rivers and steams sediment. In some cases (such as the small stream into which the Owlwood WWTS discharges its effluent the LAS load in sediment upstream of the treatment plant was higher than that downstream [Waters and Feijtel 1995] This was hypothesised to be because that the LAS contribution upstream was due to unregulated discharges of untreated wastewater to the aquatic environment while the input of the WWTS effluent actually served to dilute the concentration of this pollutant downstream. A similar case was found in the Netherlands when concentrations of LAS could be higher upstream of a WWTS due to direct discharges from storm tanks. In other locations downstream sites were found to have very slight elevations of LAS in sediment of between 0.49 and 3µg/g. A more careful policy of discharges has to be followed across the EU. A new study by NERI [NPE/DEHP in sewage sludge in Denmark, http://www.dmu.dk/beretuk98/society.htm#c4] shows that only small amounts of NPEs and DEHPs are discharged in the urban wastewater by the commercial sector and that they do not accumulate in agricultural soils treated with sewage sludge in moderate amounts. Figure d.1 shows the level of surfactants and plasticisers in soils with the depth of the soil. 182 Section 6. Case Studies Fig. d.1 Vertical distribution of various nonylphenols and phthalates in agricultural soils fertilized with large amounts of sewage sludge (17 tonnes dry matter per hectare per year). [NERI, 2000] Fate and effect of surfactants in the environment Fate of surfactants during wastewater treatment: LAS and other common surfactants have been considered to be readily biodegradable by aerobic processes, based on laboratory studies (Swisher, 1970). Figure d.2 shows schematically the fate of LAS in the environment (deWolf and Feijtel, 1998). LAS evidently undergoes nearly complete biodegradation, with 97-99% removal rates found in some wastewater treatment plants (Brunner et al., 1988; Bevia et al., 1989; De Henau et al., 1989). However, the mass loadings indicated above suggest that even at these removal rates, appreciable amounts are released to receiving waters. Ventura et al. (1989) identified LAS and a variety of other anionic, cationic and nonionic surfactants in both surface and drinking water extracts. 183 Section 6. Case Studies Figure d.2 LAS fate in the environment (after deWolf and Feijtel, 1998) Alkylphenol ethoxylates such as NPnEO are evidently less biodegradable than LAS with laboratory results ranging from 0-20% based on oxygen uptake (e.g. Swisher, 1970; Steinle, 1964; Pitter, 1968) and a wider range of removals from 0-90% based on specific analyses such as UV and IR spectroscopy (Swisher, 1970). This suggests that only partial degradation occurs, such as conversion from polyethoxylates to nonylphenol diethoxylate (NP2EO), nonylphenol monoethoxylate (NP1EO), and nonylphenol (NP). Mass balances carried out on treatment plants in Switzerland (Brunner et al., 1988) support this. The findings of Brunner et al. and other reserachers, also show that the nearly complete removal of surfactants from treated waters is not entirely due to biodegradation. Brunner et al. indicated that 19% of the surfactant load entering a treatment facility is associated with suspended solids, and other studies report levels up to 27% (Rapaport and Eckhoff, 1990), or even in excess of 50% (Bevia et al., 1989). The surfactant load linked to suspended solids is directed into sludge treatment processes via primary sedimentation. Surfactants such as LAS are not biodegraded by either mesophilic or thermophilic anaerobic digestion (McEvoy and Giger, 1985; Swisher, 1987) so a large proportion of these materials simply escapes treatment and becomes associated with sludge solids. The resulting concentrations of surfactants in sewage sludges can be substantial. LAS concentrations measured in sludges often make up between 0.5% and 1.5% of the dry solid mass, particularly for anaerobically digested sludges (McEvoy and Giger 1986; De Henau et al. 1989; Holt et al. 1989; Marcomini et al. 1989). Bevia et al. (1989) reported LAS between 2% and 4% of the sludge solids weight. In a study of 29 Swiss treatment plants, LAS concentrations averaged 4.2 and 2.1 g kg-1 respectively in anaerobic and aerobic sludges. NP exceeded 1 g kg-1 dry sludge and, in some instances NP1EO and NP2EO exceeded 0.1 g kg-1 dry sludge (Brunner et al. 1988). 184 Section 6. Case Studies Effects of surfactants on wastewater treatment As stated previously, given these surfactant concentrations and the considerable effect that surfactants can have on the properties of suspensions such as sludges, the performance of such processes as thickening, conditioning, and dewatering may be strongly influenced by these materials. For example, Bierck and Dick (1988) have shown that surface tension of sludge solids is directly related to the capillary pressure available for solids compression during the latter stages of vacuum filtration: Ps,s = v [1/R1 + 1/R2] Where, Ps,s = the pressure or effective stress producing solids shrinkage, v = the surface tension, R1 and R2 = principal radii of curvature of the solid surface. Thus the effect of surfactants, in lowering the surface tension, is to decrease the compressive dewatering by allowing gas penetration of the solids cake. Campbell et al. (1984, 1986) showed that a detergent could decrease the dewaterability of a sludge even before the compressive phase, as indicated by capillary suction time (CST) measurements. Household detergent added to anaerobically digested sludge at 0.2 and 0.3% by volume caused significant increases in the CST (poorer dewaterability) which could not be compensated for even by doubling the addition of cationic polymer used as the sludge conditioner. The implications of surfactants' influence on dewatering should not be underestimated. Costs for the sludge conditioning polymer are the greatest operating cost for dewatering at a WWTS such as Wilmington, and sludge dewatering and disposal represent up to 50% of the total cost of wastewater treatment (Evans, 1988). The biodegradation mechanism of LAS was described by Balson and Felix(1995). The mechanism of breakdown of LAS involves the degradation of the straight alkyl chain, the sulphonate group and finally the benzene ring. The breakdown of the alkyl chain starts with the oxidation of the terminal methyl group (w-oxidation) through the alcohol, aldehyde to the carboxylic acid as follows (see Fig. d.3a). The reactions are enzyme catalysed by alkane monooxygenase and two dehydrogenases. The carboxylic acid can then undergo b- oxidation and the two carbon fragment enters the tricarboxylic acid cycle as acetylCo-A. It is at this stage that problems arise with branched alkyl chains, a side chain methyl group or a gem-dimethyl-branched chain cannot undergo b-oxidation by microorganisms and must be degraded by loss of one carbon atom at a time (a-oxidation, Figure d.3b). (Scott and Jones, 2000). 185 Section 6. Case Studies Figure d.3a w-Oxidation of LAS (after Scott and Jones, 2000) The second stage in LAS breakdown is the loss of the sulphonate group. The loss of the alkyl and the sulphonate group from LAS leaves either phenylacetic or benzoic acids. Microbial oxidation of phenylacetic acid can result in fumaric and acetoacetic acids and benzene can be converted to catechol . Figure d.3b a- Oxidation of LAS ( after Scott and Jones, 2000). 186 Section 6. Case Studies Effects of the surfactants on the wider environment: The presence of surfactants in sewage sludge may have undesirable environmental effects if land application is the chosen disposal method. The surfactant molecules may leach to groundwater contributing to groundwater contamination. Federle and Pastwa (1988) studied the percolation of anionic and nonionic surfactants through a soil column. Most of the surfactant was mineralised, but this process was found to be highly dependent on the number of organisms present in the soil. A number of reports (e.g. Bevia et al., 1989; Holt et al., 1989) attribute observed decreases of LAS concentrations over time in sludge-amended soils to biodegradation, without evaluating possible migration. Marcomini et al. (1989) reported a fraction of LAS in sludge-amended soil to be resistant to biodegradation over long time periods. Table d.2 contains data on fate and persistence of surfactants in sludge amended soils. Table d.2 Fate and persistence of surfactants in sludge amended soils Application Country Surfactant/ Soil Monitoring Final Half Life Author Form derivative Concentration period Soil (days) post Conc. application -1 -1 (mg.kg ) (mg.kg ) Sludge onto SP LAS 22.4 6 months 3.1 Not Prats et al soil 12 months 0.7 reported Sludge onto CH LAS 45 12 months 5 9 Marcomini et al soil NP 4.7 0.5 Surfactant D LAS Not reported 2 months Not 5-25 Litz et al onto soil 6 months reported summer 66-117 winter Sludge onto D LAS 16 76 days 0.19 13 Figge and soil 27 106 days 0.44 26 Schoberl Surfactant USA LAS 0.05 40 days Not 1.1- Knaebel et al onto soil LAE 0.05 reported 3.6 Sludge onto SP LAS 16 90 days 0.3 26 Berna et al soil 53 170 days Not 33 reported Sludge onto UK LAS 2.6-66.4 (*) 5-6 months <1 7-22 Water et al soil Holt et al Sludge onto UK LAB 0.3-9.5 (*) 55 days 0-0.38 15 Holt and soil Bernstein Composted AUS NPE 14 000 14 weeks 1200 Not Jones and wool scour reported Westmoreland sludge (* estimated cumulative load) Not only may surfactants migrate to groundwater, but they may also carry hydrophobic organic pollutants with them. The degree of partitioning of hydrophobic organic pollutants to particles depends on the hydrophobicity of the pollutant and the amount of organic matter contained in the particle. Dissolved organic matter tends to decrease the potential for sorption by providing an additional aqueous phase to which the pollutant can partition (Enfield et al. 1989). Partitioning of surfactant to sludge particles in the sewage treatment plant would be expected to enhance the partitioning of organic pollutants to sludge. When applied to land, desorption of surfactant could lead to pollutants also being released. Kile and Chiou (1989) studied the effect of anionic, cationic and nonionic surfactants on the water solubility of DDT and trichlorobenzene. The results were extremely surprising. As would be expected, the solubility was enhanced when the surfactant was present at 187 Section 6. Case Studies concentrations greater than the critical micelle concentration. There was also a solubility enhancement at surfactant concentrations less than the critical micelle concentration. In addition to the effects of surfactants in sludge on pollution of groundwater, the surfactants may effect soil texture and water retention through processes similar to those discussed with respect to sludge dewatering. Holtzclaw and Sposito (1978) determined LAS content in a sludge amended soil to be high enough (1% of the fulvic acid fraction) that soil fertility could be affected. The fate of surfactants in sludges disposed of in landfills is somewhat surprising. Concentrations of LAS up to 1% by weight have been found in recently deposited material, with some amounts above 1 g.kg-1 even after 15-30 years (Marcomini et al., 1989). Given that landfills function in a similar manner to anaerobic digesters, the persistence of LAS is evidently due to its poor degradability in such environments. The role of surfactants in mobilizing less hydrophobic contaminants into landfill leachate is thus a relevant concern. Behaviour of nonionic surfactants: Recent studies have revealed that fish living downstream of wastewater treatment plants show oestrogenic effects [Purdom et al. 1994] as a result of alkylphenol polyethoxylates (APE) and nonylphenol (NP) present in the water. Male fish produce vitellogenin, a yolk protein which is formed under the influence of oestradiol and therefore is typically produced by females. Hermaphrodite fish species have been found as well. The decomposition products of APE, are considered as a potential cause, since their decomposition products formed in WWTS (Giger et al. 1984) show slightly oestrogenic effects (Soto et al. 1991, Jobling and Sumpter 1993). Alkylphenol polyethoxylates (APE) usually enter surface waters via WWTS, where they are degraded - but not totally - by microorganisms. In a first rapid step the ethoxylate groups are split off by hydrolysis, and the metabolites nonylphenol (NP), nonylphenol ethoxylate (NP1EO) and nonylphenol diethoxylate (NP2EO) are formed. These metabolites are more toxic than the original substances. Due to the hydrophobic properties of the aromatic group the second step of biodegradation occurs much slower. The interim products can also be biodegraded to alkylphenoxy ethoxylate carboxylic acids (APEC). The second, slower, step of biodegradation, does not always occur, and the fact that the metabolites are more lipophilic than the parent compounds can cause an accumulation of interim products in sludge and sediment. Nonylphenol, for example, was determined in digested sludge in concentrations between 0,45-2,53 g.kg-1 dry weight (Giger et al. 1984). Approximately 50 % of the APE occurring in the wastewater are estimated to reach the sludge as NP (Brunner et al. 1988). Before prohibition of APE in washing agents NP, NP1EO and NP2EO concentrations between 36-202 µg/l were found in drain channels of WWTS in Switzerland. Now NP concentrations in drain channels from WWTS, are found at concentrations between 1 and 15 µg.l-1 in Switzerland and Germany; other metabolites (NP1EO, NP2EO, NP1EC) are normally determined to be between 1 and 40 µg.l-1 (Ahel et al. 1994a, Ahel et al. 1994b, Giger 1990). Concentrations of 15 µg.l-1 in drain channels of WWTS were determined in the USA. In highly polluted streams average nonylphenol concentrations are determined to be in the range 0.3 to 3 µg.l-1 (Ahel et al. 1994a), polyphenoxy carboxylic acids products are predominant, whereas in sediment NP was the dominating degradation product. Due to their high octanol/water partition coefficient (log 4.0-4.6) nonylphenol, NP1EO and NP2EO show a tendency towards bioaccumulation in organisms. This was confirmed by residue analyses (Table d.3). The bioconcentration factor in fish is approx. 300, in one case, however it amounts to 1300. 188 Section 6. Case Studies Table d.3: Environmental concentrations of degradation products of nonionic surfactants Environmental Substance Concentration Reference compartment Sewage sludge NP 0,45 - 2,53 g kg-1* Giger et al., 1984 0,03 g kg-1* Giger and Alder, 1995 WWTS-drain NP, NP1EO, NP2EO 36 - 202 µg l-1 Stephanou and Giger, NP 1 10 µg l-1 1990 NP1EO, NP2EO 1 - 40 µg l-1 Streams NP 0,3 - 45 (2-3) µg l-1 Ahel et al., 1994b NP1EO, NP2EO < 3 - 69 µg l-1 NP1EC, NP2EC < 2 - 71 µg l-1 Stream sediment NP 0,5 - 13 mg kg-1* Ahel et al., 1994b Fish NP, NP1EO, NP2EO 0,03 - 7,0 mg kg-1* Ahel et al., 1993 Algae NP, NP1EO, NP2EO 80 mg kg-1* Ahel et al., 1993 Waterfowl (ducks) NP, NP1EO, NP2EO 0,03 - 2,1 mg kg-1* Ahel et al., 1993 * dry weight Health effects of surfactants: Prats.et.al, 1993 show significant differences between distribution of LAS homologs in water and solids (sludges, sediments, and soils), as compared to the original distribution in detergent formulations, yielding a lower LAS average molecular weight in water samples. The change observed in the homolog distribution of LAS implies a reduction in the toxicity to Daphnia, because a lower average molecular weight of LAS is less toxic. The risk assessment of LAS to terrestrial plants and animals reported by Mieure et al. (1990) also concludes that there are adequate margins of safety in the use of wastewater for the irrigation of plant species. Adverse effects on plant and animal species (earthworms Eisena foetids and Lumbricus terrestris) were observed at LAS concentration of 10 mg.l -1 , however LAS concentrations in wastewater effluents are in a range 0.09 mg l-1 to 0.9 mg l-1 . These figures give a safety margin in a range 10 to 100. The effect of surfactant on plant growth from the use of sewage sludge is difficult to assess because in general the sludge promotes plant growth. Adverse effects on plant growth were observed at 392 µg g-1 but long term monitoring at a range of 46 environmental sites gave LAS concentrations of less than 3µg.g-1 . These figures give a safety margin of 131. For terrestrial animals the limit of no adverse effects was 235 µg.g-1 giving a safety margin of 78. However, in looking at ecotoxicity from WWTS effluents the less toxic surfactant residues and surfactant catabolites must be considered and this requires analytical tests for these entities (Scott and Jones, 2000; Schoberl, 1997). Amounting to 2-4 g.kg-1 the acute mammalian (mouse, rat) toxicity of APE is low. Dermal toxicity, however, is higher (500 mg.kg-1), and eye irritation is the highest with 5-100 mg.kg-1. NP can be metabolised to a glucoronide in the body and excreted via the kidney. Nonionic surfactants are more toxic for aquatic organisms than for mammals. The toxicity of APE increases with decreasing number of ethoxylate units and increasing hydrophobic chain length. Accordingly, the toxicity of the original substances is lower than the toxicity of the metabolites NP, NP1EO and NP2EO, whereas the carboxylic acids are less toxic than the ethoxylates. For instance the LC50 (48 hours) of NP16EO is 110 mg.l-1 for fish (Oryzias latipes) and decreases to 11,2 and 1,4 mg.l-1 for NP9EO and NP, respectively (Yoshima, -1 1986). The LD50 (96 hours) for algae (Skeletonema costatum) is 27 µg.l , and the value for rainbow trouts 480 µg.l-1 (Nayler 1992). The no observed effect concentration (NOEC) for reproduction for Daphnia is in the range of 24 µg.l-1. These data show that the acute toxicity of NP is considerably high. In vitro toxicity studies with fish hepatocytes indicate that several decomposition products of APE cause weak oestrogenic effects (Jobling and Sumpter 1993, White et al. 1994). Studies 189 Section 6. Case Studies based on the vitellogenin synthesis revealed that NP, NP1EO and NP1EC have the same activity (half maximum activity: around 16 µM). The oestrogenic activity, however, is 10 times lower than that of oestradiol (Pelissero et al. 1993). Other in vitro studies give hints on potential differences between fish and mammalia regarding the binding to the oestrogen receptor (Thomas and Smith 1993). However, vitellogenin synthesis in fish hepatocytes is also induced by well-known phyto-oestrogens. Studies in the UK indicate that downstream of the drain channels of WWTS vitellogenin is formed in male fish. After 1 to 3 weeks exposure of fish in 15 drain channels of WWTS, displayed a high increase of vitellogenin synthesis (Purdom et al. 1994). It is supposed that the decomposition products of APE, especially NP, are mainly responsible for this effect. The assumption is confirmed indirectly by the results of the in vitro studies with fish hepatocytes. However, it cannot be excluded that synthetic oestrogens are also responsible for this effect. On the one hand their concentrations are lower than the usual concentrations of NP, but on the other hand their activity is some orders of magnitude higher. Experimental exposure of fish to NP or metoxychlor over 7 days induced vitellogenin synthesis in male fish (Nimrod and Benson 1995). The dose required to induce the vitellogenin synthesis was 300 times (approx. 150 mg.kg-1) higher than the necessary dose of oestradiol. Further research in this field, especially the conduction of experimental in vivo studies, is urgently required to allow for a more reliable assessment of the exposure of fish populations to oestrogenic chemicals and their potential effects. Field investigations indicate that downstream of the drain channels of WWTS oestrogenic effects may be induced in fish. The in vitro studies with fish hepatocytes seem to indicate that the oestrogenic activity of synthetic oestrogens is some orders of magnitude higher than the activity of decomposition products of APE. On the other hand the oestrogenic potency of NP, NP1EO, NP2EO and NP1EC is very similar. Consequently, all degradation products have to be taken into consideration. It seems advisable to suppose that the above chemicals have additive effects. Best environmental practice examples One of the main success stories regarding the use and fate of surfactants is linked with eco- labelling. Eco-labelling has been developed for the products used in dishwashing, laundry and cleaning detergents in Scandinavia, Germany, Austria and other European countries. Products with the ‘Swan’ and ‘Good Environmental Choice’ label do not contain LAS and Nonylphenol and have gained considerable market share. In Sweden, products with these labels accounted for more than 95% of sales by 1997 while in Finland they reached 15%. Norway and Denmark had lower sales (source, Danish Environment Agency 2000). More research is needed for the potential environmental effects of the alternatives used in these LAS-free and NPE-free surfactants. This public awareness and consumer choice, lead to the use of LAS in Sweden falling from 6300 tonnes a year to 260 tonnes a year. The Danish Environment Agency launched a public campaign against LAS in September 1999. Currently about 2500 tonnes a year of LAS are used in Denmark. During the development of the “Swan” mark, Stockholm water company identified the need to have an alternative for taking care of hazardous waste in the household rather than flushing it into the UWW collecting system. Environmental stations, or collection points were established and an extensive public information campaign was carried out about the impacts of household products on the aquatic environment [Ulmgren, 2000a, 2000b]. Detergents and cleaning agents containing alkylphenolethoxylates (APEO), such as NPE, are being gradually phased-out under various initiatives and voluntary agreements in EU. Distearyl-dimethylammonium chloride (DSDMAC), widely used in laundry softeners, was 190 Section 6. Case Studies also substituted with more degradable substances during the 1990s in Germany [Greiner, 1996] and is currently under scrutiny in the rest of the EU. The eco-labelling combined with a public awareness campaign could therefore influence consumer choice and reduce contaminant discharges in the UWW from domestic products. More research is necessary to experimentally determine the role of surfactants in sludge treatment processes and following sludge disposal in the environment. Specific effects to be investigated are · impacts on sludge thickening, conditioning, and dewatering processes and · transport and mobilization of hydrophobic organic contaminants when sludges are landfilled or land-applied. Also anticipated is an improved fundamental understanding of mechanisms by which surfactants are incorporated into sludge solids. Some important research gaps and necessary research are summarised as follows: Effects of endocrinally active chemicals have not yet been systematically investigated in amphibian and reptiles. In this field nearly no knowledge is available. · Chemical methods for the detection of traces of synthetic oestrogens and their metabolites must be elaborated, since only very few data are available on environmental concentrations, especially regarding concentrations in drain channels of wastewater treatments plants. Furthermore, data material on NP concentrations in drinking water and organisms including humans is insufficient. · The ecotoxicological relevance of vitellogenin production in male animals has to be elucidated. Which interrelations exist between the problem of vitellogenin production and further estrogenic and ecotoxicological effects of NP and other chemicals? To answer these questions in vivo experiments using histopathological, biochemical, endocrinological and reproduction biological methods have to be conducted. Furthermore, insufficient information is available about the bioaccumulation of these chemicals. In a further step the problem should be investigated by more comprehensive field studies. · The mechanisms of chronic effects of alkylphenols (modes of action) must be studied in more detail. · Finally, in vitro assays should be elaborated to identify and estimate the oestrogenic activity of existing new chemicals in fish and other organisms. 191 Section 6. Case Studies (G) Use of Polyelectrolytes; The Acrylamide Monomer in Waste Water Treatment Polyacrylamide (PAM) is a widely used flocculant in water treatment applications. Some 20,000 tonnes is used in the USA for this purpose each year. Concerns with its use are that it can degrade to the acrylamide monomer which is known to affect the central and peripheral nervous systems and is also believed to be carcinogenic. Safe levels for this chemical are said to be 300 µg l-1 over a ten-day period and 2 µg l-1 over a seven year period (EPA). PAM is used in other applications such as an aid in irrigation (Trout et al. 1995) and in pulp and paper manufacture. The EPA also notes that it is used in formulating grouting for tunnels and sewers. Effluents from a sewage works which used PAM as a flocculant in the UK were reported to be 2.3 to 17.4 µg l-1. High levels of the monomers have been reported in acrylamide manufacturing plant effluents. In this case the raw effluent contained 1100 µg l- 1 and the treated effluent 280 µgl-1 (EPA). Both PAM and its monomer are very soluble in water and the presence of PAM in soil causes leaching of microorganisms by ground or irrigation water. PAM is shown to degrade by biological action and photolytic effects (Nakamiya 1995). Experiments have shown that polyacrylamide solutions in a bottle covered with plastic film and left outside can contain significant amounts of the monomer after two weeks exposure (Smith et al. 1996). The polymer can also be degraded by turbulent shear stress in pumps and pipes (Rho et al. 1996). Once the polymer degrades the monomer is also subjected to more rapid degradation in which it is decomposed to acrylic acid and ammonia. These are non toxic as acrylic acid degrades to CO2 and water in a day in soil (Staples et al. 2000) and is thus not an ecological problem. The degradation of acrylamide under favourable conditions by pseudomonas species immobilised in calcium aliginate took one day (Nawaz et al. 1993). The EPA say that degradation of acrylamide in river water takes 4 to 12 days and is more rapid in summer than winter. Due to its solubility adhesion of the monomer on soil is unlikely, though it is reported that it is partially removed by secondary activated sludge. The general conclusion of the EPA paper is that the monomer is not an environmental hazard when released in small quantities to the aqueous environment. Tests with fish would indicate this. The monomer is relatively biodegradable within days compared to the time span of years for substances such as PCB. There are two areas where there could be some concern and these are control of effluents from acrylamide manufacturing and use in drinking water treatment. The fact that the monomer is detected in water treatment plants where the residence time is only a few hours suggests that the PAM flocculant could have significant residual amounts of monomer. There is a web site (www.fwr.org/waterg/dwi0084.htm) which quantifies some of the points mentioned. Among the points noted are: · Chlorination and the presence of potentially toxic elements can stop acrylamide degradation by passifying the bacteria present. · Degradation of the monomer is not pH dependant · 50% of PAM is removed in aerated sludge and trickle bed filters. A significant drop in the percentage of monomer present in the PAM used for flocculation has eased the likelihood of serious contamination of water. The present level of monomer present in grades of PAM used for water treatment is 0.3% and has dropped from 0.8%. However PAM used in grouting has a much higher monomer content and the use of this particular grade has caused the monomer to leak into grouted sewers. To conclude it seems that whenever water is sent into the environment it would be safe to use PAM as a coagulant. Problems may arise when it is used in a stream which is 192 Section 6. Case Studies subsequently sterilised by chlorine or by another disinfectant. This removes the bacteria that are needed to degrade the PAM and it would be a problem with using other organic polymers as well. It might be concluded that a polymer grade containing just 0.3% of monomer is very pure and improvements in purity might not be practically feasible. Therefore in treating drinking water some intermediate step may be necessary to degrade the monomer after flocculation and before adding the disinfection agent. This could involve a holding lagoon or treatment using activated charcoal. In any case the case for or against the use of PAM in flocculating drinking water lies in the conflicting aims of acrylamide degradation and product sterilization. Its use in the treatment of drinking water needs continuous monitoring and, if more stringent regulations are placed on the monomer concentration in drinking water the issue will become a concern. 193 Section 6. Case Studies (h) Landfill Leachate Introduction In the past treatment of leachates in WWTS were favoured but due to the effects of dilution in the UWW system, there is little, if any information on the elimination of persistent compounds (Alberts, 1991). In Germany, a recent requirement has been the proper preliminary treatment of leachate before discharge either directly to surface waters or indirectly to municipal WWTS. With the introduction of redrafted legal conditions in Germany, strict demands have been placed on the purification performances of leachate treatment plants. Thus, the treatment of organic substances and nitrogen compounds using nitrification and denitrification, is required prior to direct discharge to a receiving watercourse. Wastewater and leachate quality requirements in Germany Definitions of direct and indirect discharge of wastewater are as follows: · Direct discharge: To discharge wastewater directly must conform to standards of water quality in the receiving water. · Indirect discharge: Discharging wastewater directly into a public WWTS requires that the concentration of COD, BOD, NH4-N, AOX and potentially toxic elements must be reduced to the same levels found in domestic wastewater. Standard values for the composition and quality of non-domestic wastewater discharged to a public WWTS are stated in guideline/directive ATV-A 115 (worksheet for indirect dischargers). The standard concentrations for potentially toxic elements are shown in Table h.1. Table h.1: Standard concentrations for soluble and insoluble inorganic substances in wastewater from non-domestic sources [ATV-A 115, 1994]. Potentially toxic Symbol Standard element Concentrations [mg/l] Lead Pb 1 Cadmium Cd 0.5 Chromium Cr 1 Chromium (VI) Cr(VI) 0.2 Copper Cu 1 Nickel Ni 1 Mercury Hg 0.1 Zinc Zn 5 Regulation AbwV: “Requirements for discharging wastewater into watercourses” The Wastewater Regulation (AbwV) places general requirements on the introduction of wastewater into receiving watercourses. A permit for discharging wastewater into a watercourse can only be granted, when the limit values for the pollution load at the point of discharge are observed. Dilution of wastewaters in order to reach the required concentration values is not permitted. Requirements for wastewater from landfill sites are given special attention in annex 51 of the regulation. It is a requirement that the quantity and pollution load of landfill leachate must be 194 Section 6. Case Studies kept low by proper measures and operation at the landfill installation. The requirements listed in Table h.2 relate to the discharge site of leachate into watercourses. Table h.2: Requirements for wastewater quality at point of discharge (qualified sample or 2 hour mixed sample) [AbwV, annex 51, 1999]. Parameter Unit Value COD* mg/l 200 BOD mg/l 20 Ntotal** mg/l 70 Ptotal mg/l 3 Hydrocarbons, total*** mg/l 10 N02-N mg/l 2 Fish toxicity GF 2 * For wastewater with a COD value (before treatment) of more than 4000 mg/l, the COD effluent value in the qualified sample or in the 2 hours mixed sample must be reduced by 95 %. ** Sum of ammonium-, nitrite- and nitrate-nitrogen (Ntotal) or total bound nitrogen (TNb). The requirement applies to a wastewater temperature of 12 °C. A higher limit concentration of 100 mg/l is permitted when the decrease of nitrogen load amounts to at least 75 %. *** The requirement relates to the qualified sample. The requirements listed in Table h.3 below relate to leachate before mixing with other wastewaters. Table h.3: Requirements on leachate before mixing (qualified sample or 2 hours mixed sample) [AbwV, annex 51, 1999 Parameter Unit [mg/l] AOX* 0.5 Mercury 0.05 Cadmium 0.1 Chromium 0.5 Chromium VI* 0.1 Nickel 1 Lead 0.5 Copper 0.5 Zinc 2 Arsenic 0.1 Cyanide, easily released* 0.2 Sulphide* 1 * value for the qualified sample. 195 Section 6. Case Studies Leachate can be mixed with other wastewater for common biological treatment only when: · fish, indicator bacteria and Daphnia toxicity of a representative sample is not exceeded (see Table h.3). It has to be stated that exceeding the GF value is not caused by ammonia (NH3). · a DOC elimination rate of 75 % is reached. · leachate shows a COD concentration lower than 400 mg.l-1 before the common biological treatment. Table h.4: Fish, indicator bacteria and Daphnia toxicity [AbwV, annex 51, 1999] Fish toxicity GF = 2 Daphnia toxicity GD = 4 Indicator bacteria toxicity GL = 4 Landfill Leachate Formation: A substantial proportion of pollutant emissions from landfill sites enters percolating through the landfill. Rain water (and other sources of water) entering unsealed sections of the landfill undergo chemical and biological transformation in the body of the landfill to form leachates. Pollutants are taken up by solution processes or are carried in suspension. This loaded water, the so called leachate, is collected at the base of the landfill in a drainage pipeline. The quantity of leachate produced depends principally on rainfall and the state of the landfill. At new, unsealed landfill sites, the total calculated rainfall collects as leachate. During the life of the landfill leachate quantity reduces to 10 – 20 % of total rainfall, with an increase in superficial sealing. Composition: Leachate is a heterogenezous mixture, often containing a high concentration of persistent biological and toxic compounds. The type of material deposited in the landfill determines the composition of the leachate. Leachate composition and pollutant concentration are also influenced by the rate of biochemical processes in the body of the landfill. After an initial intensive phase biological and chemical reactions in the landfill slowly subside. As the age of a landfill increases, the quota of easily degradable compounds in and the COD/BOD5 ratio rises (Leonhard, Wilderer, 1992). Some of the main pollutants found in landfill leachates are organic compounds, such as alkyl phenols, chlorinated phenols, polycyclic aromatic hydrocarbons (PAH), dioxins and furans. Leachate from special waste landfills tends to have higher concentrations of inorganic substances compared to leachate from household refuse landfills; chlorides, sulphates and fluorides represent the main load. Table h.5 gives an overview of the relevant pollutant concentration in landfill leachate. 196 Section 6. Case Studies Table h.5: Mean composition of leachate from: industrial or special waste landfills, and household refuse landfills, in Germany [Ehrig et al., 1988] Parameter Units Industrial and special waste Household refuse landfill landfill Range Mean value Range Mean value pH - 5.9 – 11.6 7.7 3.5 – 9 7.5 -1 COD mg O2.l 50 – 35000 5746 500 - 60000 5000 -1 BOD5 mg O2.l 41 – 15000 2754 100 - 45000 1500 Conductivity mS.cm-1 2110 – 183000 28217 - 10000 Chloride * mg.l-1 36 – 126300 13257 100 - 15000 2000 Sulphate mg.l-1 18 – 14968 2458 50 - 3000 300 Ammonium* mg.l-1 5 – 6036 921 20 - 3000 500 Nitrite* mg.l-1 0.02 - 131 7.3 - 0.5 Nitrate* mg.l-1 0.1 - 14775 606 0 - 50 3 Total-N* mg.l-1 1 - 3892 461 20 - 4000 600 Total-P* mg.l-1 0.03 - 52 7.9 0.01 - 10 1 Fluoride mg.l-1 0.1 - 50 13.3 - - Total cyanide mg.l-1 0.007 - 15 1.3 - - Easily released mg.l-1 0.008 - 1 0.2 - - cyanide Arsenic* mg.l-1 2 - 240 51 0.1 - 1000 20 Lead* mg.l-1 4.3 - 650 155 20 - 1000 50 Cadmium mg.l-1 0.2 - 2000 144 1 - 100 5 Copper* mg.l-1 1.3 - 8000 517 10 - 1000 50 Nickel* mg.l-1 14.2 - 30000 2096 20 - 2000 200 Mercury mg.l-1 0.17 - 50 5.5 - 10 Zinc mg.l-1 20 - 272442 2936 100 - 10000 1000 Chromium (total)* mg.l-1 0.009 - 300 18.1 0.02 - 15 0.2 Iron mg.l-1 0.38 - 2700 144 1 - 1000 50 Phenol index mg.l-1 0.01 - 350 26 - 0.006 Hydrocarbons mg.l-1 0.01 - 424 30 - - AOX mg.l-1 44 - 292000 32000 320 - 3350 2000 *leachate substances not influenced by the biochemical state of the landfill matter. Treatment Practices Different methods can be used for treating leachate from landfills, consisting principally of biological, physical and chemical processes. A specific process can only treat a particular substance categories in wastewater. Because of the wide range of pollutants found, leachate treatment has to be performed using a combination of suitable processes. The choice of treatment processes depends closely on the leachate composition. A short description of processes used in Germany for treating leachate follows below. Biological Practices: Biological process can be used to degrade leachate pollutants into mineral end products. To enable degradation specialised microorganisms must be enriched in the bioreactors by proper process conditions. Nitrogen elimination can also be obtained by nitrification and denitrification. Biological processes, especially the aerobic ones, are efficient and cost- effective in comparison with the chemical-physical processes (Rudolph et al., 1988). The activated sludge process and the biofilm process are both use to biologically treat leachate from landfills. Activated sludge : In the activated sludge process micro-organisms aggregate in the form of biological sludge flocs suspended in the wastewater flow, through the treatment plant. The formation of settleable sludge is decisive for the efficient working of the activated sludge process. Leachates though are often characterized by high salt concentrations and high 197 Section 6. Case Studies concentrations of persistent organic compounds, forming a fine, dispersed sludge, which does not settle readily. So the biomass passes through the activated sludge plant without treatment. Under these conditions biological degradation of pollutants is not possible (Albers, 1991, Wilderer et al., 1989). Biofilms: Biofilm systems can be used to prevent the loss of biomass by washing-out, which may be experienced in the activated sludge process. Biomass growth is encouraged by attachment to support surfaces, in form of biofilm. SBBR, the so called sequencing batch biofilm reactors, are also used for cleaning leachate with high salt concentrations and a high percentage of persistent organic compounds. Advantages of the biofilm processes are the small space requirement and the high flexibility in service (Wilderer et al, 1989). Chemical-physical methods: Flotation, precipitation and flocculation, adsorption, reverse osmosis and thermic techniques, belong to the chemical-physical processes for treating leachate. Other chemical-physical processes are chemical oxidation and membrane filters. Flotation: Flotation is used for separating specific low density substances and suspended solid constituents or liquid substances. In a leachate treatment plant they are normally the first step of the treatment process. Precipitation, flocculation and sedimentation: In leachate treatment iron and aluminum salts are usually used to achieve precipitation and flocculation, which is then followed by sedimentation of the settleable material. Using this process, potentially toxic elements in the form of hydroxides and disperse organic substances, are separated with a removal efficiency of 40 %. Adsorption: At a leachate treatment plant, adsorption by activated carbon is always used in combination with biological pretreatment or with a chemical-physical process. Any persistent organic compounds not degraded in the pretreatment step and AOX compounds, can be separated in the back-washed carbon filters. Through adsorption processes, an agglomeration of the solute molecules takes place on the activated carbon interface. Advantages of the adsorption process are; simplicity of the technology involved; relatively low running costs; and possible thermic recycling of the exhausted carbon (Detter, 1998). Regeneration of activated carbon is problematic and expensive though. Chemical oxidation: In the oxidation stage of a leachate treatment plant non-biodegradable and inhibitory organic substances can be oxidised or reduced. In ideal conditions, given a sufficient supply of medium for oxidation, complete mineralisation can be achieved. Substances such as potentially toxic elements and neutral salts remain in solution and are not transformed (Döller, 1998). Hydrogen peroxide/UV or ozone/UV are the mainmedium used for oxidation medium in leachate treatment. In practice, leachate is enriched with ozone (O3) or hydrogen peroxide (H2O2) and afterwards conducted to the UV radiators. Thermal treatment: In thermal treatment pollutants in leachate are separated from water (stripping), concentrated (vapourising) and mineralized (combustion). Due to the different volatilities of water, organic solvent and of dissolved and suspended substances, partition by distillation can be achieved. Thus volatile hydrocarbons contained in the leachate can be separated with a stripping step. With the vapourising process, inorganic and organic residual substances are obtained separately in a chemically unchanged form. Then, the concentrated organic phases must be made inert by combustion. Proper treatment of the exhaust gases is necessary to meet air quality emission standards. During the vapourisation of critically loaded leachate, single toxic halogen organic compounds such as polychlorinated biphenyls, dibenzo-dioxins and dibenzo-furans can enter the distillate. In this case, post treatment with activated carbon is essential (Leonhard, Wilderer,. 1992). 198 Section 6. Case Studies Reverse Osmosis: In the treatment of leachate, reverse osmosis is only used for desalination and concentration of the leachate to be treated. During operation, membrane fouling caused by suspended and colloidal substances has to be prevented, which would otherwise result in a regression of the treatment performance, due to reduction of the permeate flow. At the end the accumulated concentrate must be subjected to additional treatment. The principle advantage of reverse osmosis is the low energy cost. Membrane filtration: The membrane technique has been successfully used for cleaning leachates. A biological process tank is combined with post membrane filtration (nano and ultrafiltration) for biomass retention. The activated sludge tank is in part operated by overpressure in order to reach higher oxygen solubility and with this, a better oxygen supply for the micro-organisms. The removal of treated water occurs continually over a cross-flow membrane filtration plant. The membrane modules are especially capable of finely dispersed sludge retention (Krauth,.1994). Conclusions In Germany, the discharge of wastewater into public WWTS and into receiving watercourses is strictly regulated. Thus, leachate must also be treated before discharging and legal regulations set high requirements on the performance of leachate treatment plants (ATV- A115 1994 and AbwV 1999). Prior to the discharge of non-domestic wastewater into a public WWTS, concentrations of COD, BOD, NH4-N, AOX and potentially toxic elements must be reduced to at least domestic wastewater standards. Purified leachate contributes only a relatively small proportion of the pollutant load WWTS. In ordinary analysis of treated and untreated leachate, only parameters such as COD, BOD and AOX are determined. Additional quantification of high and low volatile hydrocarbons, organic acids, phenols and single organic halogen compounds is necessary to adequately describe the potentially hazardous impact of leachate. In addition, it has to be taken into account that waste products loaded with pollutants frequently arise from leachate treatment: toxic surplus sludge results from biological treatment; charcoal is produced during adsorption processes; and polluted concentrates form during vaporisation. To minimize the problematic emission of pollutants into the environment, additional treatment of exhaust gases and proper disposal of the waste residues are required. 199 Section 6. Case Studies (i) Potentially Toxic Elements (PTE) transfers to sewage sludge Sludges from conventional sewage treatment plants are derived from primary, secondary and tertiary treatment processes. The polluting load in the raw waste water is transferred to the sludge as settled solids at the primary stage and as settled biological sludge at the secondary stage. Potentially toxic elements are also removed with the solids during the primary and secondary sedimentation stages of conventional wastewater treatment. Metal removal during primary sedimentation is a physical process, dependent on the settlement of precipitated, insoluble metal or the association of metals with settleable particulate matter. Minimal removal of dissolved metals occurs at this stage and the proportion of dissolved metal to total metal in the effluent increases as a result. The efficiency of suspended solids removal is the main process influencing the extent of metal removal during primary wastewater treatment. However, the relative solubilities of different elements present in the wastewater are also important (Table i.1). Thus, Ni shows the poorest removal (24 %) during primary treatment whereas 40 % of the Cd and Cr in raw influent is transferred to the primary sludge. Primary treatment typically removes more than 50 % of the Zn, Pb and Cu present in raw sewage. The removal of metals during secondary wastewater treatment is dependent upon the uptake of metals by the microbial biomass and the separation of the biomass during secondary sedimentation. Several mechanisms control metal removal during biological secondary treatment including: · physical trapping of precipitated metals in the sludge floc · binding of soluble metal to bacterial extracellular polymers In general the patterns in metal removal from settled sewage by secondary treatment are similar to those recorded for primary sedimentation. However, the general survey of removal efficiencies listed in Table i.1 suggests that secondary treatment (by the activated sludge process) is more efficient at removing Cr than the primary stage. Operational experience and metal removals measured by experimental pilot plant systems provide guidance on the overall likely removal and transfers to sludge of potentially toxic elements from raw sewage during conventional primary and secondary wastewater treatment. This shows that approximately 70 – 75 % of the Zn, Cu, Cd, Cr, Hg, Se, As and Mo in raw sewage is removed and transferred to the sludge (Blake, 1979) and concentrations of these elements in the final effluent would be expected to decrease by the same amount compared with the influent to the works. Lead may achieve a removal of 80 %, whereas the smallest overall reductions are obtained for Ni and approximately 40 % of this metal may be transferred to the sludge. The majority of potentially toxic elements in raw sewage are partitioned during wastewater treatment into the sewage sludge or the treated effluent. However, atmospheric volatilisation of Hg as methylmercury, formed by aerobic methylation biotransformation processes, is also suggested as a possible mechanism contributing to the removal of this element during secondary wastewater treatment by the activated sludge system (Yamada et al., 1959). Whilst it some of the Hg removal observed in activated sludge may be attributed to bacterially mediated volatilisation, it is unlikely that this is a major route of Hg loss because of the significant quantities of Hg recovered in surplus activated sludge (Lester, 1981). 200 Section 6. Case Studies Table i.1 PTE removals and transfer to sewage sludge during conventional urban wastewater treatment (Lester, 1981) PTE Removal (%) Primary(1) Secondary(2) Primary + Primary + secondary secondary(3) Zn 50 56 78 70 Cu 52 57 79 75 Ni 24 26 44 40 Cd 40 40 64 75 Pb 56 60 70 80 Cr 40 64 78 75 Hg 55 55 80 70 Se 70 As 70 Mo 70 (1)Mean removal (n = 5) from raw sewage and transfer to sludge during primary sedimentation (2)Mean removal (n = 9) in activated sludge from settled sewage (3)Blake (1979) 201 Section 6. Case Studies (J) Effect of chemical phosphate removal on potentially toxic element content in sludge The chemical treatment of wastewater to remove phosphorous is increasingly practised to control P discharges and as a measure to reduce eutrophication of sensitive water courses. This also has the advantage of increased BOD removal, reduction in polyelectrolyte coagulant consumption for sludge thickening, elimination of hydrogen sulphide in sludge digesters and reduced consumption of chemicals for exhaust gas scrubbing (Abendt, 1992). High rates of P removal can be achieved from wastewater using common precipitants such as aluminium sulphate (alum) and ferric chloride although this influences both the quality and quantity of sludge produced (Yeoman et al., 1988). Chemical precipitation also enhances the removal of potentially toxic elements from sewage effluent compared with conventional treatment practices, increasing the transfer of metals to sewage sludge and the content of metals in sludge. For example, Stones (1977) measured the reductions in metal concentrations in sewage effluents obtained after an 18 h settling period with aluminium sulphate (Al2(SO4)3) compared with sedimentation without Al salt. The removal of all the elements examined was increased by the addition of aluminium sulphate compared to the unamended control, except for Ni (Table j.1). The removal of Cu and Zn from the effluent was raised by approximately 50 % by chemical treatment compared with removals achieved by sedimentation without addition of Al. Lead removal increased by about 80 % and the largest overall increase relative to the control was obtained for Cr. In the case of Cr, precipitation with aluminium sulphate increased the recovery of this element in the sludge almost by a factor of three. Table j.1 Effect of chemical precipitation on metal removals (%) from raw sewage after 18 h sedimentation (Stone, 1977) PTE Unamended Al2((SO4))3 Removal relative control at 400 mg l-1 to control (%) Zn 50 73 48 Cu 57 90 56 Ni 19 19 0 Pb 54 96 79 Cr 22 63 193 Iron-based precipitants are marketed for use in wastewater treatment may be derived from industrial by-products of titanium oxide production. Such by-products may contain significant concentrations of potentially toxic elements (PTEs) with potentially undesirable effects on the metal content of sludge (Thiel, 1992). An example of the effects of Fe dosing with industrial by-product on the maximum potential increase in the PTE content of activated sludge is shown in Table j.2. The typical dosing -1 rates of FeSO4 are typically in the range is 15 – 30 mg Fe salt l , but may increase up to 50 mg Fe salt l-1, to comply with the discharge requirements for P in the Urban Waste Water Treatment Directive (CEC, 1991). The calculations suggest that dosing with FeSO4 may potentially increase the Cd content of activated sludge by approximately 300 % to 6 mg kg-1 ds from a typical background value of 1.5 mg Cd kg-1 ds, assuming the maximum likely dose rate of 50 mg Fe salt l-1 and that secondary sludge production is equivalent to 250 mg l-1 of total solids (UKWIR, 1997). The Ni content in activated sludge may theoretically increase by 130 % compared to sludge without Fe addition, whereas Pb and Zn concentrations may increase by about 10 % with Fe dosing. These increases in sludge content remain well within the current quality standards for agricultural use (CEC, 1986). However, the revision of the Directive on land application (CEC, 2000b) will introduce more stringent limit values for PTEs and the use of Fe salts from industrial processes could potentially penalise the acceptability of sludge for use in agriculture under the new regulatory regime. Furthermore, 202 Section 6. Case Studies the potential increase in the metal content of sewage sludge, resulting from the use of industrial-grade chemical precipitants, could also be considered as unsatisfactory because it erodes the beneficial reductions in metal inputs that have been achieved through the successful control of trade effluent discharges. The quality and metal content of low-grade chemical precipitants for use in wastewater treatment should be examined to ensure that they do not significantly increase the metal content of sludge. In Germany, for example, composition standards are recommended for Fe and Al-based coagulants used for sewage treatment and sewage sludge conditioning (Schumann and Friedrich, 1997). The use of potable water grade Fe salts should be considered for sewage treatment (Thiel, 1992) to avoid potential problems associated with contamination with potentially toxic elements. In practice, there are few published data on the effects of chemical precipitants on sludge metal contents and Fe and Al dosing. One example from the literature (Yeoman et al., 1993) showed no consistent effects of chemical treatment with Al or Fe salts on potentially toxic elements in sewage sludge from Beckton WWTS in the UK (Table j.3). However, the significance of the direct metal inputs in chemical precipitants will increase as industrial discharges are effectively controlled and as diffuse inputs from domestic sources and run-off become the predominant sources of potentially toxic elements entering the wastewater collection system. Table j.2 Metal concentrations (mg kg-1) in Fe precipitants and activated sewage sludge (UKWIR, 1997) PTE FeSO4 Increase in FeCl2 salt Increase in Activated salt activated activated sludge sludge due sludge due without Fe to FeSO4 to FeCl2 Zn 348 70 26 5.2 600 Cu 5 1.0 51 10 400 Ni 160 32 120 24 25 Cd 22 4.4 3.0 0.6 1.5 Pb 64 13 22 4.4 110 Cr 32 6.4 236 47 - Table j.3 Effect of chemical P removal on the PTE content of sludge digested sewage sludge(1) (adapted from Yeoman et al., 1993) Sludge type Salt addition Concentration (mg kg-1) ds Cd Cu Ni Pb Zn Digested None 5.4 159 24 137 231 Digested + Al Raw sludge 5.4 142 19 62 148 Digested + Al Activated sludge 4.5 254 22 168 184 Mean 4.9 198 20 115 166 Digested + Fe Raw sludge 8.5 195 40 253 300 Digested + Fe Activated sludge 4.8 95 18 155 121 Mean 6.6 145 29 204 211 (1)Sludge was collected from Beckton WWTS, sludges were digested in laboratory scale digesters (75 % raw sludge, 25 % activated sludge) Waste products from water treatment and industrial processes, incinerator ash and acid mine drainage have potential for re-use as P precipitants in wastewater treatment processes (Fowlie and Shannon, 1973). For example, Oostelbos et al. (1993) treated Fe-enriched sludge from water treatment with hydrochloric acid to convert ferric hydroxide to ferric chloride for use in sewage treatment for phosphate removal, and as a dewatering agent in 203 Section 6. Case Studies sludge conditioning. Verberne (1992) considered that the use of water treatment sludges as chemical precipitants for P removal was technically feasible and would depend on the agreement and acceptance of the approach by water and sewage treatment authorities. The recovery of Fe and Al from acid mine drainage is another source of chemical precipitants that can be used for P removal during sewage treatment (Bouchard et al., 1996). The re-use of secondary resources for precipitating P during wastewater treatment is intuitively attractive and also alleviates the environmental problems and impacts associated with disposal of those wastes. However, some product types derived from waste materials are potentially contaminated with potentially toxic elements that accumulate in the sludge (Fowlie and Shannon, 1973). Therefore, the metal content of waste derived products should be established, and the potential consequences for sludge quality determined, before a particular product is accepted for use as a chemical precipitant in wastewater treatment. 204