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Section 6. Case Studies

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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


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(a) Platinum Group Metals in the Urban Environment

Introduction

The platinum group of metals (PGMs), sometimes referred to as the platinum groupelements (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 perbillion (µg/kg) in the earth’s crust. The elements are noble chemically unreactive metals, andare found in nature as native alloys, consisting mainly of platinum.

Recently these metals have gained importance as industrial catalysts including vehicleexhaust catalysts (VECs). This use and possible implications for human health were thesubject of an earlier review undertaken by Imperial College, London for the UK Departmentof the Environment (Farago et al, 1995; 1996).

Increasing understanding of the environmental damage of vehicle emissions has led to theintroduction of stringent emission control standards throughout the western world. Since1974 all new cars imported or produced in the United States have had catalytic convertorsfitted, cutting down hydrocarbon and carbon monoxide emissions. In 1977 they were fitted toa substantial proportion of all cars sold in America, where at the time, this applicationaccounted for 32% of the total Pt usage (Herbert, et al., 1980).

Vehicle exhaust catalysts have also been used in Japan since 1974. Vehicle exhaustcatalysts were also introduced in Germany in 1985, in Australia in 1986, and into the UK atthe beginning of 1993 in response to the emission standards equivalent to the US standardswhich were introduced in the EC at that time. Other uses of PGMs are noted in latersections.

Sources

The PGMs are found in nickel, copper and iron sulphide seams (Bradford, 1988). They arecurrently mined in South Africa, Siberia and Sudbury, Ontario. World mine production of thePGMs, of which 40-50% is platinum, has steadily increased since 1970. This reflects theincreasing world-wide use of PGM vehicle catalysts (IPCS, 1991). From 1988-1992 worldmine production was essentially constant at around 255 tonnes per year (WMS, 1994). Theamount of PGMs present in the earth’s crust down to a depth of 5km, and hencetechnologically attainable, are still enormous when compared with present requirements, butonly a fraction of the pertinent ores is sufficiently rich for commercial exploitation. Of the totalof 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 tonnesrespectively for Pd 230 tonnes and 224 tonnes respectively, and for Rh 14.2 tonnes and20.9 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, withadditional major uses in the chemical industry, electrical and electronics industries,petroleum industry, the manufacture of jewellery, as a cancer treating drug in medicine, asalloys in dentistry and in the glass industry.

Demands by application for 1999 and 2000 for PGMs are shown in Table a.1. (JohnsonMatthey, 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 51000Autocatalysts: recovery -12000 -13000Jewellery 79400 83300Industrial 38400 41400Investment 5100 -1420Total Demand (Pt) 159000 146000

PALLADIUM

Autocatalysts: gross 166700 146000Autocatalysts: recovery -5530 -6520Dental 31500 24700Electronics 56100 58600Other 16600 15000Total Demand (Pd) 265000 238000

RHODIUM

Autocatalysts: gross 14400 16000Autocatalysts: recovery -1870 -2240Chemical 964 992Electronics 170 170Glass 851 1050Other 312 312Total Demand (Rh) 14900 16200

RUTHENIUM

Chemical 2440 1930Electrochemical 2040 2270Electronics 5560 6580Other 1160 1360Total Demand (Ru) 11200 12100

IRIDIUM

Automotive 964 397Chemical 198 170Electrochemical 794 680Other 936 1450Total Demand (Ir) 2890 2690


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Trends over time in platinum and palladium uses by application for Europe are shown inTable 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 17900Autocatalyst:recovery -142 -284 -567 -851 -1130

Chemical 1420 1420 1700 1700 2410Electrical 851 709 709 1280 2270

Glass 425 851 1130 709 709Investment:small 992 1276 142 142 0

Jewellery 2410 2840 3540 4540 5670Petroleum 567 709 425 425 284

Other 1560 1840 2130 2410 2840Totals (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 -425Chemical 2100 1700 1840 1840 2690

Dental 8500 7230 7230 5950 3120Electronics 5950 7230 8500 7660 7370Jewellery 992 851 851 1420 1280

Other 425 709 567 709 567Totals (Pd) 19100 25100 43200 56300 66200

Of particular interest is the increased demand for palladium in Europe, largely in response tothe introduction of Euro Stage III legislation from January 2000; palladium – rich catalystswill meet stricter emission limits for petrol models, resulting in a further move away fromplatinum technology (Johnson Matthey 2000).

Catalytic convertors

A catalytic converter is a unit about the size of a small silencer that fits into the exhaustsystem of a car. The metal catalyst is supported on a ceramic honeycomb monolith andhoused in a stainless steel box similar in shape to that of a conventional silencer. About 1-3gof PGM is contained in some vehicle exhaust catalysts, approximately 50g of PGM per cubicfoot of catalyst (Steger, 1994). Due to the commercial sensitivity of these products it isdifficult to obtain data on the exact amounts in each of the many different formulations ofcatalyst. The honeycomb made of cordierite contains 300 to 400 square channels persquare inch (6.45cm2), and is coated with an activated high surface area alumina layercalled the washcoat (Farruato, 1992) containing small amounts of the precious metals,platinum, palladium and rhodium in varying proportions. The conventional three-waycatalysts 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) andnitrous oxides (NOx) into carbon dioxide (CO2), water (H 2 O) and nitrogen (N2). Platinum isan effective oxidation catalyst for carbon monoxide and hydrocarbons, but it is moresensitive to poisoning than palladium and so can only be used in cars which use unleadedpetrol. Palladium is becoming increasingly used instead of platinum due to the higher costsof the latter. The rhodium oxidises the hydrocarbons and reduces the NOx emissions. Base


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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 laserablation and analysis by ICP-MS for 31 elements (Rauch et al, 2000). Variations incomposition were found to occur in agreement with the known fact that variations occur fromone manufacturer to another and from one year to another. An association between PGMsand Ce in road sediments was ascribed to the emission of PGMs as abraded washcoatparticles 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 isaccounted for by vehicle catalysts (IPCS, 1991). The recovery of spent autocatalysts fromvehicles at the end of their lives is regarded as important and substantial secondary sourcesof platinum as well as palladium and rhodium (Torma and Gundiler, 1989). The quantity ofspent autocatalysts greatly increased in the United States from 1984 to 1988. Thesescrapped autocatalysts present an important secondary source of the platinum groupmetals.

On current projections it is expected that 3.5 million catalysts will be available for recycling inthe UK by 2000.

Platinum group metals in the environment.

The average concentration of platinum group metals in the lithosphere is estimated to be inthe 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, 0.0001mg.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 wasreported by Lee (1983) between 1948 and 1973, it seems unlikely that this was connectedwith 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 andPd accumulation and behaviour in urban coastal sediments (Tuit et al, 2000). Increasedlevels of both metal of approximately 5 times above background concentrations wereascribed to anthropogenic activity with catalytic convertors a major source. It was concludedthat anthropogenic enrichments can significantly influence coastal marine inventories ofPGMs. The study also indicated that Pt associated with catalytic convertors is much moresoluble than expected or alternatively that there is an additional source of dissolved Pt to theharbour. 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 dueto thermal or mechanical strain and acid fume components, and are intensified byunfavourable operational conditions (misfiring, excessive heating) which may even destroythe converter (Schäfer and Puchelt, 1998).

Emission rates range between several ng and µg of Pt per km driven depending on whetherthey were measured in motor experiments or calculated on the basis of environmentalconcentrations (König et al., 1992). Platinum is mainly emitted as a metal or an oxide with


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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 etal, 1996).

Mostly inert under atmospheric conditions, the reactivity of Pt increases significantly if thesenanoparticles are brought into contact with soil components. Lustig et al., (1996)demonstrated that humic substances considerably enhance the reactivity of Pt clusters inthe nm-range under atmospheric conditions. Using road-dust from a tunnel, it wasdemonstrated that within hours Pt can be fixed to several humic acids with differentmolecular weights. The low solubility of Pt in deionized water increases significantly evenunder reducing conditions when certain anions or complexing agents are used (Nachtigall etal., 1996).

A detailed study has been undertaken in several sites in southwest Germany, selected onthe basis of traffic density and morphology, including roads in Stuttgart with 120,000 vehiclesper day and near Heidelberg with 100,000 vehicles per day (Schäfer and Puchelt, 1998). Atthese two locations, Pt concentrations in the 0-2 cm surface soil adjacent to the road rangedfrom several hundred µg/kg to local background values (≤ 1 µg.kg-1) at less than 20 m fromthe road. Maximum Pd and Rh values were 10 and 35µg.kg-1 respectively. The PGMconcentration decreased significantly with depth.

However the maximum PGM concentrations in soils at Heidelberg were only 25 percentthose at Stuttgart, even though the traffic density was only 20% lower. The authorssuggested that this could be due to frequent traffic jams at the Stuttgart site “leading toexcessive emissions due to unfavourable working conditions of the engines”.

Urban road dusts collected in Stuttgart at the same time showed concentrations of Ptranging up to 1000 µg.kg-1, 110 µg.kg-1 Rh and 100 µg.kg-1 Pd; these reflect short-terminputs of PGMs. A ratio of around 6 Pt: 1 Rh in traffic influenced soils and dusts has beenreported by Schäfer et al (1996).

Schäfer et al (1999) measured time-dependent changing PGM depositions and contents ofdusts and soils at a typical urban location at Karlsruhe in Germany. Daily deposition rates at2 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 thesteady inputs. Using data for Pt concentration in soil at a site near Pfarzheim and a dailypassage of around 15,000 Pt emitting cars per day, the authors calculated for a total numberof 11 million converter-equipped vehicles over 2 years, a total emission of at least 3,000 ngof 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 stationarymotor vehicle experiments (König et al, 1992).

A recent estimate of total Pt emission in the vicinity of roads in Germany over the period1985-2018 was 2,100 kg (using emission factors of 0.65 µg.kg-1 for highways, 0.18 µg.kg-1

for federal and national streets and 0.065 µg.kg-1 for district and city streets) (Helmers andKummerer, 1999). These different emission rates reflect the increase in Pt load of exhaustswith increasing speed of the car.

Accumulation of Pt was clearly shown in road dusts and surface soils adjacent to roads inthe UK in 1994 (Farago et al, 1996, 1998). In the heavily trafficking London Borough of


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Richmond, Pt concentrations ranged up to 33 ng.g-1 in road dusts and 8 ng.g-1 in soils. Pt inroad dusts was highest at major road intersections (mean 21 µg.g-1) compared with alongmajor roads (13 ng.g-1) and intermediate and minor roads (2ng.g-1). The local backgroundconcentration 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 ingarden 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.5and Figure a.1) with values ranging up to 298 ng.g-1 and 556ng.g-1 respectively for Pt and Pdin 1998 .

Table a.4 Summary results for Pt in garden soils (0-5cm) and road dusts fromNottingham (ng.g-1) aResidential streets with low traffic densities; b Includes major roadswith high traffic densities (from Hutchinson, 2001)

Year Sample N Range Mean Geomean MedianNottingham1982 Soil 42 0.27-1.37 0.61 - 0.591996 Soil 42 0.19-1.33 0.80 0.75 0.73

1982 Road dust 10 0.46-1.58 0.90 0.80 0.751996 Road dusta 8 0.82-6.59 2.78 2.29 2.061998 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 trafficdensities (from Hutchinson, 2001)

Year Sample N Range Mean Geomean MedianNottingham

1982 Soil 42 0.64-0.99 0.05 - 0.041996 Soil 42 0.21-1.11 0.18 - 0.10

1982 Road dust 10 0.69-4.92 1.24 - 0.221996 Road dusta 8 0.19-1.43 0.75 0.64 0.601998 Road dustb 20 5.6-556.3 92.9 40.95 35.84

An EU-funded study under the Environment and Climate Programme, CEPLACA, involvedlaboratories in Madrid, Gothenburg, Sheffield, Rome and Neuherberg. Changes in catalystmorphology over time were studied using SEM/EDX and laser induced breakdownspectrometry (LIBS) (Palacious et al, 2000). Catalysts were used up to 30,000 km in a rollerdynamoneter following a driving cycle representing urban and non-urban driving conditions.Releases of PGMs were found to decrease with time. For new petrol catalysts meanreleases were 100, 250 and 50 ng.km-1 for Pt, Pd and Rh respectively. In diesel catalysts Ptrelease ranged from 400-800 ng.km-1.

The effect of catalyst ageing was large. At 30,000 km releases were reduced to around 6-8ng/km Pt, 12-16 ng/km Pd and 3-12 ng/km Rh for petrol catalysts. In diesel catalysts, the Ptrelease ranged from 108-150 ng/km.


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The difference between diesel and petrol was ascribed to the different composition of thewashcoat and the different running conditions of diesel engines.

Soluble forms of PGMs emitted (in dilute HNO3) were significant for the fresh catalyst butless than 5% of the total amount. A previous study had reported 10% of the total Pt emissionto 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 0km. One possible explanation suggested for the relatively high amount of soluble PGMsrelated to the relatively high chloride concentration in fresh washcoat (i.e. one examplequoted of 3.4 wt %). The authors suggested that the formation of soluble PtCl6, PdCl2 , PdCl4or RhCl3 could be favoured at the high temperature and humidity that can be reached in thecatalyst. The chloride concentrations in aged catalysts are normally very much lower.However, further laboratory tests using spiked solutions showed the instability of thesechloro-complexes in the final exhaust fumes solution. It was thus thought possibly that thesoluble 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 observedbetween the labelled amount of PGMs in the different catalysts studied and the measuredamount 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 lackof an observed correlation”.

Emissions of Platinum in effluents from hospitals

Effluents from hospitals contain platinum from excreted anti-neoplastic drugs, cisplatin andcarboplatin, though workers in Germany have concluded that these are only of minorimportance to environmental inputs from other sources and in particular from the use ofcatalytic convertors (Kümmerer and Helmers, 1997). These drugs were introduced 25 yearsago to treat various tumours and are usually administered in the hospital environment. Theplatinum passes into hospital sewage which is then treated with household sewage inWWTS.

The authors monitored effluent samples from the University Hospital of Freiberg and twocommunal hospitals and found total inputs of platinum of around 330g/year from theUniversity Hospital and 12 g/year from the Community Hospital. These equated to aconsumption per bed per day of 600 µg Pt for the University Hospital and 85 µg Pt thecommunity hospital. Extrapolation on a national basis, this amounted to an upper limit for theinput in Germany of 141 kg Pt per year (c. 645,000 hospital beds (German StatisticalFederal Agency, 1994) and a lower limit of 20 kg/year, with an average calculated value of28.6 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) ReferenceCatalytic converters

emissions15 König et al., 1992

Zereini, F., personalcommunication 1996

Hospital effluents 28.6 Kümmerer and Helmers, 1997Sewage 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 Germanyaimed to provide reliable data with which to quantify sources of platinum in the environmentfrom hospitals with other sources (Kümmerer et al, 1999). This study was supported by theLIFE95/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 willtherefore end up in hospital effluents. Pt concentrations measured in the total effluent of thedifferent hospitals ranged widely from less than 10 ng.l-1 (the detection limit) in the Belgianand Italian hospitals to cca. 3,500 ng.l-1 for the Austrian and German hospitals. In all casesthe influent of the WWTS was below 10 ng.l-1 as a result of dilution within the waste watersystem.

Annual emissions by hospitals and cars in Germany, Austria and the Netherlands are listedin 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=TheNetherlands)

D 1994 D 1996 A 1996 NL 1996Total hospital beds (approx) 645000 645000 77500 60000Maximum Medical Performance 45000 45000 6500 N/APt per bed and year (mg) - maximum medical service spectrum- medium medical service spectrum

154.014.0

130.414.0

58.7N/A

22.3N/A

Pt emissions by hospitals- maximum medical service spectrum- medium medical service spectrum

6.98.4

5.88.4

0.38N/A

1.3N/A

Total emissions by hospitals (kg) 15.3 14.2 N/A N/AAll hospitals as maximum medical servicespectrum

99.3 84.1 4.6 1.3


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Table a.8 Emissions of platinum by carsD 1994 D 1996 A 1996 NL 1996

Number of cars 32 000 000 32 000 000 3 593 588 5 740 489With catalytic converter 12 800 000 19 200 000 1 607 699 3 307 300

% with catalytic converter 40.0 60.0 44.7 57.6Kilometres/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 0.5Total emission by cars (kg) 124.80 187.20 11.55 22.46

Table a.9 Platinum emissions comparison: hospitals vs. carsD 1994 D 1996 A 1996 NL 1996

Maximum medical service spectrum 5.6 3.1 3.3 6.0Medium medical service spectrum 6.7 4.5 N/A N/A

Total 12.3 7.6 3.3 6.0All hospitals calculated as maximum

medical service spectrum79.6 44.9 39.4 6.0

In this study the highest concentrations of Pt in WWTP influents were found at the beginningof rain periods and at the end of cold periods when snow was melting. It was then concludedthat the main inputs of Pt into municipal sewage were from urban and road run-off fromtraffic and other Pt emitting sources and not from hospital emitted sewage (Kümmerer et al,1999).

Emissions from other SourcesKümmerer et al (1999) further concluded that “emissions by traffic and hospitals cannotexplain the whole amount found in sewage and other sources emitting platinum directly intosewage have to be considered like glass and electronics industries or jewellerymanufacturing. For the catalytic ammonia oxidation 92 kg platinum are reported to be lostfrom the catalyst” (Beck et al., 1995). If all of this is emitted into the atmosphere andwashed off from roads and other paved areas in urban areas, which make up 11% of thetotal 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) whichuse platinum to a certain extent they might be the most important local contributor to theplatinum content of a certain municipal sewage and sewage sludge. Thus, unspecified inputdirectly from industrial processes into sewage must be taken into account. These possiblesources include jewellery manufacture, dental laboratories, electronic industries, glassmanufacturing, production of platinum-containing drugs and industrial catalysts.

Knowledge on these sources, the species involved and their environmental properties issparse if not non-existent.

A study of PGMs in sewage sludge incineration ashes from the municipal WWTS atKarlsruhe, Germany, showed Pd concentrations to have increased from 64 to 138 µg.kg-1

from 1993 to 1997, with Rh increasing from 4.8 to 6.3 µg.kg-1; Pd varied from 300 to 450µg.kg-1 with no significant trend although these concentrations were 10-fold higher than in1972 (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 samplesinfluenced by traffic emissions. They then estimated that, as more than 90% Rh is used forthe production of catalytic convertors, and as Pt and Rh are emitted in a ratio of 6:1, that thecontribution of traffic to the Pt concentration in sludge is only c. 30%, a result similar to thatfound in Munich by Laschka et al (1996). They thus suggested that the greater part of Pt insewage sludge must come from sources other than catalytic convertors, such as hospital


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and medical effluents or industrial emissions. They drew attention to the fact that “in citieswith a jewellery industry, noble metal concentrations in sewage sludge far exceeded normalvalues even before the introduction of catalytic convertors” (Lottermoser, 1994).

Laschka and Nachtwey (1997) analysed Pt on primary and secondary effluents and inprimary and digested sludge from two sewage treatment plants in Munich, where platinumpollution from industry, hospitals and traffic is considerable. Samples were taken before andafter rainfall in October 1994 and July 1995. In general Pt concentrations in effluents werehigher during rainy weather compared to dry weather. The Pt loading in secondary effluentswas lowest for the period Monday/Tuesday (i.e. after the weekend), which was consideredtypical for industrial loads; the authors concluded that during dry weather, the platinum loadoriginated mainly from industry. Comparison of the average Pt load in primary andsecondary effluents in Munich, indicated a removal rate of 74% and 70% in the treatmentplants. These elimination rates were lower than those typical for other metals such as Pband Cd, which was attributed by the authors to the stabilizing effect of chloride (>100mg.l-1 indomestic sewage, or to the low Pt content of untreated sewage (<0.1 µg.l-1).

This study confirmed the enrichment of Pt in sewage sludge, which was present in materialsfrom the 2 Munich treatment plants in concentrations ranging from 86 to 266 µg.kg-1.Sludges from other large towns and centres of industry had previously been found to contain10 to 130 µg.kg-1 Pt and from smaller rural plants <10 to 50 µg.kg-1 Pt (Lottermoser, 1994).In this earlier study, an exceptionally high value of 1070 µg.kg-1 had been found in sludge atPforzheim, a town with a jewellery industry.

The study of Laschka and Nachtwey (1997) concluded that “in a large industrial centre suchas 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 thesources, pathways and sinks of Pt in the environment. The authors noted that there was asyet not enough data to reliably investigate Pd and Rh fluxes, noting the lack of good qualityassurance for Pd analysis and the paucicity of environmental data on Rh. An examination ofarchived sewage sludge ash from Stuttgart, Germany showed a continuous increase in Ptconcentrations since 1984. With an estimated 5 x 1010 kg of sewage sludges for Germany inthe early 1990’s and c. 250 mg.kg-1 Pt in sewage sludge, this amounts to some 12,500 kg ofPt, some 2 orders of magnitude higher than the Pt flux emitted by traffic. Much smaller Ptconcentrations (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 intoaccount the amounts of Pt in effluents from hospitals being completely received by sewagesystems, the authors calculate that for Germany Pt received in the influents of WWTS fromboth these major sources amounted to 42.9 kg in 1994 and 56.4 kg in 1996. They thusconsidered 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 onagricultural land and 70% disposed of as sludge or incinerated sludge ash. Losses to theatmosphere from incineration are not yet known.

Table 10. Partition of anthropogenic Pt fluxes (in kg) within German WWTSYear Received by

the WWTSRemaining inthe sewage

sludges

Disposal withsludges or

ashes

Depositedagriculturallywith sludges

Released intofreshwater

1994 42.9 30.9 21.6 9.3 121996 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 otherEuropean countries, taking into account traffic densities, catalytic convertor policies etc., withthe qualification that “since there is no highway speed limit in Germany, highway Ptemissions 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 particlesremain bound to these and have limited mobility in the road and soil environment (Rauch etal, 2000). Experimental studies under laboratory conditions, in which ground catalysts havebeen added to soils under varying conditions of pH, chloride and sulphur concentrationshave indicated that post-deposition processes in soils and waters are of minor importanceand that “the risk of a health endangering contamination of the environment, and especiallygroundwater, 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 beenruled out and indeed Rauch et al (2000) suggest that this may occur “in the roadsideenvironment, 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 solublePGM 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 theenvironment 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 anddiesel) and subsequent runoff from road surfaces and roadside soils;

• emissions in effluents from hospitals using the anti-cancer drugs cirplatin andcarboplatin,

• 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 andPd into the environment and into WWTS is from vehicle exhaust catalysts, with hospitaleffluents accounting for some 6 to 12 per cent Pt. In large industrial centres, such asMunich, inputs from other sources (presumed industrial) may exceed those from catalyticconvertors 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 below, though reactions within the soil/dust and wastewater environments need further study.Interactions with chloride ions and humic substances may well increase solubility and thusbioavailability.

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 ispractical, studies into uptake into pasture and foodcrops are recommended. The 30 per centof Pt emitted into freshwater systems will potentially increase Pt levels in drinking watersupplies.

At present there is no evidence of health risks arising from increasing levels of PGMs in theroadside environment, in sewage sludge or in drinking water. However, as levels of usecontinue to rise, it would seem prudent to focus research into factors influencing theirsolubility and bioavailability, their uptake and input into food crops and drinking water andinto 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 discussioninvolving maintenance and quality issues. This review will provide a brief overview ofavailable techniques and structures and summarise their design characteristics. Adiscussion on performance will focus on water quality but comments on maintenance willalso be included to allow the reader to form an overall opinion. A number of source controlapplication 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 usualapproach to solving it has been a system of buried pipes designed to convey waterdownstream as soon as possible (CIRIA, 1999). Several problems in this traditionalapproach have been identified including possible flooding in downstream areas due toalteration of natural flow patterns, water quality issues that are not dealt with within the pipesystem and largely ignored amenity aspects (such as water resources, landscaping potentialand provision of varied wild habitat). These considerations have led to an effort of rethinkingsurface 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, arecollectively referred to in the UK as “sustainable urban drainage systems” (SUDS) (CIRIA,1999) or more generally “source control” (Urbonas and Stahre, 1993) or “best managementpractices” (BMPs)1 (Jefferies et al., 1999). They essentially confirm with the emergence ofAgenda 21 as a local action-planning basis for strategic and integrated approaches “to haltand reverse the effects of environmental degradation and to promote sound environmentaldevelopment” (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 thefollowing paragraphs.


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Basic design characteristics and principles of use of the most widely used of these systemswill 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 CIRIArecommendations: (a) filters and swales, (b) permeable surfaces (c) infiltration devices and(d) ponds (CIRIA, 1996; CIRIA, 1999). The overall structure of an urban catchment withsource 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 withsmooth surfaces and downhill gradient.Swales are long shallow channels while filtersare gently sloping areas of ground. Theymimic natural drainage patterns slowing andfiltering the flow and are used for the drainageof small residential areas and roads. The flowdepth should be smaller than the height of thegrass to ensure filtration. Operationalpractices include regular mowing and clearinglitter. Special care should be taken not toallow the swale to erode after heavy storms.Grass swales have been used extensively inNorth America, but have only recentlyappeared in Europe. Information on treatmentperformance comes mainly from the US (assummarised in Ellis, 1991) and indicatesremoval potential for solids, potentially toxicelements and hydrocarbons). In the UKhowever the quality improvement potential ofthe swales is ignored. A typical swalestructure can be seen in Figure b.2.

Figure b.2. Grass Swale (after CIRIA 1996)

Permeable surfaces:These include porous pavements, gravelledareas, grass areas and other types ofcontinuous surfaces with an inherent system ofvoids. The water passes through the surface tothe permeable fill, allowing for storage,transport and infiltration of water. The actualamount of water stored is dependent on thevoids ratio, the plan area and the structure’sdepth. It acts as a trap for sediment thusremoving a large number of pollutants from therunoff, but keeps them within the particular site.The principal mechanism for pollutant retentionis thought to be adsorption onto materialswithin the pavement construction (Pratt, 1989).Maintenance should ensure that the voids arenot filled by sand and silt and such anoperation may prove costly, as the surfacestructure can deteriorate under externalpressure. The US, France, Holland, Austriaand Sweden have used porous surfaces forboth traffic and pavement areas (Diniz, 1976;Hogland, 1990). A typical structure can beseen in Figure b.3.

Figure b.3. Porous Pavement (after CIRIA1994)


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Infiltration devices:Soakways and infiltration trenches are belowground and are filled with a coarse material.They drain water coming in the infiltrationdevice from a pipe or a swale directly to thesurrounding soil. Their operation is based onincreasing the natural capacity of the soil forinfiltration but effectiveness is ultimately limitedby soil permeability. The volume of storagetherefore is dependent on soil infiltrationpotential. Physical filtration can remove solids,while biochemical reactions caused bymicroorganisms growing on the fill or the soilcan degrade hydrocarbons. The level oftreatment depends on the size of the mediaand the length of the flow path (CIRIA, 1999).Extensive use of soakways in Sweden and theUS as well as in the UK generally providepositive feedback on maintenance andoperation (Pratt, 1989; CIRIA, 1996). Areasthat are drained through infiltration structures ofdifferent types include car parks, roads, roofspavements and pedestrian sidewalks. Pollutionlevels in these types of urban runoff can behowever significant and there is thereforeserious risk of introducing the pollutants to thegroundwater. Additionally the introduction ofwater to the soil may cause geotechnicalproblems. Figure b.4 describes a typicalsoakway.

Figure b.4. Soakway (after CIRIA, 1996)

(d) Basins and Ponds:These are areas of storage of surface runoffthat are free from water under dry weatherconditions. Structures can be mixed with apermanently wet area for wildlife ortreatment of runoff and an area that isusually dry to allow for flood attenuation.The ponds are normally situated near theend of the system due to detention and landprice constraints (Makropoulos et al., 1999).Flow detention would lead to settlement ofthe particles and associated pollution loads.Additionally some bacterial die-off andsoluble particle removal could be expected(CIRIA, 1994). Annual clearance of theaquatic vegetation and silt-removal everyfive to ten years should be thought of as anaverage operational practice. Figures b.5and b.6 give an idea of on and off streamdetention and the pond-wetland principlerespectively. Figure b.5. Typical on and off stream

storage ponds (after CIRIA, 1994)


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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. Researchhas been focusing on mathematical modelling of the system’s quality performance andactual data is generally not available in Europe. Scotland is a notable exception. BMPs havebeen promoted for the past five years in response to the need to combat pollution fromdiffuse sources in urban areas. To meet this need, a programme of investigations is beingundertaken into the performance of BMPs, which have been built in Scotland. An initialawareness survey by Abertay University and SEPA indicated high levels of apparentknowledge of BMPs, but subsequent investigations showed that in many instancesknowledge was very superficial and often inadequate. Jefferies et al. (1999) discussexperimental findings and theoretical considerations of that investigation and show that, inmost systems, pollutants will form sludge. This in turn must be disposed of, and indeed goodhousehold practices may be the only truly sustainable drainage practice. Pollutants removedfrom runoff in a system such as a pond may accumulate in sediments and biota. Potentiallytoxic elements and trace organics in rainfall runoff are to some extent associated with soilparticulates, as discussed earlier in this report, and will thus tend to be removed bysedimentation. The soluble fraction of pollutants will also to some extent precipitate followingchanges in pH, oxidation-reduction potential or temperature (Kiely, 1997). The activity of thepollutants however is not ended with their concentration in the sediments. Polluted sedimentmay 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 BMPsystems. Studies in the US have shown that, when disposed in soakways,organophosphates have appeared in watercourses 400 metres away only two hours afterdisposal (ENDS, 1993). The entire range of toxic pollutants identified as possible input tourban rainfall runoff may leak this way to the groundwater. A new problem has appeared inthe form of methyl butyl tertiary ether (additive to unleaded petrol), which is ten times moresoluble in water than other constituents in petrol and thus would tend to spread readily ingroundwater (Kiely, 1997). When soil is used as a filtration medium in source controlsystems (as in infiltration trenches and even grass swales), it must be regularly checked asthe adsorbed pollutants may be remobilised under various conditions. Furthermore, possibledegradation of pollutants inside the systems may give rise to hazardous by-products whichmay be more soluble or toxic than the original forms (Hallberg, 1989). Biotransformation ofTCE for example results in hazardous products such as vinyl chloride, which is a confirmedhuman carcinogen (Burmaster, 1982). Table b.1 summarises the main functions and waterquality 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 PRIMARYFUNCTION

SECONDARY WATER QUALITY ATTRIBUTES

Infiltrationpavements

Collection anddisposal of

surface water

Sediment andpollutant removal

Can remove pollutants associatedwith sediments and dissolved

pollutants but may be lead to increasein nutrient levels

Swales Conveyance ofsurface waters

Storage sediment andpollutant removal;

disposal

Can remove suspended and possiblydissolved pollutants but may be a riskto groundwater quality if not sealed

Infiltrationbasins

Disposal ofsurface water

Storage; sediment andpollutant removal

Can remove suspended and possiblydissolved pollutants but may be a risk

to groundwater qualityStorageponds

Storage ofsurface water

Sediment andpollutant removal

Can remove pollutants associatedwith sediments and provide some

biological treatmentWetlands Pollutant removal Storage Can remove and treat various

pollutants

France has also had experience in particular aspects of SUDS (for detention basins andponds). Nascimento et al. (1999) provide an overview of the French experience in detentionbasins 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 thesefacilities are increasingly popular as indicated in Deutsch et al. (1990) (Reported inNascimento et al., 1999). Recent research has focused in the quality side of performance ofthe 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 urbanwet weather discharge that have been collected in 19 catchments since 1970. The efficiencyof detention basins in reducing pollutants is the result of a large number of variablesincluding physical, chemical and biological characteristics of pollutants, precipitation regime,detention time and quality of maintenance services (Nascimento et al., 1999). Table b.2identifies 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 bythe French institution CEMAG-REF. The presented data were for basins installed inseparate drainage systems and the results indicate that such storage facilities can have areasonable performance even over quite short time scales.

Table b.2. Efficiency of detention basins (after Nascimento et al., 1999)

Yearly Inflow(kg/ha imp)

Yearly Outflow(kg/ha imp)

Reduction(%)

Reduction after2h(%)

Pb 0.893 0.054 94.0 65Zn 5.12 0.66 87.1 77Cd 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 theurban context is concerned, the outcome may be very different according to the specificcase. For the UK the high pollutant loading to urban detention basins, which has beenreported, 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


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results of measurements taken in a field scale infiltration trench. Figure b.7, summarisessome of the findings, indicating significant potentially toxic element removal efficiencyexceeding 80% after 1 year of runoff loadings.

The Technical University of Denmark (Mikkelsen, et al. 1996a and 1996b) has been involvedin a series of tests to examine the effects of stormwater infiltration on soil and groundwaterquality. They found that potentially toxic elements and PAHs present little groundwatercontamination threat, if surface infiltration systems are used. However, they express concernabout pesticides, which are much more mobile.

Recent and ongoing studies in the US have tried to identify the potential hazards from theuse of infiltration systems.

Figure b.7 Infiltration Trench removal performance and % of influent exfiltrated tosoil for a series of 4 runoff events compared to bench scale results (lab) after 1 yearof equivalent loading (after Sansalone, 1999)

In particular, a multi-year research project sponsored by the US EPA addresses the potentialproblem 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 researchfound that heavy repetitive use of mobile pesticides, such as EDB, on site with infiltrationdevices likely contaminates groundwater. Fungicides and nematocides must be mobile inorder to reach the target pest and hence, they generally have the highest contaminationpotential. Pesticide leaching depends on patterns of use, soil texture, total organic carbon


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content of the soil, pesticide persistence, and depth to the water table (Shirmohammadi andKnisel 1989). A pesticide leaches to groundwater when its residence time in the soil is lessthan the time required to remove it, or transform it to an innocuous form by chemical orbiological processes. The residence time is controlled by two factors: water applied andchemical adsorption to stationary solid surfaces. Volatilization losses of soil-appliedpesticides can be a significant removal mechanism for compounds having large Henry’sconstants (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 bysome U.S. regulatory agencies. The slower moving pesticides were recommended providedthey were used in accordance with the approved manufacture’s label instructions. Theseincluded the fungicides Iprodione and Triadimefon, the insecticides Isofenphos andChlorpyrifos and the herbicide Glyphosate. Others were recommended against, even whenused in accordance with the label’s instructions. These included the fungicides Anilazine,Benomyl, Chlorothalonil and Maneb and the herbicides Dicamba and Dacthal. Noinsecticides were on the “banned list” (Horsley et al, 1990).

In the case of potentially toxic elements, problems may appear when infiltratingstormwater using a rapid infiltration system (Crites 1985), such as a dry well. Most metalshave very low solubilities at the pHs found in most natural waters and they are readilyremoved by either sedimentation or sorption removal processes (Hampson 1986). Many arealso filtered, or otherwise sorbed, in the surface layers of soils in infiltrating devices whenusing surface infiltration. Table 3 discusses the pollutants found in stormwater that maycause groundwater contamination problems when allowed to infiltrate through infiltrationdevices.


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Table b.3. Groundwater Contamination Potential for Stormwater Pollutants (after Pitt et al., 1994)

Compounds Mobility(worst case:sandyl-1oworganic soils)

Abundancein storm-water

Fractionfilterable

Contaminationpotential forsurface infilt. andno pre-treatment

Contaminationpotential forsurface infilt. withsedimentation

Contaminationpotential for sub-surfaceinjectionwith minimal pre-treatment

Nutrients nitrates mobile low/moderate high low/moderate low/moderate low/moderate

2,4-D mobile low likely low low low lowγ-BHC (lindane) intermediate moderate likely low moderate low moderatemalathion mobile low likely low low low lowatrazine mobile low likely low low low lowchlordane intermediate moderate very low moderate low moderate

Pesticides

diazinon mobile low likely low low low lowVOCs mobile low very high low low low1,3-dichlorobenzene low high high low low highanthracene intermediate low moderate low low lowbenzo(a) anthracene intermediate moderate very low moderate low moderatebis (2-ethylhexyl)phthalate

intermediate moderate likely low moderate low? moderate

butyl benzyl phthalate low low/moderate moderate low low low/moderatefluoranthene intermediate high high moderate moderate highfluorene intermediate low likely low low low lownaphthalene low/inter. low moderate low low lowpenta- chlorophenol intermediate moderate likely low moderate low? moderatephenanthrene intermediate moderate very low moderate low moderate

Otherorganics

pyrene intermediate high high moderate moderate highnickel low high low low low highcadmium low low moderate low low lowchromium inter./very low moderate very low low/moderate low moderatelead very low moderate very low low low moderate

Potentiallytoxicelements

zinc low/very low high high low low highSalts chloride mobile Seasonally

highhigh 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 frominfiltration devices because of their potentially high concentrations of soluble potentially toxicelements, pesticides, and pathogens (Pitt et al., 1999) Similarly, all runoff frommanufacturing industrial areas should also be diverted from infiltration devices because oftheir 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 beforeinfiltration, to both minimize groundwater contamination and to prolong the life of theinfiltration device (if needed). This pre-treatment can take the form of grass filters, sedimentsumps, wet detention ponds, etc., depending on the runoff volume to be treated and othersite-specific factors. Pollution prevention can also play an important role in minimizinggroundwater contamination problems, including reducing the use of galvanized metals,pesticides, and fertilizers in critical areas. The use of specialized treatment devices can alsoplay an important role in treating runoff from critical source areas before these morecontaminated 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 underlyingsoils above the groundwater is preferable to using dry wells, trenches and especiallyinjection wells, unless the runoff water is known to be relatively free of pollutants. Surfacedevices are able to take greater advantage of natural soil pollutant removal processes.However, unless all percolation devices are carefully designed and maintained, they may notfunction properly and may lead to premature hydraulic failure or contamination of thegroundwater (Pitt et al., 1999).

It should be clear that although SUDS have great potential in both quantity and qualitycontrol in urban runoff, each case should be assessed individually, and an incrementalapproach containing both high tech and low-tech solutions is the most likely developmentscenario (Butler and Parkinson, 1997). Direct application of such methods across differentregions and countries is not always appropriate and must also include consideration of thelocal socio-economic and administrative circumstances associated with the operationaldesign, 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 andactivities. For example, in the region of Veneto (North-Eastern Italy), artisanal activitiesaccount for 20% of the regions exports. The area of Vicenza, a provincial town in the Venetoregion, represents one of the largest agglomerations of artisanal activities in Italy. With apopulation of 109,000 inhabitants, the municipality of Vicenza has a total of about 1600small to medium-scale enterprises (SMEs).

The environmental impact of artisanal activities is less clearly understood than the impactfrom industrial activities. While there are a large number of point sources, each sourcecontributes a very low wastewater flow rate, closer to the wastewater discharge fromresidential units than from industrial sites. Nevertheless wastewater from artisanal activitiesmay be dramatically different to that from residential units, both in terms of pollutantconcentration and the presence of specific pollutants.

Due to the presence of specific pollutants, wastewater discharge is regulated in the sameway as industrial wastewater (i.e. in terms of pollutant concentration limits), even thoughartisanal 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 theinterests of both artisanal workers and enterprises) sponsored a study on the origin andcontribution of pollutants to urban wastewater. The aim of the study was to assess the loadof pollutants from different artisanal activities, in comparison to the total load originating fromthe urban wastewater system. In addition, the pollutant load from artisanal activities wassubdivided into load from discharged wastewater and load from concentrated liquid wastes(which are separated and collected by external firms), to assess if wastewater segregationcould significantly affect pollutant load from artisanal activities.

Description of the studyThe 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 andtype 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 specificwastewater. Therefore, this growing number of “service” companies does not affect theconclusions 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 totalcapacity is about 137,900 p.e. (population equivalent) with a total flow rate of about 23.2million m3 year-1.


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Table c.1 capacity of the four municipal wastewater treatment plant of Vicenzamunicipality p.e.=population equivalent

WWTS Capacity (p.e.) Treated flow rate m3 year-1

CASALE 71,900 10,202,000LAGHETTO 3,500 394,000LONGARA 3,500 826,000

S. AGOSTINO 59,000 11,800,000Total 137,900 23,222,000

In this area there are about 1,579 artisanal enterprises discharging their wastewater into theUWW collecting system. Table c.2 shows the most common artisanal activities in the area ofVicenza and the number of enterprises involved in each activity. For each activity arepresentative number of enterprises was selected for further investigation. Typically onewastewater sample was drawn from each enterprise. In some cases an additional sample ofconcentrated, segregated wastewater was also drawn.

Table c.2 Main artisanal activities in the area of VicenzaType of activity No. of samples No. of enterprises in the

municipality of VicenzaFood workshops 10 53

Car-repairers 20 (14+6*) 175Ceramic and photoceramic 7(6+1*) 23

Artisanal galvanic shops 8(5+3*) 18Printing shops 21(14+7*) 140

Wood manufacturing 18(3+15*) 92Marble manufacturing 5 140

Metallurgists and mechanics 15(8+7*) 155Dental practices 21 88

Gold manufacturing shops 34 258Hairdressers 19 310

Laundrettes and dry-cleaners 25 88Textile shops 2 16

Artisanal glass manufacturing 3 23TOTAL 208 1579

*concentrated wastewater, segregated and committed to external firms

Total loadAlong with artisanal wastewater samples, influent and effluent samples from the fourmunicipal 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 largenumber of influent and effluent samples were taken and analysed. Using the specificwastewater flow rate and the influent concentration, the pollutant load was calculated foreach 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 averageconcentration 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 basedon 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

TOTAL LOAD(g/day)

POLLUTANT

AVERAGE MAX(95%) MAX

REMOVALEFFICIENCY(%)

Cd 64 75 960 40Cr(III) 636 759 13244 75Cr(VI) <LOD. <LOD <LOD

Mn 4921 4930 15623 37Ni 7092 7940 44421 30Pb 636 737 12527 60Cu 3546 3976 16061 80Zn 11221 13657 60174 70

<LOD = below limit of detection

iv) Pollutant load from specific artisanal activities

A typical example of the analytical work performed for each category of artisanal activity isreported in Tables c.4 and c.5 for car repair shops. A rough statistical analysis of the resultshas 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-repairshops

Pollutant Average Max 95%CI

Max

COD 329 547 1800Cd 0.01 0.03 0.14Cr(III) 0.1 0.2 0.8Cr(VI) 0 0 0Mn 0 0 0Ni 0 0 0Pb 0 0.1 0.4Cu 0.1 0.1 0.5Zn 4.5 11.1 55

Table c.5 Pollutant concentrations (mg l-1) in the segregated wastewater of 6 car-repair shops

Pollutant Average Max95%CI

Max

COD 10293 15346 19120Cd 0.3 0.7 1.3Cr(III) 0.2 0.3 0.5Cr(VI) 0 0 0Ni 2.5 6.1 12Pb 22.1 52 102Cu 33 74.6 144Zn 31.9 73.8 145

Pollutant load was calculated for the selected enterprises and extrapolated to the totalnumber of enterprises for each category. Even though each enterprise was equipped with itsown treatment plant, pollutant loads were calculated on the basis of concentrations in theuntreated wastewater, hypothesizing a scenario where no pre-treatment is performed and


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the wastewater is discharged directly into the UWW collecting system. Similar calculationswere performed using concentrations and volumes of segregated wastewater (spent baths).So, for the enterprises that separate wastewater for treatment by external firms (for recoveryand detoxification), two potential pollutant loads were given with reference to the twohypothesized scenarios:

• all the wastes (concentrated and diluted wastewater) are discharged into the urbanwastewater system without pre-treatment (Table c.6 D and S);

• the spent baths are treated externally, whereas wastewater is directly discharged intothe urban wastewater system (Table c.6 D only).

Car-repairers (175 shops)

The most significant pollutants are: suspended material, COD, oils, surfactants, organicsolvents, 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 concentratedwastewater (D only), the pollutant load is significantly reduced. The careful segregation ofspent baths induces a decrease of about one order of magnitude in the percentage of leadoriginating 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 theceramic and photoceramic shops is minimal due to the low wastewater flow rate. Only leadseems to represent a significant load. In the case of ceramic shops, segregation ofconcentrated 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 concentratedwastes does not appear to significantly reduce the lead load because this load originatesmainly 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 theoverall UWW flow rate (0.15%), only Cd and Cu loads from the printing activities aresignificant with respect to the total load. Segregation of concentrated wastewater reducesthe 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 bysimply segregating concentrated wastes the pollutant load to the WWTS from woodprocessing and furniture manufacturing shops is dramatically reduced. Other specific woodprocessing pollutants such as arsenic were not analysed.


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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 lowerthan in segregated wastewaters (see Table c.7) which have high average concentrations ofpollutants such as; nickel, lead, copper and zinc. Separation of concentrated wastesreduces the load to the WWTS significantly.

Goldsmiths (258 shops)

The area of Vicenza represents one of the most important districts for gold manufacturing inItaly, with up to 258 artisanal goldsmith shops. The main pollutants originating from goldmanufacture are: COD, boron, cadmium, copper, zinc, and surfactants. However goldsmithsshops are characterized by very low wastewater flow rates, with an average of about 1 cubicmeter per day per unit.In terms of contribution to the pollutant load, the 258 goldsmiths shops represent a highcontribution of Cd, Cu, and Zn. This is mainly due to the fact that goldsmith shops used toadd spent concentrated baths to the diluted wastewater in order to recover precious metalsduring 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, andsurfactants. 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, andsurfactants. The contribution of the 88 shops to the total metals pollutant load is generallylow, due to the low contribution in terms of flow rate, that is an average of about 0.07% ofthe total UWW flow rate. Mercury is an important pollutant in wastewater linked to dentalpractices, which is not considered in this particular study but is referred to in section 2.1.2 ofthis report.

Hairdressers (310 shops)

This is the category with the largest number of shops in the Vicenza’s municipality. The mainpollutants 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 Cdalthough overall the levels for this pollutant were very low .


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Table c.5 Pollutant load from specific artisanal activities. D=discharged wastewater,S=segregated wastewater, (- = not reported).

Flowrate m3

per day

Average Pollutant Load (g/day)

Cd Cr III Mn Ni Pb Cu ZnCar repair shops(D & S)

153 4 6.3 <LOD 21.9 218.1 299.7 716.1

Car repair shops(D only)

153 1 5 <LOD <LOD 25 11.3 436.8

Ceramics andphotoceramics (D & S)

33 0.1 7.2 <LOD <LOD 19.7 0.1 1.6

Ceramics andphotoceramics (D only)

33 0.1 4.9 <LOD <LOD 12.8 0.1 1.2

Galvanic shops(D & S)

72 0.3 1 - 47.8 9.8 29.8 7.9

Galvanic shops (D only) 72 0.2 <LOD - 30.1 9.4 11.4 7.9Printing Shops(D & S)

97 1.8 3.8 <LOD - 3.8 69.5 27.5

Printing Shops(D only)

97 <LOD <LOD <LOD - <LOD <LOD 26

Wood Processing(D & S)

35 <LOD 4.8 <LOD - 34.7 16.2 32.9

Wood Processing(D only)

35 <LOD <LOD <LOD - <LOD 1.5 24.5

Metallurgists andMechanics (D & S)

70 1.4 30 13 1107 163.5 187.5 978.2

Metallurgists andMechanics (D only)

70 0.2 <LOD <LOD <LOD <LOD 13.2 23.8

Goldsmiths(D only)

266 20.5 0.9 <LOD - <LOD 267.7 237.8

Food Workshops(D only)

120 <LOD <LOD <LOD <LOD <LOD 3.2 31.1

Dental Technicians(D only)

43 0.2 0.3 0.6 - <LOD 0.7 17.6

Hairdressers(D only)

289 <LOD <LOD <LOD <LOD <LOD 24.5 111.6

Laundrettes and DryCleaners (D only)

210 1.5 - <LOD 2.8 2.9 3.3 54.5

The highest values for each pollutants are highlighted in bold.NB all samples were also tested for Cr VI but they were all below the limit of detection with theexception of Galvanic shops considering both discharged and segregated wastewater (average loadCr VI14.4g/day)

Impact of all artisanal activities on urban wastewater treatment plants

Pollutant loads from all artisanal activities of Vicenza municipality were calculated, summingthe loads of each specific activity as reported in Tables c.7 and c.8.


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Table c.7 Total pollutant loads from artisanal activities in the Vicenza municipality,considering both discharged and segregated wastewater

POLLUTANT TOTAL POLLUTANT LOAD FROMARTISANAL ACTIVITIES (g/day)

PERCENTAGEOF TOTAL LOAD

TO WWTSaverage max(95%) max

B 2909 4774 38832Cd 30 66 529 46.59

Cr(III) 54 106 243 8.55Cr(VI) 14 49 43

Mn 14 19 38 0.28Ni 1180 3153 8114 16.63Pb 452 1064 2676 71.12Cu 903 1872 6387 25.48Zn 2231 3998 10589 19.88

MBAS 53404 93158 353795 25.45

flow rate m3/day 1633 2526 5559 2.57p.e. 7697 13295 46221 5.58

<LOD = below limit of detection

Table c.8 Total pollutant loads from artisanal activities in the Vicenza municipality,considering discharged wastewater only

POLLUTANT TOTAL POLLUTANT LOAD FROMARTISANAL ACTIVITIES (g/day)

PERCENTAGEOF TOTAL LOADTO WWTS

average max(95%) maxB 2909 4774 38832Cd 24 51 498 36.9Cr(III) 11 27 100 1.8Cr(VI) <LOD. <LOD. <LOD.Mn 0.6 1.7 13 0.01Ni 33 92 206 0.5Pb 50 127 491 7.9Cu 338 621 3873 9.5Zn 987 1827 6963 8.8chlorinated solvents 6400 17739 140800MBAS 50911 87609 342922 24.3flow rate m3/day 1633 2526 5559 2.6

p.e. 5867 10508 42102 4.3 <LOD = below limit of detection

In terms of flow rate the contribution of all artisanal activities is typically low (2.6%). Itremains relatively low (4%) in the worst scenario case, where the artisanal activitiesdischarge at the upper limit of the confidence interval of their cumulative wastewater flowrate, while the total UWW flow rate remains at the average value. In contrast, the percentagecontribution 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 ahypothetical centralized WWTS, for potentially toxic elements and surfactants, respectively.Figure c.1 clearly shows that only car-repairers, goldsmiths shops, metallurgists andmechanics contribute significant amounts of potentially toxic elements to UWW, with respect


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to the total metal load of the all activities. The main activities responsible for the surfactantload in UWW are; hairdressers, goldsmiths, and food workshops (Figure c.2).

Figure c.1 Heavy metal load from artisanal activities

0

1000

2000

3000

4000

5000

6000

7000

8000

9000

AL AU CE GA GR LG MA ME OD OR PR PU TS VE tot

aver

age

load

(g

/day

)

Figure c.2 Surfactant load from artisanal activities

0

10000

20000

30000

40000

50000

60000

AL AU CE GA GR LG MA ME OD OR PR PU TS VE tot

aver

age

load

(g

/day

)

AL: Food workshopsAU: Car repairersCE: CeramicsGA: galvanicGR: Printing shopsLG: Wood manufacturingMA: Marble manufacturingME: Metallurg.and mechanicsOD: Dental practicesOR: GoldsmithsPR: Hairdressers


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With the exception of surfactants and cadmium, the high pollutant loads from artisanalactivities are notably reduced when the concentrated wastes are not considered (Tablec.23). The poor effect of waste segregation on surfactant load is explained by the fact thatthe major surfactant contribution derive from activities that do not practice wastesegregation: 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 WWTSis used in agriculture, the maximum admissible pollutant concentration in sludge must beconsidered. From this value a maximum admissible pollutant concentration in the UWW maybe calculated according to the following equation:

C

C i Q f 100 %H 2O

R% Qwhere C is the maximum influent concentration permitted for sludge disposalCi is the maximum pollutant concentration allowed in sludge to land

R% is the removal efficiency (%) in the treatment plantQf is sludge flow rate in m3 day-1 calculated on the basis of the production of 1.87 l/inhabitantwith 95.5% of humidityQ 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 sludgedisposal in agriculture, considering both discharged and segregated wastewater

Pollutant Regulatory limits foragricultural use

(mg/kg dry sludge)

Admissible(*)concentration

Admissible(*)load

percentage of totaladmissible load

(%)D.Lgs 99/92 Veneto (mg/l) (g/day) average

Cd 20 10 0.01 580 5.11Cr(III) 500 0.15 9670 0.56

Ni 300 200 0.17 10549 11.18Pb 750 500 0.21 13599 3.33Cu 1000 600 0.28 17582 5.14Zn 2500 2500 0.66 42045 5.34

(*)on the basis of equation (1)


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Table c.10 Pollutant contribution of artisanal activities to the admissible load forsludge disposal in agriculture considering only discharged wastewater

Pollutant Regulatory limits foragricultural use

(mg/kg dry sludge)

Admissible(*)conc

admissible(*)load

percentage of totaladmissible load

(%)

D.Lgs 99/92 Veneto (mg.l-1) (g/day) averageCd 20 10 0.01 580 4.04

Cr(III) 500 0.15 9670 0.12Ni 300 200 0.17 10549 0.31Pb 750 500 0.21 13599 0.37Cu 1000 600 0.28 17582 1.92Zn 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 agricultureare reported. From these limits the concentration limit in the influent and the consequentadmissible load were calculated. In the last column the impact of artisanal activities on thesludge characteristics are reported in terms of percentage load from artisanal activities withrespect 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 wastesegregation does not take place. Even in this pessimistic hypothesis, the impact of artisanalactivities is typically low. The highest metal contribution is for Ni, representing 11% of theadmissible load. In the worst case scenario, considering all the maximum loads, artisanalactivities 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’ shopsin a different area of Veneto region

To validate the results obtained in the study described above, EBAV undertook an additionalstudy specifically addressed to activities of hairdressers and beauticians. This study,performed during 1999-2000, considered a group of shops representing all the hairdressersand beauticians located in the district of Valdagno (Vicenza), which discharge theirwastewater into the UWW collecting system of the municipality of Trissino (Vicenza). In thisvalidation 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 wastewatersample was drawn from each shop. A rough statistical analysis of the results has beenperformed to obtain the average concentration; then the average values have beenincreased by 10% to take into account the effects of highly polluted wastewater (due totypical products such as shampoos or dyes).

In Table c.11 the average pollutant concentrations are reported, as well as the averagewastewater flow rate per unit. In this case wastewater production was even lower than in thecase of Vicenza (600 l.day-1 vs 900 l.day-1). The contribution of hairdressers’ shops to thetotal wastewater flow rate entering the Trissino WWTS was 0.33%, whereas in Vicenza areathis was about 0.45%. From the average concentrations (with a safety increase of 10%) andfrom wastewater flow rates the pollutant load from the total 135 shops to the Trissino WWTSwas calculated. Where data was available (chromium and surfactants), this load wascompared with the total pollutant load to the WWTS. Data from Valdagno area confirm thenegligible contribution of chromium load, and other potentially toxic elements as well, to the


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WWTS. Analysis of surfactants (anionic and non-anionic) in wastewater samples showedthat the contribution of beauty shops in Valdagno district to the total surfactants load stillremains lower than the estimated contribution in Vicenza municipality, where only anionicsurfactants were considered.

The data presented above, while confirming the negligible contribution of this artisanalactivity to the total load of metallic pollutants, suggest that the extrapolation of the resultsfrom Vicenza Case Study may result in an overestimation of the contribution of artisanalactivity to the pollutant loads in the WWTS systems.

Table c.27 Pollutant concentrations and pollutant loads from hairdressers andbeauticians in the district of Valdagno (Vicenza)

POLLUTANT AVERAGECONCENTRATION

(mg/l)

POLLUTANT LOAD*(g/day)

PERCENTAGE OF LOADTO WWTS

(%)Cr(III) <0.1 <5 <0.1

Anionic surfactants 44 3915 8.2

Non ionic surfactants 51 4590 6.2

Total surfactants 99 8775 7.2Flow 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 arecommon in Italy and in other EU regions. As shown in the validation case of Valdagno, inother districts the contribution of artisanal activities to the pollutant load of the UWW systemmay be lower.

The principal pollutants originating from these artisanal shops are potentially toxic elementssuch as Cd, Ni, Pb, Cu, and Zn, and surfactants.

The main conclusion of this study was that, by segregating concentrated liquid wastes, thecontribution of artisanal activities to the pollutant load was dramatically reduced, at least forpotentially toxic elements. However, only some of the artisanal activities in this Case Studypractised wastewater segregation. One issue raised by artisanal representatives was theeconomic cost of segregation of wastewater. It is felt that the stringent environmentalrequirements concerning wastewater from Italian artisanal shops, considered industrialwastewater, do not compensate efforts for waste segregation.

Artisanal activities that did not practice segregation of concentrated liquid wastes includegoldsmiths, hairdressers and food manufacturing shops, which are also the main enterprisesresponsible for the surfactant load in wastewater. As a consequence, neglecting thecontribution of segregated liquid wastes did not significantly affect the total load ofsurfactants from artisanal shops. This load typically represents one fourth of the totalsurfactant load to the WWTS. It may be anticipated that, if careful segregation of theconcentrated liquid wastes were extended to all the artisanal activities, a dramatic decreaseof potentially toxic element and surfactant load from the artisanal activities would beobtained. Even though the wastewater flow rate from artisanal shops would not decreasesignificantly, the pollutant load could be reduced to negligible values with respect to the totalpollutant load.


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GOLD JEWELLERY PRODUCTION IN ITALY- BEST ENVIRONMENTAL PRACTICE

Introduction

In Italy there are about 6,000 small to medium gold and jewellery manufacturing shops, mostof which are concentrated in three main production districts: Arezzo (Tuscany), Vicenza(Veneto) and Valenza Po (Piemonte). In the period 1995-1999 the Italian Government andthe Association of Artisanal Activities sponsored a large research programme, carried out bythe National Research Council, in support of Craft Goldsmiths Production and Trade. Theprogramme 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 practicesfor 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 ofArezzo district may confidently be extended to draw conclusions about national practices ingoldsmiths’ shops. There is no information on the losses of Au, Ag, PGMs from the gold andjewellery shops to the wastewaters in this research programme. More data on PGM ispresented 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, chainsand medals. Hollow bars may be used to produce lighter objects. Gold or silver goods mayalso be produced using micro-casting or electrolytic processes (electro-forming). Allproduction cycles have common final steps: object assembly, polishing, finishing andcleaning.

Wastewater origin and characteristics

Different production cycles and processes generate wastewater in the gold manufacturingshops. Casting operations to prepare wire or plate do not typically require aqueoussolutions, with exception of small volumes for washing crucibles. In contrast, the preparationof 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 gypsummould and rinse waters. The “gypsum” waters are segregated from the other wastewaterproduced in the workshops and recycled after a settling step.

Wastewater from electro-forming is derived from specific activities, as well as fromoperations common to other processes. The former wastes may be highly turbid wastewater,exhaust baths and rinsing waters, mainly derived from the galvanic cycle. Wastewaterscharacterised by high turbidity are filtered to eliminate the suspended material and then re-circulated several times before collecting them with the other wastewater. Regardinggalvanic waters, acid wastes are separated from the wastes containing cyanide. Theconcentrated cyanide baths are collected by external firms, whereas the diluted rinse waterscan be either pre-treated in a separate circuit, then added to the main wastewater stream or


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re-circulated after elution through an ion exchange column (anionite). The concentratedcyanide 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 surfaceshining generally contain metal powder. In addition, the final step generates large amountsof wastewater as well as surfactants present in the spent baths.

Finally hand and floor washing waters, do not typically require pre-treatment and could besent to the sewer. However as they may contain gold and silver powder they are not directlydischarged 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 tospecialised firms for recovery of precious metals.

Wastewater may be divided into four classes:

Soapy waters contain high concentrations of detergents, along with fatty substances andmetal powder. Other pollutants such as phosphates and ammonia typically originate fromthe 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 ofpotentially toxic elements such as copper, zinc, iron and nickel.

“Gypsum” waters containing suspended gypsum particles, are recycled several times aftersedimentation of the suspended material. Then they are committed, together with thesedimented 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 beforesending them to the wastewater treatment plant.


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Wastewater management and fateAbout 40% of the goldsmiths’ workshops of Arezzo province do not declare any wastewaterproduction. 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 theirwastewater, Table c.28 reports the number (and relative percentage) of workshops thatdischarge their wastewater into the UWW collecting system or into surface waters. In termsof flow rate, the 694 workshops discharge about 121600 m3.year-1 of wastewater. Assumingan average number of 6.3 workers per unit, the average specific wastewater flow rate isabout 0.5 m3.week-1 per worker.

Table c.28 Destination of wastewater of workshops having discharge authorisation

Destination No. of workshops % of workshopsUWW collecting

system567 81.7

Surface waters 29 4.2Unknown 98 14.1

TOTAL 694 100

Survey on a representative group of goldsmiths’ shopsThe workshops were selected with the aim of choosing representative establishments interms of number of workers and in terms of type of manufacturing processes. Twelve smallto medium workshops were selected for the study. The survey included both interviews onproduction processes, wastewater flow rate, wastewater treatment, direct sampling andanalysis of treated and untreated wastewater. Typically, several wastewater samples werecollected from each workshop, in March, May and October of 1996. The chemicalcharacterisation 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.

ConclusionsIn examining the pollutant concentrations in these wastewaters (considering the resultsobtained for the 12 shops sampled), the pollutants most often detected with concentrationshigher than admissible limits for discharge are: surfactants, copper, zinc, cadmium andboron. Surfactants, copper and zinc are detected in all samples whereas boron andcadmium are present in 75 % and 60% of samples respectively. Surfactants, derivedgenerally from washing processes, are present in wastewater with an average concentrationof 34 mg.l-1 MBAS and a maximum value of 118.5 mg.l-1 MBAS. The average boronconcentration found in the wastewater of small jewellery shops examined is 13.5 mg.l-1, witha maximum value of 100 mg.l-1. The presence of boron in these wastewaters may be due toprocesses such as soldering (where boron is used as borace), voiding and washing, orthrough the use of other materials e.g. hydrofluoboric acid, detergents.

For potentially toxic elements, the average concentrations of copper and cadmium detectedin wastewater are 14.2 mg.l-1 Cu and 0.4 mg.l-1 Cd, with maximum values of 61 mg.l-1 Cuand 1 mg.l-1 Cd. Zinc concentration is highly variable, which is probably due to the differentmanufacturing steps of the shops, with an average value of 22 mg.l-1 but a maximum valueof 270 mg.l-1.


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(D) Pharmaceuticals in the Urban Environment

Introduction

Pharmaceutical substances are a group of compounds, which until recently, have not beenof major concern with regards to their environmental effects. These compounds aredeveloped for their biological effect (primarily in humans), to cure disease, fight infection orreduce symptoms. If these substances enter the environment they may have an effect onaquatic and terrestrial animals, due to these biological properties and the fact that some ofthem may bioaccumulate. Unlike other organic compounds, such as PCBs whose use hasbeen discontinued over the last 20 years, pharmaceuticals are used widely and they andtheir metabolites may easily enter the UWW system. Pharmaceuticals use is also expectedto increase in Europe with the increasing avearge age of the population.

Current research demonstrates that drugs and their metabolites entering water supplies andthe food chain may pose a real threat, both to the ecosystem and to human health, and riskassessments 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 inmixtures and at low concentrations. Furthermore, pharmaceuticals have a wide variety ofstructures and activities and that they may act synergistically (Alcock, 1999). There aremany different pharmaceuticals substances and approximately 3000 pharmaceuticalcompounds are discharged into UWW collecting systems (ENDS, 2000). Sewage sludge ispredominantly disposed of on agricultural land, as is manure from farms and both of theseproducts will contain large amounts of pharmaceutical substances. Unfortunately, very littleis known about the fate of these compounds in the environment and the potential long-termimpacts. Many pharmaceutical substances though, have the same characteristics as organiccompounds; i.e. they are lipophilic, which tends to be a requirement to be able to passmembranes, and some are designed to be persistent so that they are not inactivated beforeachieving their healing effect (Halling-Sørensen, 1998).

Sources and fate in the environmentThere are two major groups of pharmaceuticals; human and veterinary drugs, and they willenter 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 thisfigure. 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 intothe environment. However, this will depend on the chemico-physical behaviour of thepharmaceuticals in question.


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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 manurespreading 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 eatenand is deposited straight onto the sea-bed (Halling-Sørensen, 1998).


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Figure d.2: Anticipated exposure routes of veterinary and human medicinalsubstances in the environment [after Halling-Sorensen, 1998].

The compoundsFor the purpose of risk assessments, pharmaceutical compounds have been divided intofour activity groups: antibiotics, antineoplastic drugs, antiparacetic drugs, and hormonedisrupters. However, there are many other groups of pharmaceutical compounds, such aslipid regulators, which have been found ubiquitously in the environment and are highlypersistent compounds (Daughton, 1999).


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Antibiotics:Antibiotics are widely used as medicines for human and animals treatment, also used widelyas growth promoters in veterinary use. According to the Swiss environmental researchinstitute, 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 verypersistent in the environment, particularly in soils, and the most widely used growthpromoters have been shown to have no effect on invertebrates, even at relatively highconcentrations. However soil bacteria may be more sensitive (ENDS, 2000).

Veterinary drugs tend to end up in manure and so have the potential to contaminate soilswhere manure or slurry is spread. Levels of antibiotics in soil around a pig farm studiedreached up to 1400µg.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. Thisresistance 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 drugsThese anti-cancer drugs are mainly used in hospitals rather than in households. They areprimarily used for chemotherapy and are found sporadically, in a range of concentrations inthe 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 asmutagens, carcinogens, teratogens, and embryotoxins (Daughton, 1999). The most widelyused substance is cyclophocamide (CP). In Denmark, approximately 13 to 14 kg of CP isused 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 PlatinumGroup Metals.

Analgesic drugsAnalgesic drugs are used for pain relief and are probably the most commonly usedmedicines.

Endocrine-disrupting substancesThere 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 whichthey are not intended (Christensen, 1998). An endocrine-disruptor may have an ‘agonisticeffect’ 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 thenatural one from binding and producing the required response (Environment Agency ofEngland and Wales, 1998). Other effects may also occur, showing that the process is verycomplex and affects many systems in the body.

Endocrine-disrupting non pharmaceutical substances include phthalates, some PCBs, andsome pharmaceutical compounds, such as oestrogens. Many of these may be persistant insewage sludge and could enter the food chain as they are potentially taken up by plants andanimals. The effects of endocrine-disruptors were discovered about 10 years ago and mayoccur in concentration ranges of a few nanograms per litre (Jørgensen, 2000). Humans usehormones to cure diseases, as well as in contraception and hormonal replacement therapy.Table d.1 shows some categories of substances with endocrine-disrupting properties:


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Table d.1: Categories of substances with endocrine-disrupting activities [EnvironmentAgency, 1998].

Category Examples Uses Modes of actionNatural

phytoestrogens Isoflavones, lignans Present in plants Oestrogenic and anti-oestrogenic

Female sexhormones

17β-oestradiol,oestrone

Produced in animals Oestrogenic

Man-madePolychlorinated

organiccompounds

PCBs, dioxins By-products fromincineration and chemical

processes

Anti-oestrogenic

Organochlorinepesticides

DDT, dieldrin, lindane Insecticides Oestrogenic and anti-oestrogenic

Alkylphenols Nonylphenol Production of NPE andpolymers

Oestrogenic

Alklphenolethoxylates

NonylPhenolEthoxylate (NPE)

Surfactant Oestrogenic

Phthalates Dibutyl phthalate Plasticiser OestrogenicBi-phenoliccompounds

Bisphenol A In polycarbonate plasticsand epoxy resins

Oestrogenic

Synthetic steroids Ethinyl oestradiol Contraceptives Oestogenic

Of these, oestrogens are of a major concern, as they are excreted in an inactive form butare found to be reactivated in sewage effluent. Oestrogens are organic molecules derivedfrom cholesterol, which can bind to receptors and cause a physiological response(Montagnani et al., 1996). The purpose of the endocrine system is to regulate metabolicactivity, which requires a degree of interaction with the nervous system (Montagnani et al .,1996). Because oestrogen receptors are located in the cell nucleus, oestrogen-likemolecules can thus enter the cell and could potentially interact with DNA, causing damagewhich may lead to tumour formation (Montagnani et al. , 1996). Prolonged exposure to thesecompounds may induce female characteristics in males. There is increasing speculation thatthese 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: 17β-oestradiol, oestrone, and the synthetic hormone: ethinyloestradiol, from domestic sewageeffluent, developed hermaphrodite characteristics (Alcock et al., 1999). It was also found thatthese hormones were present in the biologically active form, having been transformed andreactivated 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 tobe present in wastewater effluents, the development of female characteristics in male fishpresent in WWTS lagoons is unlikely to be due to these substances, as the transformation isonly possible at a very early stage in their development (Montagnani et al., 1996). Due to thewidespread use of estrogenic substances and their entry into the environment via sludgeand effluent from WWTS, aquatic environments may be acting as a sink for thesesubstances. Natural and synthetic estrogens are extremely widely present/used. Althoughthe contribution these compounds make to oestrogenic effects is thought to be small, theyare are of concern because of their highly persistent and potent nature (ENDS, 2000).


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Consumption of pharmaceutical compounds

As already stated, it is difficult to obtain information on the quantities of pharmaceuticalsused for most countries but in Denmark consumption of the most commonly used drugs isavailable (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 305250Paracetamol Analgesic 248250

Ibuprofen Anti-rheumatic 33792Pencicillin V Antibiotic 19000Furosemide Diuretic 3744

Terbutaline Anti-asthmatic 475

Enalapril Anti-hypertensive 416

Citalopram Anti-depressant 368Diazepam Anti-depressant 207Salbutanol Anti-asthmatic 170

Bendroflumethiazide Diuretic 167Zopiclone Anti-hypertensive 144

Amlodipidine Anti-hypertensive 132Oestradiol Hormone replacement 119Nitrazepam Anti-hypertensive 116

17β-estradiol Oral contraception 45Budesonide Anti-asthmatic 39Gestodene Birth control pill 37

Cyclophosphamide Anti-cancer 20Xylometazoline Nasal decongestant 13

Digoxin Heart drug 4Desogestrel Birth control pill 3

Medicine groupsAntibiotics 37700Analgesics 28300

Hypotensiva 410Diuretica 3800

Anti-asthmatics 1700Psychleptics 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). In1994, the overall production of antibiotics in Germany was 1831 tonnes, of which Penicillincontributed 624 tonnes (Hirsch, 1999). It should be noted that the amount of antibiotics usedfor human and veterinary purposes, 37.7 tonnes and 110 tonnes respectively, is in the samerange as the amounts of certain pesticides used (Hirsch, 1999). A paper published by Gollvan (1993) estimates that if the total amount of growth promoters used in the Netherlandswas spread on their 2 million hectares of agricultural land, this would give a yearly averageof 130mg of antibiotics and metabolites/m2 (Halling-Sørensen, 1998).The pharmaceutical substances predominantly used in hospitals must also take into accountcompounds such as X-ray contrast media. Iodinated X-ray contrast media are very stablebiochemically, so they tend to be excreted unmetabolised. In Germany, 500 tonnes per yearof X-ray contrast media are used and iopromide (CAS 73334-07-3) alone accounts for 130tonnes per year (Ternes, 2000). Other X-ray contrast substances used in the EU are:


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diatrizoate (CAS 131-49-7) an ionic X-ray diagnostic drug, iopamidol (CAS 60166-93-0) andiopromide (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 diagnosticsubstances.

Detection of pharmaceutical compounds

Other biologically active compounds are common in wastewater, and there is increasingevidence that such substances are widely present in the environment. Danish research hasfound that up to 68 different drug residues can be detected in the environment. Compoundssuch 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 metabolitesin wastewater and sewage sludge. For example, clofibric acid (2-(4-chlorophenoxy)-2-methylpropionic acid), which is a breakdown product of lipid-regulating drugs, is highlypersistent 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 ofclofibric 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 inmicrograms 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 µg.l-1 range in the aquaticenvironment, 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 µg.l-1 werefound in WWTS effluents, which is comparable to the data collected by Richardson andBowron in 1985 (Hirsch, 1999). Five of the 18 compounds investigated were frequentlydetected in German WWTS effluent and rivers: erythromycin, roxithromycin, clarithromycin,sulfamethoxazole, and trimethoprim. The highest concentration was detected forerythromycin degradation products, at a median value of 2.5 µg.l-1 in WWTS effluent and amaximum value of 6 µg.l-1 (Hirsch, 1999). The other four antibiotics were only detected atlevels below 1 µg.l-1 (Hirsch, 1999). Median values for the concentrations detected in surfacewaters are one order of magnitude lower than those detected in WWTS effluent (Hirsch,1999), (see Figure d.3).


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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 effluentsor surface waters (Hirsch, 1999). Tetracyclines tend to form stable complexes with calciumand 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 β-lactam ring (Hirsch, 1999). Ground water samples were also slightly contaminated bysulphamethoxazole and sulphamethazine (not used in human medicine), due to infiltrationfrom application of contaminated sewage sludge or manure to agricultural land. The samplescollected contained concentrations of sulphonamide residues up to 0.48 µg.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 exceed1µg.l-1 during the working week but decreased at weekends as X-rays did not tend to beperformed (Ternes, 2000). Maximum concentrations detected were greater than 3 µg.l-1 (upto 15 µg.l-1 for iopamidol). Median values were around 0.25-0.75 µg.l-1, which indicates theirubiquity in German wastewater treatment effluents (Ternes, 2000). The compounds detecteddepended on the region and on the practices of particular hospitals in that area.

Antiseptics are another major group of pharmaceutical compounds commonly used both inhouseholds and in medical practices. It has been found that major antiseptics, such aschlorophene and biphenol, are present at concentrations up to 0.05 µg.l-1 in wastewater(Ternes, 1998b). However, biphenol tends to be eliminated at rates of 98% during treatmentand chlorophene at 63% (Ternes, 1998b). These compounds, particularly clorophene, aredetected in rivers at similar concentrations. This is probably due to the fact that they are alsoused in many household detergents and disinfectants, as well as in veterinary medicine onfarms, so leading to widespread contamination of the aquatic environment.

Analgesics. Salicylic acid, a major metabolite of acetylsalicylic acid, is also detectable inGerman wastewater at high concentrations (54µg.l-1 over 6 days), but treatment degradesmost of it, as the compound is no longer detectable in WWTS effluents (Ternes, 1998b).


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Oestrogens. There is increasing concern about the widespread presence of oestrogenicsubstances in wastewater and other water bodies. The daily production rate of naturaloestrogens by humans is in the microgram range, up to 400 µg of 17β-estradiol for women(Ternes, 199b). The maximum daily excretion rate is 64 µg 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 juvenilefish has become a “biomarker” for the presence of estrogenic substances in the aquaticenvironment (Larsson, 1999). In the UK, caged fish downstream of an WWTS were found toproduce vitellogenin. Two possibilities were investigated: the presence of ethinylestradioland NPE. The effluent from a Swedish WWTS showed high levels of oestrogeniccompounds (Larsson, 1999). It was found that exposure to large amounts ofethinyloestradiol 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 ofethinyloestradiol is 3.5mg/day, which is close to the concentration found in the WWTSeffluent (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 (17β-oestradiol andoestrone) 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 comparativelylow levels: 0.2 to 7 ng.l-1 (Alcock et al ., 1999). The source of these is thought to be mainlyfrom human excretion products. It was also found that the hormones were present in thebiologically active form, suggesting that they had been transformed and reactivated afterexcretion (Alcock et al., 1999).

In Germany, raw sewage was found to contain 0.015 µg.l-1 of 17β-estradiol and 0.027 µg.l-1

of oestrone and it was also found that oestrone and 17α-ethinyloestradiol were not efficientlyremoved during wastewater treatment (Ternes, 1999a). In contrast, 17β-oestradiol and 16α-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 amunicipal 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 µgl-1


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and a median value of 0.009 µgl-1. The compounds 17α-ethinyloestradiol and 16α-hydroxy-estrone were found at the detection limit of 0.001 µg l-1 (Ternes, 1999a). Oestrone was theonly compound detected in 3 of the 15 rivers sampled at concentrations between 0.7 and 1.6ng l-1 (Ternes, 1999a). Therefore, it seems that these compounds, particularly naturaloestrogens, are not degraded in the treatment system and tend to accumulate in sludge andeffluent. However, the loads entering receiving waters are quite low. DEHP is also anendocrine-disrupting compound. In Sweden, it has been found in all sewage sludge samplesanalysed, 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 ofenvironmental contamination by pharmaceuticals (see Table d.4), (Zuccato, 2000).

Table d.4: Concentrations of medicinal drugs in drinking water, river water, andsediments [Zuccato, 2000].

DRUG DRINKING WATER (ngl-1) RIVER WATER (ngl-1) RIVER SEDIMENTS (ng kg-1)

Milan Lodi* VareseLambro(Milan)*

Po (Piacenzeand,Cremona)*

Adda(Sondrio)

Lambro(Milan)

Po(Piacenze,Cremona)*

Adda(Sondrio)

Atenolol <LOD <LOD <LOD 169·9-241·9 49·5-84·3 <LOD <LOD <LOD <LOD

Bezafibrate <LOD <LOD <LOD 134·3-202·7 15·1-22·4 1·6 130 <LOD <LOD

Ceftriaxone <LOD <LOD <LOD <LOD <LOD <LOD <LOD <LOD <LOD

Clofibric acid <LOD 3·2-5·3 <LOD <LOD <LOD <LOD <LOD <LOD <LOD

Cyclophosphamide

<LOD <LOD <LOD 2·2-10·1 <LOD <LOD <LOD <LOD <LOD

Diazepam <LOD19·6-23·5

0·2 0·7-1·2 0·5-0·7 <LOD <LOD <LOD <LOD

Erythromycin <LOD <LOD <LOD <LOD-17·4 0·7-0·9 <LOD 630 400-600 10

Furosemide <LOD <LOD <LOD 85·1-88 <LOD <LOD <LOD <LOD <LOD

Ibuprofen <LOD <LOD <LOD 90·6-92·4 <LOD-4·0 1·0 220 <LOD <LOD

Lincomycin <LOD <LOD <LOD 6·8-13·8 1·2-4·6 <LOD 130 <LOD <LOD

Oleandomycin <LOD <LOD <LOD <LOD-0·8 0·4-4·8 2·7 <LOD <LOD <LOD

Ranitidine <LOD <LOD <LOD <LOD-9·4 <LOD <LOD 150 <LOD-410 <LOD

Salbutamol <LOD <LOD <LOD <LOD-3·1 <LOD-4·6 <LOD <LOD <LOD <LOD

Spiramycin <LOD <LOD <LOD 8·4-68·3 <LOD <LOD 2900 <LOD-380 380

Tilmicosin <LOD <LOD <LOD <LOD <LOD <LOD <LOD <LOD <LOD

Tylosin <LOD 0·6-1·7 <LOD <LOD-2·2 <LOD <LOD 2640 <LOD-130 <LOD

(<LOD = below the limit of detection)

It can be seen that most drugs were measurable in these media, showing widespreadcontamination. The concentrations measured could potentially give rise to human exposurein the ng day-1 range, which is 3-4 orders of magnitude lower than the concentrationscapable of producing pharmacological effects (Zuccato, 2000). Hence, acute exposure isassumed 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 theinfluent and up to 114g.day-1 for carbamazepine in the effluent (Ternes, 1998c).


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Figure d.5: Elimination of different drugs during passage through a municipal sewagetreatment 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 wereusually 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 watersmay potentially be contaminated. Subsequently, a screening programme of 49 differentGerman WWTS effluents was carried out, in addition to river sampling. Lipid regulatingagents were found in the majority of the WWTS effluents, and in many river samples but in amuch lower concentration (Ternes, 1998c). Polar metabolites of the compounds wereusually detected. For example, clofibric acid was detected at levels up to 1.6 µg.l-1 in WWTSeffluent 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 µg.l-1 in treatedeffluent effluent and 0.15 µg.l-1 in rivers (Ternes, 1998c). In the case of betablockers, thehighest median concentration was found for metoprolol at 0.73 µg.l-1 in WWTS effluent and0.45 µg.l-1 in rivers (Ternes, 1998c). β2-sympathomimetics were also present but in very lowconcentrations. Anti-cancer agents such as cyclophosphamide and ifosamide were detectedat levels of 0.02 µg.l-1 and 0.08 µg.l-1 respectively in WWTS effluent; however, they areassociated with presence of hospital effluents, and are not widespread (Ternes, 1998c).Carbamazepine, an anti-epileptic drug was widespread in the aquatic environment, with ahigh median value of 2.1 µg.l-1 in effluent and 0.25 µg.l-1 in rivers (Ternes, 1998c). Annualprescriptions of carbamazepine amount to approximately 80 tonnes per year in Germany, itthen becomes metabolised and glucuronides are excreted. However treatment ofwastewater cleaves these metabolites back to the parent compound, increasing theenvironmental concentrations (Ternes, 1998c).

Table d.6 shows specific studies on pharmaceuticals in the environment and theconcentrations found for the different substances (Halling-Sørensen, 1998).


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Ground water pollution has been detected in some instances, mainly due to leaching fromlandfill sites containing pharmaceutical wastes (Halling-Sørensen, 1998). In Berlin, clofibricacid 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 ofhuman pharmaceuticals in the river Lea in England and found that over 170 substances areused in excess of 1 tonne per year in the river’s catchment. This allowed them to predict aconcentration of at least 0.1µg.l-1 in the river water (Halling-Sørensen, 1998).


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Table d.6: Pharmaceutical compounds identified in environmental samples [afterDaughton, 1999]

Compound Use/Origin Environmental occurrenceAcetaminophen Analgesic/anti-

inflammatoryRemoved efficiently by WWTS, max. effluent 6µgl-1, not detected in surface waters

Acetylsalicylic acid Analgesic/anti-inflammatory

Ubiquitous, removal efficiency 81%, max. effluent1.5µgl-1, in surface water 0.34µgl-1.

Betaxolol betablocker Max. effluent 0.19µgl-1, in surface water 0.028µgl-1.

Benzafibrate Lipid regulator Removal efficiency 83%, max. effluent 4.6µgl-1, insurface water 3.1µgl-1.

Biphenylol Antiseptic, fungicide Extensive removal in WWTS.Bisoprolol betablocker Max. effluent 0.37µgl-1, in surface water 2.9µgl-1.Carazolol Betablocker Max. effluent 0.12µgl-1, in surface water 0.11µgl-

1.Carbamazepine Analgesic, anti-

epilepticRemoval efficiency 7%, max. effluent 6.3µgl-1, insurface waters 1.1µgl-1.

Chloroxylenol Antiseptic In influents and effluents <0.1µgl-1.Chlorophene Antiseptic Influent 0.71µgl-1, removal not very efficientClenbuterol β2-sympathomometic Max. effluent 0.08µgl-1, in surface waters 0.05µgl-

1.Clofibrate Lipid regulator River water 40ngl-1, not detected in effluent or

surface waters.Clofibric acid Metabolite of

clofibrateRemoval efficiency 51%, max. effluent 1.6µgl-1,surface waters 0.55µgl-1, up to 270ngl-1 inGerman tap waters

Cyclophosphamide antineoplastic Max. effluent 0.02µgl-1, not detected in surfacewaters, high in hospital sewage: up to 146ngl-1

Diatrizoate X-ray contrast media Resistant to biodegradation, median in Germansurface waters 0.23µgl-1, locally very highconcentrations can occur.

Diazepam Psychiatric drug Max. effluent 0.04µgl-1, not detected in surfacewaters.

Diclofenac-Na Analgesic/anti-inflammatory

Removal efficiency 69%, max. effluent 2.1µgl-1, insurface waters 1.2µgl-1.

Dimethylaminophenazone

Analgesic/anti-inflammatory

Removal efficiency 38%, max. effluent 1µgl-1, insurface waters 0.34µgl-1.

17α-ethinylestradiol Oral contraceptive Up to 7ng.l in WWTS effluent, not detected inGerman surface waters above 0.5ngl-1.

Etofibrate Lipid regulator Not detected in WWTS effluent and surfacewaters

Fenfluramine Sympathomimeticamine

No studies but is known to be an endocrine-disrupting substance

Fenofibrate Lipid regulator Efficiently removed, max. effluent 0.03µgl-1, notdetected in surface waters.

Fenofibric acid Metabolite offenofibrate


Fenoprofen Analgesic/anti-inflammatory

Not detected in WWTS effluent or surface waters

Fenoterol β2-sympathomometic Max. effluent 0.06µgl-1, in surface waters0.061µgl-1.

Flurorquinolonecarboxylic acids

Antibiotics Ubiquitous, led to resistance in pathogenicbacteria, strongly sorbs to soil.

Fluoxetine Antidepressant No studiesFluvoxamine Antidepressant No studies


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Gemfibrozil Lipid regulator Removal efficiency 69%, max. effluent 1.5µgl-1, insurface waters 0.51µgl-1.

Gentisic acid Metabolite ofacetylsalicylic acid

Efficiently removed by WWTS, max. effluent0.59µgl-1, in surface waters 1.2µgl-1.

o-hydroxyhippuricacid

Metabolite ofacetylsalicylic acid

Efficiently removed by WWTS, not detected ineffluent or surface waters.

Ibuprofen Analgesic/anti-inflammatory


Ifosamide Antineoplastic Max. effluent 2.9µgl-1, not detected in surfacewaters, hospital sewage 24ngl-1, totally refractoryto removal by WWTS.

Indomethacine Analgesic/anti-inflammatory

Removal efficiency 75%, max. effluent 0.60µgl-1,in surface waters 0.2µgl-1.

Iohexol X-ray contrast media Very low aquatic toxicity.Iopamidol X-ray contrast media Max. effluent 15µgl-1, median 0.49µgl-1.Iopromide X-ray contrast media Resistant to biodegradation, yields refractory,

unidentified metabolites, max. effluent 11µgl-1.Iotrolan X-ray contrast media Very low aquatic toxicity.Ketoprofen Analgesic/anti-

inflammatoryMax. effluent 0.38µgl-1, in surface waters 0.12µgl-1.

Meclofenamic acid Analgesic/anti-inflammatory

Not detected in WWTS effluent or surface waters.

Metoprolol betablocker Removal efficiency 83%, max. effluent 2.2µgl-1, insurface waters 2.2µgl-1.

Nadolol Betablocker Max. effluent 0.06µgl-1, not detected in surfacewaters.

Naproxen Analgesic/anti-inflammatory

Removal efficiency 66%, max. effluent 0.52µgl-1,in surface waters 0.39µgl-1.

Paroxetine antidepressant No studiesPhenazone Analgesic Removal efficiency 33%, max. effluent 0.41µgl-1,

in surface waters 0.95µgl-1.Propranolol Betablocker Removal efficiency 96%, max. effluent 0.29µgl-1,

in surface waters 0.59µgl-1.Propyphenazone Analgesic/anti-

inflammatoryPrevalent in Berlin waters.

Salbutamolalbuterol

β2-sympathomometic Max. WWTS influent 0.17µgl-1, in surface waters0.035µgl-1.

Salicylic acid Metabolite ofacetylsalicylic acid

Up to 54µgl-1 in WWTS effluent but efficientlyremoved in effluent, average in effluent 0.5µgl-1,in surface waters 4.1µgl-1.

Sulfonamides Antibiotics Present inn landfill leachatesTerbutaline β2-sympathomometic Max. effluent 0.12µgl-1, not detected in surface

waters.3,4,5,6-tetrabromo-o-cresol

Antiseptic, fungicide Found in influents and effluents in Germany<0.1µgl-1.

Timolol Betablocker Max. effluent 0.07µgl-1, in surface waters 0.01µgl-1.

Tolfenamic acid Analgesic/anti-inflammatory

Not detected in WWTS effluent or surface waters.

Triclosan Antiseptic 0.05-0.15µgl-1 in water, very widely used.Verapamil Cardiac drug No occurrence data


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Table d.6 summarises the occurrence of certain pharmaceuticals in the environment. It canbe seen that some pharmaceuticals, such as lipid regulators, X-ray contrast media,antibiotics etc., are ubiquitous and extremely persistent in the environment, some are evenpresent in drinking water: clofibric acid for example has been found at concentrations up to0.27 µg.l-1 in some German waters. This would breach EC regulations if the compoundswere classed as pesticides (ENDS, 2000). However only a fraction of the drugs on themarket, 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 liverthrough oxidation, reduction, or hydrolysis to “phase I” metabolites, which tend to be moretoxic than the parent compound. Other reactions, such as conjugation, metabolise thecompounds 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 theparent compounds, either during treatment of wastewater and sewage sludge or in theenvironment. 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 theinvestigation of metabolites as well as parent compounds.

For example, 17β-oestradiol, is administered orally and mainly undergoes first-pass hepaticmetabolism, being transformed to oestrone and oestriol, which are less potent (Christensen,1998). Other metabolites are also formed but to a lesser extent. Experiments using thediluted slurry of activated sludge from a WWTS, were undertaken to investigate thepersistence of natural oestrogens and contraceptives under aerobic conditions (Ternes,1999b). The natural oestrogen 17β-oestradiol, was oxidised to oestrone, which is thenlinearly removed with time. Rapid elimination also occurred for16α-hydroxy-oestrone.However, the contraceptive 17α-ethinyloestradiol was persistent and highly stable underenvironmental conditions (Ternes, 1999b). Two glucuronides of 17β-oestradiol were cleavedto their parent compounds and 17β-oestradiol was re-released in an activated form (Ternes,1999b). This indicates that the microorganisms present have the ability to deconjugateoestrogen glucuronides. It is interesting to note that glucuronide conjugates are the mainoestrogen metabolites excreted by humans, so during wastewater treatment, theconcentration of free oestrogen increases due to the cleavage of the glucuronide moietiesfrom the compounds. As a result, the predominant presence of oestrone in WWTS effluentsand rivers is due to; its high stability during treatment; the cleavage of glucuronideconjugates from oestrone and 17β-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 thatup to 40% of penicillin V is excreted unchanged (Christensen, 1998).

Cyclophosphamide, an anti-cancer drug is administered intravenously or orally. It is notactive in itself but undergoes activation in the body when transformed to phosphoramidemustard 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 (seeTable d.7) (Hirsch, 1999) and it has been determined that up to 90% of the parentcompounds are excreted unchanged (ENDS, 2000). These active products can be excretedeither as unchanged compounds or as conjugates; 30-90% of administered antibiotics areexcreted via urine as active substances (Alcock et al., 1999). This introduces the problem at


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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].

Excretion (%)Antibiotic Amount prescribed(t/a) Unchanged Other

MetabolitesAmoxicillin 25.5-127.5 80-90 10-20Ampicillin 1.8-3.6 30-60 20-30

Penicillin V 40 40 60Penicillin G 1.8-3.6 50-70 30-50

Sulphamethoxasole 16.6-76 15Trimethoprim 3.3-15 60Erythromycin 3.9-19.8 >60Roxithromycin 3.1-6.2 >60Clarithromycin 1.3-2.6 >60

Minocycline 0.8-1.6 60 40Doxycycline 8-16 >70

A study looking at the amounts of antibiotics in human faeces found trimethoprim anddoxycycline at concentrations between 3-40 mg.kg-1, and erythromycin at concentrationsaround 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 removedefficiently because elimination is mainly due to adsorption on activated sludge, which ismediated through hydrophobic interactions (Hirsch, 1999). As a consequence, receivingwaters and other environmental media may become contaminated. Furthermore,erythromycin and other drugs such as naproxen and sulphasalazine, have survived in theenvironment for over a year (Zuccato, 2000). Clofibric acid was also found to survive for 21years 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 asorganic compounds, such as persistence and lipophilicity; much less is known though abouttheir entry into the environment and their subsequent fate (Alcock et al., 1999). Over 30% ofall 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 threegroups:

• 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 duringwastewater 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 studiedwith regards to their biodegradability. Chinolones and nitromidazoles possess differentchemical structures, actvity spectra and modes of action. The study found low rates ofbiodegradation for ciprofloxacin, ofloxacin, and metronidazole; it also found that thegenotoxicity of these compounds remained unaffected during treatment (Kummerer, 2000).


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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 treatmentand 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 fishfarms, are transferred into the environment (Halling-Sorensen, 1998).

Table d.6 gives an overview of the present knowledge on the environmental fate of specificpharmaceuticals. It can be seen that most hormones, such as oestrogen, are persistent in allareas and that most of the antibiotics used for human treatment are not biodegradable. Themajority of other compounds used for human treatment are also non-biodegradable, with theexception of the following: paracetamol, nicotinamide, ibuprofen, caffeine, and aspirin. Thecompounds used in veterinary treatment tend to be more biodegradable than humanpharmaceuticals, although the speed of degradation will depend on environmentalconditions, such as pH, and temperature.

It has also been determined that iodinated X-ray contrast media are not degraded duringwastewater 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, soare not readily biodegradable. Ninety percent of X-ray contrast media are excretedunmetabolised (Ternes, 2000); hence, receiving waters will also tend to be contaminated.Concentrations up to 0.49 µg.l-1 for iopamidol were detected in receiving rivers (Ternes,2000). It appears that groundwater may also become contaminated, as concentrations of upto 2.4 µg.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 lowerthan that detected in wastewater effluent; however, due to the high persistence of suchmedia in the environment, this reduction seems to be less important than for otherpharmaceutical compounds (Ternes, 2000). This shows that pharmaceutical compoundshave 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 ofpharmaceuticals, which renders the task of studying the fate of these compounds inwastewater very difficult. The list of pharmaceutical substances could be exhaustive andprioritisation is necessary.

Certain physical processes occur that may be used to degrade these contaminants: sorptionto solids, volatilisation, chemical degradation, and biodegradation. The effectiveness ofsorption 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 notadsorb 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.• if Hc is greater than 1x10-4 and Hc/Kow is greater than 1x10-9, the compound is thought

to have a high volatilisation potential,• if Hc is less than1x10-4 and Hc/Kow is less than 1x10-9 the compound is thought to

have a low volatilisation potential.


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Several models, using Kow, Koc, and Hc, have been developed in order to take all thesecharacteristics into account in order to prioritise pollutants. This has enabled an assessmentto be made of the exposure risk to such pollutants, through consuming food derived fromsludge-amended soil. Two parameters are important when trying to determine movement ofcontaminants through the food chain: persistence, and non-polarity. Easily metabolised orpolar compounds do not move through the food chain. As bioconcentration increases withlipophilicity, 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, itbecomes difficult to eliminate or prioritise pharmaceuticals using this screening process. Inaddition, 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 sewagesludge (Alcock et al., 1999). In the absence of detailed knowledge, it can be presumed thatmost pharmaceutical substances have the same properties as pesticides and other organicpollutants (Halling-Sorensen, 1998). Also, as mentioned already, many of these compoundshave lipophilic characteristics, so they are likely to accumulate. Some pharmaceuticals mayeven 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 metaboliccharacteristics leads to incomplete removal in WWTS (Daughton, 1999). They tend to benon-volatile, so transport and movement through the environment will occur via aquaticmedia. In fact, their polarity and non-volatile characteristics will often prevent them fromleaving the aquatic environment (Daughton, 1999). As it has been seen earlier, metabolitesalso tend to be cleaved to the parent compound during wastewater treatment and thenreleased afterwards.

Nutraceuticals/Herbal Remedies

During the last several years, the popularity of nutritional supplements was codified by thecreation of a new term for the subclass of highly bioactive food supplements callednutraceuticals (Daughton and Ternes, 1999) also referred to as nutriceuticals.Nutraceuticals are a rapidly growing commercial class of bioactive compounds, usuallybotanicals, intended as supplements to the diet. Nutraceuticals and many herbal remediescan have potent physiologic effects. These are a mainstay of alternative medicine and haveenjoyed 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 activitybut are not classified as drugs by the FDA, primarily because a given botanical usually hasnot one but an array of distinct compounds whose assemblage elicits the putative effect andbecause these arrays cannot be easily standardized. As such, they are not regulated andare available over the counter (heavily promoted via the Internet). Even in those cases inwhich 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 activeingredient in the approved prescription drug Mevacor used to lower cholesterol levels), arecent 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 Officeof Dietary Supplements (ODS) via the DSHEA in 1995 under the National Institutes ofHealth (NIH) (DSHEA, 1994). The ODS maintains a searchable database (InternationalBibliographic Information on Dietary Supplements [IBIDS]) of published scientific literatureon dietary supplements (NIH Office of Dietary Supplements, 1999).


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Although these substances are readily available off the counter, not always in acharacterized/standardized forms, an effort is underway to patent various nutraceuticals bystandardizing the extracts and thereby making them available only by prescription. Thepatenting of hundreds of multiple-molecule nutraceuticals for therapeutic purposes couldlead to more widespread use of these substances.

As an example, a recent addition to this class is a substance called huperzine A, an alkaloidextracted from a Chinese moss, which has been documented to improve memory. It istherefore experiencing strong demand for treating Alzheimer's disease and has captured theattention of those who follow the nutraceutical market because of its true pharmaceuticalqualities. The significance of this particular compound is that it possesses acute biologicactivity as a cholinesterase inhibitor identical to that of organophosphorus and carbamateinsecticides. It is so effective that the medical community is concerned about itsabuse/misuse, especially since it is legal. While huperzine A, and alkaloids in general(compounds with heterocyclic nitrogen, proton-accepting group, and strong bioactivity), arenaturally occurring compounds, their susceptibility to biodegradation in WWTS or in surfacewaters is unknown. This is the case for almost all nutraceuticals, therefore more research isneeded.

Another example is Kava, which is prepared from the root of Piper methysticum, used of itsmild narcotic effect among other effects. The active ingredients in Kava are believed to be asuite of lipophilic lactones comprising substituted -pyrones (methysticin, kavain, yangonin,and others) (Shao, et.al.,1998). These compounds display a host of effects in humans, butnothing is known about their effects on other organisms or fate in WWTS.

There are many nutraceuticals, both new and ‘traditional’, experiencing increasedconsumption. These few examples illustrate the unknowns regarding whether thesecompounds are being excreted, surviving WWT, and then having possible effects on aquaticorganisms. Nutraceuticals and herbal remedies would have the same potential fate in theenvironment as pharmaceuticals, with the added dimension that their usage rates could bemuch higher, as they are readily available and taken without the controls of prescriptionmedication. However, because these compounds are natural products, they would beexpected to biodegrade more easily .

Legislation and policy for risk assessment

At the beginning of the 1980s, environmental risk assessment was introduced for newchemicals 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 andthose used in veterinary practice. For several years legislation has been implemented forveterinary medicines. The EU Directive 81/852/EEC, (Amended 1993), introduced therequirement for a tiered environmental risk assessment of new veterinary products, andattempts are being made to implement this for a review of existing substances (ENDS,2000). Currently, environmental risk assessment consists of examining the likelyenvironmental sectors and if levels of pharmaceutical compounds exceed the trigger valuesset, such as 100 µg.l-1 in manure, further data is required (ENDS, 2000). The technicaldirective [Directive 81/852/EEC, amended 1993] concerning veterinary medical productsoutlines the basic requirements for conducting an environmental risk assessment (Halling-Sorensen, 1998). The technical directive [Directive 75/318/EEC, amended 1993] concerninghuman medical products does not refer to any ecotoxicology or ecotoxicity tests and noguidance is given on how to carry out an environmental risk assessment for drugs used byhumans. However, a draft directive for human pharmaceuticals is currently being devised,proposing that risk assessment should be part of the approved procedure of new medicalsubstances (Halling-Sorensen, 1998). The EC is proposing a similar programme for human


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medicines: if drug concentrations in surface waters are predicted to exceed 0.01 µg.l-1,toxicity testing is required to find the no effect level (NOEC) (ENDS, 2000). Althoughenvironmental assessment of the potential impacts of newly developed drugs has beenexpected in the EU since the 1st of January 1995 (Christensen, 1998), it is should be notedthat the end point of human exposure is not usually investigated. Also, assessment ofindividual compounds is usually based on a limited number of tests but pharmaceuticals inthe environment may affect a large number of different organisms and species, so this abilityshould be reflected in the tests carried out (Stuer-Laurisden, 2000). Pharmaceuticals maynot 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 onestandard ecotoxicological test. The most sensitive reaction however, was to a non-standardtest incorporating relevant end points for the pharmaceuticals (Henschel, 1977), so provingthe limitations of standard tests.

Risk assessment of pharmaceuticals

Risk assessment for pharmaceuticals in the environment has not usually been carried outdue to the lack of data and the need for more precise and sensitive measures in theenvironmental sectors. Some high consumption compounds, such as antibiotics and clofibricacid, are being released into the environment and have been found to be widely present inaquatic environments, sediments and soils. Although the liver often metabolisespharmaceutical compounds to more easily hydrolysed compounds, the metabolites can becleaved back to the more hydrophobic, active parent compounds by bacteria, which canthen persist and bioaccumulate.

There is a strong opinion that there are more pressing environmental problems thanpharmaceuticals and that these compounds do not pose a large risk because they arepresent 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 favouredby low concentration exposure and compounds such as antineoplastic agents and hormoneshave 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 lowlevels (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 continuouslyinto 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 thelong-term, chronic impacts that may be of concern are genotoxicity and reproductionimpairment. 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 becauseof toxic effects in the food organisms (ENDS, 2000). The effects of continuous exposure toeven low levels of pharmaceuticals in the environment are very complex and affect manydifferent organisms. More studies are necessary with regards to the long-term impacts andthe potential synergistic effects of exposure to a mixture of drugs.

Exposure route is very important in determining environmental loading, as the dose andduration of exposure are important parameters in risk assessment. Drugs tend to bereleased in low concentrations, although local discharges, such as those coming fromhospitals, may have higher concentrations (Jorgensen, 2000).


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A study carried out in Italy (Zuccato, 2000), found that most drugs were present in drinkingwater, 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 beobserved. The study concluded that risk seems negligible but possible long-term exposuresand impacts still need investigation. The same was found with X-ray contrast media, whichwere ubiquitous in the aquatic environment and highly persistent. No acute toxic effectswere 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 programEUSES, which was developed as a support to the technical guidance document for riskassessment on new and existing substances (Christensen, 1998). The three compoundsassessed were; the synthetic estrogen, 17α-ethinylestradiol; the antibiotic, penicillin V; andthe antineoplastic agent, cyclophosphamide. This program estimated environmental fate andhuman exposure based on worst-case scenarios, using data on the physico-chemicalproperties of the compounds and amounts consumed. The results indicated that for all threethere was negligible human risk. However, the author stressed the point that manyuncertainties are associated with this method and that the drugs, although seeminglyinsignificant, still contribute to the total toxic load in the environment, and that interactionsmay have ecotoxicological impacts.

Research in Denmark attempted to carry out a risk assessment for the 25 most highly useddrugs in the primary health sector in Denmark, including furosemide, paracetamol,ibuprofen, and estradiol (Stuer-Laurisden et al., 2000). Different parameters are used tocalculate environmental exposure: biodegradation, bioaccumulation, and bioavailability arethree of the most important. Nevertheless, it has been seen that biodegradation ofpharmaceuticals does not happen very often. Bioaccumulation in the human body does nothappen for drugs, as they are metabolised to more polar compounds that can then easily beexcreted. The bioavailability of drugs is different if it is bound to solids, adsorbed, ordissolved. However, as seen above, the octanol-water partition coefficient can be used todetermine bioaccumulation, and other parameters can be used to determine bioavailability.However, this information on the properties of the compounds is not readily available. Theyfound that ecotoxicology data was available only for 6 of the 25 compounds, andbiodegradation 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 itis anticipated that the highest values would be found, i.e. aquatic ecosystems and sewagesludge (Jorgensen, 2000). In order to do so, modelling could be a useful tool; however, notmany 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 concentrationsfor the 6 compounds were calculated and all exceeded 0.001µg.l-1, which is the cut off valuein EU legislation for carrying out more investigations. The majority of the PECs are between1-100ng.l-1, it is only for the top 5 compounds that these reach the µg.l-1 range (Stuer-Laurisden, 2000). The predicted no effect concentration (PNEC) is based onecotoxicological 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, preventingcomplete risk assessments. Nevertheless, it was concluded, with this data, that ibuprofen,aspirin, and paracetamol may pose a risk; hence, contradicting other studied that hadconcluded these were efficiently removed by treatment and did nor reach hazardous levelsin the environment (ENDS, 2000). The efforts are hampered by the fact that concentrationsmeasured in sludge and effluent vary extensively, and furthermore comparisons of predictedconcentrations in sludge based on Kow, sludge-water partition coefficients (Kd), or acid-baseconstants (pKa) also reveal large variations (Stuer-Laurisden, 2000).


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The cost/benefit stage of risk assessment is extremely important: the indirect and directeffects of the drug on the environment and the human body must be known in order to makean informed decision (Jorgensen, 2000 and Halling-Sørenson,1998). This can allowselection 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 ofrisks.

Examples of good environmental practice

In France, and the UK, there are procedures for returning prescribed but unusedpharmaceuticals. In France, these plans are encouraged by the ADEME "RETOUR" initiative(ADEME, 1997a), which also encourages the distributor to include the costs of collection andtreatment into the product's selling price. This strategy is useful for reducing the amount ofpolluted domestic and artisanal (laboratories, photographic shops etc) wastewater throughspecial collections for specific pollutants such as thermometers, medicines, and paintleftovers. Most areas in France have implemented such programmes and they aresuccessful.

There are also possibilities of reformulating and substituting certain pharmaceuticalcompounds with substances incurring fewer impacts. It has been found thatcyclophosphamide and ifosfamide, the very widely used antineoplastic drugs, act through anactive metabolite, which is highly unstable. German researchers have detected anothercompound that has the same therapeutic activity but that is much more readily biodegraded(ENDS, 200). However, in order to research more environmentally suitable drugs, thecharacteristics of the existing ones must be known, and, as yet, there is still very little dataavailable.

If the assessment of a drug gives a high risk, the response may not have to be the phasingout of the drug, but maybe just the collection and specific treatment of the faeces and urinecontaining this compounds (Jorgensen, 2000). Environmental risk assessment should bepart of the development of all new drugs and could be used as a marketing tool, as publicconcern for the environment is increasing (Jorgensen, 2000).

Analysis tools and their sensitivity must be improved for the determination of the very lowconcentrations of drugs. A method obtaining detection limits within the ngl-1 range forvarious pharmaceutical compounds, particularly neutral basic drugs such as betablockers,has been developed, using advanced solid phase extraction, modified derivatisationprocedures and LC-electrospray-MS/MS detection (Ternes, 1998a). This allows detectiondown to 10ng.l-1, in different aqueous matrices. In another study, determination limits downto 5ng.l-1 were achieved for phenolic compounds and other acidic drugs, such as lipidregulators and acetylsalicylic acid metabolites, using solid phase extraction and methylationor acetylation of the carboxylic and phenolic hydroxyl groups, followed by detection byGC/MS (Ternes, 1998b).

In Sweden, a pharmaceutical company, AstraB, was discharging very toxic effluentscontaining a large amount of persistent organic pollutants and phosphorus, which hadcaused operational problems at the WWTS (Rosen, 1998). There were large variations inthe composition of their effluent over time, as the drugs tend to be produced in discontinuousbatches. Hence, a broad and flexible treatment method had to be introduced to treat thewastewater at source. The wastewater was investigated and it was found that the maincontribution came from the treatment of packages containing non-approved liquidpharmaceutical preparations (Rosen,et.al. 1998). The washing water had a very high toxicityand could not be treated biologically. This effluent was removed and incinerated and theremaining effluent still too toxic for discharge into the UWWT system is now treated using a


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multi-stage biofilm process removing all organic matter and the toxicity is no longermeasurable (Rosen, et.al. 1998).

Conclusion and Recommendations

Pharmaceutical compounds must become priority substances in the same way as persistentorganic pollutants are. Judgements on the relative priorities are based on the knowledge atthe time and the priority list will obviously change over time as more studies are carried outand more data is gathered. Pharmaceuticals are widely used and mainly disposed ofthrough the sewerage system, allowing their entry into the environment continually, asremoval rates can be compensated by replacement rates (Daughton and Ternes, 1999).They are concerning because they are biologically active and are usually lipophilic andpotentially bioaccumulating. Many resist biodegradation, within the WWTS and in theenvironment, and can end up in surface and ground waters, as well as sediment and soilsand are found to be highly stable under environmental conditions. Furthermore, metabolitestend 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 haveunpredicted and unknown side effects particularly after long-term exposure to lowconcentrations. Aquatic ecosystems are the most vulnerable, as this is the mainenvironmental compartment where pharmaceuticals are found ubiquitously.

Analytical methods must be improved to detect pharmaceuticals at very low levels andsampling procedures must be of very high quality so as not to cause contamination. Moreinformation on the physicochemical, ecotoxicity, and ecotoxicological characteristics, usingappropriate tests that better accommodate subtle end points, of drugs and their metabolitesshould 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 thewhole process. Furthermore, use patterns of drugs in all countries is still very limited andshould be determined, as it is essential for the elaboration of amounts released into theenvironment.

Screening of high use drugs should be carried out and samples with high potential shouldthen be subject to more analyses (Daughton and Ternes, 1999). Furthermore, a moreprecautionary view on the potential impacts of the drugs should be adopted and morestudies are required to elucidate these effects at the concentrations observed and alsoinvestigating additive and synergistic effects of mixtures (Daughton and Ternes, 1999).

Risk assessment for pharmaceuticals should include an assessment of their biodegradabilityand environmental fate and potential impact as occurs for other discharged substances,such as detergent residues. Furthermore, the disposal of unwanted drugs into thewastewater system from domestic sources should be discouraged by encouraging collectionof these wastes.


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(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 havingintended end uses, primarily on the human body (products not intended for ingestion) or inthe household. In general, these chemicals alter odour, appearance, touch, or taste withoutdisplaying significant biochemical activity (Daughton and Ternes, 1999). Most of thesechemicals are used as the active ingredients or preservatives in cosmetics, toiletries andfragrances. They are not used for treatment of disease, but some may be intended toprevent diseases (e.g., sunscreen agents). In contrast to drugs, almost no attention hasbeen given to the environmental fate or effects of personal care products, the focus hastraditionally been on the effects from intended use on human health.

Personal care products differ from pharmaceuticals in that large amounts can be directlyintroduced to the environment and unlike medicinal compounds, there are rarelyrecommended doses. These products can be released directly into recreational waters orvolatilised into the air. Because of this direct release they can bypass possible degradationin UWWT. Also, in contrast to pharmaceuticals, less is known about the effects of this broadand diverse class of chemicals on non-target organisms, such as aquatic organisms. Dataare also limited on the potential adverse effects on humans. For example, commonsunscreen 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. Forexample, 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 careproducts that are ubiquitous pollutants, which may possess varying degrees of bioactivity.

Table e.1 Personal care and fragrances produced in Germany (1993)

Product category TonnesBath additives 162 300Shampoos, hair tonic 103 900Skin care products 75 500Hair sprays, hair dyes, setting lotions 71 000Oral hygiene products 69 300Soaps 62 600Sun screens 7 900Perfumes, aftershaves 6 600TOTAL 559 100

• Preservatives

Parabens (alkyl-p-hydroxybenzoates) are one of the most widely and heavily used types ofantimicrobial preservatives in cosmetics (skin creams, tanning lotions, etc.), toiletries,pharmaceuticals, and even foodstuffs (up to 0.1% wt/wt). Although the acute toxicity of thesecompounds is very low, Routledge et al.[1998] report that these compounds (methyl throughbutyl homologs), display weak oestrogenic activity. Although the risk from dermal applicationin humans is unknown, the probable continual introduction of these benzoates intowastewater treatment systems and directly to recreational waters from the skin, leads to thequestion of risk to aquatic organisms. Butylparaben showed the most competitive binding tothe rat oestrogen receptor at concentrations one to two orders of magnitude higher than thatof nonylphenol and showed oestrogenic activity in a yeast oestrogen screen at 10-6 M .


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• Disinfectants/Antiseptics

Triclosan, a chlorinated diphenyl ether: 2,4,4´-trichloro-2´-hydroxydiphenyl ether, is anantiseptic agent that has been widely used for almost 30 years in a vast array of consumerproducts. Its use as a preservative and disinfectant continues to grow; for example, it isincorporated at < 1% in Colgate's “Total” toothpaste, the first toothpaste approved by theFDA to fight gingivitis. While triclosan is registered with the U.S. EPA as a pesticide, it isfreely 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 isincorporated into a wide variety of plastic products from children's toys to kitchen utensilssuch as cutting boards. Many of these uses can result in direct discharge of triclosan toUWW collecting systems, and as such this compound can find its way into receiving watersdepending on its resistance to microbial degradation. Okumura and Nishikawa [1996] foundtraces of triclosan ranging from 0.05 to 0.15 µg.l-1 in water. Although triclosan has long beenregarded 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 actuallyacts as an antibacterial, having particular enzymatic targets (lipid synthesis). As such,bacteria could develop resistance to triclosan. As with all antibiotics in the environment, thiscould 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 byhouseholds and livestock breeders. These compounds are often substituted phenolics aswell as other substances, such as triclosan. Biphenylol, 4-chlorocresol, chlorophene,bromophene, 4-chloroxylenol, and tetrabromo-o-cresol [Ternes et al 1998] are some of theactive 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 2.6µg/L for biphenylol and up to 0.71 µg.l-1 for chlorophene, and effluents. The removal ofchlorophene from the effluent was less extensive than for biphenylol, with surface watershaving concentrations similar to that of the effluents.

• Sunscreen Agents

The occurrence of sunscreen agents (UV filters) in the German lake Meerfelder Maar wasinvestigated by Nagtegaal et al. [1997]. The combined concentrations of six sunscreenagents (SSAs) identified in perch (Perca fluviatilis) in the summer of 1991 were as high as2.0 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 Germanlakes. These lipophilic SSAs seem to occur widely in fish from small lakes used forrecreational swimming. Both fish species had body burdens of SSA on par with PCBs andDDT. The bioaccumulation factor, calculated as quotient of the MBC concentration in thewhole fish (21 µg.kg-1) versus that in the water (0.004 µg l-1), exceeded 5,200, indicating highlipophilicity. The fact that SSAs (e.g., 2-hydroxy-4-methoxybenzophenone [oxybenzone] and2-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 andbioconcentration in aquatic species. No data have been published on newer SSAs such asavobenzene (1-[4-(1,1-dimethylethyl)phenyl]-3(4-methoxyphenyl)-1,3-propanedione).

• Perfume IngredientsThe raw ingredients in perfumes include essential oils, plant extracts and animal secretions,and synthetic or semi synthetic (natural material that has undergone some chemicalmodification) compounds. Thousands of these substances can be blended to createperfumes. These can be used directly as perfumes or as scents in other products, for


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example in cosmetics, cleaning agents and air fresheners. Perfume ingredients may enterthe urban wastewater system directly from domestic sources, such as in the washing agentsor 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 concerninclude

- nitro-musk compound,- polycyclic musk compound,- solvents and fixatives- other fragrances.

Details of the physical properties of these compounds are included in Appendix B . Healthand environmental effects of the compounds discussed are also briefly introduced within thiscase study.

This case study will focus on musk compounds. Fragrances (musks) are ubiquitous,persistent, bioaccumulative pollutants that are sometimes highly toxic; amino musktransformation products are toxicologically significant.

Synthetic musks comprise a series of structurally similar chemicals (which emulate the odourof the more expensive, natural product, from the Asian musk deer), used in a broadspectrum of fragranced consumer items, both as fragrance and as fixative. Included are theolder, synthetic nitro musks (e.g., ambrette, musk ketone, musk xylene, and the lesserknown musks moskene and tibetene) and a variety of newer, synthetic polycyclic musks thatare 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. Thesesubstances are used in nearly every commercial fragrance formulation (cosmetics,detergents, toiletries) and most other personal care products with fragrance; they are alsoused 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 persistenceand possible adverse environmental impacts and therefore are beginning to be phased outin some countries. Musk xylene has proved carcinogenic in a rodent bioassay and issignificantly absorbed through human skin; from exposure to combined sources, a personcould absorb 240 µg/day [Bronaugh et al 1998]. The human lipid concentration of variousmusks 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 whichwere 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 muskketone 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), andoysters (Crassosterea gigas). This was followed by a number of studies in Europe, some ofwhich are summarised in table e.2.


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Table e.2 Concentrations of various musk compounds in environmental samples.Location Tissue/Substance Product Concentration ReferenceNorth Germany Freshwater Fish

(Fillet)Musk XyleneMusk Ketone

10-350µg.kg-1

10-380µg.kg-1Geyer et al, 1994Geyer et al, 1994

Ruhr River,Germany

Bream and Perch(Fatty tissue)

Galaxolide,Tonalide andCelestolide

Averageconcentrations

between 2.5 and4.6 mg.kg-1 (ppm)

Eschke et al 1998

Berlin, Germany Surface waters Galaxolide,Tonalide andCelestolide

Maximumconcentrationsabove 10µ L-1

Herberer et al1999

Elbe River,Germany

Particulate matterfrom river samples

Musk ketoneGalaxolideTonalide

4-22 ng/g148-736 ng/g194-770 ng/g

Winkler et al. 1998

Italy Freshwater Fish(Fatty tissue)

Galaxolide,Tonalide

4 ng.g-1 –1054 ng.g-1

Draisci et al 1998

Musks are refractory to biodegradation (other than reduction of nitro musks to aminoderivatives), which explains why they have been detected in water bodies throughout theworld [Gattermann et al 1998]. They also are very lipophilic [octanol-water partitioncoefficients are similar to those for DDT and hexachlorocyclohexane , Winkler et al, 1998]and therefore can bioaccumulate, leading to very high concentrations being measured insome studies.

The values for the three most prevalent musks in the Elbe river study (table e.2) were withinthe same order of magnitude as those for 15 polycyclic aromatic hydrocarbons (PAHs) andexceeded those for 14 common polychlorinated organic pollutants (only hexachlorobiphenyl[HCB] and p,p´-DDT were of similar concentration). Also, all the 31 water samples containedmusk ketone (2-10 ng.l -1), Galaxolide (36-152 ng.l -1), and Tonalide (24-88 ng.l-1); Celestolidewas found only at 2-8 ng.l-1. These higher values exceeded those for all the polychlorinatedorganics and the PAHs. The occurrences of individual musks are sometimes correlated as aresult of their use as mixtures in commercial products. In Germany, the nitro musks arebeing replaced by the polycyclic musks, therefore resulting in lower concentrations for muskketone [Winkler et al, 1998].

Although the significance of the aquatic toxicity of the nitro and polycyclic musks isdebatable (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 reducedderivatives are undoubtedly created under the anaerobic conditions of sewage sludgedigestion. Behecti et al. [1998] tested the acute toxicity of four reduced analogs of muskxylene on Daphnia magna. The p-aminodinitro compound exhibited the greatest toxicity ofthe four, with extremely low median effective concentration (EC50) values averaging 0.25µg.l-1 (0.25 ppb).

Recently, the amino transformation products of nitro musks were identified in wastewatertreatment effluent and in the Elbe River, Germany. Gatermann et al. [1998] identified muskxylene and musk ketone together with their amino derivatives 4- and 2-amino musk xylenesand 2-amino musk ketone. In wastewater entering treatment plants, the concentrations ofmusk xylene and musk ketone were 150 and 550 ng.l-1, respectively. In the effluent, theirconcentrations dropped to 10 and 6 ng.l-1, respectively. In contrast, although the aminoderivatives could not be detected in the influent, their concentrations in the effluentsdramatically increased, showing extensive transformation of the parent nitro musks: 2-aminomusk xylene (10 ng.l-1), 4-amino musk xylene (34 ng.l-1), and 2-amino musk ketone (250ng.l-1). It was concluded that the amino derivatives could be expected in wastewater effluentat concentrations more than an order of magnitude higher than the parent nitro musks. In the


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Elbe, 4-amino musk xylene was found at higher concentrations (1-9 ng.l-1) than the parentcompound.

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 theenvironment almost exclusively via treated wastewater; and are persistent andbioconcentratable, they are prime candidates for monitoring in both water and biota asindicators for the presence of other personal care chemicals. Their analysis, especially inbiota, 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 website2000] 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 adequatetoxicologically 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 primarilyin domestic situations or in buildings connected to the UWW collecting system, implies thattheir presence in surface waters occurs via municipal wastewater treatment plants. Themean concentrations found in biologically clarified wastewater from 25 German municipalWWTP were,

o for musk-xylene: 0.12 µg.l-1 (concentration range: 0.03 – 0.31 µg.l-1) ando 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 thecity’s main WWTP, during 1999. Concentrations of musk compounds in WWTP influent andeffluent are shown in Table e.3.


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Table e.3: Musk compound concentrations in influent and effluent of the pilot WWTPSimmering, Austria in 1999 [Hohenblum et al., 2000].n: number of samples analysed.

LOD: limit of detectionCompound(µg.l-1)

Type ofsample

Samplenumber >

LOD

Range(µg.l-1)

MeanValue(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 thesediment of UWW collecting systems in German commercial and residential areas, arepresented in Table e.4. In all samples noticeably high musk scent concentrations weredetected. Sediment samples from the UWW collecting system for the residential area hadslightly higher concentrations of the three polycyclic compounds ADBI (celestolide), HHCB(galoxolide) and AHTN (tonalide). This is probably due to the more frequent use of perfumesand detergents in domestic areas.

Table e.4: Concentration of different musk scents in sewage sludge and UWWcollecting system sediment in mg/kg DS, Germany [Antusch, 1999].

N = number of samples analysed. N>LOD number of samples over the limit of detection

Sediment:

industrial area

(n=17)

Sediment:

residential area

(n=2)

Sewage sludge

(n=2)

Compound N>

LOD

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


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Musk Compounds in Watercourses:

The pollution of the river Ruhr (Germany) with musk compounds was found to be relativelylow, 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 theirmuscle 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 foundin German receiving waters at concentrations up to the µg.l-1 level. In the Wuhle, a smallstream consisting almost totally of wastewater effluent, maximum concentrations were 12.5µg.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 watersample [Heberer et al., 1999].

The concentration of the three compounds tonalide, celestolide and galaxolide, weremeasured in different watercourses in Germany. The results of this study are shown in Tablee.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 theländer Sachsen and Sachsen-Anhalt, Germany [Lagois, 1996].<LOD: below limit of

detectionNitro-musk compounds

ADBI (celestolide)

[µg.l-1]

HHCB (galaxolide)

[µg.l-1]

AHTN (tonalide)

[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 <LOD 0.057 0.092 0.062 0.116

Bank filtration, Torgau <LOD <LOD <LOD <0.03 <LOD <0.03

Pure water, Torgau-East <LOD <LOD <LOD <0.03 <LOD <0.03

Dam, Rappdode <LOD <LOD <0.03 <0.03 <0.03 <0.03

Pure water, Wienrode <LOD <LOD <LOD <0.03 <LOD <0.03

Ground water, Kossa <LOD <LOD <LOD <LOD <LOD <LOD

Pure water, Kossa <LOD <LOD <LOD. <LOD <LOD <LOD

ConclusionsThere is very limited information on the health and environmental effects of personal careproducts, such as the musk compounds found in perfumes. Many of these compounds havethe potential to bioaccumulate, which is why there is concern about their presence inwastewater. Though these products may be used in large quantities there is insufficient datathough to establish whether the presence of these compounds in UWW could cause anydetrimental environmental or health effects.


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(f) Surfactants in Urban Wastewaters and Sewage Sludge

Introduction

Surfactants are the largest class of anthropogenic organic compounds present in rawdomestic wastewater. They are used in household and commercial laundry and cleaningoperations. Surfactants can be classified [Ullman’s Encyclopaedia of Industrial Chemistry,2000] into:

• Anionic surfactants are anion-active, amphiphilic compounds in which thehydrophobic residues carry anionic groups with small-sized counter-ions, such assodium, potassium or ammonium ions. These counter-ions have only a slightinfluence on the surface active properties. Examples include soaps, alkylbenzenesulphonates (ABS), alkylsulphates (AS), and alkylphosphates (AP).

• Non-ionic surfactants (NIS) – are amphiphilic compounds that are unable todissociate into ions in aqueous solutions, for example, alkyl- and alkylphenylpolyethylene 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 thehydrophobic groups exist as cations with counter-ions such as chloride, sulphate oracetate. Examples include tetraalkyl ammonium chloride, N-alkylpyridinium chlorideand others.

• Amphoteric surfactants have zwitterionic* hydrophilic groups (*electrically neutralions 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 fordomestic 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, sulphosuccinatesof 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 isreduced), mercerising (cotton treatment that requires wetting agents), and finishing. Woolwashing is done with NIS, while cotton is washed with anionic surfactants. The leatherindustry uses non-ionic and cationic surfactants as wetting and cleansing agents, and alsofor leather conditioning. Surfactants are also used as emulsifiers and dispersants, and alsoas food additives (natural substances only, such as glycerides, fatty acid salts, etc.).Pharmaceuticals manufacturing uses surfactants, and they are also used in agriculturalapplications such as crop-protection and pest control agents. Metal working and machining,petroleum extraction and processing, ore flotation and dressing, mineral oil industry, roadconstruction and maintenance work, cement industry, plastics production, pulp and paperindustry and printing, electroplating, adhesives manufacturing, all use surfactants for theirunique properties.


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Non-ionic surfactants are detergents which possess specific physicochemical properties,including relative ionic insensitivity and sorptive behaviour [deVoogt et.al, 1997] whichmakes them particularly suited for use wherever interfacial effects of detergents, foaming-defoaming, de-emulsification, dispersion or solubilisation can enhance product or processperformance. The major part of the non ionic surfactants group consists of alcoholethoxylates (AE) and alkylphenol ethoxylates (APE) of which, nonylphenol ethoxylate (NPE)is the main representative. Because of the formation of persistent metabolites in theenvironment, OSPARCOM member states have decided to phase out the use of NPE and toreplace the APEs with AEs. The production of non-ionic surfactants in the USA and EUamounts 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 hydrophilicgroups. They thus tend to concentrate at interfaces of the aqueous system including air, oilymaterial and particles. The hydrophobic group is generally a hydrocarbon radical (R) of 10 to20 carbon atoms. The hydrophilic portion may ionise or it may not. Ionic surfactants may beeither anionic or cationic. Ionic surfactants constitute approximately two-thirds of thesurfactants used. Cationic surfactants constitute less than 10% of the ionics and are used forfabric softening, disinfection and other specialized applications. The predominant class ofanionic surfactants is linear alkylbenzene sulphonates (LAS).

The concentrations of linear alkylbenzene sulphonates (LAS) in raw wastewater range from3 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 readilybiodegradable by aerobic processes, much of the surfactant load into a treatment facility(reportedly 20-50%) is associated with suspended solids and thus escapes aerobictreatment processes, being directed via primary sedimentation into sludge managementprocesses. Because LAS is not biodegraded by anaerobic biological processes usuallyemployed in sludge stabilization (McEvoy and Giger, 1985; Swisher, 1987), it may be foundin the gram per kilogram range in anaerobic sludges. Given these concentrations and themajor effects of surfactants in particle surface modification, deflocculation, and surfacetension reduction, it seems clear that the performance of certain treatment processes asthickening, conditioning, and dewatering may be strongly influenced by these materials.

Thus, surfactants may induce significant extra costs in sludge handling. Increased watercontent in landfilled sludges represents an additional possible impact, adding to the difficultyof proper landfill leachate control. Surfactants may also mobilise otherwise insoluble organicpollutants within the landfill. Similar implications exist for land application of surfactant-ladensludges.

Feijtel et al [1995] examined five WWTS across Europe and found the influent of LAS toWWTS in the UK to be higher than in the other plants (see Table d.1).


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Table d.1 LAS in influent and effluent in the UK compared with other regions [Feijtel etal 1995]

Country Influent mg/lMean (+/- 95% CIs)

Country Effluent mg/lMean (+/- 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 influentconcentrations than in Germany, Italy and the Netherlands (Spain had a very small data setand therefore there is uncertainty in these results, which is reflected in the large confidenceintervals, which overlap with those of the UK). Having a high influent of LAS was unrelatedto the concentration in the effluent and the UK samples had a lower concentration of LAS inits effluent than Germany, Italy or Spain. The level of biodegradation of LAS during thewastewater treatment process is high and also varied between the plants, with the UKWWTS having the highest level of biodegradation breaking down greater than 99.9 percentof the LAS.

The study above, used UK data from Holt et al [1995] in which the levels of LAS in theinfluent 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 Yorkshireregion, found much lower levels of LAS than expected. In no cases did the concentrations ofLAS reaching the plant approach the level predicted from consumer usage. The previousstudy had been carried out in March while the second was carried out in a ‘warm drysummer’. This study suggested there was significant biodegradation of LAS under certainconditions in the UWW collecting systems. Even in the relatively short residence time in theYorkshire UWW collecting system, for a couple of hours, up to 60% of the LAS was removedprior to the wastewater treatment plant. This was then followed by removal of between 70 to99 percent of the remaining LAS in the treatment plants.

Given the right treatment conditions, LAS are biodegradable and more research isnecessary 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 riversand steams sediment. In some cases (such as the small stream into which the OwlwoodWWTS discharges its effluent the LAS load in sediment upstream of the treatment plant washigher than that downstream [Waters and Feijtel 1995] This was hypothesised to bebecause that the LAS contribution upstream was due to unregulated discharges of untreatedwastewater to the aquatic environment while the input of the WWTS effluent actually servedto dilute the concentration of this pollutant downstream. A similar case was found in theNetherlands when concentrations of LAS could be higher upstream of a WWTS due to directdischarges from storm tanks. In other locations downstream sites were found to have veryslight elevations of LAS in sediment of between 0.49 and 3µg/g. A more careful policy ofdischarges 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 andDEHPs are discharged in the urban wastewater by the commercial sector and that they donot accumulate in agricultural soils treated with sewage sludge in moderate amounts. Figured.1 shows the level of surfactants and plasticisers in soils with the depth of the soil.


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Fig. d.1 Vertical distribution of various nonylphenols and phthalates in agriculturalsoils fertilized with large amounts of sewage sludge (17 tonnes dry matter per hectareper 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 byaerobic processes, based on laboratory studies (Swisher, 1970). Figure d.2 showsschematically the fate of LAS in the environment (deWolf and Feijtel, 1998). LAS evidentlyundergoes nearly complete biodegradation, with 97-99% removal rates found in somewastewater 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 removalrates, appreciable amounts are released to receiving waters. Ventura et al. (1989) identifiedLAS and a variety of other anionic, cationic and nonionic surfactants in both surface anddrinking water extracts.


184

Figure d.2 LAS fate in the environment (after deWolf and Feijtel, 1998)

Alkylphenol ethoxylates such as NPnEO are evidently less biodegradable than LAS withlaboratory 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 analysessuch as UV and IR spectroscopy (Swisher, 1970). This suggests that only partialdegradation occurs, such as conversion from polyethoxylates to nonylphenol diethoxylate(NP2EO), nonylphenol monoethoxylate (NP1EO), and nonylphenol (NP). Mass balancescarried 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 completeremoval of surfactants from treated waters is not entirely due to biodegradation. Brunner etal. indicated that 19% of the surfactant load entering a treatment facility is associated withsuspended 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 solidsis directed into sludge treatment processes via primary sedimentation. Surfactants such asLAS are not biodegraded by either mesophilic or thermophilic anaerobic digestion (McEvoyand Giger, 1985; Swisher, 1987) so a large proportion of these materials simply escapestreatment and becomes associated with sludge solids.

The resulting concentrations of surfactants in sewage sludges can be substantial. LASconcentrations measured in sludges often make up between 0.5% and 1.5% of the dry solidmass, particularly for anaerobically digested sludges (McEvoy and Giger 1986; De Henau etal. 1989; Holt et al. 1989; Marcomini et al. 1989). Bevia et al. (1989) reported LAS between2% and 4% of the sludge solids weight. In a study of 29 Swiss treatment plants, LASconcentrations 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 0.1g kg-1 dry sludge (Brunner et al. 1988).


185

Effects of surfactants on wastewater treatment

As stated previously, given these surfactant concentrations and the considerable effect thatsurfactants can have on the properties of suspensions such as sludges, the performance ofsuch processes as thickening, conditioning, and dewatering may be strongly influenced bythese materials. For example, Bierck and Dick (1988) have shown that surface tension ofsludge solids is directly related to the capillary pressure available for solids compressionduring 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 thecompressive 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 evenbefore 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 volumecaused significant increases in the CST (poorer dewaterability) which could not becompensated for even by doubling the addition of cationic polymer used as the sludgeconditioner.

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 aWWTS such as Wilmington, and sludge dewatering and disposal represent up to 50% of thetotal cost of wastewater treatment (Evans, 1988).

The biodegradation mechanism of LAS was described by Balson and Felix(1995). Themechanism of breakdown of LAS involves the degradation of the straight alkyl chain, thesulphonate group and finally the benzene ring. The breakdown of the alkyl chain starts withthe oxidation of the terminal methyl group (w-oxidation) through the alcohol, aldehyde to thecarboxylic acid as follows (see Fig. d.3a). The reactions are enzyme catalysed by alkanemonooxygenase 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 isat this stage that problems arise with branched alkyl chains, a side chain methyl group or agem-dimethyl-branched chain cannot undergo b-oxidation by microorganisms and must bedegraded by loss of one carbon atom at a time (a-oxidation, Figure d.3b). (Scott and Jones,2000).


186

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 thealkyl and the sulphonate group from LAS leaves either phenylacetic or benzoic acids.Microbial oxidation of phenylacetic acid can result in fumaric and acetoacetic acids andbenzene can be converted to catechol .

Figure d.3b a- Oxidation of LAS ( after Scott and Jones, 2000).


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Effects of the surfactants on the wider environment:

The presence of surfactants in sewage sludge may have undesirable environmental effectsif land application is the chosen disposal method. The surfactant molecules may leach togroundwater contributing to groundwater contamination. Federle and Pastwa (1988) studiedthe percolation of anionic and nonionic surfactants through a soil column. Most of thesurfactant was mineralised, but this process was found to be highly dependent on thenumber of organisms present in the soil. A number of reports (e.g. Bevia et al., 1989; Holt etal., 1989) attribute observed decreases of LAS concentrations over time in sludge-amendedsoils 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 longtime periods. Table d.2 contains data on fate and persistence of surfactants in sludgeamended soils.

Table d.2 Fate and persistence of surfactants in sludge amended soils

ApplicationForm

Country Surfactant/derivative

SoilConcentrationpostapplication(mg.kg-1)

Monitoringperiod

FinalSoilConc.(mg.kg-1)

Half Life(days)

Author

Sludge ontosoil

SP LAS 22.4 6 months12 months

3.10.7

Notreported

Prats et al

Sludge ontosoil

CH LASNP

454.7

12 months 50.5

9 Marcomini et al

Surfactantonto soil

D LAS Not reported 2 months6 months

Notreported

5-25summer66-117winter

Litz et al

Sludge ontosoil

D LAS 1627

76 days106 days

0.190.44

1326

Figge andSchoberl

Surfactantonto soil

USA LASLAE

0.050.05

40 days Notreported

1.1-3.6

Knaebel et al

Sludge ontosoil

SP LAS 1653

90 days170 days

0.3Notreported

2633

Berna et al

Sludge ontosoil

UK LAS 2.6-66.4 (*) 5-6 months <1 7-22 Water et alHolt et al

Sludge ontosoil

UK LAB 0.3-9.5 (*) 55 days 0-0.38 15 Holt andBernstein

Compostedwool scoursludge

AUS NPE 14 000 14 weeks 1200 Notreported

Jones andWestmoreland

(* estimated cumulative load)

Not only may surfactants migrate to groundwater, but they may also carry hydrophobicorganic pollutants with them. The degree of partitioning of hydrophobic organic pollutants toparticles depends on the hydrophobicity of the pollutant and the amount of organic mattercontained in the particle. Dissolved organic matter tends to decrease the potential forsorption 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 treatmentplant would be expected to enhance the partitioning of organic pollutants to sludge. Whenapplied to land, desorption of surfactant could lead to pollutants also being released. Kileand Chiou (1989) studied the effect of anionic, cationic and nonionic surfactants on thewater solubility of DDT and trichlorobenzene. The results were extremely surprising. Aswould be expected, the solubility was enhanced when the surfactant was present at


188

concentrations greater than the critical micelle concentration. There was also a solubilityenhancement at surfactant concentrations less than the critical micelle concentration.

In addition to the effects of surfactants in sludge on pollution of groundwater, the surfactantsmay effect soil texture and water retention through processes similar to those discussed withrespect to sludge dewatering. Holtzclaw and Sposito (1978) determined LAS content in asludge amended soil to be high enough (1% of the fulvic acid fraction) that soil fertility couldbe 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). Giventhat landfills function in a similar manner to anaerobic digesters, the persistence of LAS isevidently due to its poor degradability in such environments. The role of surfactants inmobilizing 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 plantsshow 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 yolkprotein which is formed under the influence of oestradiol and therefore is typically producedby females. Hermaphrodite fish species have been found as well. The decompositionproducts of APE, are considered as a potential cause, since their decomposition productsformed 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 theyare degraded - but not totally - by microorganisms. In a first rapid step the ethoxylate groupsare split off by hydrolysis, and the metabolites nonylphenol (NP), nonylphenol ethoxylate(NP1EO) and nonylphenol diethoxylate (NP2EO) are formed. These metabolites are moretoxic than the original substances. Due to the hydrophobic properties of the aromatic groupthe second step of biodegradation occurs much slower. The interim products can also bebiodegraded to alkylphenoxy ethoxylate carboxylic acids (APEC). The second, slower,step of biodegradation, does not always occur, and the fact that the metabolites are morelipophilic than the parent compounds can cause an accumulation of interim products insludge and sediment. Nonylphenol, for example, was determined in digested sludge inconcentrations 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 etal. 1988). Before prohibition of APE in washing agents NP, NP1EO and NP2EOconcentrations 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 between1 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 theUSA. In highly polluted streams average nonylphenol concentrations are determined to be inthe range 0.3 to 3 µg.l-1 (Ahel et al. 1994a), polyphenoxy carboxylic acids products arepredominant, whereas in sediment NP was the dominating degradation product. Due to theirhigh octanol/water partition coefficient (log 4.0-4.6) nonylphenol, NP1EO and NP2EO showa 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 itamounts to 1300.


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Table d.3: Environmental concentrations of degradation products of nonionicsurfactantsEnvironmentalcompartment

Substance Concentration Reference

Sewage sludge NP 0,45 - 2,53 g kg-1*0,03 g kg-1*

Giger et al., 1984Giger and Alder, 1995

WWTS-drain NP, NP1EO, NP2EONP 1NP1EO, NP2EO

36 - 202 µg l-1

10 µg l-1

1 - 40 µg l-1

Stephanou and Giger,1990

Streams NPNP1EO, NP2EONP1EC, NP2EC

0,3 - 45 (2-3) µg l-1

< 3 - 69 µg l-1

< 2 - 71 µg l-1

Ahel et al., 1994b

Stream sediment NP 0,5 - 13 mg kg-1* Ahel et al., 1994bFish NP, NP1EO, NP2EO 0,03 - 7,0 mg kg-1* Ahel et al., 1993Algae NP, NP1EO, NP2EO 80 mg kg-1* Ahel et al., 1993Waterfowl (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 waterand solids (sludges, sediments, and soils), as compared to the original distribution indetergent 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 toDaphnia, because a lower average molecular weight of LAS is less toxic. The riskassessment of LAS to terrestrial plants and animals reported by Mieure et al. (1990) alsoconcludes that there are adequate margins of safety in the use of wastewater for theirrigation of plant species. Adverse effects on plant and animal species (earthworms Eisenafoetids and Lumbricus terrestris) were observed at LAS concentration of 10 mg.l -1 , howeverLAS concentrations in wastewater effluents are in a range 0.09 mg l-1 to 0.9 mg l-1 . Thesefigures give a safety margin in a range 10 to 100. The effect of surfactant on plant growthfrom the use of sewage sludge is difficult to assess because in general the sludge promotesplant growth. Adverse effects on plant growth were observed at 392 µg g-1 but long termmonitoring 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 adverseeffects was 235 µg.g-1 giving a safety margin of 78. However, in looking at ecotoxicity fromWWTS effluents the less toxic surfactant residues and surfactant catabolites must beconsidered 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. Dermaltoxicity, 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. Nonionicsurfactants are more toxic for aquatic organisms than for mammals. The toxicity of APEincreases with decreasing number of ethoxylate units and increasing hydrophobic chainlength. Accordingly, the toxicity of the original substances is lower than the toxicity of themetabolites NP, NP1EO and NP2EO, whereas the carboxylic acids are less toxic than theethoxylates. For instance the LC50 (48 hours) of NP16EO is 110 mg.l-1 for fish (Oryziaslatipes) and decreases to 11,2 and 1,4 mg.l-1 for NP9EO and NP, respectively (Yoshima,1986). The LD50 (96 hours) for algae (Skeletonema costatum) is 27 µg.l-1, and the value forrainbow trouts 480 µg.l-1 (Nayler 1992). The no observed effect concentration (NOEC) forreproduction for Daphnia is in the range of 24 µg.l-1. These data show that the acute toxicityof NP is considerably high.

In vitro toxicity studies with fish hepatocytes indicate that several decomposition products ofAPE cause weak oestrogenic effects (Jobling and Sumpter 1993, White et al. 1994). Studies


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based on the vitellogenin synthesis revealed that NP, NP1EO and NP1EC have the sameactivity (half maximum activity: around 16 µM). The oestrogenic activity, however, is 10times lower than that of oestradiol (Pelissero et al. 1993). Other in vitro studies give hints onpotential differences between fish and mammalia regarding the binding to the oestrogenreceptor (Thomas and Smith 1993). However, vitellogenin synthesis in fish hepatocytes isalso induced by well-known phyto-oestrogens. Studies in the UK indicate that downstream ofthe drain channels of WWTS vitellogenin is formed in male fish. After 1 to 3 weeks exposureof 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 ofthe in vitro studies with fish hepatocytes. However, it cannot be excluded that syntheticoestrogens are also responsible for this effect. On the one hand their concentrations arelower than the usual concentrations of NP, but on the other hand their activity is some ordersof magnitude higher. Experimental exposure of fish to NP or metoxychlor over 7 daysinduced vitellogenin synthesis in male fish (Nimrod and Benson 1995). The dose required toinduce the vitellogenin synthesis was 300 times (approx. 150 mg.kg-1) higher than thenecessary dose of oestradiol.

Further research in this field, especially the conduction of experimental in vivo studies, isurgently required to allow for a more reliable assessment of the exposure of fish populationsto oestrogenic chemicals and their potential effects.

Field investigations indicate that downstream of the drain channels of WWTS oestrogeniceffects may be induced in fish. The in vitro studies with fish hepatocytes seem to indicatethat the oestrogenic activity of synthetic oestrogens is some orders of magnitude higher thanthe activity of decomposition products of APE. On the other hand the oestrogenic potency ofNP, NP1EO, NP2EO and NP1EC is very similar. Consequently, all degradation productshave to be taken into consideration. It seems advisable to suppose that the above chemicalshave additive effects.

Best environmental practice examplesOne 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, laundryand cleaning detergents in Scandinavia, Germany, Austria and other European countries.Products with the ‘Swan’ and ‘Good Environmental Choice’ label do not contain LAS andNonylphenol and have gained considerable market share. In Sweden, products with theselabels 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). Moreresearch is needed for the potential environmental effects of the alternatives used in theseLAS-free and NPE-free surfactants.

This public awareness and consumer choice, lead to the use of LAS in Sweden falling from6300 tonnes a year to 260 tonnes a year. The Danish Environment Agency launched apublic campaign against LAS in September 1999. Currently about 2500 tonnes a year ofLAS are used in Denmark.

During the development of the “Swan” mark, Stockholm water company identified the needto have an alternative for taking care of hazardous waste in the household rather thanflushing it into the UWW collecting system. Environmental stations, or collection points wereestablished and an extensive public information campaign was carried out about the impactsof 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


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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 influenceconsumer choice and reduce contaminant discharges in the UWW from domestic products.

More research is necessary to experimentally determine the role of surfactants in sludgetreatment processes and following sludge disposal in the environment. Specific effects to beinvestigated 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 understandingof 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 inamphibian and reptiles. In this field nearly no knowledge is available.

• Chemical methods for the detection of traces of synthetic oestrogens and theirmetabolites must be elaborated, since only very few data are available onenvironmental concentrations, especially regarding concentrations in drain channelsof wastewater treatments plants. Furthermore, data material on NP concentrations indrinking water and organisms including humans is insufficient.

• The ecotoxicological relevance of vitellogenin production in male animals has to beelucidated. Which interrelations exist between the problem of vitellogenin productionand further estrogenic and ecotoxicological effects of NP and other chemicals? Toanswer 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 thesechemicals. In a further step the problem should be investigated by morecomprehensive field studies.

• The mechanisms of chronic effects of alkylphenols (modes of action) must be studiedin more detail.

• Finally, in vitro assays should be elaborated to identify and estimate the oestrogenicactivity of existing new chemicals in fish and other organisms.


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(G) Use of Polyelectrolytes; The Acrylamide Monomer in Waste Water Treatment

Polyacrylamide (PAM) is a widely used flocculant in water treatment applications. Some20,000 tonnes is used in the USA for this purpose each year. Concerns with its use are thatit can degrade to the acrylamide monomer which is known to affect the central andperipheral nervous systems and is also believed to be carcinogenic. Safe levels for thischemical 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) andin pulp and paper manufacture. The EPA also notes that it is used in formulating grouting fortunnels and sewers. Effluents from a sewage works which used PAM as a flocculant in theUK were reported to be 2.3 to 17.4 µg l-1. High levels of the monomers have been reportedin 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 inwater and the presence of PAM in soil causes leaching of microorganisms by ground orirrigation 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 filmand 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 pumpsand pipes (Rho et al. 1996). Once the polymer degrades the monomer is also subjected tomore rapid degradation in which it is decomposed to acrylic acid and ammonia. These arenon toxic as acrylic acid degrades to CO2 and water in a day in soil (Staples et al. 2000) andis thus not an ecological problem. The degradation of acrylamide under favourableconditions by pseudomonas species immobilised in calcium aliginate took one day (Nawazet al . 1993). The EPA say that degradation of acrylamide in river water takes 4 to 12 daysand is more rapid in summer than winter. Due to its solubility adhesion of the monomer onsoil is unlikely, though it is reported that it is partially removed by secondary activatedsludge.

The general conclusion of the EPA paper is that the monomer is not an environmentalhazard when released in small quantities to the aqueous environment. Tests with fish wouldindicate this. The monomer is relatively biodegradable within days compared to the timespan of years for substances such as PCB. There are two areas where there could be someconcern and these are control of effluents from acrylamide manufacturing and use indrinking water treatment. The fact that the monomer is detected in water treatment plantswhere the residence time is only a few hours suggests that the PAM flocculant could havesignificant residual amounts of monomer.

There is a web site (www.fwr.org/waterg/dwi0084.htm) which quantifies some of the pointsmentioned. Among the points noted are:

• Chlorination and the presence of potentially toxic elements can stop acrylamidedegradation 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 flocculationhas eased the likelihood of serious contamination of water. The present level of monomerpresent 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 thisparticular 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 touse PAM as a coagulant. Problems may arise when it is used in a stream which is


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subsequently sterilised by chlorine or by another disinfectant. This removes the bacteria thatare needed to degrade the PAM and it would be a problem with using other organicpolymers as well. It might be concluded that a polymer grade containing just 0.3% ofmonomer is very pure and improvements in purity might not be practically feasible.

Therefore in treating drinking water some intermediate step may be necessary to degradethe monomer after flocculation and before adding the disinfection agent. This could involve aholding lagoon or treatment using activated charcoal. In any case the case for or against theuse of PAM in flocculating drinking water lies in the conflicting aims of acrylamidedegradation and product sterilization. Its use in the treatment of drinking water needscontinuous monitoring and, if more stringent regulations are placed on the monomerconcentration in drinking water the issue will become a concern.


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(h) Landfill Leachate

Introduction

In the past treatment of leachates in WWTS were favoured but due to the effects of dilutionin the UWW system, there is little, if any information on the elimination of persistentcompounds (Alberts, 1991). In Germany, a recent requirement has been the properpreliminary treatment of leachate before discharge either directly to surface waters orindirectly to municipal WWTS.

With the introduction of redrafted legal conditions in Germany, strict demands have beenplaced on the purification performances of leachate treatment plants. Thus, the treatment oforganic substances and nitrogen compounds using nitrification and denitrification, is requiredprior 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 ofwater quality in the receiving water.

• Indirect discharge: Discharging wastewater directly into a public WWTS requires thatthe concentration of COD, BOD, NH4-N, AOX and potentially toxic elements must bereduced to the same levels found in domestic wastewater.

Standard values for the composition and quality of non-domestic wastewater discharged to apublic WWTS are stated in guideline/directive ATV-A 115 (worksheet for indirectdischargers). The standard concentrations for potentially toxic elements are shown in Tableh.1.

Table h.1: Standard concentrations for soluble and insoluble inorganic substances inwastewater from non-domestic sources [ATV-A 115, 1994].

Potentially toxicelement

Symbol StandardConcentrations

[mg/l]Lead Pb 1

Cadmium Cd 0.5Chromium Cr 1

Chromium (VI) Cr(VI) 0.2Copper Cu 1Nickel Ni 1

Mercury Hg 0.1Zinc Zn 5

Regulation AbwV: “Requirements for discharging wastewater into watercourses”

The Wastewater Regulation (AbwV) places general requirements on the introduction ofwastewater into receiving watercourses. A permit for discharging wastewater into awatercourse can only be granted, when the limit values for the pollution load at the point ofdischarge are observed. Dilution of wastewaters in order to reach the required concentrationvalues is not permitted.

Requirements for wastewater from landfill sites are given special attention in annex 51 of theregulation. It is a requirement that the quantity and pollution load of landfill leachate must be


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kept low by proper measures and operation at the landfill installation. The requirementslisted 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 sampleor 2 hour mixed sample) [AbwV, annex 51, 1999].

Parameter Unit ValueCOD* mg/l 200BOD mg/l 20

Ntotal** mg/l 70Ptotal mg/l 3

Hydrocarbons, total*** mg/l 10N02-N mg/l 2

Fish toxicity GF 2* For wastewater with a COD value (before treatment) of more than 4000 mg/l, the CODeffluent 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). Therequirement applies to a wastewater temperature of 12 °C. A higher limit concentration of 100 mg/l ispermitted 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 otherwastewaters.

Table h.3: Requirements on leachate before mixing (qualified sample or 2 hours mixedsample) [AbwV, annex 51, 1999

Parameter Unit[mg/l]

AOX* 0.5Mercury 0.05

Cadmium 0.1Chromium 0.5

Chromium VI* 0.1Nickel 1Lead 0.5

Copper 0.5Zinc 2

Arsenic 0.1Cyanide, easily released* 0.2

Sulphide* 1* value for the qualified sample.


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Leachate can be mixed with other wastewater for common biological treatment only when:

• fish, indicator bacteria and Daphnia toxicity of a representative sample is notexceeded (see Table h.3). It has to be stated that exceeding the GF value is notcaused 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 = 2Daphnia toxicity GD = 4Indicator bacteria toxicity GL = 4

Landfill Leachate

Formation:A substantial proportion of pollutant emissions from landfill sites enters percolating throughthe landfill. Rain water (and other sources of water) entering unsealed sections of the landfillundergo 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 loadedwater, 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 lifeof the landfill leachate quantity reduces to 10 – 20 % of total rainfall, with an increase insuperficial sealing.

Composition:Leachate is a heterogenezous mixture, often containing a high concentration of persistentbiological and toxic compounds. The type of material deposited in the landfill determines thecomposition of the leachate. Leachate composition and pollutant concentration are alsoinfluenced by the rate of biochemical processes in the body of the landfill. After an initialintensive phase biological and chemical reactions in the landfill slowly subside. As the age ofa landfill increases, the quota of easily degradable compounds in and the COD/BOD5 ratiorises (Leonhard, Wilderer, 1992).

Some of the main pollutants found in landfill leachates are organic compounds, such as alkylphenols, chlorinated phenols, polycyclic aromatic hydrocarbons (PAH), dioxins and furans.Leachate from special waste landfills tends to have higher concentrations of inorganicsubstances compared to leachate from household refuse landfills; chlorides, sulphates andfluorides represent the main load.

Table h.5 gives an overview of the relevant pollutant concentration in landfill leachate.


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Table h.5: Mean composition of leachate from: industrial or special waste landfills,and household refuse landfills, in Germany [Ehrig et al., 1988]

Industrial and special wastelandfill

Household refuse landfillParameter Units

Range Mean value Range Mean valuepH - 5.9 – 11.6 7.7 3.5 – 9 7.5COD mg O2.l

-1 50 – 35000 5746 500 - 60000 5000BOD5 mg O2.l

-1 41 – 15000 2754 100 - 45000 1500Conductivity mS.cm-1 2110 – 183000 28217 - 10000Chloride * mg.l-1 36 – 126300 13257 100 - 15000 2000Sulphate mg.l-1 18 – 14968 2458 50 - 3000 300Ammonium* mg.l-1 5 – 6036 921 20 - 3000 500Nitrite* mg.l-1 0.02 - 131 7.3 - 0.5Nitrate* mg.l-1 0.1 - 14775 606 0 - 50 3Total-N* mg.l-1 1 - 3892 461 20 - 4000 600Total-P* mg.l-1 0.03 - 52 7.9 0.01 - 10 1Fluoride mg.l-1 0.1 - 50 13.3 - -Total cyanide mg.l-1 0.007 - 15 1.3 - -Easily releasedcyanide

mg.l-1 0.008 - 1 0.2 - -

Arsenic* mg.l-1 2 - 240 51 0.1 - 1000 20Lead* mg.l-1 4.3 - 650 155 20 - 1000 50Cadmium mg.l-1 0.2 - 2000 144 1 - 100 5Copper* mg.l-1 1.3 - 8000 517 10 - 1000 50Nickel* mg.l-1 14.2 - 30000 2096 20 - 2000 200Mercury mg.l-1 0.17 - 50 5.5 - 10Zinc mg.l-1 20 - 272442 2936 100 - 10000 1000Chromium (total)* mg.l-1 0.009 - 300 18.1 0.02 - 15 0.2Iron mg.l-1 0.38 - 2700 144 1 - 1000 50Phenol index mg.l-1 0.01 - 350 26 - 0.006Hydrocarbons 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 PracticesDifferent methods can be used for treating leachate from landfills, consisting principally ofbiological, physical and chemical processes. A specific process can only treat a particularsubstance categories in wastewater. Because of the wide range of pollutants found,leachate treatment has to be performed using a combination of suitable processes. Thechoice of treatment processes depends closely on the leachate composition. A shortdescription 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. Toenable degradation specialised microorganisms must be enriched in the bioreactors byproper process conditions. Nitrogen elimination can also be obtained by nitrification anddenitrification. Biological processes, especially the aerobic ones, are efficient and cost-effective in comparison with the chemical-physical processes (Rudolph et al., 1988). Theactivated sludge process and the biofilm process are both use to biologically treat leachatefrom landfills.

Activated sludge : In the activated sludge process micro-organisms aggregate in the formof biological sludge flocs suspended in the wastewater flow, through the treatment plant. Theformation of settleable sludge is decisive for the efficient working of the activated sludgeprocess. Leachates though are often characterized by high salt concentrations and high


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concentrations of persistent organic compounds, forming a fine, dispersed sludge, whichdoes not settle readily. So the biomass passes through the activated sludge plant withouttreatment. 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, whichmay be experienced in the activated sludge process. Biomass growth is encouraged byattachment to support surfaces, in form of biofilm. SBBR, the so called sequencing batchbiofilm reactors, are also used for cleaning leachate with high salt concentrations and a highpercentage of persistent organic compounds. Advantages of the biofilm processes are thesmall 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-physicalprocesses are chemical oxidation and membrane filters.

Flotation: Flotation is used for separating specific low density substances and suspendedsolid constituents or liquid substances. In a leachate treatment plant they are normally thefirst step of the treatment process.

Precipitation, flocculation and sedimentation: In leachate treatment iron and aluminumsalts are usually used to achieve precipitation and flocculation, which is then followed bysedimentation of the settleable material. Using this process, potentially toxic elements in theform of hydroxides and disperse organic substances, are separated with a removalefficiency of 40 %.

Adsorption: At a leachate treatment plant, adsorption by activated carbon is always used incombination with biological pretreatment or with a chemical-physical process. Any persistentorganic compounds not degraded in the pretreatment step and AOX compounds, can beseparated in the back-washed carbon filters. Through adsorption processes, anagglomeration of the solute molecules takes place on the activated carbon interface.Advantages of the adsorption process are; simplicity of the technology involved; relativelylow 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-biodegradableand inhibitory organic substances can be oxidised or reduced. In ideal conditions, given asufficient supply of medium for oxidation, complete mineralisation can be achieved.Substances such as potentially toxic elements and neutral salts remain in solution and arenot transformed (Döller, 1998). Hydrogen peroxide/UV or ozone/UV are the mainmediumused for oxidation medium in leachate treatment. In practice, leachate is enriched withozone (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 differentvolatilities of water, organic solvent and of dissolved and suspended substances, partition bydistillation can be achieved. Thus volatile hydrocarbons contained in the leachate can beseparated with a stripping step. With the vapourising process, inorganic and organic residualsubstances are obtained separately in a chemically unchanged form. Then, the concentratedorganic phases must be made inert by combustion. Proper treatment of the exhaust gasesis necessary to meet air quality emission standards. During the vapourisation of criticallyloaded leachate, single toxic halogen organic compounds such as polychlorinated biphenyls,dibenzo-dioxins and dibenzo-furans can enter the distillate. In this case, post treatment withactivated carbon is essential (Leonhard, Wilderer,. 1992).


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Reverse Osmosis: In the treatment of leachate, reverse osmosis is only used fordesalination and concentration of the leachate to be treated. During operation, membranefouling caused by suspended and colloidal substances has to be prevented, which wouldotherwise result in a regression of the treatment performance, due to reduction of thepermeate flow. At the end the accumulated concentrate must be subjected to additionaltreatment. The principle advantage of reverse osmosis is the low energy cost.

Membrane filtration: The membrane technique has been successfully used for cleaningleachates. A biological process tank is combined with post membrane filtration (nano andultrafiltration) for biomass retention. The activated sludge tank is in part operated byoverpressure in order to reach higher oxygen solubility and with this, a better oxygen supplyfor the micro-organisms. The removal of treated water occurs continually over a cross-flowmembrane filtration plant. The membrane modules are especially capable of finely dispersedsludge retention (Krauth,.1994).

Conclusions

In Germany, the discharge of wastewater into public WWTS and into receiving watercoursesis strictly regulated. Thus, leachate must also be treated before discharging and legalregulations 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 ofCOD, BOD, NH4-N, AOX and potentially toxic elements must be reduced to at leastdomestic wastewater standards. Purified leachate contributes only a relatively smallproportion of the pollutant load WWTS.

In ordinary analysis of treated and untreated leachate, only parameters such as COD, BODand AOX are determined. Additional quantification of high and low volatile hydrocarbons,organic acids, phenols and single organic halogen compounds is necessary to adequatelydescribe the potentially hazardous impact of leachate.

In addition, it has to be taken into account that waste products loaded with pollutantsfrequently arise from leachate treatment: toxic surplus sludge results from biologicaltreatment; charcoal is produced during adsorption processes; and polluted concentratesform during vaporisation. To minimize the problematic emission of pollutants into theenvironment, additional treatment of exhaust gases and proper disposal of the wasteresidues are required.


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(i) Potentially Toxic Elements (PTE) transfers to sewage sludge

Sludges from conventional sewage treatment plants are derived from primary, secondaryand tertiary treatment processes. The polluting load in the raw waste water is transferred tothe sludge as settled solids at the primary stage and as settled biological sludge at thesecondary stage. Potentially toxic elements are also removed with the solids during theprimary and secondary sedimentation stages of conventional wastewater treatment. Metalremoval during primary sedimentation is a physical process, dependent on the settlement ofprecipitated, insoluble metal or the association of metals with settleable particulate matter.Minimal removal of dissolved metals occurs at this stage and the proportion of dissolvedmetal to total metal in the effluent increases as a result. The efficiency of suspended solidsremoval is the main process influencing the extent of metal removal during primarywastewater treatment. However, the relative solubilities of different elements present in thewastewater are also important (Table i.1). Thus, Ni shows the poorest removal (24 %) duringprimary treatment whereas 40 % of the Cd and Cr in raw influent is transferred to the primarysludge. Primary treatment typically removes more than 50 % of the Zn, Pb and Cu present inraw sewage.

The removal of metals during secondary wastewater treatment is dependent upon theuptake of metals by the microbial biomass and the separation of the biomass duringsecondary sedimentation. Several mechanisms control metal removal during biologicalsecondary 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 aresimilar to those recorded for primary sedimentation. However, the general survey of removalefficiencies listed in Table i.1 suggests that secondary treatment (by the activated sludgeprocess) is more efficient at removing Cr than the primary stage. Operational experienceand metal removals measured by experimental pilot plant systems provide guidance on theoverall likely removal and transfers to sludge of potentially toxic elements from raw sewageduring conventional primary and secondary wastewater treatment. This shows thatapproximately 70 – 75 % of the Zn, Cu, Cd, Cr, Hg, Se, As and Mo in raw sewage isremoved and transferred to the sludge (Blake, 1979) and concentrations of these elementsin the final effluent would be expected to decrease by the same amount compared with theinfluent to the works. Lead may achieve a removal of 80 %, whereas the smallest overallreductions are obtained for Ni and approximately 40 % of this metal may be transferred tothe sludge.

The majority of potentially toxic elements in raw sewage are partitioned during wastewatertreatment into the sewage sludge or the treated effluent. However, atmospheric volatilisationof Hg as methylmercury, formed by aerobic methylation biotransformation processes, is alsosuggested as a possible mechanism contributing to the removal of this element duringsecondary 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 tobacterially mediated volatilisation, it is unlikely that this is a major route of Hg loss becauseof the significant quantities of Hg recovered in surplus activated sludge (Lester, 1981).


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Table i.1 PTE removals and transfer to sewage sludge during conventional urbanwastewater treatment (Lester, 1981)

PTE Removal (%)Primary(1) Secondary(2) Primary +

secondaryPrimary +

secondary(3)

Zn 50 56 78 70Cu 52 57 79 75Ni 24 26 44 40Cd 40 40 64 75Pb 56 60 70 80Cr 40 64 78 75Hg 55 55 80 70Se 70As 70Mo 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)


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(J) Effect of chemical phosphate removal on potentially toxic element content insludge

The chemical treatment of wastewater to remove phosphorous is increasingly practised tocontrol P discharges and as a measure to reduce eutrophication of sensitive water courses.This also has the advantage of increased BOD removal, reduction in polyelectrolytecoagulant consumption for sludge thickening, elimination of hydrogen sulphide in sludgedigesters and reduced consumption of chemicals for exhaust gas scrubbing (Abendt, 1992).High rates of P removal can be achieved from wastewater using common precipitants suchas aluminium sulphate (alum) and ferric chloride although this influences both the qualityand quantity of sludge produced (Yeoman et al., 1988). Chemical precipitation alsoenhances the removal of potentially toxic elements from sewage effluent compared withconventional treatment practices, increasing the transfer of metals to sewage sludge and thecontent of metals in sludge. For example, Stones (1977) measured the reductions in metalconcentrations in sewage effluents obtained after an 18 h settling period with aluminiumsulphate (Al2(SO4)3) compared with sedimentation without Al salt. The removal of all theelements examined was increased by the addition of aluminium sulphate compared to theunamended control, except for Ni (Table j.1). The removal of Cu and Zn from the effluentwas raised by approximately 50 % by chemical treatment compared with removals achievedby sedimentation without addition of Al. Lead removal increased by about 80 % and thelargest 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 sludgealmost by a factor of three.

Table j.1 Effect of chemical precipitation on metal removals (%) from raw sewage after18 h sedimentation (Stone, 1977)

PTE Unamendedcontrol

Al2((SO4))3

at 400 mg l-1Removal relativeto control (%)

Zn 50 73 48Cu 57 90 56Ni 19 19 0Pb 54 96 79Cr 22 63 193

Iron-based precipitants are marketed for use in wastewater treatment may be derived fromindustrial by-products of titanium oxide production. Such by-products may contain significantconcentrations of potentially toxic elements (PTEs) with potentially undesirable effects onthe metal content of sludge (Thiel, 1992).An example of the effects of Fe dosing with industrial by-product on the maximum potentialincrease in the PTE content of activated sludge is shown in Table j.2. The typical dosingrates of FeSO4 are typically in the range is 15 – 30 mg Fe salt l-1, but may increase up to 50mg Fe salt l-1, to comply with the discharge requirements for P in the Urban Waste WaterTreatment Directive (CEC, 1991). The calculations suggest that dosing with FeSO4 maypotentially 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 doserate of 50 mg Fe salt l-1 and that secondary sludge production is equivalent to 250 mg l-1 oftotal solids (UKWIR, 1997). The Ni content in activated sludge may theoretically increase by130 % compared to sludge without Fe addition, whereas Pb and Zn concentrations mayincrease by about 10 % with Fe dosing. These increases in sludge content remain wellwithin the current quality standards for agricultural use (CEC, 1986). However, the revisionof the Directive on land application (CEC, 2000b) will introduce more stringent limit valuesfor PTEs and the use of Fe salts from industrial processes could potentially penalise theacceptability of sludge for use in agriculture under the new regulatory regime. Furthermore,


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the potential increase in the metal content of sewage sludge, resulting from the use ofindustrial-grade chemical precipitants, could also be considered as unsatisfactory because iterodes the beneficial reductions in metal inputs that have been achieved through thesuccessful control of trade effluent discharges.

The quality and metal content of low-grade chemical precipitants for use in wastewatertreatment should be examined to ensure that they do not significantly increase the metalcontent of sludge. In Germany, for example, composition standards are recommended forFe 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 beconsidered for sewage treatment (Thiel, 1992) to avoid potential problems associated withcontamination with potentially toxic elements. In practice, there are few published data onthe effects of chemical precipitants on sludge metal contents and Fe and Al dosing. Oneexample from the literature (Yeoman et al., 1993) showed no consistent effects of chemicaltreatment with Al or Fe salts on potentially toxic elements in sewage sludge from BecktonWWTS in the UK (Table j.3). However, the significance of the direct metal inputs in chemicalprecipitants will increase as industrial discharges are effectively controlled and as diffuseinputs from domestic sources and run-off become the predominant sources of potentiallytoxic elements entering the wastewater collection system.

Table j.2 Metal concentrations (mg kg-1) in Fe precipitants and activated sewagesludge (UKWIR, 1997)

PTE FeSO4

saltIncrease inactivated

sludge dueto FeSO4

FeCl2 salt Increase inactivated

sludge dueto FeCl2

Activatedsludge

without Fe

Zn 348 70 26 5.2 600Cu 5 1.0 51 10 400Ni 160 32 120 24 25Cd 22 4.4 3.0 0.6 1.5Pb 64 13 22 4.4 110Cr 32 6.4 236 47 -

Table j.3 Effect of chemical P removal on the PTE content of sludge digested sewagesludge(1) (adapted from Yeoman et al., 1993)

Concentration (mg kg-1) dsSludge type Salt additionCd Cu Ni Pb Zn

Digested None 5.4 159 24 137 231Digested + Al Raw sludge 5.4 142 19 62 148Digested + Al Activated sludge 4.5 254 22 168 184Mean 4.9 198 20 115 166Digested + Fe Raw sludge 8.5 195 40 253 300Digested + Fe Activated sludge 4.8 95 18 155 121Mean 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 acidmine 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-enrichedsludge from water treatment with hydrochloric acid to convert ferric hydroxide to ferricchloride for use in sewage treatment for phosphate removal, and as a dewatering agent in


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sludge conditioning. Verberne (1992) considered that the use of water treatment sludges aschemical precipitants for P removal was technically feasible and would depend on theagreement and acceptance of the approach by water and sewage treatment authorities. Therecovery of Fe and Al from acid mine drainage is another source of chemical precipitantsthat can be used for P removal during sewage treatment (Bouchard et al., 1996). The re-useof secondary resources for precipitating P during wastewater treatment is intuitivelyattractive and also alleviates the environmental problems and impacts associated withdisposal of those wastes. However, some product types derived from waste materials arepotentially contaminated with potentially toxic elements that accumulate in the sludge(Fowlie and Shannon, 1973). Therefore, the metal content of waste derived products shouldbe established, and the potential consequences for sludge quality determined, before aparticular product is accepted for use as a chemical precipitant in wastewater treatment.

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