Sludge Pollutants

273
European Commission Pollutants in urban waste water and sewage sludge

Transcript of Sludge Pollutants

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Pollutants in urban wastewater and sewage sludge

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Executive Summary

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Authors

I C Consultants Ltd London

Professor Iain Thornton

(scientific co-ordinator)

Dr David Butler

Paul Docx

Martin Hession

Christos Makropoulos

Madeleine McMullen

Dr Mark Nieuwenhuijsen

Adrienne Pitman

Dr Radu Rautiu

Richard Sawyer

Dr Steve Smith

Dr David White

Technical University Munich

Professor Peter Wilderer

Stefania Paris

IRSA Rome

Dr Dario Marani

Dr Camilla Braguglia

ECA Barcelona

Dr Juan Palerm

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Table of Contents

EXECUTIVE SUMMARY 5

1. INTRODUCTION 9

1.1 INTRODUCTION TO POLLUTANTS IN URBAN WASTEWATER (UWW) AND

SEWAGE SLUDGE (SS)

1.2 OBJECTIVES AND GOALS

2. POTENTIALLY TOXIC ELEMENTS, SOURCES, PATHWAYS, AND FATE THROUGH

URBAN WASTEWATER TREATMENT SYSTEMS 13

2.1.SOURCES AND PATHWAYS OF POTENTIALLY TOXIC ELEMENTS IN UWW AND SS

2.1.1 DOMESTIC SOURCES

2.1.2 COMMERCIAL SOURCES

2.1.3 URBAN RUNOFF

2.2 INFLUENCE OF VARIOUS TREATMENT PROCESSES ON THE FATE OF

POTENTIALLY TOXIC ELEMENTS THROUGH WASTEWATER TREATMENT SYSTEMS

(WWTS) AND SEWAGE SLUDGE TREATMENT (SST)

2.3 QUANTITATIVE ASSESSMENT OF POTENTIALLY TOXIC ELEMENTS IN

UNTREATED UWW, TREATED UWW AND TREATED SS

3. ORGANIC POLLUTANTS: SOURCES, PATHWAYS, AND FATE THROUGH URBAN

WASTEWATER TREATMENT SYSTEMS 64

3.1. SOURCES AND PATHWAYS OF ORGANIC POLLUTANTS IN UWW AND SS

3.1.1 DOMESTIC AND COMMERCIAL

3.1.2 URBAN RUNOFF

3.2 INFLUENCE OF VARIOUS TREATMENT PROCESSES ON THE FATE OF ORGANIC

POLLUTANTS THROUGH WWTS AND SS

3.3 QUANTITATIVE ASSESSMENT OF ORGANIC POLLUTANTS IN UNTREATED UWW,

TREATED UWW AND TREATED SS

4. HEALTH AND ENVIRONMENTAL EFFECTS OF POLLUTANTS IN UWW AND SS 94

4.1 POTENTIALLY TOXIC ELEMENTS

4.2 ORGANIC POLLUTANTS

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5. A REVIEW OF EU AND NATIONAL MEASURES TO REDUCE THE POTENTIALLY

TOXIC ELEMENTS AND ORGANIC COMPOUNDS CONTAMINATION OF UWW AND SS

102

6. CASE STUDIES 113

(A) PLATINUM GROUP METALS IN URBAN ENVIRONMENT

(B) SUSTAINABLE URBAN DRAINAGE

(C) POLLUTANT SOURCES AND LOAD FROM ARTISANAL ACTIVITIES IN URBAN

WASTEWATER (THE MUNICIPALITY OF VICENZA, INCL. GOLD JEWELLERY SHOPS)

(D) PHARMACEUTICALS IN THE URBAN ENVIRONMENT

(E) PERFUME COMPOUNDS IN WASTEWATER AND SEWAGE SLUDGE

(F) SURFACTANTS IN URBAN WASTEWATERS AND SEWAGE SLUDGE

(G) USE OF POLYELECTROLYTES; THE ACRYLAMIDE MONOMER IN WATER

TREATMENT

(H) CASE STUDY: LANDFILL LEACHATE

(I) PTE (POTENTIALLY TOXIC ELEMENTS) TRANSFERS TO SEWAGE SLUDGE

(J) EFFECT OF CHEMICAL PHOSPHATE REMOVAL ON POTENTIALLY TOXIC

ELEMENT CONTENT IN SLUDGE

7. REPORT SYNOPSIS, DISCUSSIONS AND CONCLUSIONS 205

7.1 COMMENTS, CHALLENGES AND STRATEGIES FOR THE NEXT FIVE TO TEN

YEARS

7.2 IDENTIFICATION OF GAPS IN THE AVAILABLE INFORMATION,

7.3 RECOMMENDATIONS FOR FURTHER RESEARCH

7.4 SUGGESTIONS

APPENDICES 232

APPENDIX A - URBAN WASTEWATER TREATMENT SYSTEMS (WWTS) AND SEWAGE SLUDGE

TREATMENT (SST) - EU AND REGIONAL ASPECTS

APPENDIX B - PHYSICO-CHEMICAL PROPERTIES OF SELECTED POLLUTANTS

DATABASES, REFERENCES

GLOSSARY AND ABBREVIATIONS

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Executive Summary

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POLLUTANTS IN URBAN WASTE WATER AND SEWAGE SLUDGE

EXECUTIVE SUMMARY

Water policy in the European Union is aiming to promote sustainable water use and a majorobjective of the new Water Framework Directive (2000/60/EC) is the long-term progressivereduction of contaminant discharges to the aquatic environment in urban wastewater(UWW). Sewage sludge is also a product of wastewater treatment and the Urban WasteWater Treatment Directive (91/271/EEC) aims to encourage the use of sludge wheneverappropriate. Potentially toxic elements and hydrophobic organic contaminants largelytransfer to the sewage sludge during waste water treatment with potential implications for theuse of sludge although some may be emitted with the effluent water.

Inputs of metals and organic contaminants to the urban wastewater system (WWTS) occurfrom three generic sources: domestic, commercial and urban runoff. A review of availableliterature has quantified the extent and importance of these various sources and the inputsfrom different sectors. In general, urban runoff is not a major contributor of potentially toxicelements to UWW. Inputs from paved surfaces due to vehicle road abrasion and tyre andbrake-lining wear have been identified and losses from Pb painted surfaces and Pb and Znfrom roofing materials represent localised sources of these elements.

Platinum and Pd are components of vehicle catalytic converters and emissions occur as theautocatalyst deteriorates. Catalytic converters are the main source of these metals emittedto the environment and releases have increased with the expansion in use of autocatalysts.Platinum group metals (PGMs) potentially enter UWW in runoff and transfer to sewagesludge in a similar way to other potentially toxic elements. The Pt content in sludge istypically in the range 0.1 – 0.3 mg kg-1 (ds) and the background value for soil is 1 µg kg-1.PGMs are inactive and immobile in soil.

In contrast to potentially toxic elements, inputs of the main persistent organic pollutants ofconcern, including: PAHs, PCBs and PCDD/Fs, to UWW are principally from atmosphericdeposition onto paved surfaces and runoff. Combustion from traffic and commercial sourcesaccounts for the major PAH release to the environment, although inputs from foodpreparation sources also represent an important and often under-estimated contribution ofcertain PAH congeners. PCDD/Fs are released during waste incineration and also by coalcombustion. Soil acts as a long-term repository for these contaminant types andremobilisation by volatilisation from soil is an important mechanism responsible for recyclingand redistributing them in the environment. For example, the industrial use of PCBs wasphased out in Europe during the 1980s-1990s, but 90 % of the contemporary emissions ofPCBs are volatilised from soil. Since emission controls are already in place for the mainpoint sources and PAHs, PCDD/Fs or PCBs enter UWW principally from diffuse atmosphericdeposition and environmental cycling, there is probably little scope, from source control, tofurther reduce inputs and concentrations of these persistent organic substances in UWW orsewage sludge.

Being strongly hydrophobic these organic pollutants are efficiently removed during urbanwastewater treatment (WWTS) and bind to the sludge solids. However, the increasing bodyof scientific evidence has not identified a potential harmful impact of these substances on theenvironment in the context of the urban wastewater system. Therefore, on balance, theimportance of these contaminants in UWW and sewage sludge has significantly diminishedand there may be little practical or environmental benefit gained from adopting limits orcontrols for PAHs, PCBs or PCDD/Fs in UWW or sewage sludge. This is emphasised furtherby the high cost and specialist analytical requirements of quantifying these compounds insludge and effluent.

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Potentially toxic element contamination of urban wastewater and sewage sludge is usuallyattributed to discharges from major commercial premises. However, significant progress hasoccurred in eliminating these sources and this is reflected in the significant reductions inpotentially toxic element concentration in sewage sludge and surface waters reported in allEuropean countries where temporal data on sludge and water quality have been collected.However, potentially toxic element concentrations remain higher in sludge from large urbanwastewater treatment plant (WWTP) compared with small WWTP and they are also greaterin sludges from industrial catchments compared with rural locations. These patterns insludge metal content suggest that commercial sources may still contribute significantly to thetotal metal load entering UWW. Indeed, recent regional surveys of metal emissions fromcommercial premises confirm that further reductions in most elements could be achievedfrom this sector. The primary targets for source control include health establishments, smallmanufacturing industries (particularly metal and vehicle related activities) and hotel/cateringenterprises, as 30 % of medical centres and 20 % of the other types of activity could bedischarging significant amounts of potentially toxic elements in UWW. Mercury is a specificcase where compulsory use of dental amalgam separators, and substituting Hg withalternative thermoreactive materials in thermometers, may be effective in reducingdischarges of this element to the WWTS wastewater treatment system .

Faeces contribute 60 – 70 % of the load of Cd, Zn, Cu and Ni in domestic wastewater and>20 % of the input of these elements in mixed wastewater from domestic and industrialpremises. Faecal matter typically contains 250 mg Zn kg-1, 70 mg Cu kg-1, 5 mg Ni kg-1, 2 mgCd kg-1 and 10 mg Pb kg-1 (ds). The other principal sources of metals in domesticwastewater are body care products, pharmaceuticals, cleaning products and liquid wastes.Plumbing is the main source of Cu in hard water areas, contributing >50 % of the Cu loadand Pb inputs equivalent to 25 % of the total load of this element have been reported indistricts with extensive networks of Pb pipework for water conveyance. Adjusting waterhardness in order to reduce metal solubilisation from plumbing is technically feasible, but islikely to be impractical at the regional scale necessary to significantly reduce metalconcentrations in UWW and sludge and may be unpopular with consumers in hard waterareas. The gradual replacement of Pb water pipes can be achieved during building renewaland renovation programmes.

Reductions in domestic discharges of metals may be possible through increased publicawareness of appropriate liquid waste disposal practices and the provision of accessibleliquid waste disposal facilities. It may be impractical to eliminate the use of metals in bodycare products when they are an important active ingredient, but advice and labelling couldbe improved to minimise excessive use. Cadmium may be a contaminant present inphosphatic minerals and removing phosphate from detergent formulations can reduceassociated potential discharges of Cd from domestic sources.

Detergent residues (e.g.nonyl phenol, NP), surfactants (e.g. linear alkyl benzenesulphonates, LAS), plasticising agents (e.g. di-(2-ethylhexyl)phthalate, DEHP) andpolyacrylamide compounds, added to sludge to aid dewatering, are quantitatively amongstthe most abundant organic contaminants present in UWW and/or sewage sludge.Dewatering agents based on polyacrylamide may contain traces of the potentially toxicacrylamide monomer, but this is rapidly degraded and polyacrylamide itself is biologicallyinactive. Detergent residues and DEHP are primarily of domestic origin and they areeffectively degraded during aerobic wastewater treatment and are not considered torepresent a potential environmental problem from the discharge of treated effluents tosurface waters. Anaerobic digestion is the principal method employed for stabilising sewagesludge, but NP accumulates during anaerobic digestion, DEHP is not removed by thisconventional process and, although a significant amount of LAS is biodegraded, residues ofthis substance remain because of the large concentrations initially present in raw sludge.The inability to degrade detergent residues anaerobically and the large concentrationspresent in sludge and UWW have prompted ecolabelling initiatives in a number of Europeancountries to influence consumer choice away from detergents containing these surfactants to

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alternative products. This has been successful when supported by extensive publicawareness campaigning. For example, the market share for ecolabelled detergents inSweden increased to 95 % and the consumption of LAS has decreased to a similar extent.Surfactant residues and plasticisers degrade quickly when added to aerobic soils. Theoestrogenic activity of NP is however, a principal concern and measures are proposed toeliminate the discharge of this substance to UWW.

Natural and synthetic oestrogens are degraded in WWT, but trace amounts remain andrepresent the main source of oestrogenic activity in treated effluents. Further work isnecessary to link these substances to oestrogenic responses in aquatic life, but it may benecessary in future to consider the requirement for tertiary treatment processes (e.g.ozonation) to eliminate these substances from treated effluents.

A number of other groups of organic compound are identified as being potentially resistant towastewater and sewage sludge treatment and the most significant of these are brominateddiphenyl ethers (PBDEs) and chlorinated paraffins. Further research is warranted, inparticular to assess the persistence and potential environmental significance of thesecompounds. Synthetic nitro musks are used in perfumed products and traces may bepresent in UWW and sludge. Little is known about the environmental fate of thesecompounds, but effects on human health from this route seem unlikely given that the mainexposure route is through direct contact.

The degree of removal and biodegradation of pharmaceutical compounds during WWTvaries considerably, although many common analgesic drugs rapidly biodegrade. They aresoluble and transfer to sludge is only of minor concern. Significant amounts of prescribeddrugs are excreted from the body and controlling these inputs from the general populationwould be impractical. However, the disposal of unused drugs into UWW should be reviewedand alternative methods of disposal should be encouraged. The potential significance ofpharmaceuticals in the environment should be assessed in context of the major inputs andpresence arising from widespread veterinary administration of drugs to livestock and farmwaste disposal to land.

A general recommendation to protect the water and soil environment is that a hazard,biodegradability and fate assessment should be required for all new synthetic chemicals,irrespective of their purpose or end-use, to determine the potential from them to transfer toUWW or sewage sludge and the subsequent implications for the environment. Specifiedcriteria regarding toxicity and biodegradation could be set for compounds that exhibit apropensity to enter the WWTS and restrictions could be enforced regarding production anduse if these were not met. These decisions would need to balanced against the potentialbenefits to health derived from the administration of pharmaceutical drugs.

Strategies aimed at controlling pollutant discharges can only focus on those sources that canbe identified and quantified. Published mass balance calculations indicate there is a highdegree of uncertainty regarding inputs of potentially toxic elements entering the WWTS.Indeed, unidentified sources may contribute as much as 30 - 60 % of the total metal loadentering the WWTS, although more than 80 % of the Cd discharged is from identified inputs.This apparent discrepancy could be related to difficulties in measuring the highly variableinputs of metals in urban runoff and the underestimation of discharges from commercialpremises that have not been subjected to trade effluent control.

The European Commission has proposed a list of 32 priority and 11 hazardous substances(COM/2001/17) with the aim of progressively reducing emissions and discharges of thesechemicals to the environment. Current developments also suggest that Zn, Cu and LAS maybe the most limiting constituents in sludge if the proposed maximum permissibleconcentrations for these substances in soil (Zn and Cu) and sludge (LAS) are carriedthrough in the revised of Directive 86/278/EEC, but they are not listed as priority substances.Consideration should be given to designating Zn, Cu and LAS as priority substances to

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minimise their to UWW as far as is practicable and to ensure there is a consistent link andapproach to defining the environmental quality standards for sludge with those forsustainable water use and contaminant discharge reduction.

The main identified priorities for future research relating to contaminant sources, fate andbehaviour in the WWTS are:

• To reduce the uncertainty in quantifying contaminant discharges to UWW by identifyingand surveying specific sources to determine the potential for controlling inputsparticularly from small commercial sources and medical establishments;

• To establish the extent and variability of contaminant entry into UWW by catchmentinvestigations in relation to precipitation frequency and changes in sludge quality;

• To critically and independently review the fate, behaviour, degradability, toxicity andenvironmental consequences of alternative surfactant and plasticing compounds, incollaboration with the related chemical manufacturing industries, to inform decisions ofthe benefits and disadvantages of product substitution in detergent formulations andplastics manufacture;

• To determine the extent of volatilisation-deposition cycling of persistent organicpollutants in the environment, identifying the processes controlling the extent andmagnitude of diffuse inputs of these substances to UWW and to provide long-termpredictions of changes in release patterns and the consequences for UWW and sludge;

• To develop a consistent statistical and reporting protocol for national chemicalcomposition data presented in surveys of sewage sludge quality.

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1. Introduction

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1. INTRODUCTION

The primary objective of this study is to determine the sources of pollution in urbanwastewater (UWW) treated in wastewater treatment systems (WWTS). This includes thepollutants introduced into the UWW collecting system with run-off rainwater, from domesticand small commercial sources. The pollutant contents in urban wastewater and sewagesludge has been evaluated by review of the existing literature, in order that measures maybe proposed to reduce pollution at source.

1.1 Introduction to pollutants in urban wastewater

The pollutants of interest can be divided into two main groups;

• potentially toxic elements (PTEs) including cadmium (Cd), chromium (Cr III and CrVI), copper (Cu), mercury (Hg), nickel (Ni), lead (Pb) and zinc (Zn),

• organic pollutants including PAHs, PCBs, DEHP, LAS, NPE, dioxins (PCDD) andfurans (PCDF). Over 6,000 organic compounds have been detected in raw watersources most of which are due to human activities. While some of these are highlypersistent, others are easily biodegradable in WWTS.

Other pollutants of interest are the metalloids, arsenic and selenium and the metal silver.Platinum group metals (PGMs), and pharmaceuticals are covered in detail in case studies.

The sources of metal pollution in the wastewater system can be classified into three maincategories:

• Domestic,• Light industrial (connected to the WWTS) and commercial,• Urban runoff (which also encompass lithospheric and atmospheric sources).

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Figure 1.1: Sources of pollutants in wastewater [after Lester, 1987]

A summary of the various inputs, outputs and pathways followed by water and associatedcontaminants from both natural and anthropogenic sources encountered in urbanenvironments is shown in Figure 1.1. It depicts the drainage area as an open system [Ellis,1986]. A more detailed urban catchment figure is included in Appendix A.

Wastewater contains many constituents and impurities arising from diffuse and pointsources. Large point sources are easily quantifiable and result from specific activities in thearea that are connected to UWW collecting systems. The contribution from small pointsources, such as households and small businesses, is much more difficult to identify andquantify, compared to point sources which are usually regulated. UWW is also vulnerable toillegal discharges of pollutants.

Diffuse sources, such as atmospheric deposition and road runoff have also beencharacterised and this study will attempt to present an overview of the available informationin this area. Different methods have been used to estimate point sources and diffuse (non-point) sources contributions to the pollution load [Vink, 1999]. Inventories of point and diffusesources, can link observed water quality trends to changes in socio-economic activities.

Atmosphere

Lithosphere

deposition

wet & drydeposition

INDUSTRY

products

DOMESTIC

Wastes

Wastes Product wastes RUNOFF

UWW COLLECTINGSYSTEMS

COMBINED UWWCOLLECTING SYSTEMS

STORM UWWCOLLECTING SYSTEMS

WASTEWATERTREATMENT WORKS

ReceivingWater

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The type of pollutants and the magnitude of the outfall loadings are a complex function of:

• size and type of conurbation (commercial, residential, mixed)• plumbing and heating infrastructure• atmospheric quality, for example long range transport of pollutants• factors affecting deposition of pollutants such as precipitation• activity and intensity surface composition and condition• urban land use• traffic type and density• urban street cleaning• maintenance practices and stormwater controls• specific characteristics of storm events• accidental releases

A review of the sources and pathways of potentially toxic element pollutants in urbanwastewater is presented in Section 2.1 and for organic pollutants in Section 3.1

1.2 Objectives and Goals

The main goals of the study were:

• To determine the sources of potentially toxic elements and organic pollutants indomestic, commercial, and urban run-off wastewater, which end up in the UWWcollecting system.

• To make a qualitative and quantitative assessment of the pollutants in urbanwastewater and runoff rainwater on the basis of the available data in the literature.

• To evaluate the percentage of inorganic and organic pollutants concentrated insewage sludge and the percentage of pollutants released in the environment with thetreated effluents.

• To review wastewater and sewage sludge treatment processes and possiblemeasures to prevent pollution at source. The most important practices to treatwastewater and sewage sludge in Europe will be closely examined.

• Based on an overall assessment of the existing data from various sources, to identifyfurther research directions in those areas with insufficient data.

This report presents a thorough literature review and is primarily based on the analysis andpresentation of case studies from a wide variety of sources and test catchments acrossEurope, covering a time range from 1975 to date. Databases used during this project arelisted in the reference section. As theoretical approaches, such as modelling of pollutantsources and predicted concentrations, are scarce the report attempts to summarise themonitoring, sampling and measurement of numerous studies, thus providing a conciseoverview of pollution source types and concentration ranges. The reader must keep in mindthat there are significant differences between the experiments (in duration, location,measurement methods, measurement targets and initial conditions), and thus conclusionson mean or extreme values of pollutants will have to be drawn carefully.

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2. Potentially Toxic Elements

Pollutants in Urban Waste Water and Sewage Sludge 12

2. POTENTIALLY TOXIC ELEMENTS: SOURCES, PATHWAYS, AND FATETHROUGH URBAN WASTEWATER TREATMENT SYSTEMS

The aim is to reduce inputs of pollutants entering the wastewater system to backgroundlevels because this represents the minimum potential extent of contamination that can beachieved. Potentially toxic elements are of concern because of their potential for long-termaccumulation in soils and sediments.

The majority of metals transfer to sewage sludge (see Fig 2.1). However, 20% may be lost inthe treated effluent, depending on the solubility and this may be as high as 40% - 60% forthe most soluble metal, Ni. Although the use of sludge on agricultural land is largely dictatedby nutrient content (nitrogen and phosphorus), the accumulation of potentially toxic elementsin sewage sludge is an important aspect of sludge quality, which should be considered interms of the long-term sustainable use sludge on land. Application of sludge to agriculturalland is the largest outlet for its beneficial use and this is consistent with EC policy of wasterecycling, recovery and use. This is a critical issue due to the increasing amount of sludgeproduced, the increasingly stringent controls on landfilling, the public opposition toincineration (a potential source of further atmospheric pollution), and the ban on disposal atsea. Consequently sludge quality must be protected and improved in order to secure theagricultural outlet as the most cost effective and sustainable option.

Figure 2.1: Origin and fate of metals during treatment of wastewater [from ADEME,1995]

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2.1. Sources and pathways of potentially toxic elements in UWW

The average concentrations of potentially toxic elements in domestic and commercialwastewater are given in Table 2.1. The maximum concentrations of potentially toxicelements found in commercial wastewater are generally greater than those in domesticwastewater. This is supported by Scandinavian studies [SFT-1997a, 1997b, 1999]considering all urban sectors together, which judged that commercial and light industrialsectors contributed larger loads of potentially toxic elements to urban wastewater thanhousehold sources.

Table 2.1 Concentrations of metals in domestic and commercial wastewater[Wilderer and Kolb, 1997 in Munich, Germany]

Element DomesticWastewater [mg.l-1]

CommercialWastewater [mg.l-1]

Pb 0.1 ≤ 13Cu 0.2 0.04-26Zn 0.1-1.0 0.03-133Cd <0.03 0.003-1.3Cr 0.03 ≤20Ni 0.04 ≤7.3

Table 2.2 Potentially toxic elements in UWW from various sources(% of the total measured in the UWW)

Pollutant Country DomesticWastewater

CommercialWastewater

UrbanRunoff

NotIdentified

Reference

France 20 61 3 16 ADEME, 1995

Norway 40 SFT report 97/28Cd

UK 30 29 41 WRc, 1994

France 62 3 6 29 ADEME, 1995

Norway 30 SFT report 97/28Cu

UK 75 21 4 WRc, 1994

France 2 35 2 61 ADEME, 1995

Norway 20 SFT report 97/28Cr

UK 18 60 22 WRc, 1994

Hg France 4 58 1 37 ADEME, 1995

France 26 2 29 43 ADEME, 1995

Norway 80 SFT report 97/28Pb

UK 43 24 33 WRc, 1994

France 17 27 9 47 ADEME, 1995

Norway 10 SFT report 97/28Ni

UK 50 34 16 WRc, 1994

France 28 5 10 57 ADEME, 1995

Norway 50 SFT report 97/28Zn

UK 49 35 16 WRc, 1994

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Cd distribution

DomesticStorm eventsCommercialNon Identified

Cu distribution

DomesticStorm eventsCommercialNon Identified

Cr distribution

DomesticStorm eventsCommercialNon Identified

Hg distribution

DomesticStorm eventsCommercialNon Identified

Pb distribution

DomesticStorm eventsCommercialNon Identified

Ni distribution

DomesticStorm eventsCommercialNon Identified

Zn distribution

DomesticStorm eventsCommercialNon Identified

Figure 2.2 Pie charts showing the breakdown of potentially toxic elements enteringUWW from different sources in France (ADEME 1995) This uses the French data in Table2.2 but is included to give a clearer visual representation of the source breakdown for thedifferent metals.

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The data in Table 2.2 and Figure 2.2 show that for some elements over 50% of thepotentially toxic elements in wastewater are unaccounted for. This is in line with findings byCritchley & Agy [1994] Better source inventory data is essential in order to effectively targetreductions in emissions from all the different sources. It may be that identification of some ofthe industrial sources will require increased trade effluent discharge controls ifconcentrations of pollutants are to be reduced. Domestic and urban run-off sources mayrequire different types of action, such as changes in products used.

Emissions of potentially toxic elements from industrial point sources were the major sourcesof pollution to urban wastewater. However, stringent and more widespread limits applied toindustrial users has reduced the levels of potentially toxic elements emitted by industry intourban wastewater considerably. This continues a general decline of potentially toxicelements from industrial sources since the 1960s, due to factors such as cleaner industrialprocesses, trade effluent controls and heavy industry recession. For example, the liquidsused in metal finishing typically contain 3-5 mg.l-1 of copper, 5-10 mg.l-1 of chromium, 3-5mg.l-1 of zinc, 5-10 mg.l-1 of zinc, 1-5 mg.l-1 of cyanide, and 10-50 mg.l-1 of suspended solids[Barnes, 1987]. However, metal finishing industries are now required to pre-treat theseliquids before disposal, reducing toxic discharges by 80-90%.

In the Netherlands, a survey of potentially toxic element load in UWW influent [SPEED,1993], also made estimations for 1995 and forecasts up to 2010. The overall prevalence ofpotentially toxic elements in the UWW system is expected to decrease, mainly due to adecrease in runoff and industrial sources, while the potentially toxic elements share inWWTS loads from households was expected to increase. As industrial sources of potentiallytoxic elements in UWW decline, the relevant importance of diffuse sources will increase.

Wiart and Reveillere [1995] carried out studies at the Achères WWTS in France. Theirstudies showed a significant decrease (50-90%) in the potentially toxic element content ofsewage sludge since 1978, following the application of the "at-source discharge reduction"policy [Bebin, 1997]. However, the main concern is now with organic pollutants, and currentregulations require monitoring of the influent, in order to set up a baseline database fromwhich limits may then be devised.

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2.1.1 Domestic sources

Domestic sources of potentially toxic elements in wastewater are rarely quantified due to thedifficulty in isolating them. Domestic sources include the potentially toxic elementsdischarged from the household to UWW collecting systems and, in addition, corrosion frommaterials used in distribution and plumbing networks, tap water and detergents.

A study by RIVM (Dutch Institute of Public Health and the Environment) in the Netherlands[SPEED, 1993], quantified the waterborne emissions of potentially toxic elements fromhousehold sources, dentistry and utility buildings in the urban environment. Table 2.3 showsthe data of waterborne potentially toxic elements emissions in tonnes per annum.

Table 2.3 Emissions by Dutch households of potentially toxic elements[adapted from SPEED, 1993].

Gross waterborne emissions* tonnes.y-1 to surfacewater (1993)

Potentiallytoxic element

Householdsources

Dentistry Utility buildings

Copper 94 0.6 27Zinc 118 - 26Lead 13 - 3.1Cadmium 0.7 - 0.2Nickel 7.3 - 0.9Chromium 2.9 - 0.3Mercury 0.3 2.3 0.01

* 96 % of the waterborne emissions are expected to go to the UWW collecting systems, with 4% goingdirectly to surface waters.

Domestic products containing potentially toxic elements used on a regular basis at homeand/or at work, are also reviewed by Lewis [1999]. The following lists the principal PTEs andproducts containing them that may enter urban wastewater;

Cadmium: is predominantly found in rechargeable batteries for domestic use (Ni-Cdbatteries), in paints and photography. The main sources in urban wastewater are fromdiffuse sources such as food products, detergents and bodycare products, storm water[Ulmgren, 2000a and Ulmgren, 2000b].

Copper: comes mainly from corrosion and leaching of plumbing, fungicides (cuprouschloride), pigments, wood preservatives, larvicides (copper acetoarsenite) and antifoulingpaints.

Mercury: most mercury compounds and uses are now banned or about to be banned,however, mercury is still used in thermometers (in some EU countries) and dentalamalgams. Also, mercury can still be found as an additive in old paints for water proofingand marine antifouling (mercuric arsenate), in old pesticides (mercuric chloride in fungicides,insecticides), in wood preservatives (mercuric chloride), in embalming fluids (mercuricchloride), in germicidal soaps and antibacterial products (mercuric chloride and mercuriccyanide), as mercury-silver-tin alloys and for "silver mirrors".

Nickel: can be found in alloys used in food processing and sanitary installations; inrechargeable batteries (Ni-Cd), and protective coatings.

Lead: The main source of lead is from old lead piping in the water distribution system. It canbe found in old paint pigments (as oxides, carbonates), solder, pool cue chalk (ascarbonate), in certain cosmetics, glazes on ceramic dishes and porcelain (it is banned now

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for uses in glazes), also in "crystal glass". Lead has also been found in wines, possibly fromthe lead-tin capsules used on bottles and from old wine processing installations.

Zinc: comes from corrosion and leaching of plumbing, water-proofing products (zincformate, zinc oxide), anti-pest products (zinc arsenate - in insecticides, zinc dithioamine asfungicide, rat poison, rabbit and deer repellents, zinc fluorosilicate as anti-moth agent), woodpreservatives (as zinc arsenate), deodorants and cosmetics (as zinc chloride and zincoxide), medicines and ointments (zinc chloride and oxide as astringent and antiseptic, zincformate as antiseptic), paints and pigments (zinc oxide, zinc carbonate, zinc sulphide),printing inks and artists paints (zinc oxide and carbonate), colouring agent in variousformulations (zinc oxide), a UV absorbent agent in various formulations (zinc oxide), "healthsupplements" (as zinc ascorbate or zinc oxide).

Silver: originates mainly from small scale photography, household products such aspolishes, domestic water treatment devices, etc. [Shafer, et.al, 1998, Adams and Kramer,1999]

Arsenic and Selenium: are among the potentially toxic metalloids found in urbanwastewaters. These are of importance due to their potential effects on human/animal health.Only a limited number of studies have taken these into account. Arsenic inputs come fromnatural background sources and from household products such as washing products,medicines, garden products, wood preservatives, old paints and pigments. Selenium comesfrom food products and food supplements, shampoos and other cosmetics, old paints andpigments. Arsenic is present mainly as DMAA (dimethylarsinic acid) and as As (III) (arsenite)in urban effluents and sewage sludge [Carbonell-Barrachina et.al., 2000].

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Household products

Household products were investigated as potential sources of PTE pollution entering theWWTS. Table 2.4 shows metal concentrations in various household products in UKTable 2.4 Metal concentrations in household products[Comber and Gunn, 1996, WRc report, 1994].

Product Zinc(µg g-1)

Copper(µg g-1)

Cadmium(µg g-1)

Nickel(µg g-1)

Washing Powders‘Big Box’

abc

37.935.93.3

1.4<0.5<0.5

74.3136.0

6.6

<0.5<0.5<0.5

Washing Powders ‘Ultra’ abc

<0.12.31.0

<0.51.401.38

24.010.611.8

<0.5<0.5<0.5

Fabric Conditioners abc

0.1<0.10.1

<0.5<0.5<0.5

9.49.0

10.7

0.6<0.5<0.5

Hair Conditioners abcd

<0.11.01.70.5

<0.51.4

<0.51.4

16.817.28.6

68.0

<0.5<0.5<0.51.0

Cleaners ab

0.3<0.1

2.8<0.5

26.017.8

<0.5<0.5

Shampoo (medicated) a 4900 1.4 17.4 <0.5Washing Up Liquid a 0.2 1.1 11.0 0.8Bubble bath a

b0.2

<0.1<0.51.4

13.610.4

<0.5

As can be seen from Table 2.4 there is a great deal of variability between products and alsobetween types of the same products in terms of potentially toxic element content.

The high variability of cadmium concentrations found in the big box washing powders can beexplained by the differences in the composition of phosphate ores used in their production.Cadmium impurities in these phosphate ores have been shown to vary greatly depending onmining source [Hutton et al reported in WRc report 1994]. Reducing the amount ofphosphate in washing powders, or choosing phosphate ores with low Cd concentration couldlead to a reduction in Cd in wastewater from diffuse sources. In Sweden the amount ofcadmium in sewage sludge was reduced from 2 mg kg-1 ds to 0.75 mg kg-1 ds [Ulmgren,1999], and cadmium discharges from households in the Netherlands have been substantiallyreduced due to the switch to phosphate-free detergents [SPEED, 1993]. The 'Ultra' washingpowders, usually phosphate-free, have smaller potentially toxic element contents than thetraditional powders, and are designed to be used in smaller quantities. A shift to these newerproducts will reduce the overall metal load from this source.

The products with the highest metal contents are shown in bold in Table 2.4. The medicated(anti-dandruff) shampoos contain zinc pyrithione and the high zinc concentrations will thusraise the zinc inputs to the UWW collecting system. In 1991 these shampoos were estimatedto represent 26% of the market [*BLA Group 1991- reported in Comber and Gunn 1996 andWRc 1994]. Cosmetics are not included here but they may also contain high levels of zincand several of these products are likely, at least in part, to enter in the waste water system.One study in France [ADEME, 1995a] identified that the main sources of potentially toxicelements in domestic wastewater came from cosmetic products, medicines, cleaningproducts and liquid wastes (including paint), which were directly discharged from thehousehold sink.

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Table 2.5 provides a general picture of some of the potentially toxic elements in variousdomestic products including food products [after Lester, 1987 and WRc report 1994].Sources for each metal are marked with a tick. In addition to the main metals considered inthis study, cadmium, chromium (III and VI), copper, mercury, nickel, lead and zinc, silver,arsenic, selenium and cobalt are also included. Other metals and metalloids for which moreinformation is necessary include manganese, molybdenum, vanadium, antimony and tin.

TABLE 2.5 Domestic sources of potentially toxic elements in urban wastewater[modified from Lester, 1987, and WRc, 1994]Product type Ag As Cd Co Cr Cu Hg Ni Pb Se ZnAmalgam fillingsand thermometers

Cleaning products √ √Cosmetics,shampoos

√ √ √ √ √ √ √

Disinfectants √Fire extinguishers √Fuels √ √ √ √Inks √ √Lubricants √ √ √Medicines andOintments

√ √ √ √ √

Health supplements √ √ √ √ √Food products √ √ √ √ √Oils and lubricants √ √ √ √Paints andpigments

√ √ √ √ √ √ √ √ √ √

Photographic(hobby)

√ √ √

Polish √ √ √Pesticides andgardening products

√ √ √ √ √

Washing powders √ √ √Wood-preservatives

√ √ √

Other sources

Faeces and Urine √ √ √ √ √ √ √ √ √Tap Water √ √ √ √ √Water treatmentand heatingsystems

√ √ √ √ √

Domestic activities

The main domestic sources of potentially toxic elements in wastewater were estimated byWRc [1994] to be (in order of importance):cadmium: faeces > bath water > laundry > tap water > kitchenchromium: laundry > kitchen > faeces > bath water > tap watercopper: faeces > plumbing >tap water > laundry > kitchenlead: plumbing > bath water > tap water > laundry > faeces > kitchennickel: faeces > bath water > laundry > tap water > kitchenzinc: faeces > plumbing > tap water > laundry > kitchen.

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Estimates of the mean potentially toxic element inputs to UWW collecting systems fromdomestic activities are presented in Table 2.6. The results show that for the particular UK(hard water) catchment studied in 1994, the domestic inputs of copper and zinc are majorcontributors to the overall level of potentially toxic elements reaching the WWTS. Most of thezinc is derived from faeces and household activities such as washing and cleaning.Chromium, lead and cadmium were also found to be mainly from domestic activities ratherthan from plumbing.

Table 2.6 Potentially toxic element loads to the UWW collecting systems fromdomestic activities [adapted from Comber and Gunn, 1996]

Load (µg.person-1.day-1)Activity (study in a hardwater catchment area) Zn Cu Pb Cd Ni CrWashing Machine Input Water

Washing6624452

6859977

36.0515

0.611

2752

4238

Dishwashing(machine)

Input WaterWashing

3942

698

2.96

0.031.3

22

0.310

Dishwashing(hand)

Input WaterWashing

5911010

6125<20

3246

0.57.8

24138

3.7136.7

Bathing Input WaterBathing

11401095

1065167

4645

1.013.1

409

5.97.4

Toilet Input WaterFaeces

253111400

80822104

63121

2.048.0

77284

8.551.5

Miscellaneous Input Water 1453 6951 62.7 1.2 54 7.0

Predicted Total 24416 41894 978 86 710 464.2Measured mean fromhousing estates

CatchmentPopulation

50 000

15314 46772 1237 71 925 686

Predicted load to UWWcollecting systems fromdomestic sources (kg/day)

1.2 2.1 0.05 0.01 0.04 0.02

Measured mean total load tothe WWTS kg per day

2.6 3.3 0.3 N/A 0.1 0.15

% of potentially toxicelements from domesticsources in the UK

46.0 64.2 16.9 N/A 25.7 15.3

Based on the above results, changes in population behaviour, such as a shift to dishwasheruse rather than washing up by hand, would reduce potentially toxic element input into theWWTS.

It is noted that, while the quantities of potentially toxic elements dissolved in water fromplumbing will vary across the Europe they will make up a significant proportion of thepotentially toxic element loading going to any WWTS.

Table 2.7 summarises the percentages of the domestic inputs at the Shrewsbury WWTS, inthe UK. As can be seen over a fifth of the copper, zinc, cadmium and nickel entering thewastewater treatment plant from domestic sources are from faeces. This emphasises thefact that faeces are an important source of potentially toxic elements pollution. This source isalso very difficult to reduce. The percentage inputs of chromium and lead from this sourceare much lower.

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Table 2.7 Potentially toxic elements entering wastewater, breakdown by source [WRc,1994]

Cu Zn Cd Ni Pb CrPercentages of total load

Break down of domestic sources as percentage of total metal entering WWTSBathing 12.4 1.2 0.3 1.6 4.4 0.1Toilet 8.9 2.1 0.5 4.3 8.1 0.2WashingMachine

8.0 0.7 0.2 0.9 3.2 0.1

Miscellaneous 8.0 1.2 0.3 2.4 6.1 0.2

Plumbing Input(% of total entering WWTS)

Dishwashing 7.1 0.6 0.2 0.8 2.7 0.1Faeces 20.6 28.0 20.0 23.6 3.1 2.1WashingMachine

9.53 10.8 4.6 4.4 18.6 9.3

Bathing 0.2 2.6 0.9 0.8 1.1 0.3

Activities(% of total metal enteringWWTS)

Dishwashing 2.6 4.6 11.2 1.1 5.1

Human faeces contain high concentrations of potentially toxic elements from normal dietarysources and this represents a principal input of metals to domestic wastewater and sludge ofdomestic origin. The normal dietary contribution of metals represents the background metalconcentration represents the background metal concentration and is the minimumachievable in waste water and sludge.

Concentrations are expected to vary with the intake of metals in the diet, drinking water andmedication and may also be influenced by the increasing prevalence of mineralsupplementation of food, for example with zinc, iron, selenium, and manganese. One studyin France (ADEME 1997) found the following concentrations of potentially toxic elements infaeces (as dry matter): Zn: 250mg kg-1, Cu: 68mg kg-1, Pb: 11mg kg-1, Ni: 4.7mg kg-1, andCd: 2mg kg-1. Differences can also occur due to geographical variations in the dietary habits.For some elements, such as Cd, the weighted average concentration in sludge (e.g. 3.3 mgkg-1 in the UK) is typical of the amount originating naturally in faeces from the normal traceamounts of this element ingested in food (typically 18.8 µg d-1). Other differences in reportedconcentrations of Cd in different EU member states is discussed in section 2.3.

Domestic water and heating systems

Studies in the USA [Isaac et.al, 1997], and Europe [WRc 1994] show that corrosion of thedistribution-plumbing-heating networks contribute major inputs of Pb, Cu and Zn. Leadconcentration for instance can vary between 14 µg.l-1 at the household input and 150 µg.l-1 atthe output.

It has been found that concentrations of copper in sewage sludge are directly proportional towater hardness [Comber et al 1996]. Hard water (high pH) is potentially more aggressive tocopper and zinc plumbing, increasing leaching. However, the opposite is true for lead in thatit dissolves more readily in than soft, acidic water. The high lead levels in drinking water inScotland due to its soft waters are a major concern.

Reductions in the amounts of copper and lead in wastewater have been reported by pHadjustment of tap water and addition of sodium silicate. The addition of alkali agents to waterat the treatment stage and the replacement of much lead piping has led to reductions in leadconcentration [Comber et al., 1996]. Adjusting the pH of tap water may be limited bypractical and economic factors.

Zinc in domestic plumbing comes from galvanised iron used in hot water tanks but is lessproblematic than lead and copper because the amount decreases with the ageing of theinstallations. Copper corrosion and dissolution is also greater in hot water than in cold water

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supplies [Comber et al 1996]. The 'first draw' (initial flow of water in the morning) has higheramounts of copper and lead compared to subsequent draws [Isaac et.al., 1997]. The Cucontent was found to be between 73.7 and 1430 µg l-1, and Pb content between 8.3 and22.3 µg l-1, much greater than in the average effluent from households. Water treatmentwould be recommended for certain water domestic uses, such as boilers and heatingsystems, in order to reduce the metal corrosion.

The type of housing was also found to be important by the WRc report [1994]. Table 2.8gives an average concentration of effluent from two types of estates, "1960s residential" and"1990s residential" from daily bulk- and flow-weighted samples. The larger copper levelsfrom the "modern estate" can be explained by the newer plumbing system. In manycountries copper is the major element used in plumbing. In the UK it has been estimated thatleaching from copper plumbing accounts for over 80% of the copper entering domesticwastewater [Comber et al 1996]. The higher lead level found in the newer housing did notcorrelate with similar studies comparing old and new housing and could not be explainedsatisfactorily; as in general older houses in the UK contain more lead plumbing. Lead fromsolders in the piping system may also be an important source. In other regions of the EUsteel and zinc galvanised iron are used widely. This may explain why zinc in sludge isproportionally greater than Cu in other member states compared with the UK.

Table 2.8 Mean concentration of potentially toxic elements in the effluent fromhouseholds in two types of residential areas [WRc Report, 1994]

Zn Cu Ni Cd Pb Cr HgPotentially toxic element concentration µg.l-1

" 1960s residential" 74.3 219.4 5.2 0.71 9.02 5.65 0.114" 1990s residential" 147.0 458.3 7.56 0.34 90.61 3.3 0.088

In summary, potentially toxic elements entering UWW collecting systems from domesticsources are related to:

• household water consumption• the plumbing and heating system in the household• the concentrations of potentially toxic elements in the products used in the household

and quantities of the products used• any grey water recycling schemes• how much of the products are discharged into wastewater.

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2.1.2 COMMERCIAL SOURCES

Limited data is available for the potentially toxic element contribution from commercialsources and health care inputs (such as hospital and clinical wastes). Inputs from artisanalsources are looked at in more detail in a separate Case Study in Section 6.

Cadmium could originate from laundrettes, small electroplating and coating shops, plasticmanufacture, and also used in alloys, solders, pigments, enamels, paints, photography,batteries, glazes, artisanal shops, engraving, and car repair shops. Data from ADEME[1995], estimated that worldwide, 16000 tonnes of cadmium were consumed each year; 50-60% of this in the manufacture of batteries and 20-25% in the production of colouredpigments.

Chromium is present in alloys and is discharged from diffuse sources and products such aspreservatives, dying, and tanning activities. Chromium III is widely used as a tanning agentin leather processing. Chromium VI uses are now restricted and there are few commercialsources.

Copper is used in electronics, plating, paper, textile, rubber, fungicides, printing, plastic, andbrass and other alloy industries and it can also be emitted from various small commercialactivities and warehouses, as well as buildings with commercial heating systems.

Lead, as well as being used as a fuel additive (now greatly reduced or banned in the EU) itis also used in batteries, pigments, solder, roofing, cable covering, lead jointed waste pipesand PVC pipes (as an impurity), ammunition, chimney cases, fishing weights (in somecountries), yacht keels and other sources.

Mercury is used in the production of electrical equipment and is also used as a catalyst inchlor-alkali processes for chlorine and caustic soda production. The main sources in effluentare from dental practices, clinical thermometers, glass mirrors, electrical equipment andtraces in disinfectant products (bleach) and caustic soda solutions.

Nickel is used in the production of alloys, electroplating, catalysts and nickel-cadmiumbatteries. The main emission of nickels are from corrosion of equipment from launderettes,small electroplating shops and jewellery shops, from old pigments and paints. It also occursin used waters from hydrogenation of vegetable oils (catalysts).

Zinc is used in galvanisation processes, brass and bronze alloy production, tyres, batteries,paints, plastics, rubber, fungicides, paper, textiles, taxidermy (zinc chloride), embalming fluid(zinc chloride), building materials and special cements (zinc oxide, zinc fluorosilicate),dentistry (zinc oxide), and also in cosmetics and pharmaceuticals. The current trend towardselectrolytic production of zinc which, in contrast to thermally produced zinc, has virtually nocadmium contamination. This means that cadmium pollution to UWW due to the corrosion ofgalvanised steel will in time become negligible. [SPEED 1993].

Platinum and platinum group metals (PGMs) such as palladium and osmium can enterUWW from medical and clinical uses, mainly as anti-neoplastic drugs. The amount inhospital/clinical effluent has been estimated to be between 115 and 125 ng l-1 [Kümmererand Helmers, 1997, Kümmerer et.al., 1999] giving a total emission of 84-99 kg per annumfrom hospitals in Germany. Other sources of platinum metals in the environment related tocommercial activities come from catalysts used in petroleum/ammonia processing andwastewaters, from the small electronic shops, jewellery shops, laboratories and glassmanufacturing. Section 6 contains a detailed Case Study (a) on PGMs in urban waste waterand sewage sludge.

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Silver could potentially be emitted from photographic and printing shops, from jewellerymanufacturers and repairers, plating and craft shops, glass mirror producers and small-scale water filters.

Studies in Spain showed the presence of elevated concentrations of Cd, Cu, Hg, Pb and Znin urban wastewater and in the coastal environment [Castro, et.al., 1996], with largeconcentrations of copper and zinc possibly due to the use of fungicides in glass-houses.

A summary of concentrations of metals found in effluent from commercial sources indifferent regions in Europe is given in Table 2.9.

Table 2.9 Summary of potentially toxic elements in UWW from commercial sources( g l-1)Element Country Industry Industrial Effluent

(µg l-1)Reference

Cd GermanyGreece

All sectorsPetroleum industries

3-1,250300-400

Wilderer et al 1997NTUA, 1985

Cu GermanyGreeceItaly

All sectorsMetal and electrical industriesArtisanal galvanic shopsGoldsmiths and jewelleryshops

37-26,0005,000-10,000

20,500700-1,900

(max.13,300)

Wilderer et al 1997NTUA, 1985EBAV, 1996

Cr GermanyGreece

Italy

All sectorsMetal and electrical industriesTanneriesArtisanal galvanic shops

<10-20,100500-13,000

100-7,000,00016,000

Wilderer et al 1997NTUA, 1985

EBAV, 1996Pb Germany

GreeceItaly

All sectorsMetal and electrical industriesCeramics and photoceramicsshops

<50-13,400500

6,000

Wilderer et al 1997NTUA, 1985EBAV, 1996

Ni GermanyGreeceItaly

All sectorsMetal and electrical industriesArtisanal galvanic shops

<10-7,300500-14,500

19,700

Wilderer et al 1997NTUA, 1985EBAV, 1996

Zn GermanyGreeceItaly

All sectorsMetal and electrical industriesGoldsmiths and jewelleryshops

30-133,00060-2,830

1,000 (max. 7,000)

Wilderer et al 1997NTUA, 1985EBAV, 1996

As Spain Paper mills 3.4 Navarro, et.al1993

In 1999, a project carried out by Anjou Recherche [LIFE, 1999], attempted to classify allcommercial sources of wastewater pollution based on a matrix of 73 main pollutantsincluding many potentially toxic elements and certain organic pollutants. The UWWcollecting system of Louviers, for example, had listed 1054 establishments of which 39%were capable of emitting at least one of these pollutants. Ten classes of activities wererecorded, of which health and social action (33.5%), manufacturing industry (20%), hotelsand restaurants (17.8%), and collection services (10%) appeared most often as potentialpolluters. It was found on average, that between a third and a half of the activities emittedpollutants. In the Louviers area, it was found that 53 urban businesses and institutions couldpotentially emit cadmium, 168 chromium, 147 copper, 35 nickel, 167 mercury, 50 lead, and63 zinc. This suggests that more can be done to reduce trade effluent discharges.

Between 1990 and 1992, Stockholm Water Company investigated measures to reducedischarges of potentially toxic elements into the UWW collecting system. This programmeinvolved the following groups: the city council, neighbouring municipalities, small businessesand industry, professional associations and NGOs as well as local households. Efforts toreduce pollution entering the wastewater system were divided among all the major sourcesi.e., small businesses and industry, wastewater from urban runoff, household wastewaterand storm water.

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Collaborative projects were developed both for research, product development andeducational programmes. Local commercial organisations (particularly the Swedish DentalFederation) co-operated in the project; new technologies were developed and an evaluationof alternative products was carried out. Pollution limits were imposed that were determinedto be appropriate to encourage the purchase of the endorsed environmental products. Thisprogramme of research, and earlier work during the 1980s led to a reduction of between 50and 80% of potentially toxic elements in sewage sludge [Ulmgren 2000a].

The results of some more specific investigations into sources of potentially toxic elements inUWW are outlined below.

Motor industry - vehicle washing

Scandinavian studies [SFT-1997a, 1997b, 1999] showed that the motor industry, followed byvehicle workshops contribute most to the potentially toxic element load in UWW. Vehiclewashing, particularly heavy goods vehicles (HGVs), was found to be an important source ofpotentially toxic element contamination.

In Sweden, oil separators are commonly used in vehicle washing and motor industriesbefore discharging effluent to UWW collecting systems. Most facilities in Sweden arereported to be equipped with combined oil separators and sludge traps where the dispersedoil and sludge should be retained. However, tests at one of the light vehicle (LV) washingfacilities showed that this equipment was ineffective with practically no difference betweenthe influent (before the separator) and the effluent (after the separator). This was due to theformation of stable emulsions in the wastewater caused by the detergents in themicroemulsion formulations used for vehicle washing [Paxéus, 1996b].

A study by the Norwegian Pollution Control Authority [SFT, 1999] examining potentially toxicelement pollutants in Norway found that out of six petrol stations investigated, only one hadan oil separator/sand trap that worked effectively. Although designed to reduce thecontamination of urban wastewater, oil separating devices are generally ineffective atreducing pollutant emissions from vehicle washing and motor industry facilities.

Dental practices and healthcare (mercury)

In the late 1980s, the high concentration of mercury in sewage sludge (SS) at HenrikdalWWTS, Stockholm, prompted an investigation to identify potential sources (Table 2.10). Itwas concluded that the high mercury content of sludge was attributable to dental practicesand the use of mercury in dental amalgams. Amalgam separators were ineffective atretaining mercury and new legislation was introduced to combat this [Ulmgren, 2000a].Recent reduction or bans on the use of mercury in various products, such as batteries andthermometers, has led to a reduction of mercury input into UWW [Ulmgren, 2000a andUlmgren, 2000b]. Mercury recycling schemes have also proved to be successful, and couldbe extended to other countries and activities.

Other discharges from dental technicians shops are covered in detail in Section 6, CaseStudies.

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Table 2.10 Sources of mercury in urban wastewater in Sweden[Table adapted from text, Ulmgren 2000a]

Source CommentsHeavy Industry no Ruled out as a source of mercury to UWW and SS as these were

not connected to the UWW collecting systemSmall andMediumEnterprises(18 Companies)

no Ruled out as they were operated in such a way that nocontamination of the wastewater was likely

Storm Drains yes 20% of the total mercury load came from storm drains. This waslargely traced to the deposition of particulates emitted fromcrematoria which are estimated to be about 50 kg of mercury a year

HouseholdWastewater

yes About 15% of the mercury entered the waste water system throughthe use of mercury thermometers in the home, and also from smallamounts of mercury in food and amalgam fillings in teeth

Dentists andDentalTechnicians

yes High mercury content of sludge was largely attributable to dentalpractices and the use of mercury in dental amalgams

Hospitals yes Samples indicated that hospitals emit 10% of the mercury loadingOld SewagePipes

yes Investigations in the last few years have found many sources ofmercury in old pipes

Similar findings have been reported in other countries. A WRc report [1994] established thatin the UK mercury emissions are much higher from commercial, rather than domesticsources, mainly due to dental practices. In France, it is estimated [Agence de l'Eau, 1992]that between 73 and 80 % of the mercury in UWW is from dental practices and amalgamfillings corrosion. In the Louviers area of France, the analysis of wastewater and sewagesludge showed that out of the total load of mercury, 50% was lost from medical practices,13% from dentistry practices, 28% from medical auxiliaries (nurses etc.), 4% from hospitalactivities, 4% from veterinary activities, and 1% from ambulance activities. Thus, in thisinstance, targeting medical/dental practices may help reduce pollution from mercury [LIFE,1999].

Other sources of mercuryIn 1993, the amount of mercury entering France was 209 tonnes, the amount leaving Francewas 87 tonnes, and hence 122 tonnes were entering the environment in the form of waste[AGHTM, 1999-2000] (see Table 2.11).

Table 2.11: Mercury contained in waste[from Dossier sur les dechets mercuriels en France: AGHTM, 1999].

Activity Amount (tonnes) % Treated and recycledZinc and lead metallurgy 18 4.8

Thermometers 9 5.6

Dentistry amalgams 9 11.4Batteries 6.8 7.4Laboratories 0.9 0.0Fluorescent tubes 0.8 12.5Barometers 0.4 10.0High intensity lamps 0.2 10.0Chlorine production 77 ?

Table 2.11 highlights the amount of mercury waste produced by zinc/lead metallurgy andchlorine production. In Galicia, north-western Spain, high mercury levels in UWW andsewage sludge are attributed to chlor-alkali production in the Pontevedra area [Cela et.al.1992]. In Portugal too, the presence of mercury in treated wastewater, sewage sludge andthe lagoons of Aveira is linked with chlor-alkali production [Lucas et.al, 1986 and Pereira

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et.al, 1998]. There appears to be potential for improved control and recycling of mercurywaste associated with these activities.

Sources of chromium

Mine production of chromium in Finland has increased from 348 thousand tonnes to justover a million tonnes in 1990 [Mukherjee 1998], representing just under 10% of worldproduction. The main sources of chromium in wastewater are from the metal, chemical andleather industries (Table 2.12). As can be seen, the chemical industry contributes over halfof the total emissions to UWW and surface waters in this region.

Table 2.12 Chromium emissions to water in Finland [adapted from Mukherjee, 1998]Emissions to water (not exclusively UWW)Source Categorytonnes per annum % of total contribution

Chemical Industry 14.3 58.1Paint Manufacture 0.01 <0.1Electroplating 0.1 0.4Ferro-chrome and Stainless SteelPlants

4.6 18.7

Leather Processing 5.5 22.5Total 24.6 100.0

Mukherjee [1998] reports that in Scandinavia, chromium compounds are also used in woodpreservatives, along with arsenic compounds (As2O5) and an oily mixture of organicchemicals (phenol and creosol).

All wet-textile processing in Finland discharges its wastewater, containing chromium andother metals, to the WWTS. The textile companies studied [Kalliala, 2000] producedbetween 50 and 500 litres of wastewater per kg of textile produced. Wastewater analyseswere carried out at six major Finnish textile companies (two of these include analysis forpotentially toxic elements (Table 2.13)):

Table 2.13 Potentially toxic elements in wastewater from textile processing in Finland[Kalliala, 2000]

Wastewater analysis (µg l-1)Company 2 Company 4

Lead 0.11 -Chromium 0.03 60Copper 0.4 80Zinc - 20

There is a very high variation in the process emissions between these two plants. Company2 was noted to use cellulose blends while company 4 was noted to have mainly polyesterand polyester blends.

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Sources of lead

Data from ADEME [1995] showed that worldwide consumption of lead is around 5.4 milliontonnes per year. In a Swedish study [Palm, Östlund, 1996] in the Stockholm area the totalamount of lead used in products such as those listed previously, was estimated at between44,000 and 47,000 tonnes per annum. Clearly the potential for lead entering UWW fromthese sources will vary greatly. The largest amount of lead that finds its way to the WWTS islikely to be contributed by piping. Estimates for the amount of lead used are 8,000 tonnes inlead jointed water pipes used inside buildings, followed by 2,000 tonnes used in lead jointedwater pipes used outdoors (higher replacing rate), and 120 tonnes used in PVC piping.

In the case of Finland, the Ministry of the Environment report that the drinking waterpipelines are predominantly plastic (85% PVC and PEH), with 11% cast iron; no lead is usedfor pipes conveying water. The wastewater pipes for the UWW collecting system are 57%concrete and 41% plastic.

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2.1.3 URBAN RUNOFF

Runoff to UWW collecting systems and waterways has been intensely studied due to itspotentially high loading of potentially toxic elements [WRc, 1994]. Atmospheric inputs to theurban runoff depend on the nature of surrounding industries, on the proximity of majoremission sources such as smelters and coal fired power stations and the direction of theprevailing wind. Potentially toxic element loads can be five fold greater in runoff nearcommercial activities, than in residential areas far from industrial emitters. Roof runoff andbuilding runoff also contribute to the total runoff loading and may be a source of considerableamounts of potentially toxic elements such as zinc, lead, copper and cadmium. Road androof runoff sources are particularly important during storm events, which will allow flushing ofpotentially toxic elements and other pollutants from surfaces. Furthermore, it is important tonote that the metal species released are usually in a freely dissolved, bioavailable form.Nevertheless, these sources are very variable, as every event is different and depends ontraffic, material and age of roofs and other surfaces, and meteorological and environmentalconditions.

Although a number of studies had focused on the effects of urban land use in thequantification of precipitation runoff, it was not until the 1950s that the first qualitative1

studies were undertaken [Palmer, 1950 and 1963; Wilkinson, 1956]. Table 2.14 providesconcentrations for a number of potentially toxic elements in urban runoff, as a summary ofvarious investigations from 1975 to 1978. It is important to note that the measuredconcentrations differ considerably. The main sources of pollution in urban precipitation runoffcan be summarised as follows [based on Mitchell, 1985]:

• Road and vehicle related pollution• Degradation of roofing materials• Construction• Litter, vegetation and associated human activities• Erosion of soil

Table 2.14 Maximum and mean concentrations of potentially toxic elements (mg l-1)in urban precipitation runoff pollutants [after Mitchell, 1985]

Droste and Hartt,1975

Mance and Harman,1978

Mattraw andSherwood, 1977Pollutant

Mean Max Mean Max Mean MaxLead 0.205 3.7 0.21

Iron 0.317 4.2 5.3

Copper 0.028 0.35

Manganese 0.11 1.7

Zinc 0.271 1.63

A comparative analysis between the different sources of pollution in urban precipitationrunoff is subject to variation, dependent on catchment characteristics, time of year andmeasurement procedures. Urban precipitation runoff pollution has received far less attentionthan other forms of urban pollution (i.e. atmospheric). Table 2.15, gives an indication of thegeographic and temporal variability involved.

1 Focusing on pollutant loads in rainfall runoff

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Table 2.15 Mean concentrations in rainwater runoff (in µg/l)

Site Cd Cu Pb Zn ReferenceUrban (Netherlands) 0.9 8 20 31 Van Daalen, 1991

Central Paris 2.4 60 140 Granier, 1991

Central Paris 0.11 6 13.7 38.8 Garnaud et al., 1996-1997

Garnaud et.al., [1999] attempted a comparative study between the main sources of pollutionin urban precipitation in Paris. In this recent study of individual rain events, bulk sampleswere collected within four gutter pipes (roof runoff), three yard-drainage pipes (yard runoff),six gullies (street runoff) and one combined UWW collecting system, at the catchment outlet.The results can be seen in Figure 2.3. The values are median values.

Figure 2.3 Comparison between main elements contributing to precipitation runoff inrespect to potentially toxic elements pollution load [after Garnaud et al., 1999]

A comparison of samples from consecutive phases of precipitation runoff (precipitation toroof runoff to urban runoff to catchment outlet) indicated that the potentially toxic elementconcentration increased by a factor of 12, 30, 30 and 60 for cadmium, copper, lead and zinc,respectively. Corrosion of roof and urban surfaces as well as human activities contributes tothis contamination, including corrosion or emission from vehicles and commercial activities.Bulk metal concentrations were similar within all urban runoff samples except for zinc andlead, which were particularly concentrated in roof runoff samples, due to their contaminationby corrosion of roof materials. At all sites it appeared that bulk metal concentrations could beranked as: Cd << Cu < Pb << Zn.

Differences in potentially toxic elements concentration in precipitation water and runoff fromroofs and streets have also been found in a German study (Table 2.16). In this case it canbe seen that runoff from streets has the highest potentially toxic element content.

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Table 2.16 Concentration changes of certain contaminants in precipitation water andrunoff from different outflow paths, Germany.[Xanthopoulos and Hahn, 1993] LOD Limit of detection

Pollutant Precipitation[µg/l]

Run-off fromroofs [µg l-1]

Run-off fromstreets [µg l-1]

Pb 5 104 311Cd 1 1 6.4Zn 5 24 603Cu 1.5 35 108Ni 5 <LOD 57

A study carried out by Rougemaille (1994) analysed the wastewaters of the treatment plantin Achères in the Paris region and found that lead concentrations varied between 0.05 and0.5mg l -1 and that the average concentration was 0.1mg l-1. These wastewaters come fromfour different urban areas. The study showed that lead concentrations were 3 times largerduring wet weather than during dry weather, hence proving the importance of the runoffsources.

A study carried out around the region of Nantes in France in 1999, analysed road runoff froma major highway for a year showing that lead and zinc are the main pollutants present inrunoff waters (Table 2.17).

Table 2.17: Analysis of raw runoff waters [Legret, 1999]

PAH(ng l-1)

Pb( g l-1)

Cu( g l-1)

Cd( g l-1)

Zn( g l-1)

Mean <96 58 45 1 356Median <74 43 33 0.74 254Range <11-474 14-188 11-146 0.2-4.2 104-1544SD 76 44 27 0.86 288

The Swedish study mentioned in previous sections [Palm, Östlund, 1996], estimated thetotal amount of zinc passing into wastewater at 6,300 tonnes per annum. Most of the zincpresent in the urban environment was generated by urban runoff and rain; from roofs andbuilding surfaces (1,600 tonnes), from cars excluding tyres (1,500 tonnes), from tyres (200tonnes), and from lampposts and street furniture (an estimated 1,142 tonnes). Zinc was alsoattributed to water pipe couplings (1,000 tonnes).

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A Road and vehicle contribution

Roads are a major source of pollution in urban environments and contribute to wastewaterpollution both directly and indirectly (airborne pollutants generation). Sartor and Boyd[USEPA, 1972] determined the major constituents of road related runoff to be inorganicmatter, but the total mass of inorganic matter present seemed to increase as the antecedentdry period (ADP) increased. Sources of the organic and inorganic fraction of road-producedpollutants are summarised as follows:

• Vehicle lubrication systems losses• Vehicle exhaust emissions• Degradation of automobile tyres and brakes• Road maintenance• Road surface degradation• Load losses from vehicles (accidental spillages)• Precipitation (wet deposition)• Atmospheric deposition (dry deposition)

Potentially toxic elements in runoff occur from motor fuel combustion, brake linings, tyrewear and road surface wear. Motor fuel combustion was the largest source of lead to runoffbut it is on the decrease due to the gradual phasing out of leaded fuel in the EU. Othermetals emitted from exhausts are zinc, chromium and more recently tin from thereplacement anti-knock compounds in petrol. The presence of Zn and Cd in road surfacesediments can also be explained by the addition of Zn dithiophosphate in the manufacturingof lubricating oil, Cd being present as an impurity of the original Zn. Brake lining wearcontributes copper, nickel, chromium and lead to runoff. Tyre abrasion contributes to theload of zinc, lead, chromium and nickel due to the soot and metal oxides constituents.Cadmium in car tyres is attributed to zinc-diethylcarbonate, which is used during thevulcanisation process. Road surface wear contributes to emissions of nickel, chromium,lead, zinc and copper.

Legret and Pagotto (1999) produced estimates of potentially toxic element content fromvehicle related pollution sources (Table 2.18), and contributions to road runoff (Table 2.19).

Table 2.18 Potentially toxic element contents in vehicle and road materials (mg kg-1)[after Legret and Pagotto, 1999]

Sources Pb Cu Cd Zn

Leaded petrol 200 - - -

Unleaded petrol 17 - - -

Brake linings 3900 142000 2.7 21800

Tyre rubber 6.3 1.8 2.6 10250

De-icing agent 3.3 0.5 0.2 0.5

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Table 2.19 Emissions fluxes in precipitation runoff (kg km-1 annum) and percentageremoved in drainage waters [after Legret and Pagotto, 1999]

Sources Solids Pb Cu Cd ZnTyre wear 314 0.002 0.0006 0.0008 3.22Brakelinings

100 0.390 14.2 0.0003 2.17Vehicles

Petrol - 13.0 - - -Safetyfence

- 0.002 0.0002 0.0002 0.95Road

De-icingagent

130 0.015 0.002 0.0007 0.002

Air deposition 86 0.014 0.015 0.0009 0.21

Drainage water %2 235 5 2 313 37

The percentages of these potentially toxic elements entering the drainage systems show thata large proportion of the pollutants released do not end up in the runoff waters but areprobably emitted into the atmosphere. The Zn content in urban runoff remains highly variabledepending on the use of safety barriers made from galvanized steel or from an alternativematerial. All parameters are affected by traffic and by variables connected to themeteorological and environmental conditions of the sites concerned, which make thecomparison highly uncertain (Montrejaud-Vignoles et al., 1996).

The following mechanisms summarise the way the pollutants are transported (pathways)from the catchment over the roads and finally in the drainage network:

- Soluble contaminants dissolved in the runoff water- Insoluble particles acting as sorbents for potentially toxic elements and organic

contaminants which are transported by the runoff water [Ellis, 1976; Sylvesterand DeWalle, 1972]

- Removal by air-dispersal involving transfer of the surface contaminants to theatmosphere either as dry particles or dissolved in surface water.

Note that the first two mechanisms above result in pollutants being transferred ultimately intoreceiving waters or sewage sludge, while the last mechanism removes pollutants prior totheir ultimate disposal. The general conclusion, however, is that the majority of road surfacerunoff contamination is sediment associated. In particular, the prime transport mechanismsand pathways with respect to road-runoff sediment transport are [Sartor and Boyd, 1972]:

- Particle entrapment- Cross surface transport by overland sheet-like flow to the gutter as turbulent

heavy fluid transport process- Linear transport of the particle material parallel to the kerb line into the road gully.

The transportation of surface particulates is sporadic in nature [Mitchell, 1985]. Metal levelstend to fall after periods of rain, whilst elevated concentrations have been recorded afterprolonged dry periods. Also introduced in the movement patterns are local storage andresidence effects due to the intrinsic configurations and micro-topography of the roadsurface [Harrop, 1983]. The fact that contaminants move through this sequence is wellknown, however, the relationship between the contaminants and the various mechanismsinvolved are poorly understood. Comparison of metal distribution during particle transport,

2 For value >100% the research identified larger concentrations in the samples than expected from the sources that were

taken into account. This is particularly true for Cd concentrations. Additional sources of Cd (such as lubricating oils, asdiscussed earlier) may account for this underestimation.

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from atmosphere to receiving water body, clearly demonstrates a metal mobility evolution(Garnaud et al., 1999). Extremely labile (i.e. hydrosoluble or exchangeable), within dryatmospheric deposits Cd, Pb and Zn, become bio-available within street runoff and stablewithin UWW collecting system or river sediment. In conclusion particulate metal mobility maybe classified as: Cu<<Cd<Pb<Zn<Fe.

Potentially toxic elements are predominantly associated with inorganic particles andresearch indicates that potentially toxic element concentrations increase as particle sizedecreases (Sansalone and Buchberger, 1997; Lloyd and Wong, 1999). In particularColandini and Legret (1997), found a bimodal distribution with the highest concentrations ofCd, Cu, Pb and Zn being associated with particles of less than 40 µm in size. Table 2.18indicates the potentially toxic elements distribution across the particle size distribution for anumber of potentially toxic elements.

Table 2.20 Approximate concentration of potentially toxic elements associated withparticles [after Colandini and Legret, 1997]

Approximate concentration of potentially toxic elementsassociated with particles (mg kg-1)

Inorganicparticle sizefraction (µm) Zinc Lead Copper Cadmium

<40 900 920 240 2400040-63 275 100 100 500063-80 300 100 125 500080-125 350 150 175 5000125-250 400 200 200 5000250-500 450 175 300 3000

500-1000 240 225 30 3000

Table 2.21 summarises measurements from 1975 to 1982 on concentrations of potentiallytoxic element pollutants in road runoff in three German towns:

Table 2.21 Summary of pollutant loads in urban runoff caused by road related sourcesfrom 1975 to 1982 [after Klein, 1982]

Test catchmentsPollutant meanconcentrations(mg/l)

Pleidelsheim Obereisesheim Ulm / West

Cd 0.0059 0.0059 0.0028Cr 0.0096 0.02 0.0052Cu 0.097 0.117 0.058Fe 3.42 5.16 2.18Pb 0.202 0.245 0.163Zn 0.36 0.62 0.32

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Vehicle emissions:Prior to the lead ban in fuel3, research had identified exhaust lead emissions to account formore than 90% of atmospheric lead pollution. Although the main source of lead in theaverage urban atmosphere is lead from fuel additives, it appears that only 5% of the leadcan be traced in runoff water. The greatest part may, therefore, disperse in the atmosphereor settle on the soil by the roadside (Hewitt and Rashed, 1990). The actual rate and form oflead emission is crucially dependent on driving conditions. High engines speeds and rapidacceleration can increase emission levels due to reactivation of particles deposited on theexhaust system. Total lead emissions can double if the engine and/or weather is cold. Thedeposition from petrol to the road surface was estimated to be 0.049g Pb.vehicle-1. km -1. Inthe case of lead-containing fuel, approximately 75% of lead is discharged to the atmosphere(Hewitt and Rashed, 1990) and a further 20% is retained in engine sump oil (Wilson, 1982).

Concentration of lead added to petrol has rapidly declined in the EU during the 1980s due tolegislation. In Austria for example, the lead emissions between 1985 and 1995 have beenreduced by approximately 88 % [UBA, 1999]. Figure 2.4 is indicative of this period of policychange (between 1972 and 1992). Note the high correlation between lead in petrol and leadair pollution despite the large rise in traffic flow experienced in the same period.

Figure 2.4 Reduction in lead concentrations in air and petrol between 1972-1992in the EU [after Montague and Luker, 1994]

Platinum group metals4 (PGMs) are increasingly used in vehicle exhaust catalysts (VECs)and can be traced in the urban environment [Farago et al., 1995]. These are covered indetail in Section 6, case studies.

3 Amid mounting concern that the lead dispersed in the environment was causing damage to humans and theenvironment, a series of regulations and directives have been adopted in Europe since the 1980s, in order to phase outthe use of leaded petrol. Lead content of petrol has been reduced from 0.4g l-1 to 0.15gl-1, following the Lead In PetrolDirective 85/210/EEC. Further regulations on air quality have now come into force limiting the levels of lead in air, with anattainment date of 2005. Some countries, namely: Austria, Denmark, Finland, Germany, the Netherlands, Norway,Sweden, and the UK have already phased out its use.

However, in some countries in Eastern Europe, higher levels of lead are permitted, up to 0.4g l-1. In most of the newlyindependent states and the Russian Federation, permitted lead levels are 0.15-0.37 g l-1. Various strategies for reductionof lead in petrol have been made in these countries, though it is expected that some will have difficulty in achieving thesetargets.4 Platinum, palladium in vehicle exhaust catalysts and rhodium in three-way catalysts

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Vehicle degradation:

Tyre wear releases lead, zinc and hydrocarbons, either in particulate form or in larger piecesas a result of tyre failure. The deposition of Zn on the road surface from tyres has beendetermined to be 0.03g vehicle km-1 (Mitchell, 1985). Metal particles, especially copper andnickel, are released by wear of clutch and brake linings. Additionally, the presence of Ni andCr in storm runoff can result from the degradation of car bumpers and window sealingswhere both metals are used in the manufacture of chrome plating. Copper is a commonconstituent of piping and other components of the engine and chassis work. The contributionto road surface material of vehicle tyres and brake linings from different road types has beensummarised in Table 2.22.

Table 2.22 Input from tyre wear and brake lining degradation for each road category[after Muschack, 1990]

Individual elements(g ha of road-1 annum-1)

TotalAbrasion(kg ha ofroad-1) Pb Cr Cu Ni Zn

Typeof street

Tyrewear

Brakelining

Tyrewear

Brakelining

Tyrewear

Brakelining

Tyrewear

Brakelining

Tyrewear

Brakelining

Tyrewear

Brakelining

Residentialway 137 6.8 60 7.3 10 13.7 12 210 10 51 35 0.9

Residentialstreet 62 8.5 76 9.0 13 17.0 17 260 12 63 43 1.7

Distributorroad

72 12.4 112 16.6 19 24.9 26 381 18 93 63 2.4

Maindistributorroad

109 19.1 172 20.4 29 38.3 39 586 27 143 96 2.6

Main road 127 30.3 266 32.2 44 60.5 60 926 42 226 149 2.1Dualcarriageway 120 43.4 382 46.3 64 86.9 72 1329 61 324 214 5.8

Motorway 328 82.1 572 87.6 120 164 164 2513 115 613 405 11.0

These findings are also supported by evidence in non-EU countries. Drapper et al. (2000) inan experimental site in Australia concluded that brake pad and tyre wear, caused by rapidvehicle deceleration, contributes to the concentrations of copper and zinc in road runoff.Laser particle sizing indicated that the median size (by volume) of the sediment found was100 µm, but have settling times of around 24 hours under laboratory conditions. Theexplanation offered to this unexpected behaviour is the presence of potentially toxicelements. One of the most important findings of this study (which took into account bothAustralian and US research), was that traffic volume cannot account for more than 20-30%of the variability of pollutant load variations and therefore a traffic volume criterion onwhether or not the precipitation runoff should be treated may not be acceptable. Drapper etal. (2000) also stated that precipitation intensity and preceding dry days could be asignificant factor influencing actual pollutant concentrations. It can be argued that a moresuitable criterion for treatment need assessment would be the environmental significance ofthe receiving waters.

Road related sources:Pollution in precipitation runoff from road-related sources stems mainly from maintenancepractices including de-icing, road surface degradation and re-surfacing operations.

De-icing agents include large amounts of sodium chloride, and may also contain smallerconcentrations of iron, nickel, lead, zinc chromium and cyanide as contaminants. Their usecan greatly increase corrosion rates in vehicles and metal structures leading to increasedmetal deposition. Salt in solution can also create conditions that allow the release of toxicmetals such as mercury from silts and sludge [WRc, 1993]. It is likely that de-icing is a majorsource (at least in the winter) of bromide, nickel and chromium [Hedley and Lockley, 1975]

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and it is suggested it may affect the solubility and mobility of other metals, notably of lead,which may precipitate more readily in the presence of sodium [Laxen and Harrison, 1977].

The use of NaCl as a de-icing agent may change the behaviour of the accumulatedcontaminants in roadside soils. In soils exposed to high Na concentrations with asubsequent supply of lower salinity water, as in snowmelt periods and storm flow events,there is a risk of colloid dispersion and mobilization [Norrstrom and Jacks, 1998]. Soilcolumn leaching experiments with NaCl and low-electrolyte water have provided evidencefor the mobilisation of organic colloids and Fe-oxides, suggesting that potentially toxicelements may reach the groundwater via colloid-assisted transport [Amrhein et al., 1992,1993]. Moreover, the use of road-salt may result in the increased mobilisation of potentiallytoxic elements due to complexation with chloride ions [Doner, 1978; Lumsdon et al., 1995].

Complexed cyanide ion (in the form of sodium ferrocyanide) is added as anti-cakingagent to de-icing agents, and compounds containing phosphorus may also be added asrust inhibitors. Novotny et al. (1998) argue that although ferrocyanide is non-toxic in itsoriginal form, its instability under predominant natural surface waters conditions, results indecomposition to free cyanide which is toxic (free cyanide↔HCN(aq)+CN-(aq)). The initialcyanide form is only stable within the pH range 8 to 14 and zero to –600mV redox potential(Eh). The rate of decomposition is estimated around 10.2µg l-1 h-1 in salt water. Meeusen etal. (1992) estimated similar rates of decomposition.

Road surface degradation is likely to release various substances: bitumen and aromatichydrocarbons, tar and emulsifiers, carbonate and metals depending on the road constructiontechniques and materials used. The following table provides some indication of theconcentrations of potentially toxic elements released from road surface abrasion, classifiedby type of road encountered in urban areas.

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Table 2.23 Emissions from abrasion of urban streets surface material[after Muschack, 1990]

Individual elements(g ha of road-1annum-1)

Typeof street

Total abrasion(kg ha of road-1

annum-1) Pb Cr Cu Ni ZnResidential way 1734 177 619 88 2030 285

Residential street 2148 219 767 110 2513 352

Distributor road 3152 322 1125 161 3688 517

Main distributorroad

4850 495 1731 247 5674 795

Main road 7665 782 2736 391 8965 1257

Dual carriageway 11000 1124 3927 561 12870 1804

Motorway 10000 1020 3570 510 11700 1640

Accidental discharges:Although spillages can be considered a minor add-in in terms of total pollutant loading, theycan be one of the most serious sources of contaminants in urban areas. They can rangefrom minor losses of fuel to major losses from fractured tanker vehicles. The resulting impacton wastewater treatment plants or directly to water receptors is hard to estimate due to therandom nature and the unpredictability to both the extent of the spill and its position relativeto the precipitation runoff system.

In the case of chemical accidents, water or foam medium are used for road cleaningpurposes or for fire fighting. The compositions of a typical and unusual load of the surfacerun-off are compared in Table 2.24.

Table 2.24 Example of a typical and unusual load in run-off water from a Germanmotorway [Ascherl, 1997, Krauth and Klein, 1982]

Parameter Run-off rain water- mean value[mg l-1]

Water for firefighting- [mg l-1]

Cd 0.0059 0.057Cr 0.0096 0.053Cu 0.097 0.203Fe 3.42 4.0Pb 0.202 0.439Zn 0.36 4.7

Information relating to the frequency of accidental spillages is presented in Table 2.25,based on data extracted from NRA (Thames Region, UK), indicating pollution incidentsregistered from 1988 to 1993 (July):

Table 2.25 Accidental spillages in Thames Region (UK)

Year 1988 1989 1990 1991 1992 1993(July)Number of incidents reported 2811 3613 3444 3417 3598 1979Road incidents reported 125 177 175 136 212 104Road incidents where pollutionsubstantiated

88 63 64 42 58 36

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B Roof runoff

Runoff from roof surfaces constitutes a significant fraction of the total sealed surface runoffand is often regarded as unpolluted. Assessment of roof runoff quality found in literaturegives rather contradictory results. Some authors conclude that rain runoff from roof surfacesis polluted (e.g. Zillich 1991; Good 1993); others found that there is a low pollution potentialassociated with roof runoff (Shinoda 1990). For the specific case of potentially toxicelements pollution however, the literature seems to agree that roof runoff can be at least aspolluted as road runoff (Herrmann et al., 1994; Förster, 1999). The pollution effect is muchgreater when the source of pollutants is the roof material itself. Förster (1999) found that zincconcentrations in runoff from zinc sheet roof were actually two or three orders of magnitudeabove those measured in runoff from roofs without any metal components (i.e. fibrouscement roof).

However, even in the case of normal rooftops (non-metal dominated) there is aconcentration of potentially toxic elements to be expected due to various metal components(gutters, downspouts, fittings etc.). The main pollutants in this case are zinc and copper. Therunoff rate of zinc was proven to be considerably lower than its corrosion rate, varyingbetween 50±90% for zinc and 20±50% for copper during exposures up to five and two years,respectively (Wallinder et al., 2000). Similar to its corrosion rate, runoff rates of zinc arestrongly related to the atmospheric SO2 concentration and are, as such, different for a rural,compared to an urban or an industrial, environment. Observed lead pollution from roof runoff(which can be considered significant in many cases compared to other sources) can beexplained by the use of lead in window frames and rooftops (the case of slate roofs in Figure2.5) and the use of lead sheeting, particularly in the UK. The lead emission factor from leadflashing and roofing in Denmark is estimated at 5.10-4 kg kg-1year-1 (Jøergensen andWillems, 1987]. The presence of cadmium can be explained by the fact that cadmium is aminor contaminant of zinc products.

A study in Calais, France, and its surrounding area found that concentrations of cadmium,mercury, and lead in roof sludge were: 34.5 mg kg-1, 4.4 mg kg-1, and 4mg kg-1 respectivelyin the urban areas [ADEME, 1997]. A study carried out in Nancy in 1996 analysed wetweather runoff and determined the main metal-contributing sources. For zinc, roof runoff wasthe largest source as it contributed 72%, and road runoff only contributed 12%. For lead, themain source was again roof runoff with 33% contribution and then road runoff with 32%[Autugelle et al., AGHTM, 1996]. However, these data may not be typical, as they representone storm event.

Relationships between rooftop material and runoff pollutant load can be observed in Figure2.5 and Table 2.27

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Figure 2.5 and Table 2.27 Rooftop material and runoff pollutant load [after Gromaire-Mertz et al., 1999]

Roof Name CoveringMaterial

GutterMaterial

Tile 1 InterlockingClay Tiles

Zinc

Tile 2 Flat clay tiles(70%)+ zincsheets

Copper

Zinc Zinc Sheets Zinc

Slate Slate Zinc

Table 2.28 provides some indicative ranges of potentially toxic element pollutionconcentrations from roof runoff. For comparison purposes the ranges of concentrations ofthe same pollutants from other sources in the same study, are also included. The medianvalue is used, instead of the mean, to filter out the effect of isolated extreme events.

Table 2.28 Potentially toxic element pollution concentrations from roof runoff[after Gromaire – Mertz et al., 1999]

Roof runoff Yard runoff Street runoffMin. Max. Median Min. Max. Median Min. Max. Median

Cd (µg l-1) 0.1 32 1.3 0.2 1.3 0.8 0.3 1.8 0.6

Cu (µg l-1) 3 247 37 13 50 23 27 191 61

Pb (µg l-1) 16 2764 493 49 225 107 71 523 133

Zn (µg l-1) 802 38061 3422 57 1359 563 246 3839 550

Lead pollution from roof degradation is strongly particle-bound (measurements indicatedmedian values of 87%), whereas the distribution between particle and dissolved pollutionfluctuates for the rest of the potentially toxic elements. Although solids are the main vector ofpollution in street and yard runoff, literature agrees that in the case of roof runoff thedissolved phase is of primary importance [Gromaire – Mertz et al., 1999; Förster, 1996;Förster, 1999; Wallinder et al., 2000]. The very high concentrations of dissolved potentiallytoxic elements in roof runoff make its infiltration hazardous and highly dependent on roofmaterials. Low settling velocities for roof runoff particles in conjunction with high percentageof dissolved pollutants make settling alone an insufficient technique for treatment.Uncontrolled local infiltration practices (infiltration trenches etc) may lead to serious local soilcontamination and present a threat to groundwater quality.

In conclusion it can be said that roof runoff pollution is influenced by roof material, airpollution, the precipitation event (intensity, antecedent dry period), the meteorology (season,wind speed and direction) and the pollutants’ physico-chemical properties. It can undercertain conditions by far exceed threshold toxicity values. The following suggestions(Förster, 1999) summarise a number of research areas that should be considered tominimise roof runoff pollution.

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• First flush diversion valves and their automated control• Sorbents for potentially toxic elements• Durable coating for potentially toxic element surfaces• Alternative materials for gutters and downpipes (eg plastics or carbon fibre based

materials, not metals or PVC)• Roof runoff quality database sufficient for predictive modelling.

C Construction and building maintenance

Contaminants from paints:While lead concentrations in consumer products (i.e. petrol) continue to decrease, thereseems to be enough residual material from historic lead use to cause high leadconcentrations in the environment. In a recent study by Davis and Burns (1998) in the US,lead runoff from painted structures in an urban setting was assessed. Although constructionpractices in the US can be considered different from those in EU, the conclusions of theDavis and Burns (1998) study should be taken into account due to the variability of thestructures investigated. In many cases, high lead concentrations were found. Leadconcentrations (100 ml over 1600 cm) from washes of 169 different structures followed theorder (geometric mean, median, Q10±Q90): wood (40, 49, 2.6±380 mg.l-1)>brick (22, 16,3.3±240 mg.l-1)>block (9.7, 8.0, <2±110 mg.l-1). Lead concentration depended strongly onpaint age and condition. Lead levels from washes of older paints were much higher thanfrom freshly painted surfaces, which were demonstrated quantitatively as: paint age [>10 y](77, 88, 6.9±590 mg.l-1)>>[5±10 y] (22, 16, <2±240 mg.l-1)>[0±5 y] (8.4, 8.1, <2±64 mg.l-1).Lead from surface washes was found to be 70% or greater in particulate lead form,suggesting the release of lead pigments from weathered paints. High intensity washes werefound to liberate more particulate lead than lower intensities. It can be concluded that oldsurface paints can contribute high masses of lead into a watershed, targeting thesestructures for source preventive actions to curtail future lead input into the environment.

Contaminants from concrete leaching:Concrete is one of the main materials used in building and road construction and issystematically in contact with precipitation, much of which ends up in the UWW collectionsystems. Recent work (PCA, 1992) identified As, Be, Cd, Cr, Hg, Ni, Pb, Sb, Se and Th inconcrete in detectable concentrations. Hillier et al. (1999) discuss the importance of theconcentrations of these pollutants in leachate from Portland cement. They concluded thatleaching of well-cured Portland cement produces undetectable concentrations of toxicmetals (such as the ones outlined in 80/778/EEC for water fit for human consumption). Inpoorly cured (1 day) Portland cement there were detectable concentrations of vanadium(reaching concentrations of 61.7 ppb). However, even this cannot be considered verysignificant, as the leaching was restricted to the surface of the samples. Furthermore, thewater-to-cement ratio has no significant impact on the leaching potential of the cement. Thisstudy suggests that concrete leaching is not a major source of metals to UWW.

D Wet and dry deposition

Rainwater can add its own absorbed and dissolved pollutants to the loads generated fromother sources. This was shown in Figure 2.3, which shows the initial potentially toxic elementloading of precipitation and the subsequent additional loading from roof, pavement and roadsurfaces. Both traffic density and location of industry have a strong influence on thedeposition of potentially toxic elements in precipitation.

The composition of atmospheric loaded precipitation water in urban regions in Germany,contaminated with atmospheric wash-out, is represented in Table 2.29.

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Table 2.29 Composition of atmospheric loaded run-off water in German urban regions[Freitag et al. 1987, Göttle 1988, Hahn 1995].

PTE Values Munich(Pullach/ Harlaching)

Mean value[mg l-1]

Extreme values[mg l-1]

1988[mg l-1]

Zn 0.05-0.15 0.02-1.9 0.0945Cu 0.007-0.2 <0.06 0.0355Pb 0.03-0.11 <0.24 0.0121Cd 0.001-0.003 <0.13 0.0014Mn 0.05 <0.1 -Cr 0.002 <0.08 -

Currently, vehicle traffic, steel and glass production, and combustion processes representthe main sources of lead in the atmosphere. Data by ADEME studies have shown thatatmospheric fallout contributes 87-536 g ha-1year-1 of lead, and is particularly high in urbanareas. These studies determined that 70-80% of lead present in UWW comes from runoff,15-20% from commercial sources and 5% from domestic sources. The main cadmiumsources are from combustion processes (vehicle traffic, waste combustion in incinerationplants). Combustion processes and chlorine and steel production represent the mainsources of the atmospheric mercury emissions in the Central Region. Table 2.30 shows thereduction in Cd, Hg and Pb emissions in Austria between 1985 and 1995 as a result oflegislative controls.

Table 2.30 Emissions and predicted yearly emissions of potentially toxic elements inAustria, 1985-2010 [UBA, 1999].

PTEs 1985[t y-1]

1990[t y-1]

1995[t y-1]

Prediction2005[t y-1]

Prediction2010[t y-1]

Cadmium 4.80 3.10 1.80 1.20 1.30Mercury 4.30 2.70 1.50 1.20 1.10Lead 320.00 202.00 39.00 24.00 21.00

The presence of impermeable surfaces in urban areas prevents natural filtration processesin the soil zone from removing the pollution. The actual quality of the rain can varydependent on pollution sources other than urban traffic. In coastal areas the increasedconcentrations of sodium and chloride in precipitation will increase the susceptibility ofvehicles to corrosion. A road drainage study in Sweden noted precipitation pH of between3.8 and 4.9 [Morrison et al. 1988].

Atmospheric sources contribute significantly to the mass of contaminant available on animpervious urban surface for transport by surface runoff. This deposition may occur during astorm event or as dust fall during dry periods. Cattell and White (1989) [as reported in Ball etal. 2000] in their study in Sydney reported that the geometric mean of the total phosphorousconcentration is 29 µg l-1. If the mean annual precipitation for Sydney of approximately 1600mm is considered, then the likely annual phosphorus load is approximately 0.5kg ha-1y-1,which is more than half the likely annual phosphorus load from an urban catchment of 0.7kgha-1y-1 [Lawrence and Goyen, 1987] to 1.1kg ha-1y-1 [Cullen, 1995].

The second of the two atmospheric sources of contamination is dry deposition in the periodbetween storm events. For dry deposition, removal by stormwater runoff is not the solemechanism responsible for depletion of the contaminant store, which is developed on thecatchment surface. As well as removal by storm events, removal also can occur throughstreet sweeping, through the local turbulence arising from the motion of vehicles on roads,and through wind events where the wind has sufficient capacity to entrain the contaminantand to move it from the surface of the urban catchment [Ball et al., 2000].

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Table 2.31 Dry Weather Build-up of Contaminants on Road Surfaces[after Ball et al., 1998]

Constituent Load (mg/m of gutter)Constituent Minimum Maximum MeanPb 1.4 7.4 3.7Zn 0.7 3.9 1.8Cu 0.3 1.8 0.9Cr 0.02 0.61 0.14P 0.23 1.9 0.8

Dry atmospheric fallout can be responsible for large proportions of road dust but estimatesvary according to atmospheric and weather patterns, as well as with different sampling andestimation methods. Nevertheless, atmospheric deposition can be very significant proportionof the total particulate/sediment input to a highway and since much of it is inert, inorganicand often calcareous in nature (depending upon local geology and building stone), it canprove beneficial in neutralising acid emissions, reducing potentially toxic element solubilityand assisting in the binding of organic and other pollutants [Luker and Montague, 1994].

Snowfall has an increased pollutant scavenging efficiency because of the large snowflakesurface area and the increased ionic strength of the melt water provides enhanced metalexchange capabilities. Trees in urban settings are also effective scavengers of metals, whichare deposited on surfaces at leaf fall.

Atmospheric contaminants are deposited during the early stages of a precipitation event andtherefore the resulting impermeable surface loading tends to be independent of bothprecipitation volume and intensity, with the pollutant load tending to increase with the lengthof antecedent dry period and with local traffic density.

E Sewer cleaningSewer cleaning is carried out for a number of reasons, but principally to deal with blockagesor to remove sediment in order to restore hydraulic capacity and limit pollutant accumulation.A number of cleaning techniques and methods is in use, depending particularly on locationand severity, including rodding, winching, jetting, flushing and hand excavation (Lester &Gale, 1979). A combination of more than one method may well be used in any particularlocality.

The frequency with which sewers are cleaned will depend on the maintenance strategy inoperation. If reactive maintenance is used, problems are dealt with on a corrective basis asthey arise (i.e. after failure). This approach will always be required to a certain extent, asproblems and emergencies occur from time-to-time in every system. However, its frequencycannot be predicted. In planned maintenance, potential problems are dealt with prior tofailure. Unlike reactive maintenance, planned maintenance is proactive and has the objectiveof reducing the frequency or risk of failure (Butler & Davies, 2000). Planned maintenancediffers from routine maintenance (operations at standard intervals, regardless of need), andinvolves identifying elements that require maintenance and then determining the optimumfrequency of attention. Frequency of cleaning will thus vary from place to place but is usuallyless than once per year even in heavily affected sewers.

Sediment removed from sewers during cleaning is normally classified as a controlled waste,and is therefore disposed of to licensed landfill sites.

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2.2 Influence of various treatment processes on the fate of potentially toxic elementsthrough WWTS and SST

PTE transfers to sewage sludgeSludges 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. Thus, nickel 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 2.32 suggests that secondary treatment (by the activated sludgeprocess) is more efficient at removing Cr than the primary stage. Operational experience andmetal 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 some of the Hg removal observed in activated sludge may be attributed to bacteriallymediated volatilisation, it is unlikely that this is a major route of Hg loss because of thesignificant quantities of Hg recovered in surplus activated sludge (Lester, 1981).

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Table 2.32 PTE removals and transfer to sewage sludge during conventionalurban wastewater 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)

Predicting the fate of potentially toxic elements during WWT and sludge treatmentChemical contaminants present in UWW are associated principally with the organic andmineral fractions in sewage. The microbial biomass in secondary treatment is also effectiveat scavenging potentially toxic elements from the settled sewage transferring metals to theactivated sludge. About 60 - 80 % of the wastewater load of potentially toxic elements suchas Cu and Pb is transferred into sludge (Table 2.32) and the sludge contains approximately1000 times (mg kg-1 ds basis) the concentration of these metals present in the wastewater(mg l -1). The relationship between metal concentrations in wastewater and in sludge can bedetermined empirically as follows (after Blake 1979):

S = P / ((Ws/1x106) x (100/T)) mg kg-1 ds

Where P = Concentration of contaminant in wastewater mg l-1

S = Concentration of contaminant in sludge mg kg-1 dsWs = Dry solids content of wastewater mg l-1

T = Transfer efficiency of contaminant from wastewater to sludge during sewage treatment %

Ds = Dry substance

For example, a wastewater containing 1.3 mg l-1 of Zn (P) and 450 mg l-1 of dry solids (Ws)would produce sludge with a Zn concentration of 2000 mg kg-1, assuming a transfer factor of70 % (T):

S = 1.3 / ((450/1x106) x (100/70)) mg kg-1

= 2000 mg kg-1

This relationship can also be used to estimate the concentrations of metals in the rawsewage effluent, based on analytical data for the metal content in sludge and the dry solidsin the raw wastewater:

P = S x ((Ws/1x106) x (100/T)) mg l-1

The concentration in the final effluent, F mg l-1, can therefore be estimated from:

F = (1-T/100) x (S x ((Ws/1x106) x (100/T))) mg l-1

Based on the above assumptions, the Zn content in the final effluent in this example isestimated to be:

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F = (1-70/100) x (2000 x ((450/1x106) x (100/70))) mg l-1

= 0.4 mg l-1

A more mechanistic approach to predicting the fate of metals during wastewater treatment isdescribed by Monteith et al. (1993) using a computer-based mathematical model. Themechanistic model, TOXCHEM (Bell et al., 1989), includes both steady state and dynamicmodels to simulate operation of a conventional activated sludge plant, incorporating gritremoval, primary clarification, aeration and secondary clarification unit processes.Biodegradation, surface volatilisation, air stripping, precipitation and sorption removalmechanisms are determined for trace metallic and organic contaminants in the model’sdatabase. A sensitivity analysis feature allows the user to investigate the impact of designand operating conditions and the contaminant’s physical/chemical properties on removal byunit process and by mechanism.

Inputs to the TOXCHEM model include influent wastewater flow rate, influent metalsconcentrations, dimensions of individual process units, plant operation conditions (e.g. mixedliquor suspended solids concentration), and raw wastewater, primary effluent, andsecondary effluent suspended solids concentrations. The model then predicts metalconcentrations in primary sludge, return activated sludge, surplus activated sludge, andsecondary settler effluent, based on mass balance calculations and removals throughprecipitation and sorption. The steady-state model accounts for removal by sorption ofsoluble metals onto settleable solids and precipitation of the metals into a settleable form.The following assumptions were used to develop the model equations:

1, The system is at equilibrium with regard to sorption and desorption;2, Sorption follows a linear isotherm;3, Precipitation and dissolution are instantaneous;4, Precipitated metals are integrated into the biomass and are removed at the same

efficiency as the bulk solids during primary and secondary settlement.

In both the primary settlement and aeration tanks, the model calculates the metalconcentrations in the soluble and solid phases:

Ct = Cs + Cx

Cx = Cp + KpXCs

Where:

Ct = total metal concentration, mg l-1

Cs = soluble metal concentration, mg l-1

Cx = solid-phase metal concentration, mg l-1

Cp = concentration of precipitated metal, mg l-1

Kp = linear sorption coefficient, l g-1

X = mixed liquor volatile suspended solids (MLVSS) concentration, g l-1

The values of Kp and metal solubilities are experimentally defined. Solubility was determinedfrom dosing studies in which metal concentrations were measured in filtered and unfilteredwastewater samples at equilibrium.The soluble, sorbed and precipitated metal fractions are calculated by the expression:

Ctest = CT,0/(1 + KP.1Xo)Where

CT,0 = influent total metal concentrationKP.1 = primary clarifier sorption coefficientXo = influent volatile suspended solids (VSS) concentration.

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Fractional removal of metals in the primary and secondary settlers is calculated by:FRm = (FRsolidsCx)/CT,0

WhereFRm = fractional removal of metal in settler, andFRsolid = fractional removal of solids in settler.

Fractional removal of the solids is input by the user based on site-specific experience. Theoutput for the primary and secondary settlers is calculated by:

Ct,out = (1-FRm) CT,0

Where Ct,out = total outlet metal concentration.

The TOXCHEM model was validated against operational data on effluent metal contentscollected from a full-scale wastewater treatment plant (Table 2.33). The predictions werecompared to observed values in the final effluent using linear regression analysis and the r2

statistic. Copper exhibited the highest correlation between predicted and observed effluentconcentrations of the elements tested and the r2 value for this metal was 0.96. The modelpredicted that 70 % of this element would be transferred to the sewage sludge. Zinc alsogave relatively good agreement between predicted and observed effluent concentrations andthe results showed that 50 – 60 % of Zn in wastewater would be recovered in the sludge.The poorest correlation of modelled and actual values was obtained for Ni and the r2 was0.41 in this case. The model underestimated the removal of Ni by the wastewater treatmentplant and the predicted value was 26 % compared with the observed removal of 37 %. Leadand Cr have low solubilities and in both cases, the TOXCHEM model predicted much largerremovals of approximately 70 % for these metals than was observed in practice at thesewage treatment works. However, the model predictions for Pb and Cr were comparablewith operational experience at other wastewater treatment plant (Table 2.32). The slopes ofthe regressions lines fell within the range 0.58 – 1.96 bracketing the ideal slope of 1.0. Themodel provided good representations of the concentrations of Cu, Ni and Zn in the finaleffluent, but was less satisfactory for Pb or Cr. Cadmium was not subject to the verificationexercise because concentrations in the effluent were below the analytical limit of detection.The modelling and experimental studies showed that Cd and Ni were the most solublemetals, and the most poorly removed, while Pb and Cr were the least soluble. The majorityof metals, with the exception of Ni, were maintained in the mixed liquor, and the masscollected in effluent was small in comparison to the total influent mass. At low metalconcentrations (0.02 – 0.1 mg l-1), the soluble fraction is controlled by sorption to solids. Atlarger influent metal concentrations, precipitation is the main process controlling solublemetal fractions. In consequence of this differential fate of metals in WWTS, nickel may needtighter regulation at source to reduce the amount in the effluent.

The mechanistic approach to model development incorporates the main chemical andphysical processes that are recognised as being important in removing metals fromwastewater. However, in some circumstances, there may be specific properties of awastewater at a particular treatment works that influence metal behaviour and removalefficiency that are not accounted for by these basic assumptions. For example, recoveries ofinsoluble elements such as Cr or Pb in sewage sludge are usually relatively high and of theorder >75 % (Table 2.33). This contrasts with the recoveries of these elements measured byMonteith et al. (1993), which were much lower than those normally observed in practice.Operational and mechanistic models have the capability to predict partitioning anddistribution of potentially toxic elements from raw sewage during conventional wastewatertreatment. They provide insights into the physico-chemical processes controlling metalremoval in terms of solubility, sorption and precipitation mechanisms, and calculate the massbalance and partitioning of metals into sewage sludge and the final effluent. The metalconcentration in sludge can be calculated using the estimated partitioning coefficients andthe mass of primary and biological sludge produced by primary sedimentation andsecondary wastewater treatment.

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Table 2.33 Predicted metal concentrations in sewage effluents estimated by theTOXCHEM model relative to observed values at a full-scale sewage treatment plant(adapted from Monteith et al., 1993)

PTE Mean concentration (mg l-1) Removal (%)Influent Observed

effluentPredictedeffluent

r2 forpredicted

vs.observed

Observed Predicted

Zn 0.120 0.051 0.057 0.61 58 53Cu 0.120 0.041 0.034 0.96 66 72Ni 0.027 0.017 0.020 0.41 37 26Pb 0.069 0.051 0.022 0.66 26 68Cr 0.026 0.016 0.007 0.63 39 73

Effect of stabilization processes on metal concentrations in sewage sludgePotentially toxic element concentrations in sewage sludge are influenced significantly by thetype of stabilization process operated at a WWTS. For example, the destruction of volatilesolids by microbial decomposition of putrescible organic matter in sludge by mesophilicanaerobic digestion or aerobic composting processes increase the metal content in directrelation to the extent of volatile solids removal by microbial decomposition.

Unstabilised, co-settled primary + activated sludge typically contains 75 % of volatile matteron a dry solids basis. During anaerobic digestion, 40 % of the volatile matter is destroyedreducing the volatile solids concentration in digested sludge by about 50 % compared to theundigested product. Potentially toxic elements are conserved and are retained in the sludgeduring the digestion process. As a consequence of the microbial decomposition of organicmatter, metal concentrations increase in direct proportion to the loss of solids and aretypically raised by approximately 40 % compared with undigested sludge.

Volatile matter is also lost during aerobic composting of sewage sludge (Figure 2.6).Composting dewatered sewage sludge cake requires a bulking agent, such as straw orwoodchips, to increase porosity and reduce moisture content to values that will supportaerobic microbial decomposition processes. Straw is often used for this purpose and may bemixed with sludge cake at 25 % ds at a rate of 5 – 10 % on a fresh weight basis (Smith andHall, 1991). The addition of bulking agent to sludge reduces the initial metal concentration.However, metal concentrations in the composting material increase proportionally in linearrelation to the extent of volatile matter destruction by microbial activity. In windrowcomposting trials (Smith and Hall, 1991), for example, the volatile solids concentration ofcomposting sewage sludge and straw decreased by 21 % during a period of 100 days(Figure 2.6). During that period, the metal concentration in the composted product wasincreased above the value in the original sludge (Figure 2.7).

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Figure 2.6 Volatile solids content of sludge-straw compost at different rates of strawaddition (fresh weight) to sludge cake (25 % ds) over time (Smith and Hall, 1991)

Sludge stabilisation processes reduce the fermentability of the sludge and the potentialodour nuisance and health hazards associated with its use on agricultural land. However,the loss of volatile matter during microbial decomposition also significantly increases itsmetal content by as much as 40 % compared with material that has not been subject to abiological stabilisation process. More than 75 % of the sludge produced in the EU is treatedby anaerobic or aerobic digestion. Despite the wide adoption of biological stabilisationprocesses for sludge throughout the EU, which generally increase the metal content ofsludge, reported potentially toxic element concentrations in sewage sludge continue todecline (e.g. Figure 2.7).

This further emphasizes the major and significant improvements in sludge quality, in terms ofpotentially toxic element content, that have been achieved in the past 20 years throughimproved industrial practices and the successful and effective implementation of tradeeffluent controls.

50

55

60

65

70

75

0 20 40 60 80 100 120

5.0%7.0%8.5%

Time (days)

VolatileSolidsContent (%)

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Figure 2.7 (a) Cu and (b) Zn concentrations in sewage sludge-straw compost inrelation to volatile solids content (Smith and Hall, 1991)

500

600700

800900

1000

11001200

1300

50 55 60 65 70 75

Cu

(mg

kg-1 d

s)

0200400600800

1000120014001600

50 55 60 65 70 75

Zn

(mg

kg-1 d

s)

Volatile Solids Content (%)

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Figure 2.8 Reduction in (a) zinc and (b) cadmium concentrations (untransformed andlog10 transformed data, respectively) in sewage sludge from Nottingham STW, UKduring the period 1978 – 1999

Tot

al Z

n (m

g kg

-1 d

s)

400

600

800

1000

1200

1400

1600

1800

2000

1978 1983 1988 1993 1998

(a) Zinc (b) Cadmium

Tot

al C

d (L

og m

g kg

-1 d

s)

0

0.2

0.4

0.6

0.8

1

1.2

1.4

1.6

1.8

2

1978 1983 1988 1993 1998

YearYear

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2.3 Quantitative assessment of potentially toxic elements in untreated UWW, treatedUWW and treated SS

There is limited information available on pollutant concentrations in influent and effluent fromWWTS. Concentrations may also be less than the detection limit of the analytical techniquesused. It must be noted that the variability of influent concentrations within a WWTS can bevery high due to season, precipitation, and levels of industrial and domestic activities. Thismeans that comparisons based on small numbers of samples must be made with caution.

Table 2.34 Concentrations of potentially toxic elements found in influents andeffluents of WWTS

WWTSPTE Country DomesticWastewater

( g l-1)

UrbanRunoff( g l-1) Influent

( g l-1)Effluent( g l-1)

Reference

Austria 30 (<20-60) 70 (<20-170)

Hohenblum et al.,2000

France0.2-4.2

6-85 ADEME, 1995Legret,1999

Germany 0.5<5

80* 0.4 0.1 Raach et al., 1999Wilderer et al 1997

Greece <1 <1 Greek report, 2000Italy:CentralNorthern

< 5n.a

Braguglia, et.al., 2000

Cd

Sweden* 200-800 100(domestic)

Adamsson et al 1998

Austria 120 (90-130)

120 (<70-190)

Hohenblum et al.,2000

Germany 150 Wilderer et al 1997Greece <100 <100 Greek report, 2000Sweden* 350000-

80000050000

(domestic)Adamsson et al 1998

Italy:CentralNorthern

20-9002-25

Braguglia, et.al., 2000

Cu

France 0.011-0.146

Legret,1999

Austria 7100(6200-7900)

3900(<900-5600)

Hohenblum et al.,2000

Germany 30 Wilderer et al 1997Italy:CentralNorthern

< 200.5-32

Braguglia, et.al., 2000

Cr

Sweden* 4000-14000 3000(domestic)

Adamsson et al 1998

(Table 2.34 continued)Austria <10 <10 Hohenblum et al., 2000France 1-8 ADEME, 1995Germany 0.4 1* 0.6 0.1 Raach et al., 1999Greece <1-3 <1-9 Greek report, 2000

Hg

Italy:CentralNorthern

< 1n.a

Braguglia, et.al., 2000

Pb Austria 30 (<20-60) 70 (<20-160)

Hohenblum et al., 2000

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France 14-188 51-630 ADEME, 1995Legret,1999

Germany 100100

110* 81 31 Raach et al., 1999Wilderer et al 1997

Greece <1-5000 <1-17 Greek report, 2000Italy:CentralNorthern

< 200.3-9

Braguglia, et.al., 2000

Spain 10-64 Cabrera, et.al,1995Sweden* 4000-23000 2000

(domestic)

Adamsson et al 1998

Austria <40-170 240(<40-620)

Hohenblum et al., 2000

Germany 40 Wilderer et al 1997Italy:CentralNorthern

< 500.5-95

Braguglia, et.al., 2000

Ni

Sweden* 3000-10000 3000-5000

(domestic)

Adamsson* et al 1998

Austria 1000(<20-3700)

2500(20-

5000)

Hohenblum et al., 2000

France 104-1544 Legret,1999Germany 100-1000 Wilderer et al 1997Greece 450-3200 <20-900 Greek report, 2000Italy:CentralNorthern

100-90012-185

Braguglia, et.al., 2000

Zn

*Sweden 150000-1300000

50000(domesti

c)

Adamsson et al 1998

As Spain 2.2 Navarro, et.al. 1993Se Spain:

Se (IV)Se (VI)

0.40.3

Diaz, et.al., 1996

Fe France 60-999000 ADEME, 1995*The data from Sweden are from a very small aquaculture wastewater plant, which only serves a fewhouses and may not be typical of other WWTS.

Organotin compounds were the subject of monitoring in Switzerland by Fent and Müller[1991], and Fent [1996]. Monobutyltin (MBT), dibutyltin (DBT), and tributyltin (TBT) weremonitored in raw wastewater (influent), effluent and sewage sludge (Table 2.35). Phenyltincompounds were also investigated.

Table 2.35 Organotin compounds in Zürich WWTS

Unit MBT DBT TBTRaw wastewater ng l-1 140 to 560 130 to 1,030 60 to 220Digested sludge mg kg-1 ds 0.3 to 0.8 0.5 to 1.0 0.3 to 1.0

The potentially toxic element loading per inhabitant, per year has been determined in twoGerman studies (Table 2.36). There is a good correlation between these studies for zinc,copper, lead and cadmium concentrations in the influent and effluent of WWTS.

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Table 2.36 Potentially toxic element loading per member of population in two Germantowns [Raach et al.,1999; DAF, 1995]

Flow Element Raach et al.,1999[g inhabitant-1 a-1)]

DAF, 1995[g inhabitant-1 a-1)]

Zn 69 75Cu 19 28Pb 14 32

Influent

Cd 0.1 0.5Zn 26 39Cu 6.3 6.9Pb 5.1 5.0

Effluent

Cd 0.03 0.07

In Norway and Sweden; emissions of potentially toxic elements from WWTS serving morethan 20,000 persons account for almost 80 percent of the wastewater. The discharges ofmetals from these WWTS in Sweden and Norway are summarised in Table 2.37. This showsthat, for Sweden, the amount of potentially toxic elements discharged with urban wastewaterdecreased over the period 1992 to 1998 for all metals, except for copper. The statistical datado not take into account the WWTS size and also the potentially toxic elements reported arefor the treated sewage sludge. There is only limited data available on potentially toxicelement content in the treated wastewater (effluent), due to the low concentration, oftenbelow detection limits.

Table 2.37 Emissions of potentially toxic elements from WWTS in Sweden and Norway1998

Sweden1992

Sweden1998

Norway1998

Element

(kg.annum-1)Cadmium 325 137 150Chromium 5420 3308 3000Copper 14060 15377 15000Mercury 530* 304 300Nickel 8165 7603 8000Lead 2960 1464 1500Zinc 37420 32346 32000

* data for 1995 [Statistika centralbyrån Sweden, 1998; SFT-1999]

Power et al. [1999] monitored water in the Thames, UK and found statistically significantreductions in the concentrations of Cd, Cu, Hg, Ni and Zn over the period 1980 to 1997. Forlead, the initial improvements were reversed by drought in the period 1990-1997 resulting ina slight, though statistically significant rise in Pb concentrations. However; Pb had fallen priorto this and was still found to have a statistically significant decrease overall in the period1980-1996 (Table 2.38).

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Table 2.38 Achieved concentration reductions of potentially toxic elements in watersin the Thames estuary, compared with other European estuaries

PTE Thames1986/7µg.l-1

Thames1995µg.l-1

%reduction

Elbe1983µg.l-1

Rhine1984µg.l-1

Humber1984µg.l-1

Severn1988µg.l-1

Cd 0.43 0.32 24.1 0.10 0.30 0.31 0.25Cu 31.30 10.70 65.8 2.00 5.50 2.17 5.00Hg 0.24 0.09 63.0 - - - -Ni 17.30 6.30 63.4 3.25 3.20 - 4.50Pb 16.30 9.90 39.4 1.90 - 1.24 <0.03Zn 92.00 29.10 68.4 - 21.80 - 17.50

[from Power et al 1999, data included from *Mart 1985, *Nolting 1986, *Balls 1985, *Apte 1990]

The overall reductions in most metal concentrations in the Thames was greater than 50percent over the years 1986-1995, with lower reductions being experienced for lead andcadmium. This complements evidence for sludge quality that overall PTE emissions havemarkedly improved.

In spite of the reductions achieved by 1995 in the River Thames, potentially toxic elementconcentrations are still high compared to the levels found in the 1980s in the other rivers inthe study. It is concluded that while much progress has been made in reducing theanthropogenic sources of potentially toxic element pollution discharged into the Thames,improvements are still needed if water is to approximate to background levels. For cadmiumand mercury; the year 2000 levels are forecast to be about twice estimated backgroundlevels [Power et al 1999].

A study carried out in the Rhine region in France [Commission Internationale pour laProtection du Rhin, 1999] examined all pollutant sources entering the river over a period of10 years. The study showed that the decrease in certain pollutants was due to pre-treatment, cleaner technologies and more care in the handling of the priority substances.Nevertheless, in 1996, the study showed that diffuse and domestic (communal) sources areimportant contributors, particularly storm runoff for mercury, lead, and copper.

The study further divided the sources for each metal and other pollutants into greater detailfor each of the riverine countries in the region (France, Switzerland, Belgium, Luxembourg,Netherlands, Germany). An estimate into the source breakdown of metal pollution into theRhine is included in Figure 2.9.

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0%

10%

20%

30%

40%

50%

60%

70%

80%

90%

100%

Hg Cd Cu Zn Pb Cr Ni

Percentage contribution of heavy metals for the different sources in the Rhine

Diffuse

Communal

Industrial

Figure 2.9 The percentage of potentially toxic elements in the Rhine from differentsources (France, 1999)

The communal sources of wastewater into the river Rhine (in France), include wastewaterfrom the UWW collecting system, hence showing the widespread presence of thesepollutants (Table 2.39). For mercury, lead, and copper more than half is due to storm runoff.Erosion and drainage are also important sources, contributing 18% of the mercury and 55%of the chromium. Atmospheric deposition generally tends to contribute around 8% of the totaldiffuse metal contribution, particularly for cadmium and zinc.

Table 2.39 : Potentially toxic element pollution contribution by different sources in kga-1. [from Commission Internationale pour la Protection du Rhin, 1999]

Hg Cd Cr Cu Zn Pb NiDiffuse sources

kg.annum-1

Erosion 18 53 5144 3547 10642 4789 5853Surface runoff 1 10 474 229 1013 67 62Drainage 7 286 429 2145 28600 2145 1430Atmospheric deposition 8 75 189 1131 9425 1508 566Separating system 8 57 475 1900 7600 1520 855Storm discharge 32 126 630 3780 17640 3780 1890Untreated wastewater 16 62 310 1860 8680 1860 930Houses not connected toUWW collecting systems

2 7 36 216 1008 216 108

Point sourcesIndustry 74 242 5548 11190 41100 3120 14300

Communal sources 40 200 1400 12470 30000 2700 3500

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A comparison of potentially toxic element concentrations in sewage sludge applied tofarmland in different countries within the EU (Table 2.40) indicates there is some variationapparent in the metal contents of sludges used in agriculture. For example the reportedaverage concentrations of Cd in German and UK sludges in 1996 were 1.5 and 3.5 mg kg-1,respectively (CEC, 2000a). This could indicate differences in the amount of Cd discharged tosewer in these countries from industrial, domestic and diffuse inputs or the adoption ofdifferent sludge treatment practices and regulatory procedures influencing metal content.However, these issues are difficult to reconcile given the policies on preventing industrialdischarges of Cd followed in both countries and that both states practice extensive sludgestabilisation treatment. A possible explanation may be related to the statistical characteristicsof metal concentration data and how data on metal contents in sludge are reported. Forexample, CEC (2000a) does not state whether arithmetic averages or weighted averagesare given for metal contents. The UK figures are weighted according to works size andprovide a conservative estimate of sludge metal content because sludge from largetreatment works usually have larger metal concentrations compared with smaller works (EA,1999). For example, the median Cd concentration in sludge used on to farmland from largeworks (2.9 mg Cd kg-1, pe>150000) in the UK was more than twice that from small works(1.3 mg Cd kg-1, pe<10000) in 1996/97. This trend could be interpreted as being the result ofhigher industrial inputs of metals to the large works, although it may also be explained by thegreater interception of atmospheric deposition of metals by paved areas in urban centresthat are served by the largest sewage treatment works. The majority of sludge recycled toland in the UK is produced by 55 large works (160,000 t ds y-1) whereas approximately 840small works generate sludge (45,000 t ds y-1) for agricultural use. Therefore, an arithmeticmean would indicate a significantly lower concentration was apparent for sludge because allworks would have equal weighting. Indeed, the median concentrations recorded for UKsludge are comparable to the values reported for Germany.

Table 2.40 Comparison of potentially toxic element concentrations (mg kg-1) insewage sludge applied to agricultural land in Germany and the United Kingdom in1996 (CEC, 2000)

Potentially toxicelement

Germany(average)

United Kingdom(weighted average)

United Kingdom(median)

Zn 776 792 559Cu 305 568 373Ni 24 57 20Cd 1.45 3.3 1.6Pb 57 221 99Cr 40 157 24Hg 1.35 2.4 1.5

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Table 2.41 Survey of potentially toxic elements in sewage sludge: values in mg.kg-1

DS-Cadmium

CADMIUM Country Mean Median Min. Max. Survey Year/sAustria 1.5 1.2 0.4 3.4 1994/95 (11)Germany 1.5 1995-97 (4)Denmark 1.4 1995-97 (4)France 4.1 1995/97 (4)Finland 1.0 1995-97 (4)Greece (a) (b)

1.61.4

1996 (6)1997 (1)

Italy 0.8 23 1998/99 (10)Ireland 2.8 1997 (4)Luxembourg 3.8 1997 (4)Netherlands 3 1990 (16)Sweden 1.5 1995/96 (4)UK 3.5 1995/96 (4)Norway 0.97 1998 (12)Poland 9.93 13.5 0.8 15.3 1999 (3)EU 4.0

2.21992 (9)1994-98

USA 38.125

8.95 1988 (13)1992 (2)

Limits AgriculturalSoils

Sewage Sludge

EU 1-3 10 1* (4)WHO 7 (5)USEPA 39 (14)

Table 2.41b Survey of potentially toxic elements in sewage sludge: values in mg.kg-1

DS-ChromiumCHROMIUM Country Mean Median Min. Max. Survey Year/s

Austria 62 54 25 130 1994/95 (11)Germany 50 52 46 52 1995-97 (4)Denmark 33 34 24.8 40.3 1995-97 (4)France 69.4 58.8 80 1995/97 (4)Finland 85.7 84 82 91 1995-97 (4)Greece (a) (b)

885.843.8

1996 (6)1997 (1)

Italy 14.8 1400 1998/99 (10)Ireland 165 1997 (4)Luxembourg 51 1997 (4)Netherlands 64 1990 (16)Sweden 38.4 37.7 39 1995/96 (4)UK 159.5 157 162 1995/96 (4)Norway 28.5 1998 (12)Poland 144.2 136.5 6.8 289.0 1999 (3)EU 145

741992 (9)1994-98

USA 589178

150 1988 (13)1992 (2)

Limits AgriculturalSoils

Sewage Sludge

EU 30 –100(proposed)

1000 600* (8)

USEPA 1200 (14)

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Table 2.41c Survey of potentially toxic elements in sewage sludge: values in mg.kg-1

DS-CopperCOPPER Country Mean Median Min. Max. Survey Year/s

Austria 264 240 170 540 1994/95 (11)Germany 275 1995-97 (4)Denmark 284 1995-97 (4)France 322 1995/97 (4)Finland 288 1995-97 (4)Greece (a) (b)

302103

1996 (6)1997 (1)

Italy 160 373 1998/99 (10)Ireland 641 1997 (4)Luxembourg 206 1997 (4)Netherlands 190 1990 (16)Sweden 522 1995/96 (4)UK 562 1995/96 (4)Norway 287.1 1998 (12)Poland 237.5 183.4 24.1 592.0 1999 (3)EU 380

3651992 (9)1994-98

USA 639616

444 1988 (13)1992 (2)

Limits AgriculturalSoils

SewageSludge

EU 50-140 1000-50* (4)USEPA 1500 (14)

Table 2.41d Survey of potentially toxic elements in sewage sludge: values in mg.kg-1

DS-Mercury

MERCURY Country Mean Median Min. Max. Year/s of SurveyAustria 5.1 2.1 1.0 48.0 1994/95 (11)Germany 1.2 1995-97 (4)Denmark 1.29 1995-97 (4)France 2.85 1995/97 (4)Finland 1.4 1995-97 (4)Greece 4.1 1996 (6)Italy 0.46 5 1998/99 (10)Ireland 0.6 1997 (4)Luxembourg 1.9 1997 (4)Netherlands 1.8 1990 (16)Sweden 1.85 1995/96 (4)UK 2.50 1995/96 (4)Norway 1.34 1998 (12)

EU 2.72.0

1992 (9)1994-98

USA 3.242.3

2.3 1988 (13)1992 (2)

Limits AgriculturalSoils

SewageSludge

EU 1-1.5 10 0.5* (4)WHO 5 (15)USEPA 17 (14)

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Table 2.41e Survey of potentially toxic elements in sewage sludge: values in mg.kg-1

DS-Nickel

NICKEL Country Mean Median Min. Max. Year/s of SurveyAustria 39 35 14 94 1994/95 (11)Germany 23.3 1995-97 (4)Denmark 22.8 1995-97 (4)France 35.5 1995/97 (4)Finland 41 1995-97 (4)Greece (a) (b)

6723.6

1996 (6)1997 (1)

Italy 25 182.5 1998/99 (10)Ireland 54 1997 (4)Luxembourg 24 1997 (4)Netherlands 37 1990 (16)Sweden 19.3 1995/96 (4)UK 58.5 1995/96 (4)Norway 15.4 1998 (12)Poland 41.1 36.7 8.4 78.8 1999 (3)EU 44

331992 (9)1994-98

USA 90.671

46.5 1988 (13)1992 (2)

Limits AgriculturalSoils

Sewage Sludge

EU 30-75 300 50* (4)WHO 850 (5)USEPA 420 (14)

Table 2.41f Survey of potentially toxic elements in sewage sludge: values in mg.kg-1

DS-Lead

LEAD Country Mean Median Min. Max. Year/s of SurveyAustria 109 100 40 290 1994/95 (11)Germany 67.7 1995-97 (4)Denmark 59.9 1995-97 (4)France 119.9 1995/97 (4)Finland 43 1995-97 (4)Greece (a) (b)

283140.6

1996 (6)1997 (1)

Italy 41 560 1998/99 (10)Ireland 150 1997 (4)Luxembourg 128 1997 (4)Netherlands 145 1990 (16)Sweden 48.2 1995/96 (4)UK 221.5 1995/96 (4)Norway 21.7 1998 (12)Poland 211.8 190.6 27.7 456.5 1999 (3)EU 97 1994-98

USA 204170

152 1988 (13)1992 (2)

Limits AgriculturalSoils

Sewage Sludge

EU 50-300 750-70* (4)WHO 150 (5)USEPA 300 (14)

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Table 2.41g Survey of potentially toxic elements in sewage sludge: values in mg.kg-1

DS-Zinc

ZINC Country Mean Median Min. Max. Year/s of SurveyAustria 1188 1250 700 1700 1994/95 (11)Germany 834 1995-97 (4)Denmark 777.2 1995-97 (4)France 837.6 1995/97 (4)Finland 606 1995-97 (4)Greece (a) (b)

27521236

1996 (6)1997 (1)

Italy 391 4213 1998/99 (10)Ireland 562 1997 (4)Luxembourg 1628 1997 (4)Netherlands 1320 1990 (16)Sweden 620.5 1995/96 (4)UK 778 1995/96 (4)Norway 340 1998 (12)Poland 3641 2948 546 7961 1999 (3)EU 1000

8171992 (9)1994-98

USA 14901285

970 1988 (13)1992 (2)

Limits AgriculturalSoils

Sewage Sludge

EU 150-300 2500-150* (4)USEPA 2800 (14)

Table 2.41h Survey of potentially toxic elements in sewage sludge: values in mg.kg-1

DS-Arsenic

ARSENIC Country Mean Median Min. Max. Year/s of SurveyItaly 1.1 1.8 1998/99 (10)UK 2.5 1996/97 (7)USA 11.0

4.96.7 1988 (13)

1992 (2)

Limits AgriculturalSoils

Sewage Sludge

WHO 9 (5)USEPA 41 (14)

Table 2.41i Survey of potentially toxic elements in sewage sludge: values in mg.kg-1

DS-Selenium

SELENIUM Country Mean Median Min. Max. Year/s of SurveyUK 1.6 1996/97 (7)USA 6.14

6.04.5 1988 (13)

1992 (2)

Limits AgriculturalSoils

SewageSludge

WHO 140 (5)USEPA 36 (14)

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Table 2.41j Survey of potentially toxic elements in sewage sludge: values in mg.kg-1

DS-Silver

SILVER Country Mean Median Min. Max. Year/s of SurveyUSA 48.2 852 1988 (15)

Limits AgriculturalSoils

Sewage Sludge

WHO 3 (5)Notes: *For sludge applied to soil lower and higher limits are allowed for soil with pH in the range 5-6 and >7respectively (Commission of the European Communities (2000) working document on Sludge: 3rd draft.ENV.E.3.LM, 27 April, BrusselsEU 1992 is for B, DK, F, D, EL, IRL, I, L, NL, P, ESP, UKEU 1994-98 is derived from table values for AT, D, DK, F, FI, IRL, L, SE, UK and NO.UK and EU is sludge used in agricultureUSA is for all sludge: detection limit set at minimum levelGreece: (a) values are specific to Athens WWTS;(b) values are average of two rural WWTS.

References

Agelidis M.O. et al, 1997Bastian, R.K. 1997Bodzek, B. et al, 1999.CEC, 1999.Chang, A.G et al,Cristoulas, D.G et al, 1997.Environment Agency of England and Wales, 1999European Union, 2000Hall, J.E. et al, 1994Braguglia et al 2000Scharf, S et al, 1997SFT, 2000USEPA 1992USEPA 1993USEPA 1999Wiart, J. et al, 1995.

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In the particular case of Greece, Voutsa et al. [1996] have compared the potentially toxicelement content of municipal and industrial sludge for Thessaloniki. The results areillustrated in Figure 2.10. Sludge A is municipal sludge from Thessaloniki’s main WWTS andsludge B is from a WWTS treating partially treated industrial wastes from the greaterThessaloniki region.

Figure 2.10. Potentially toxic element concentrations in sewage sludge from biologicaltreatment of municipal and industrial wastewater (Municipal sludge: Dark, Industrialsludge: Light Grey)

As shown concentrations of most metals are 2-9 fold higher in municipal sludge than inindustrial (except for Cd and Cu where the contents are similar). This interesting fact, duepossibly to pre-treatment of the most toxic industrial wastewater on site, is not adequatelyexplained by the authors.

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3. Organic Pollutants: sources, pathways, and fate through urbanwastewater treatment systems

3.1 Sources and pathways of organic pollutants in UWW

There are a large number of organic pollutants from a wide range of sources which mayenter UWW. Paxéus (1996a) identified over 137 organic compounds in the influent of themunicipal wastewater plants in Stockholm. The physical and chemical properties of some ofthese organic pollutants are outlined in Appendix B. The main categories of organicpollutants detailed in this report are:

Polycyclic Aromatic Hydrocarbons: Polycyclic aromatic hydrocarbons (PAHs) arise fromincomplete combustion or pyrolysis of organic substances such as wood, carbon or mineraloil. Such combustion processes include food preparation in households and food shops;discharge of certain petroleum products (from garages, vehicle washing and maintenance,fuel stations); discharge of storm runoff with PAHs from car exhaust particles and roadrunoff; and also from incomplete combustion processes in urban landfills.

The most frequent anthropogenic sources of PAHs are: house fires, heat and energy powerstations, vehicle traffic, waste incineration and industrial plants (cement works, metalsmelting, aluminium production). Forest fires represent natural sources. PAHs concentrate insewage sludge due to their low biodegradability.

Polychlorinated Biphenyls (PCBs): There are two main sources of PCBs:• Directly manufactured PCBs (by chlorination of biphenyls), used as hydraulic liquids

(hydraulic oils), emollients for synthetic materials, lubricants, impregnating agents forwood and paper, flame protective substances, carrier substances for insecticides andin transformers and condensers. The EU1996 PCB Disposal Directive 96/59/ECrequires the phasing out of all PCBs by 2010 or by 1999 under internationalagreement by the North Sea States. Existing transformers and other electricalequipment which contain 50-500 mg.kg-1 PCB may be retained in service until theend of their useful life.

• The other main source of PCBs in the environment are combustion processes, fromwaste incineration plants, fossil fuel burning and to other incomplete combustionprocesses.

PCBs are adsorbed by solids and therefore they accumulate in sewage sludge. The highlysubstituted (high chlorine content) PCBs are the main representatives potentially present insewage sludge, while they amount to just 35% of the total technical PCBs. Recycling ofPCBs in the environment is very important and remediating historical pollution would benecessary if the background levels found are to be reduced.

Di-(2-ethyhexyl)phthalate (DEHP):DEHP is used as emollient in synthetic materials. In Germany, 90 % of DEHP is used inPVC and about 10% in laquers and paints. It is common to use DEHP as antifoaming agentin paper production, as an emulsifier for cosmetics, in perfumes and pesticides, they aid inthe production of different synthetic materials such as dielectric in condensers, andsubstitute for substances such as PCBs and pump oil. DEHP specific emissions fromvarious human activities have been identified by Bürgermann [1988] as follows:

• cellulose/paper production• DEHP production• plastisol-coating process• PVC production and processing, leaching from PVC products• leaching from waste in landfills• waste incineration and uncontrolled combustion

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DEHP is found regularly in municipal wastewater and, because of its lipophilic properties, itconcentrates in sewage sludge.

Anionic and Non-ionic Surfactants:

Surfactants are contained as the main active agents in all washing and cleaning agents.These compounds are covered in detail in Case Study (f).

Polychlorinated Dibenzo-p-dioxins and Dibenzofurans (PCDD/PCDF):

The generic term "dioxins" represents a mixture of 219 different polychlorinated dibenzo-p-dioxins and furans. The most well known and hazardous dioxin, is the tetrachlorodibenzo-p-dioxin (TCDD). Dioxin concentrations are calculated as sum of the toxicity equivalents (TEQ)relevative to the most toxic dioxin [TCDD].

The three main sources of polychlorinated dibenzo-p-dioxins and dibenzofurans are asfollows [Mahnke, 1997, Horstmann, 1995]:

• Chemical reactions or chemical reaction processes: Dioxins arise as unwantedby-products from the production or use of many organo-chlorine compounds, such aschlorine bleaching of cellulose in paper production and chlorine alkali electrolysis. Inthese cases the formation mechanism can be explained by substitution,condensation or cyclisation reactions.

• Combustion processes or thermal processes: Dioxins arise by thermal processesand are released into the atmosphere. The dioxin formation results from a de-novosynthesis. Important thermal sources are:

o waste incineration plants and incomplete combustion processes in landfills;o combustion plants;o iron smelting;o sinter plants, non-ferrous smelting and recycling plants;o petrol and diesel engines.

• Dioxins can also arise from all incomplete combustion processes involvingchlorine. This explains the ubiquitous dioxins occurrence in the environment.Anthropic production of dioxins has predominated since the introduction oforganochlorine compounds in industrial applications (1920). With the improvement ofthe catalysts in waste incineration plants and other measures for reducing the dioxinsemission, the fraction of anthropic dioxins has been declining since 1970. Dioxinscan be formed and released into the atmosphere also by natural events, e.g. forestfires. Dioxins can also be generated by the biochemical transformation of precursorcompounds (for example during degradation of chlorophenols).

Dioxins speciation in household wastewater and laundry wastewater is similar to those in thesediments of UWW collecting systems and sewage sludge. A mass balance indicates that 2-7 times more dioxins in sewage sludge originates from households than from urban runoff.Washing machine effluent is a major source of dioxins in household wastewater. Dioxinswere also detected in shower water, and in urban run-off from various human activities[Horstmann 1993, 1995]. These results suggest that the importance of householdwastewater as a dioxin source has been underestimated [Horstmann et.al., 1993,Horstmann, 1995].

Sources of other potential organic pollutants are listed below:

Organic pollutants can originate from food and household related products, such as longchain fatty acids and their methyl and ethyl esters, originating from faeces, soaps and foodoils. Being relatively hydrophobic these compounds are attached to particles, theconcentration of fatty acids and esters in the unfiltered influent is more than 500 µg/l. Otherorganic pollutants from domestic origin are the sterols from animal foods and faeces andindol from faeces. Caffeine is also found from discharges from coffee processing.

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Plasticisers and flame retardants are still used in many products for household andindustrial applications. Among the organic pollutants present are benzenesulphonamides,adipates (esthers of hexandioic acid), phthalates (esters of phthalic acid, among whichDEHP is the most common), and several phosphate esters. (2-chloroethanol phosphate) andTBP (tri-n-butyl phosphate) are used in flame-retardant compositions in textiles, plastics aswell as in other products.

Preservatives and antioxidants are constituents of household and industrial products, andamong the organic pollutants linked with these compounds are parabens (esters ofhydroxybenzoic acid), and also substituted phenols and quinones are among theconstituents.

Solvents both chlorinated and non-chlorinated (alcohols, ethers, ketones) are present in alarge range of products such as car shampoos and degreasing products, householdcleaners and degreasing agents from vehicle maintenance and production. Chlorinatedsolvents, such as trichloroethylene and trichloroethane, are in increasingly wide use: theamounts consumed in France per year are 24,000 and 28,000 tonnes, respectively. Theprincipal sources of diffuse pollution from chlorinated solvents are due to artisanal activitiessuch as metal finishing activities and dry cleaners. Nevertheless, domestic sources fromaerosols and other agents are not negligible. Pollution by metal cleaning activities is usuallyconsidered as diffuse discharges as they are usually from small firms with only fewemployees. Garages consumed around 15,000 tonnes of solvents in 1988, about 60% ofwhich is lost to the atmosphere and the rest as waste. Of the 6,000 tonnes, of waste solventsome will be discharged into the UWW collecting system [Agences de l'Eau, 1993]. Metalfinishing used 50,000 tonnes of solvents in 1991 and their aqueous wastes are dischargedinto UWW collecting systems, although these are usually in low levels. Dry cleaningconsumed around 19,500 tonnes of solvents in 1988 and it has been determined that0.3x10-3 kg of solvent/100 kg of clothes cleaned ended up in wastewater.

Fragrances from households, beauticians and hairdressers, generate mixtures of terpenesand synthetic musks (galaxolides), and are also found in industrial detergents. These arecovered in more detail in Case Study E, Section 6.

Pesticides and herbicides are also a common component of the urban wastewaters andthey result from road and rail weed treatment, and from gardens, parks and urban woodlandareas. They include the triazine group, the phenyl urea group (e.g. chlorotoluron, isoproturonand diuron), the phenoxy acid group (eg. Mecoprop and 2,4-D) and glyphosate [Revitt et al.,1999].

An enormous quantitiy of pharmaceutical products are prescribed every year: 100 tonnesof human drugs were prescribed in 1995 in Germany [Ternes, 1998]. Pharmaceuticals in theUrban Environment are discussed in Case Study (d), Section 6.

Triclosan (2,4,4’trichloro-2’hydroxydiphenyl ester) has been used in soaps, shampoo andfabrics, as an antimicrobial agent. While these compunds are regarded as low toxicity their2-hydroxy isomers have been shown to undergo thermal and photochemical ring closure toform polychlorinated dibenzo-p-dioxins which are highly toxic. (Okumura et al 1995).

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3.1.1 Domestic and Commercial Sources

A study carried out in France in 1995 by ADEME, showed the sources of the main organicmicropollutants in sludge from WWTS were mainly domestic and commercially related (seeTable 3.1). Another study, by SFT (in collaboration with the wider Norwegian governmentenvironmental study programme and the A/S Sentralrenseanlegget RA-2 WWTS),investigated sources of PAH, PCB, phthalates, LAS and NPE. This study found that sewagefrom domestic sources, in this instance from an isolated housing estate with a separatesewage and stormwater drainage system, does make a significant contribution of the aboveorganic pollutants to urban wastewater [SFT report 98/43].

Table 3.1 Principal sources of organic micropollutants in urban wastewater treatmentworks [ADEME, 1995] +++ very likely, ++ likely, + less likely present

POLLUTANT ORIGIN Domesticusage

Stormrunoff

Commercialeffluent

Aliphatichydrocarbons

Fuel ++ ++ ++

Monocyclic aromatichydrocarbons

Solvents, phenols + + ++

PAHs By-products of petroltransformation and

insecticides

+ + +

Halogens Solvents, plastics,chlorination

++ + ++

Chlorophenols andChlorobenzenes

Solvents, pesticides + + ++

Chlorinated PAHs PCB, hydraulic fluids (+) + ++Pesticides + + ++

Phthalate esters Plastifier + + ++

Detergents ++ + ++

Nitrosamines Industrial by-products(rubber)

0 + ++

Soil is also a major repository of organic matter and the soluble fractions can leach/run-off into water courses, especially in upland areas where measures to remove colour andformation of trihalomethanes during drinking water treament is important.

A. PAHs and PCBs

Table 3.2 shows that the PAH concentration profiles for three Swedish WWTS varies. Thismay in part be due to differences in the catchment areas, with the sources of the pollutantscoming from different local industries. Most of these PAHs are expected to derive fromdiffuse commercial activities and traffic but PAHs such as pyrene, which is believed to bederived from at least 50% domestic sources, is present in all the samples at more consistentconcentrations than some of the other compounds.

Mattson et al (1991) referenced in Paxéus (1996a) found that PAHs from food, an oftenoverlooked source of this pollutant, from households can reach 50-60 % of the total UWWcollecting system load for pyrene and phenanthrene. This is an important observation ashousehold sources of PAHs are likely to be more difficult to control than commercialsources.

Another source of PAHs from domestic and commercial activities is the use of phenol andcreosol in products such as wood preservatives. In Finland, 430 tonnes of woodpreservatives were used in 1995 [Finnish Environmental Institute, 1997]. PAHs may enterUWW as a result of spillages or as surface runoff from rainwater.

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Table 3.2 PAHs concentrations in urban waste waters in Sweden [Paxéus 1996a]

WWTSPAHsHSTµg/l

GRYAABµg/l

SSWµg/l

Naphthalene, dimethyl 1 0.5 <LODNaphthalene, methylpropyl 3 <LOD <LOD

1,1’- Biphenyl, dimethyl 2 0.5 <LOD1,1’- Biphenyl, ethyl 1 <LOD <LOD

Anthracene/Phenanthrene 1 <LOD 0.5Methyldibenzothiophene <LOD <LOD 0.5

2,8-Dimethyldibenzothioprene 5 <LOD <LODAnthracene/Phenanthrene methyl (different

isomers)2 <LOD 3

Anthracene/Phenanthrene dimethyl (differentisomers)

1 <LOD 1

Retene <LOD <LOD 0.5Pyrene 3 <0.5 2.5

Pyrene, methyl (different isomers) 2 <LOD 1Pyrene, methyl, methylethyl or tetramethyl 1 <LOD 1

1,1-Diphenylethane <LOD 0.5 <LOD

1, (H)- Indene, 1-phenylmethylene 0.5 <LOD <LOD9H-Flouren-9-one 0.5 <LOD <LOD

2-Anthracenaemine <LOD <LOD 9Acridine, 9-methyl-Dibenz(b,f) azepine <LOD <LOD 0.5

Octahydrophenanthrene, dimethyl-, isopropyl 0.5 <LOD <LODTotal PAH 23.5 <2 19.5

HST = Henriksdal Sewage Treatment Plant, GRYAAB = Gothenburg Regional SewageWorks, SSW = Sjölunda Sewage Works (<LOD = below limit of detection)

A study carried out in the Rhine region of France, [Commission Internationale pour laProtection du Rhin, 1999], showed that control of organic pollutants from point sources hasbeen effective at reducing levels of contamination in the Rhine. Between 1985 and 1996, thepollution from PAHs and PCBs had decreased by over 90%. In 1985, 1,075 kg of PCBs weredischarged, which was reduced to 250 kg in 1992, and to 3 kg in 1996, all of which werefrom industrial sources. For trichloromethane, 9,000 kg were discharged in 1985, 2,300 kg in1992, and 2,210 kg in 1996; of these 600 kg were from industry and 1,610 kg fromcommunal sources.

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B. DEHP

The Danish Ministry of the Environment and Energy [Danish Report, 1999] have estimatedthe annual consumption of phthalates in Denmark to be approximately

o 10,000 tonnes in 1992 (about 90% of this used in soft PVC)o 11,000 tonnes in 1995

In Germany, the total production of DEHP in 1988 was 234,000 tonnes. Of this, 1% wasdischarged to surface and groundwater [Brüggermann, 1988].

The vast majority of phthalate emissions to the environment occur, not during themanufacture, but during the use of the finished products. While in some cases this is acommercial setting (such as vehicle washing, which will be examined subsequently), thereare also major sources in the domestic environment. Mattson et al. (1991) mentionedpreviously regarding domestic sources of PAHs, estimated the household contribution ofphthalates and adipates to the Gothenburg sewage works as 70% of the total load (thisfigure emphasises the ubiquity of compounds and difficulty of control). Two major sources ofdomestic releases to wastewater (shown in bold in Table 3.3) are floor and wall coveringsand textiles with PVC prints.

Table 3.3 DEHP emissions in Denmark[adapted from Appendix 1 Danish Ministry of Environment and Energy Report, 1999]

Product Phthalate use (t y-1) Emission to airduring production

(t y-1)

Emission to airduring use

(t y-1)

Release towastewaterduring use

(t y-1)Cars 1000 - 0.1-1 2-10

Floor and WallCoverings

2000 - 0.2 1-5

Textiles withPVC prints

5-15 - - 2-13

C Dioxins and furans (PCDD/Fs)

The Environment Agency of England and Wales [1998] estimates dioxin emissions fromindustrial Part A processes to UWW collecting systems in the UK as 4.5 µg (TEQ), whereasemissions to air from these processes was estimated to be 1.1kg. Routes of these pollutantsinto wastewater via deposition or industrial process (i.e. washing of air pollution cleaningequipment), are not discussed. Actions taken to reduce dioxin emissions continue to ensureIPC authorisations are met.

Recent research at the University of California, Berkeley, reports that deposition of dioxins tosoil is 6 to 70 times greater than estimated emissions [Eduljee 1999]. This suggests thateither not all sources of dioxin are known and/or the contributions from these sources maynot be accurately characterised.

Table 3.4 shows the dioxin emissions for the years 1994-1998 in Austria. There was little orno change in the dioxin emissions in Austria over this period, but slight reductions, wereachieved in some sectors. The main reason for the emission reduction in 1998 is due to theair hold ordinance, which limited dioxin emissions from waste combustion as well as fromsteam-boiler plants.

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Table 3.4 Dioxin emissions in the time period 1994-1998, Austria[Federal Environmental Agency, UNECE/CLRTAP, 1999].

1994 1995 1996 1997 1998Issuer groups Dioxin emissions (tonnes per annum)

Small consumer(household, trade,

administration)16,820 18,160 18,400 16,780 16,260

Industry (burningand processes)

3,470 3,730 3,880 3,980 3,910

Industryprocesses

8,170 8,900 7,990 8,550 8,380

Waste handlingand landfills

180 180 180 180 180

Total 28,640 30,970 30,450 29,500 28,740

In Spain, concentrations of dioxins are reported for recent samples (1999) of sewage sludgeand for archived samples (from 1979 to 1987) [Eljarrat, et.al., 1999]. Results are shown inTable 3.5. It is estimated that the current concentrations of dioxin in sludge have droppedsince the 1970s-80s. This is expected to be due to the source reduction of pollutants, fromcombustion and incineration processes, and from certain pesticides contamination andemphasises the success that controls on use of compounds and trade effluent discharge inreducing pollutant levels.

Table 3.5 Concentrations of PCDD/F in sewage sludge in Spain [Eljarrat, et.al., 1999]

Type of sewagesludge

Range ofconcentrations

(pg.g-1 DW as I-TEQ)

Mean value(pg.g-1 DW as I-TEQ)

Fresh [1999] 7 to 160 55Archived [1979-1987] 29 to 8,300 620

E. Other organic compounds

Adsorbable organo-halogen compounds (AOX) resulting from bleach products and fromchlorine use, were reported in studies done in Portugal, in Ria Formosa lagooned sewage[Bebianno, 1995] and in Italy in the city of Parma [Schowanek, et.al, 1996]. The averageAOX concentration in sewage was reported as 37 µg.l-1.

Sterols were reported in sewage sludge and around discharge wastewater points inPortugal, in Faro, Tavira and Olhao [Mudge et al., 1997, 1998 and 1999]. Concentrationsranged between 0.1 to 27.8 µg.g-1 sterols of dry weight of sludge. Hospital wastewater maycontain high phenol concentrations, up to 20,000 µg.l-1, plus other compounds such as LAS,NPE, PCBs and pharmaceuticals.

F. Vehicle washing

A specific activity identified as a source of a number of organic pollutants in urbanwastewater is vehicle washing, which consists of two distinct phases:

o Actual cleaning, involving the removal of oily dirt, which, on a quantitative basiswould be expected to be similar to the type of oily dirt (asphalt and vehicle exhaustparticles) which is in road runoff. However, this would also involve the use ofdegreasing solvents and surfactants which can enter the wastewater treatmentprocess.

o Vehicle Treatment, involves the use of protective treatments, often coatings usingdifferent types of wax against corrosion, dust and dirt.

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The effluent is usually discharged to the UWW collecting system. Several studies of theeffluents from vehicle washing facilities have been undertaken [Paxéus 1996a, 1996b,Paxéus and Schröder, 1996, Ulmgren 2000a]. In Sweden, an environmental standard for carwashing detergents was established in Göteborg in 1992 [EHPA, 1992], based on thePrecautionary Principle and Substitution Principle in the Chemical Products Act. In generalCOD values found at the effluents of vehicle washes are in the range of typical untreatedindustrial petrochemical wastewaters [Huber, 1988].

In Gothenburg, an important site for vehicle manufacture, vehicle washing was estimated tocorrespond to 0.5 % of the total wastewater at the Gothenburg WWTS, which was concludedto have a very small effect on the total load of organic pollutants at the plant. The majorcomponents of the effluents were aliphatic hydrocarbons and alkylbenzenes, originating frompetroleum base degreasing solvents and the oily dirt on the vehicles themselves (asphalt,vehicle exhaust particles). Low aromatic products reduce the potential environmentalassociated with detergent use in car washing facilities. These are produced byhydrogenation of petroleum-based solvents where substituted benzenes and naphthalenesare converted to corresponding naphthenes and decalins. The formation and discharge ofpolyaromatic compounds is negligible for detergents that come from low aromaticmicroemulsions.

Table 3.6 summarises the results of a study on washing both of light vehicles (LV) andheavy vehicles (HV) [Paxéus 1996]. As can be seen, HVs tend to contribute larger organicpollutant loads than LVs.

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Table 3.6 Concentration of organic pollutants in car wash effluents in mg l -1

[after Paxeus, 1996]

Conventionalparameters

LV HV

Mean Median Range Mean Median RangeTotal oil 291 242 10-1750 550 460 65-1200

COD 1263 1180 120-4200 4600 4500 1700-7500

Aliphatic hydrocarbonsC8-C16 29 22 1-139 103.86 76.72 41-220C17-C30 0.6 0.4 <0.001 1.84 1.87 0.9-3.0

Aromatic hydrocarbonsBenzene 0.01 0.01 <0.01-0.2 0.02 0.02 0.02-0.03Toluene 0.08 0.05 <0.01-0.6 0.10 0.08 0.03-0.2

Naphthalene 0.17 0.13 <0.001-0.7

1.1 0.75 0.3-3

Biphenyl 0.015 0.005 <0.001-0.1

0.12 0.11 0.04-0.2

Dibenzofuran 0.001 0.002 <0.001-0.03

0.011 0.011 0.009-0.012

Phenathrene 0.005 <LOD <0.001-0.03

0.021 <LOD 0.005-0.03

Pyrene 0.003 <LOD <0.001-0.01

0.009 <LOD 0.01-0 .02

Fluoranthene 0.003 <LOD <0.001-0.01

0.004 <LOD 0.002-0.006

PlasticizersDiethyl phthalate 0.005 0.01 2E-3-0.06 0.01 0.01 0.01-0.02Dihexyl phthalate 0.05 0.03 <0.001-

0.150.3 0.21 <0.001-

0.7DEHP 0.52 0.38 0.03 - 4.1 1.50 1.30 0.4 - 3

Washing agentsp-nonylphenol 0.60 0.26 0.01-4 0.43 0.41 0.1-0.8

2-Botoxyethanol 25 15 <0.001-270

15 17 <0.001-27

It is not known if this area is representative of the Scandinavian region as a whole in terms ofthe car washing input. However, it does seem that car washing is also an important sourceof pollutants in Norway [SFT, 1998a, 1998b]. In Norway 41 businesses were reported on assources of hazardous organic pollutants, PAHs, phthalates (DBP, BBP, DEHP),nonylphenols (nonylphenol, nonylphenol mono- and di-ethoxylates). The studies found thehighest pollutant loads in the effluents from motor vehicle workshops to urban wastewatercame from petrol stations with car washes, long haul transport depots with ‘car washes’commercial laundries, paint spraying workshop and chemical businesses [SFT, 1998a,1998b].

There are two main types of washing agent available and the choice of these would result insignificant differences in wastewater quality:

• Water-based formulations (microemulsions) containing 10-30% hydrocarbons butincreased surfactants (10-30%);

• Petroleum-based degreasing formulations containing 95-99% of hydrocarbons and3% surfactants.

Plasticisers found in the effluents from vehicle cleaning included phthalates, althoughanalysis of the cleaning and washing chemicals showed that they themselves contribute verylittle to the discharge of plasticisers.

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3.1.2 Urban runoff

A significant proportion of organic contaminants in wastewater are derived from urban runoff.These organic compounds include aliphatic and aromatic hydrocarbons, PAHs, fatty acids,ketones, phthalate esters, plasticisers and other polar compounds. Solvent extractableorganics are dominated by petroleum hydrocarbons, which arise from motor oil and tyresfrom road surfaces. Organic pollutant sources have not received the extent of researchattention that potentially toxic element pollution has. For example, in the case of PAHs whichare combustion by-products and enter wastewater principally through atmosphericdeposition and urban runoff, the sources can be stationary (industrial sources, power andheat generation, residential heating, incineration and open fires) and mobile (petrol anddiesel engine automobile) [Sharma et al.,1994]. Different PAH species are associated witheach one of these sources.

A. Road and vehicle related pollution

The main sources of road and vehicle related metals pollution have been outlined in Section2.1.3. Table 3.7, shows some of the road and vehicle related sources of organic pollutants.

Table 3.7 Qualitative classification of road related sources of organic pollutants[after Montague and Luker, 1994].

Traffic Maintenance AccidentsPetrol

(PAHs and MTBE)Tar and bitumen Petrol

Oil Oil OilGrease Grease Grease

Antifreeze Solvents SolventsHydraulic fluid PAHs

AsphaltPCBs

Pesticides andherbicides

Table 3.8 summarises the results from three experimental catchments from 1975 to 1982 onmean concentrations of PAH.

Table 3.8 Summary of pollutant concentrations in urban runoff caused by road relatedsources [after Klein, 1982]

Test catchmentsPollutant meanconcentrations (mg.l-1) Pleidelsheim Obereisesheim Ulm / West

PAH 2.61 2.97 2.51

The necessary conditions for PAH formation is the presence of benzene and a highconcentration of radical intermediates, which then form stable compounds. Multiple ringsystems are autocatalytic and promote further ring condensations. Fuel aromatic content hasbeen shown to influence particle-associated PAH emissions almost linearly [Pedersen et al.,1980; Nunnermann, 1983; Egeback and Bertilsson, 1983]. However, the relationshipbetween the aromatic content of petrol and PAH formation is not fully understood.

PAHs are produced by unburned fuel, exhaust gases and vapour, lead compounds (frompetrol additives) and hydrocarbon losses from fuel, lubrication and hydraulic systems.Volatile solids will be added to the total suspended solids loading of rainfall runoff and canalso act as carriers for both potentially toxic elements and hydrocarbons. Some road dustshave been found to contain 8.5 µg g-1 of PAHs [Colwill et al., 1984 as reported in Luker and

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Montague, 1994]. The introduction of the catalyst technology for motor vehicles lowered theemissions of PCDD/F in Germany to about 98% [UBA, 1999].

Tyre wear releases hydrocarbons either in particulate form or in larger pieces as a result oftyre failure. A tyre loses about 10 to 20 per cent of its weight in a lifetime. Annually it isestimated an average of 140 g of tyre-derived particles are eroded per metre of road[Environment Agency of England and Wales, 1999].

Plasticisers (such as diethyl phthalate and dihexyl phthalate) are also considered animportant parameter of organic pollution load in urban runoff. Cary et al. [1989], stated thatplasticisers, especially phthalates, represent the major pollutants found in urban storm water.The concentrations found for 8 plasticisers were recorded. Of these DEHP was found in thegreater concentrations than the other seven plasticisers combined. The main sources ofplasticisers are traffic grime and dirt, associated with the degradation of plastic componentsof the vehicles.

B. Roof Runoff

Regarding roof runoff as an interface between atmospheric boundary layer and the runoffreceiving system, Förster (1993) investigated the role of roofs as source and sink of organicpollutants. The trace organics analysed included PAH, chlorinated hydrocarbons and nitrophenols. The research indicated that the insecticide HCH was primarily introduced to theroof runoff system by wet deposition, while the amount of adsorbed PAHs (pyrene;benzo[a]pyrene=BaP) in roof runoff exceeded the input by rain with events during coldertimes of the year where fossil fuel heating systems constitutes additional source for thispollutant. The concentration profiles for a number of PAHs are illustrated in Figures 3.1 and3.2 below.

Figure 3.1 PAH in runoff from zinc sheet roof [after Förster, 1993]

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Figure 3.2 PAH in runoff from tar roof [after Förster, 1993]

As can be seen, the concentrations of PAH in roof runoff from zinc roofs was found to beabout ten-fold higher than for tar roofs. There is a difference in the pattern of distribution forPAH concentration at different precipitation flow rates. For tar roofs PAH concentration ishighest at the lower and higher precipitation flows and lower at intermediate events, whereasfor zinc roofs it tended to be higher at lower precipitation flows. Therefore, concentrations of**pollutants in roof runoff can be considered variable depending on the characteristics of theroof material itself as well as on the characteristics of the precipitation event.

A number of hydrocarbons are present in urban rainfall runoff, particularly those associatedwith motor vehicles, such as petrol, fuel oils and lubricants. In an unmodified form theseliquids are insoluble in, and lighter than, water. Typically, 70-75% of hydrocarbon oils show astrong attachment to suspended solids [Luker and Montague, 1994]. PAHs have an evengreater affinity. In contrast, Methyl-tertiary-butyl-ether (MTBEs) the new additive to unleadedfuel is significantly more soluble in water than all other hydrocarbons in rainfall runoff.Hydrocarbons, even in low concentrations, can give rise to surface sheens and thusadversely affect surface waters. Most hydrocarbons eventually degrade by a combination ofmicrobial and oxidative processes; degradation though is slow, so the increase in oxygendemand in watercourses and wastewater is likely to be marginal and not a principalenvironmental impact.

C. Urban vegetation control practices

Herbicides and pesticides are used in road maintenance operations to control weeds andpests on the roadsides and verges. The triazine group of herbicides, including atrazine andsimazine, has been used extensively for roadside weed clearance and is more soluble andmobile than their organo-chlorine predecessors. Combined levels of atrazine and simazineabove 1µg l-1 are not uncommon in watercourses near highways (Ellis, 1991). Collins andRidgeway (1980), report that half of pesticides in urban runoff are associated with particles<63 µm, although these particles are less than 6% of the total suspended solids load.In urban areas, pesticides in general, and herbicides in particular, are becoming an integralpart of the control of unwanted vegetation by local and municipal authorities, rail and airportoperators. The main herbicides used in the UK are of the triazine group, the phenyl ureagroup (e.g. chlorotoluron, isoproturon and diuron), the phenoxy acid group (e.g. Mecopropand 2,4-D) and glyphosate (Revitt et al., 1999). Of the phenyl urea compounds, only diuron

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has been widely used in the urban environment and in 1989 this herbicide accounted for13% of the total 550 tonnes of active ingredient used in the UK (Department of theEnvironment, 1991). The comparable use of triazines was 39% but following the introductionof restrictions for the non-agricultural use of these herbicides in 1992, many users convertedto the use of diuron and glyphosate for the control of vegetation in urban environments(White and Pinkstone, 1995). The removal of herbicides by rainfall runoff is influenced byrainfall characteristics, the time interval between herbicide application, the precipitation eventand the properties of the herbicide. However, the full range of factors that influence herbiciderelease from sites of application and the mechanisms governing the transport to, and fate ofherbicides in the aquatic environment are not fully understood [Davies et al., 1995; Heatherand Carter, 1996]. The principal herbicide sources in urban catchments include [Revitt et al.,1999]:

• Urban parks and private gardens• Road maintenance (to road kerbstones and backwalls)• Railway system maintenance.

Concentrations in receiving waters, reported by Revitt et al., (1999) in the UK, wereconsistently above the drinking water limit of 0.1 µg l-1 recommended for simazine anddiuron; the mean concentrations of which reached 0.34 and 0.45 µg l-1, respectively. InFrance [Farrugia et al., 1999], the average application rates for pesticides on the mostconsuming urban land uses are reported as 900 g ha-1 for roads and streets, 4000 g ha-1 forcemeteries and 500 to 800 g ha-1 for parks and sport yards. Householders may also uselarge amounts of herbicides and other pesticides but information on the quantities applied isnot available in published literature. However, there was considerable variation in the extentof water contamination with herbicides between catchments. Farrugia et al, (1999), reportedthe average concentration of diurons in water receiving urban runoff was 5 µg l-1, andattributed this entirely to use in urban situations.

It is to be noted that the hydrological characteristics of hard urban surfaces provide the idealconditions for the efficient transport of herbicides (particularly diuron, see also Farrugia et al.,1999) into UWW collecting systems. This, combined with the existence of inert physico-chemical environments involving neutral pH, low nutrient and total organic carbon levels,absence of absorption sites and low bacterial populations, allow the application of herbicidesin urban areas (although in low use), to be an important potential source of contamination ofwaste water.

D. Wet and dry deposition

The main repository of PCBs, PAHs and PCCD/Fs is soil. Volatilisation from soil, then furtheratmospheric transport and deposition of PAHs, PCBs and PCDD/Fs is considered to be oneof the main contemporary sources of these contaminants in the environment Wild et al.,.[1995]. PAHs are difficult to control because they are a combustion product.

The Austrian Federal Environment Agency (UBA) analysed PAHs in several media (surfaceand wastewater, sediment, soil, sewage sludge, compost, plants, street dusts and ambientair) between 1989 and 1998 [Gans, et.al., 1999]. Only 10 % of samples were above thedetection limit for PAHs of between 2.6 and 20.3 ng l-1 and these were all taken during winterand spring, suggesting that PAH originates from the emissions of heating systems during thecold period.

Once released (by the sources mentioned in the previous paragraphs), airborne PAHs aretransported by the prevailing meteorology before being removed from the atmospherethrough various scavenging mechanisms. As with other airborne pollutants the majormechanisms of removal of PAHs from the atmosphere are wet deposition, such as rain,sleet, snow, hail, and dry deposition to the surface. The wet removal of gaseous compoundsis better understood than particulate PAH removal [Ligocki et al., 1985]. The extent of in-

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cloud or below cloud scavenging, collection efficiency of falling precipitation, solubility andsize particles has been examined in the literature [McVeety, 1986 as reported in Sharma etal., 1994].

Dry removal is a function of atmospheric conditions and the surface level concentration ofPAHs. PAHs adsorbed to particles greater than 20 µm have higher settling velocities andthus will settle in the vicinity of the source. However, this mechanism will only account for aminor percentage of removal, as PAH are mostly adsorbed on particles less than 10 µm indiameter.

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3.2 INFLUENCE OF VARIOUS TREATMENT PROCESSES ON THE FATE OF ORGANICPOLLUTANTS THROUGH WASTEWATER TREATMENT AND SEWAGE SLUDGETREATMENT

3.2.1 PARTITIONING OF ORGANIC POLLUTANTS IN WASTEWATER TREATMENTPROCESSES.

The general effect of wastewater treatment processes is to concentrate the organicpollutants in the sewage sludge and the extent of this removal depends on the properties ofthe organic species. The overall result of this process is to discharge a treated wastewaterrelatively free of organic and inorganic contaminants and a sewage sludge that containsmost of the organic contamination present in the feed wastewater. The main complication ofthis general study arises from the large number of possible organic species that could bepresent in the feed stream and the complex chemistry sorbtion mechanisms on the solids.

During the treatment cycle, some organic materials can degrade to a certain extent,especially in aerobic environments and organic material of biological origin is easy todegrade. Indeed, some common organic pollutants such as LAS, are specifically added todetergents because they are aerobically biodegradable. A considerable body of literatureexists on this aspect and a variety of oxidants have been proposed. The main aim of thistype of work has concentrated on reducing the organic pollutant content in sewage sludgeprior to land disposal. Advanced oxidation processes might be used in tertiary treatmentespecially if the final effluent is to be used for drinking water. However, use of theseprocesses; regardless of the power of the oxidant, cannot be expected, a priori, to degradeall types of organic pollutants within a reasonably short time scale. Indeed, the presence oforganics in final effluents is an obstacle in expanding the recycling of wastewater.

3.2.2 Wastewater Treatment

Traditionally, wastewater treatment is supposed to begin at the head of a WWTS at the inletscreens used to remove large objects such as wood plastics and paper. However, in realitywastewater conditioning starts in the sewer, in large conurbations the wastewater can havequite a significant residence time in a sewer. However, it is suggested that dilution ofsewage with runoff water is likely to have an adverse effect on the efficiencies of thedownstream treatment processes (Dorussen et al., 1997).

Primary treatment is installed to enable sedimentation of the feed wastewater. This processis used to settle, retain and concentrate most of the particulate material to the bottom of thetank as primary sludge. The process is affected by temperature and the solids content of thesupernatant or primary overflow is significantly higher if the temperature is low, as it is inwinter. Though simple, primary sedimentation is a widespread process in Europe, althoughnot practised in all WWTS. In some cases primary sedimentation is not installed and in otherplants flocculation, by addition of flocculants, is carried out in the primary sedimentation tank(Hahn et al., 1999).

The objective of secondary treatment is to contact the primary overflow (settled sewage) withair in the presence of aerobic bacteria and other micro-organisms, which convert the organicmatter to carbon dioxide and water to a variable extent. There are two types of plantcommonly used for this process: bio filters and activated sludge. Most WWTS use primaryand secondary processes. However some plants may have tertiary treatment which, caninvolve coagulation, flocculation and rapid gravity filtration.

A novel process for secondary treatment is the lagoon (Salter et al., 1999). This is large unitseveral meters deep and can be stirred gently and aerated. Aquatic life including fish cansurvive in some lagoons. The residence time in the lagoon is long and they can be used totreat the more contaminated municipal wastes. In addition secondary pre-treatment can becarried out using magnetic flocs. In this process the organic contaminants present are

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loaded on to the magnetic flocs at a low pH and washed off in a high pH medium (Booker etal., 1996).

There is some concern about the use of iron coagulants, which is of direct relevance to thisstudy. Some iron reagents used in wastewater treatment are made as a by-product oftitanium oxide production. The titanium ore contains traces of vanadium and uranium. Twoother tertiary methods often cited are activated carbon and membrane filtration. Bothhowever are rather expensive. Activated carbon is a very efficient means of removal oforganic pollutants and the technique is widely used in small domestic plants used to polishdrinking water. Membrane filtration is also very effective in removing particulate materialfrom water. However, the membranes are expensive and fouling can occur.

In order to estimate organic and inorganic pollutant removal in wastewater treatmentprocesses models are required to simulate them. In such models physical properties of thepollutants are used to determine the likelihood that they will be removed by the process.More work is needed on modelling the fate of organic pollutants through WWTS and theirtransformation throughout the different treatment methods.

It is clear that the regular screening of priority organic pollutants on a day-to-day basis wouldbe complex and uneconomical. It has been suggested that determination of adsorbableorganic halogens (AOX) be used as an indicator for these priority substances (Hahn et al .,1999). AOX determination is a relatively easy technique to use (Korner, 2000). Thesesubstances are sorbed from the water on charcoal, which is subsequently pyrolysed. Thehydroxyhalides produced are sorbed and analysed by titration. Another general testmentioned in the literature (Ono et al., 1996) is the bacterial umu-test, which measuresdamage caused by organic pollutants on DNA.

3.2.3 Properties of Organic PollutantsOctanol-water partition coefficient and solubilityThe octanol- water partition coefficient is the ratio of a compound’s concentration in octanolto that in water at equilibrium.

Kow =

Kow is dimensionless and values vary over the range of at least 10-3 to 107 and are usuallyexpressed logarithmically. Large Kow values are characteristic of large hydrophobicmolecules which tend to be associated with solid organic matter while smaller hydrophilicmolecules have low Kow values. Octanol-water partition coefficients can be measured directlyby using conventional “shake flask” methods (Leo and Hansch, 1971). This experimentalapproach is restricted to compounds of low-to-medium hydrophobicity, since for compoundswith high hydrophobicity, the concentration in the aqueous phase is too low to be measuredaccurately.

Kow can also be correlated with various environmental parameters, such as solubility. Bydefinition, the partition coefficient expresses the concentration ratio at equilibrium of anorganic chemical partitioned between an organic liquid and water. This partitioning is, inessence, equivalent to partitioning the organic chemical between itself and water. One wouldexpect that a correlation would exist between the partition coefficient and solubility. Lyman etal. (1990) presented the following correlation between solubility based on 156 compounds:

978.0log339.11

log oww

KS

=

where Sw is the solubility expressed in mol l-1. This correlation was obtained empirically andthe correlation coefficient was found to be 0.874.

Concentration of compound in octanol

Concentration of compound in water

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Organic carbon-water partition coefficient, KOC

The tendency of a compound to sorb to the organic matter such as humic substances in soil or sewage sludge particles can be assessed using the organic carbon-water partitioncoefficient. It is defined as the ratio between the concentration of the organic compound onorganic carbon (mg.g-1) and its concentration in water (mg.l-1), at equilibrium.

Koc =

The likelihood of the leaching of a compound through soil or adsorption onto soil organiccarbon can be assessed from Koc values. Generally, organic compounds with high Koc valueswill tend to adsorb onto organic carbon whilst compounds with low values have a greatertendency to be leached. Koc values can be estimated from the octanol-water coefficient orwater solubility. Karickhoff et al. (1981) found the following correlation:

Log Koc = 0.82 log Kow + 0.14

when he examined sorbtion data for a variety of aromatic hydrocarbons, chlorinatedhydrocarbons, chloro-S-triazines and phenyl ureas. The correlation coefficient was 0.93.

In this study a specific list of organic pollutants has been defined and it can be seen thattheir solubilities are very low but the Koc values are very high in the region of 105 indicatingthat the sorbtion would be very favourable. From the Koc values and the weight fraction oforganic carbon species present in the feed (f) an estimate of the removal of organic speciescan be made. The amount left in the supernatant water as a percentage left (L) is given by:

L = 100

fK oc

1

Thus if f = 10-3 and Koc = 10-5, the percentage left would be 1%. There is limited dataavailable or actual results but figures for L are generally much higher. (Pham et al., 1997)report that 30% of PCB and only 25% of the PAHs were removed from a specific treatmentplant.

3.2.4 Modelling

Understanding the processes involved in wastewater treatment is likely to provide a basis forunderstanding the pathways and partitioning of pollutants in these processes. The way to dothis is to develop models of the processes and to simulate plants using computers. Anexample of such a comprehensive model has been published (Gabaldon et al., 1998). Themodel does not specifically include large molecular weight organics.

It is of some of interest to note that there is some work on processes that occur in a sewer.One study aims to model the emissions of volatile organic compounds in cocurrent air flow inopen and closed sewers (Olsen et al., 1998). Another study measures the removal of CODand proteins within a sewer (Raunkjaer et al., 1995) and found that there were quitenoticeable losses in a sewer. In another study the sewer pipe was considered to consist of asediment above which was a bio-film and above that the water phase (Fronteau et al., 1997).

O’Brien et al. [1995] and Mann et al.[1997] present a first order model for a wastewaterplant. In the secondary section aeration for stripping, biodegradation and sorption on to aPAC (Powdered Activated Carbon) were considered. PCB, PCDD/F or PAH were notincluded but the methodology presented in this paper could be applicable to the study of thefate of these high molecular weight pollutants in secondary treatment. Work has been doneon modelling trickling filter-beds (Shandalor et al., 1997). This predicts the drop of solidsloading in the water as it trickles down the bed. On the more specific case of organic

Concentration of compound on organic carbon

Concentration of compound in water

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pollutant removal, a detailed paper has been published on the removal of volatile organiccontaminants in a wastewater plant (Melcer et al., 1994). However, again no specificmention of PAH or similar organics was made.

3.2.5 Organic Degradation in WastewaterAmong the organic pollutants being studied in this report, LAS is somewhat unusual as it isadded to water in detergents. Studies in this subject (Holt et al., 1998 and Prats et al., 1997)report very similar LAS degradation levels of over 90%. Although the removal of LAS inWWTS is quite effective some 16% of the feed LAS is taken out in the sewage sludge (Fieldet al., 1995). In this sorbed form it is more difficult to degrade. Some studies of LAS in riversediments (Tabor et al., 1996) show that this compound is sorbed on to the solids and onlyslowly biodegradable. Thus there would be an amount of non degraded LAS in the solidresidue.

There have been a number of studies on the degradation rate of PCDD, PCDF and PCBs,which have been reported in a review article (Sinkkonen et al., 2000). The experiments wereconducted in laboratory rigs and the data reported as half-life analogous to radioactivedecay. The mean half-life quoted is given in Table 3.09.

Table 3.9: Half-lives of PCDD, PCDF and PCB in water

Substance Half life in water(years)

PCDD 2.6PCDF 5.0PCB 9.3

This study seems to indicate that these organics will not be degraded in a WWTS. Thesehalf-lives are considerably longer than the residence time in a sewage treatment plant orsewer. As the authors point out the experiments were conducted near ideal conditions and,in practice, the half lives are believed to be longer than the figures quoted in the table,especially if the temperature is low.

PAH compounds are believed to be persistent in the environment. There is some work thatpresents evidence that some of these compounds can be degraded in periods of 12-80hours (McNally et al., 1998). Compared with PCDDs this time period is rapid. However,these experiments on biological degradation of PAH were carried out under ideal conditions.There was a constant temperature (20oC), specially adapted bacteria were used andnutrients were added. In a practical case where low temperature and few nutrients arepresent, the actual degradation times would be much longer (in the region of 80-600 hours)so PAH compounds are unlikely to be degraded in a conventional wastewater treatmentplant. Research in Greece by Samara et.al. [1995] and Manoli et al. [1999], shows that thelower-molecular mass PAHs are removed effectively in Thessaloniki's WWTS, whereas thehigher molecular mass PAHs are resistant to the biological treatment. The heavy molecularmass PAHs are partially removed by adsorption, whereas the lower molecular mass PAHsare removed by volatilisation and/or biodegradation.

Work on oestrogenic compounds, analysing 17β-oestradiol equivalent concentrations, foundthat the load of oestrogenic activity in the wastewater was reduced by about 90% in thesewage plant. Less than 3% of the oestrogenic activities was found in the sludge (Korner etal 2000).

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3.2.6 Removal of OrganicsCoagulants such as aluminium and ferric salts are used in water treatment to removeparticulate matter. However, soluble organics may also be removed by coagulation bymechanisms such as specific adsorption to floc particles and co-precipitation (Semmens andOcanas, 1977). Sridhan and Lee (1972) studied the removal of phenol, citric acid andglycine from lake waters by co-precipitation with iron. Though these results were reasonable,excessive concentrations of coagulant (300-1500 mg.l-1) were required. Other workers madesimilar findings. Semmens and Ocanas (1977) examined the removal of dihyroxybenzoicacid (DHBA) and resorcinol from distilled water by coagulation with ferric sulphate. Resultsindicated that the extent of organic removal increased as coagulant dosage increased.Maximum percentage removals were 35% for DHBA and 8% for resorcinol. Semmens andAyers (1985) examined the effectiveness of alum and ferric sulphate in removing octanoicacid, salicylic acid, phenol and benzoic acid from Mississippi river water and water samplesfree of organic matter. These compounds were generally poorly removed by coagulation andin most cases the extent of removal did not depend strongly on coagulant dosage. Removalsranged between 3-20%. Salicylic acid was most efficiently removed and benzoic acid wasmost poorly removed. Generally, better removal of the organic compounds occurred whennatural organics were not present.

The general consensus of the work done to date indicates that the use of coagulants forremoving organics is feasible. However it is impractical as the excessive addition ofcoagulants is necessary.

Humic substances account for around 50% of the dissolved organic matter in natural water(Vik and Eikebrokk, 1989). They are formed easily from waste material and there is evidencethat they will sorb organic matter by binding with them. Activated carbon is widely used as ameans of removing organic compounds from water. The presence of humic acid reduces therate of organics uptake (Kilduff et al., 1988). The capacity of activated carbon fortrichloroethylene (Summers et al., 1989, Wilmanki and Breeman, 1990), trichlorophenol(Najm et al., 1996) and lindane decreased in the presence of humic substances. Othersorption media such as organoclays (Dentel et al., 1998, Zhoa and Vance, 1998) and anorganic polymer resin (Frimmel et al., 1999) are not so badly affected by the presence ofhumic substances. Ying et al., (1988) studied the effects of iron precipitation on the removalof natural organic compounds like tannic acid and humic acid, and toxic organic compoundslike chlorendic acid (HET), polychlorobiphenyls (PCBs) and organochlorine pesticides.Freshly formed ferric hydroxide flocs were very effective in removing humic acid and tannicacid and it was found that the presence of humic acid enhanced significantly the removals ofPCBs and many of the organochlorine pesticides by ferrous and ferric hydroxideprecipitates. Removals were achieved by a combined mechanism of complexation,adsorption and co-precipitation. This evidence suggests that humic substances are capableof sorbing organic material.

A process was devised in which organic contaminants were removed by adding humic acidand a coagulant such as ferric hydroxide (Rebhun et al., 1998). This showed good recoveryfor the organics tested. The results of this work suggest that humic acid might be added in atertiary cycle. The humic acid could be made by composting grass cuttings, potato peelingand other waste feeds. Such material could be added to the final effluent of a wastewatertreatment plant followed by contact and flocculation.

3.2.7 Conclusions – removal of organics in wastewaterThe practical issue of the removal of organics in wastewater treatment is not welldocumented in the literature. Modelling work reviewed here, has shown that the work hasconcentrated on the removal and degradation of organic matter of biological origin and thatsynthetic organic pollutants have been largely neglected. Clearly modelling work forpollutants should be promoted.

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This data in turn relies upon analysis of these organic pollutants. Present methods usingGC/MS are extremely complex and not suitable for routine plant use. Lack of easier methodsfor their analysis will hinder the development of simple processes to remove these organicmaterials. It could be argued that identification of a specific pollutant is not crucial for plantdevelopment and that a generic test would be suitable. One of the most important aspects offuture work is the development or identification of simple tests for WWTP analysis. Theproblem is not confined to treatment plants alone but rapid treatment methods could be usedto detect sources of heavy organic chemicals.

One of the results of the difficulty in doing analyses is that there is little data available onpartitioning process. There is a suggestion that around half the organics fed to a wastewatertreatment plant remain in the supernatant stream. This might well be a surprising result giventhe measured property values in the region of 105 (dimensionless) would suggest that therewould be a good binding between organic pollutants and the settled sludge. It is possiblethat there is some competition for sorbtion sites in the organic matter from the moreconcentrated organic compounds.

With more rapid analysis techniques in place, there would be the opportunity to makeprocess changes to reduce the amounts of organics present in the final effluents. It is feltthat advanced oxidative techniques such as the use of ozone would not be applicable in thepresent context as the organics have a very small concentration in solution and have a lowreactivity. One interesting possibility is in situ treatment in sewers such as adding activatedcarbon to contaminated streams. As humic substances are efficient scavengers for organicpollutants, humic acid derived from composting food waste could be added in a tertiary stageto strip organics from the final effluent. It is a matter of policy, to see if such ideas should bepromoted further but initial work could start before the rapid analysis methods had beenagreed.

Transfer and partitioning of organic contaminants to the sludge matrixThe sorption of organic contaminants onto the sludge solids is determined by physico-chemical processes and can be predicted for individual compounds by the octanol-waterpartition coefficient (Kow). During primary sedimentation, hydrophobic contaminants maypartition onto settled primary sludge solids and compounds can be grouped according totheir sorption behaviour based on the Kow value as follows (Rogers, 1996):

Log Kow < 2.5 low sorption potentialLog Kow > 2.5 and < 4.0 medium sorption potentialLog Kow > 4.0 high sorption potential

Volatilisation and thermal degradationMany sludge organics are lipophilic compounds that adsorb to the sludge matrix and thismechanism limits the potential losses in the aqueous phase in the final effluent. A proportionof the volatile organics in raw sludge including: benzene, toluene and the dichlorobenzenesmay be lost by volatilisation during wastewater and sludge treatment at thickening,particularly if the sludge is aerated or agitated, and by dewatering. Volatilisation is used todescribe the passive loss of organic compounds to the atmosphere from the surface of opentanks such as clarifiers. The majority of volatilisation, however, occurs through air stripping inaerated process vessels. As a general guide, compounds with a Henry’s Law constant >10-3

atm (mol -1 m -3) can be removed by volatilisation (Petrasek et al., 1983). The significance ofvolatilisation losses of specific organic compounds during sewage treatment can bepredicted based on Henry’s constant (Hc) and Kow (Rogers, 1996):

Hc > 1 x 10-4 and Hc/Kow > 1 x 10-9 high volatilisation potentialHc < 1 x 10-4 and Hc/Kow < 1 x 10-9 low volatilisation potential

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However, more recent studies (Melcer et al., 1992) suggest that the stripping of volatilesmay not be as significant as was initially thought and biodegradation during secondarybiological wastewater treatment may be the main mechanism of loss of the potentiallyvolatile compound types (Table 3.10). For example, Melcer et al. (1992) reported thatbiodegradation processes removed ≥90 % of the dichoromethane, 1,1,1-trichoromethane,trichloroethylene, toluene and xylene from a municipal wastewater. Volatilisation was only asignificant mechanism of removal for 1,4-dichlorobenzene (20 %) and tetrachloroethylene(60 %). The fate and behaviour of volatile organic compounds in wastewater treatment planthave been modelled numerically by the TOXCHEM computer-based model that incorporatesfour removal mechanisms including: volatilisation, stripping, biodegradation and sorption onto solids (Melcer et al., 1992).

Table 3.10 Observed and predicted (TOXCHEM) removals of volatile organiccontaminants during wastewater treatment by stripping and biodegradation (Melcer etal., 1992)Compound Air stripping (%) Biotransformation (%)

Observed Predicted Observed Predicted

Dichloromethane 2.6 3.2 92.4 91.9Chloroform 7.4 7.8 73.6 71.91,1,1-Trichloroethane 10.5 6.0 79.7 89.1Trichloroethylene 10.7 3.1 82.7 91.3Tetrachloroethylene 58.7 64.2 15.8 0.01,4-Dichlorobenzene 19.1 17.2 54.7 54.8Toluene 1.2 0.4 98.6 98.3p- and m-Xylene 1.3 0.6 98.1 97.9

High temperature treatment of sludge by disinfection processes at 70 oC for 30 minutes canenhance the loss of volatile compounds. Mono- and two-ringed aromatic compounds(benzene, toluene, xylene, naphthalene, dichlorobenzene etc) may be partially lost underthese conditions (Wild and Jones, 1989). Other more persistent hydrophobic compounds, eglesser chlorinated PCBs, and the three-ringed PAHs, may also be susceptible tovolatilisation. Thermal drying is being introduced as an enhanced treatment process toproduce sanitised biosolids for unrestricted use and for improved handling and bulkreduction. This process is potentially the most effective at removing volatile substances fromsludge because the solids are exposed to high temperatures (400 oC) and the sludge isdried to >90 % ds. Thermal degradation may also be an important mechanism for theremoval of organic contaminants from sewage sludge during heat treatment (Wild andJones, 1989). Volatile organic compounds in sewage sludge are not regarded as a potentialrisk to human health or the environment when sludge is used in agriculture (Wilson et al.,1994).

Destruction by sludge stabilisation processesMesophilic anaerobic digestion is the principal sludge stabilisation process adopted in mostEuropean countries, where approximately 50 % of sludge production is treated by thismethods. Volatile compounds are generally lost to the atmosphere or transferred to thesupernatant during digestion, whereas PAHs and phthalate acid esters are conserved (Bridleand Webber, 1982).Many organic contaminants are biodegraded under anaerobic conditions and this isenhanced by increasing retention time and digestion temperature. Five characteristicbehaviour patterns (Figure 3.3) of decay are observed for organic contaminants in anaerobicdigestion systems based on net gas (total CH4 + CO2) production (Battersby and Wilson,1989):

• Easily degradable (eg ethylene glycol, diethylene glycol, triethylene glycol, sodiumstearate, ethanol);

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• Degradable after a lag phase (eg phenol, 2-aminophenol, 3- and 4-cresol, catechol,sodium benzoate, 3 and 4-aminobenzoic acid, 3-chlorobenzoic acid, phthalic acid,dimethyl phthalate, di-n-butyl phthalate, pyridine and quinoline);

• No degradation or gas production (3- and 4-aminophenol, 2-chlorophenol, 2-cresol,2-nitrophenol, 2- and 4-chlorobenzoic acid, bis (2-ethylhexyl)phthalate, hexyleneglycol, neopentyl glycol, n-undecane, n-hexadecane, 2,4-D, dieldrin, cis- and trans-permethrin, tetrahydrofuran, furan, pyrrole, N-methylpyrrole, thiophene, benzene,pyrimidine, 1-naphthoic acid);

• Inhibitory in the initial phase of incubation (eg 3- and 4-chlorophenol, 2,4- and 2,6-dichlorophenol, 2,4,6-trichlorophenol, 3- and 4- nitrophenol, 2-phenylphenol, 2-, 3-and 4-nitrobenzoic acid, CTAB, MCPA, MCPP, lindane, naphthalene,anthraquinone);

• Inhibitory throughout incubation (eg 3,5-dichlorophenol, pentachlorophenol, 2,4- and2,5-dinitrophenol, 4-nonylphenol, sodium dodecylbenzene sulfonate, sodium 4-octylbenzene sulphonate, 2,4,5-T, butyltin trichloride, dibutyltin dichloride, tributyltinchloride).

Degradation is generally aided by carboxyl and hydroxyl groups, whereas chloro or nitrogroups tend to inhibit anaerobic biodegradation and gas production.

Figure 3.3 Typical patterns of net gas production (CH4 + CO2) from organic chemicalsincubated anaerobically with diluted primary digested sewage sludge.1, Easily degradable; 2, Degradable after a lag period; 3, little effect on gas production; 4,inhibitory in initial phase of incubation; 5, inhibitory throughout incubation(Battersby and Wilson, 1989).

Biodegradation during anaerobic digestion may virtually eliminate certain organiccontaminants from sewage sludge, but in general the destruction achieved is typically in therange of 15 – 35 % (WRc, 1994). Aromatic surfactants including linear alkyl benzenesulphonates (LAS) and 4-nonylphenol polyethoxylates (NPnEO) occur in sludge in largeconcentrations. These compounds are not fully degraded during sewage treatment and thereis significant accumulation in digested sludge. For example, mass balance calculationssuggest that approximately 80 % of LAS is biodegraded during the activated sludge processand 15-20 % is transferred to the raw sludge (Brunner et al., 1988). Approximately 20 % ofthe LAS in raw sludge may be destroyed by mesophilic anaerobic digestion sludge. Thecompounds, nonylphenol monoethoxylate (NP1EO) and nonylphenol diethoxylate (NP2EO)are formed during sewage treatment from the microbial degradation of NPnEO. These

Net

gas

pro

duct

ion

(% th

eore

tica

l)100

-100 5

4

}3

21

Time

Inhi

biti

on

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metabolites are relatively lipophilic and accumulate in the sludge and are also dischargedwith the treated sewage effluent. One of the most important consequences of anaerobicdigestion, however, is the production of nonylphenol (NP), which accumulates in digestedsludge. Approximately 50 % of the NPnEO in raw sewage is transformed to NP in digestedsewage sludge. The loadings of LAS and NP to soil in sewage sludge used on farmland aresignificantly larger than for most of the other organic contaminants present in sludge andthere is concern about their potential environmental effects. This is particularly the case forNP in sludge due to its potential oestrogenic activity (UKWIR, 1997). However, in the aerobicsoil environment, these compounds provide substrates for microbial activity and are rapidlydegraded so there is minimal potential risk to the environment or transfer to the humanfoodchain. For example, LAS has a short half-life in soil in the range 7 – 22 days intemperate field conditions (Holt et al., 1989) and the half-life for NP is <10 days (UKWIR,1997). Current studies at Imperial College, funded by the Food Standards Agency in the UK,are investigating the potential for plant uptake of NP into staple food crops from sludge-treated soil.

Another class of organic chemicals, the phthalate acid esters, are also an abundant group ofcompounds present in sewage because of their extensive use as plasticising agents. Thephthalates are also suspected as being potential environmental oestrogens (UKWIR, 1997).Shelton et al. (1984) reported the complete degradation of the lower molecular weightphthalate esters, and of butyl benzyl phthalate, within 7 days in laboratory scale anaerobicdigesters operated at 35 oC. Therefore, these phthalate compounds should generally beremoved by most municipal anaerobic digesters at the normal mean retention timesoperated in practice (>12 days). The extent and rate of biodegradation during anaerobicdigestion is apparently related to the size of the alkyl side chain and compounds with largerC-8 group are much more resistant to microbial attack. Therefore, di-n-octyl and di-(2-ethylhexyl)phthalate (DEHP) are considerably more persistent to anaerobic microbialmineralisation and are generally not removed by conventional anaerobic stabilisationprocesses. However, phthalate esters are rapidly destroyed under aerobic conditions,usually achieving >90% removal in 24 h in activated sludge wastewater treatment systems.In soil, the reported half-life is <50 d (UKWIR, 1997).

Composting is a thermophilic aerobic stabilisation process and usually involves blendingdewatered sludge at approximately 25 % ds with a bulking agent, such as straw or woodchips, to increase porosity of the sludge to facilitate microbial activitiy. The biodegradation ofrelatively persistent organic compounds such has been reported for composted sludge (Wildand Jones, 1989). For example, PAHs may be partially degraded by composting sludge andaverage removals of 13 % and 50 % have been measured for benzo(a)pyrene andanthracene, respectively, although phenanthrene persisted unchanged in laboratorycomposting trials (Martens, 1982; Racke and Frink, 1989).

Thermophilic aerobic digestion processes and sludge storage for three months can achievesimilar overall removal rates for organic contaminants as those obtained with mesophilicanaerobic digestion (WRc, 1994). Thermal hydrolysis conditioning of sludge prior toconventional anaerobic stabilisation may have a significant influence on the removal oforganic contaminants from sludge, but this is a comparatively new enhanced treatmentprocess and effects on the destruction of organic contaminants have yet to be investigated.

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3.3 Quantitative assessment of organic pollutants in untreated UWW, treated UWWand treated SS

For the main list of organic pollutants considered in this report there is little available data ofthe concentrations in the influent to the wastewater treatment plant. Paxéus and Schröder[1996] looked at over 50 organic compounds, in the influents and effluents of theGothenburg wastewater treatment plant. The high cost of testing explains the lack of data ondioxins in urban wastewater.

Most of these compounds were reduced to below the limit of detection during the treatmentprocess. Some of the organic compounds, such as caffeine were reduced from a level of37µg.l-1 to 4µg.l-1. Some of the phosphorus containing compounds were not reduced duringthe treatment process (although the influents and effluents were quite low at 1µg.l-1). Theoverall toxicity of the influent and the effluent were also measured and found to havedecreased by approximately 50% during the treatment process.

Figure 3.4 Dioxin content of archived samples of sewage sludge form Mogden WWTS,UKIt can be seen (Figure 3.4) that there has been a significant reduction in the concentration ofdioxins since the 1950s and 1960s in sludge over recent years.

The concentrations of other organic contaminants in sludge, including, PCBs and PAHs,have also declined significantly in sludge in the UK. This is due to the control of primarysources of these substances. In 1984, McIntyre and Lester (1984) measured median and99th percentile concentrations for PCBs in sludge (444 samples from UK sewage treatmentworks) of 0.14 and 2.5 mg kg-1, respectively. Ten years later, Alcock and Jones (1993)reported the total PCB content of 12 UK sludges from rural, urban and industrial sewagetreatment works ranged between 0.106 to 0.712 mg kg-1, with a mean value of 0.292 mg kg-

1. These results indicate that overall PCB concentrations in UK sludges have declinedmarkedly in response to the ban on industrial production, use and discharge of thesesubstances. Similar trends are apparent in Germany (Table 3.11). In effect, this means thatthe chemical composition of sewage sludge is already subject to stringent, albeit indirect,controls that have been effective in minimising industrial sources and inputs of persistentorganic contaminants.

Year

Dio

xin

co

nce

ntr

atio

n (

ng

TE

Qkg

-1 d

s)

0

50

100

150

200

250

300

350

400

450

1944 1949 1953 1956 1958 1960 1998

German limit =100 ng TEQ kg-1 ds

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Table 3.11 Mean concentrations of organic contaminants in German sewage sludge in1988/89 relative to data collected until 1996 (Leschber, 1997)

Contaminant 1988/89 1991/96Adsorbable organo-halogens mg kg-1 ds 250-350 140-280Polychlorinated biphenyls(1) mg kg-1 ds <0.1 0.01-0.04

Polycyclic aromatic hydrocarbons(1) mg kg-1

ds0.25-0.75 0.1-0.6

Di(2-ethylhexyl)phthalate mg kg-1 ds 50-130 20-60Nonylphenol mg kg-1 ds 60-120 -

Dioxins and furans (ng TEQ kg-1 ds) <50 15-45(1)Single congeners

Table 3.12 Survey of organic pollutants in UWW and WWTS (µg.l-1)

WWTSCompound Country

Influent(µg.l-1)

Effluent(µg.l-1)

Reference

Austria:Total PAHs - EPA15 147-625 20-70 Gans et al.,1999

Germany:Total PAHs

Benzo(a)pyreneBenzo(k)fluoranthene

0.790.080.05

Hagenmaier et al,1986

Greece:Benzo(a)pyreneFluoroanthene

Indeno (1,2,3-cd) pyrene

0.0220.240.015

0.0050.0290.005

Manoli et al, 1999

France 0.05-0.44 0.02-0.09 ADEME, 1995Germany 33 Koch et al, 1989 &

Balzer et al 1991

PAHs

UK 51.8 (5.6 to349)

30.8 (2.4-147)

Morris et al, 1994

Austria 4.4 0.3 Hohenblum et al.,2000

DEHP

Germany 122 (7-232) 15 (5.6-184) Faltin, 1985Anionic

SurfactantsItaly 290-4800 - Braguglia et al,

2000Detergents France 1-26 0.1-2.7 ADEME, 1995

Austria 400-3500 11-55 Scharf et al., 1995Germany 5400 67 Feijtel et al 1995Greece 129 (35-

325)Kilikidis et al. 1994

Italy 4600 43 Feijtel et al 1995Netherlands 4000 9 Feijtel et al 1995

Spain 9600 140 Feijtel et al 1995

LAS

UK 15100 10 Feijtel et al 1995Austria:

Nonylphenol-monoethoxylate

Nonylphenol-diethoxylate

2,096,000

13,093,000

363,000

639,000

Hohenblum et al.,2000

Italy:NP

NPEONPEC

427145

Di Corcia et al.1994

Sweden 0.5-6.0 Paxéus 1996a

NPE

Germany 0.02 0.002 Koppe et al, 1993

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Other organicpollutants:

ChlorophenolsChlorinated

organicsPesticides

VOCs

Iodinated X-Raycontrast

substances:iopamidoldiatrizoate

iothalamic acidiomeprol

iopromide

France

Germany

0.1-0.43001.1510

4.33.30.180.171.67.5

<0.1-0.5 ADEME, 1995

Ternes et al, 2000

Total Phenols Italy 2.5-300 Italian RegionalEnvironmental

Protection AgencyDioxins Italy: 0.024-16.9 Italian Regional

EnvironmentalProtection Agency

PAHs: wastewater from 8 different sewage treatment influents was investigated in 1996 bythe UBA [Gans, et.al., 1999]. Similar PAHs content were determined, except for the influentfrom a chemical plant, which had an approximately 1,000 times larger concentration. PAHsespecially, with a low molecular weight were found in high concentrations. Apart from thehigher PAH content of the wastewater from the chemical plant, no significant differencescould be detected between municipal and industrial influents. The PAHs content of theeffluent was about 10 times smaller than the influents [Gans et.al.1999].

The Danish regulation of the application of waste products [Ministry of the Environment andEnergy 1996] sets certain cut off values for the maximum concentrations of organiccontaminants in sludge to be distributed on agricultural land as shown in Table 3.13.Concentrations of PAHs in Danish sewage sludge are also shown in Table 3.14 The PAHconcentration of the nine selected compounds were all found to have mean concentrationsabove the concentrations permissible for use on agricultural land in Denmark.

Table 3.15 Danish standards for maximum concentrations of organic contaminants insewage sludge (Danish Ministry of the Environment and Energy, 1996)

Danish Standards 1997 - cut off valuesmg.kg-1 DS

2000 - cut off valuesmg.kg-1 DS

LAS 2,600 1,300nonylphenol (including nonylphenol

ethoxylates)50 10

PAHs* 6 3DEHP 100 50

*(total concentration of nine selected PAHs) Acenaphthylene, Fluorene, Phenanthrene, Fluoranthene,Pyrene, Benzo(b,j,k)fluoranthene, Benzo(a)pyrene, Benzo(g,h,i)perylene and Indeno(1,2,3,-cd)pyrene

The values in bold are difficult to achieve, as they are far below current sludgeconcentrations. If 50% of pyrene and phenathrene is from food sources and gives sludgeconcentrations of > 300mg.kg-1 ds, then this emphasises how difficult these standards are toachieve.

The mean concentrations of LAS, NPE and DEHP were found to be within the Danish limitsfor use on agricultural land but the range of concentrations in all cases went over the cut offlimits; therefore many of the sludges would not be allowed to be used on agricultural land.

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The concentrations of some of the organic contaminants in the sludge were found to dependstrongly on the wastewater treatment process [Danish EPA]. The concentrations of LAS, NPand NPE were significantly lower (P<0.005) following activated sludge treatment than inmixed activated and digested sludge treatment, presumably due to extended aeration.

It can be seen that government and other institutions are trying to introduce limits for certainpollutants and that concern for wastewater pollution reduction is increasing. Nevertheless, itis noted that important discrepancies exist in analysis techniques, even within a country,hence slowing the determination of limits, particularly for PAHs and PCBs. Due to theexpected increase in sludge production and the reinforcing of the legislation in relation to theconcentration limits for potentially toxic elements and organic pollutants, it seems necessarythroughout Europe to harmonise analysis techniques and the pollutants targeted in thecontrol of wastewater and sludge quality. Discharge standards to UWW collecting systemsfor industries and possibly reformulation of certain domestic products should be determinedin order to reduce pollution entry into the systems.

Table 3.14a) Survey of organic pollutants in sewage sludge: mg kg-1 DS (a)PAHsPAHs Country Mean Median Min. Max. Year/s of

SurveyB[A]P

I[1,2,3-cd]pAustria 0.30

0.270.220.21

0.090.07

0.670.58

1994/95(24)

Σ PAHs*B[a]p

Fl.theneI[1,2,3-cd]p

B[a]p

Germany(municipal)

6.40.351.20.30.5 0.4

2.60.10.60.10.1

15.31.12.70.83.4

1996 (10)1996 (10)1996 (10)

(10)1995 (14)

B[a]pFl.thene

I[1,2,3-cd]p

Denmark 0.150.30.67

0.070.10.23

<0.01<0.01<0.01

1.43.30.63

(28)

Fl.thene Spain 3.4 1.1 6.0 (19)B[a]p

Fl.theneFrance 0.04

0.151131

1994 (6)1994 (6)

B[a]pFl.thene

I[1,2,3-cd]p

Greece 0.241.10.11

0.241.30.12

0.10.380.05

0.361.40.15

(15)

Σ PAHsFl.thene

Sweden 1.2-2.2 0.7-1.40.01 0.7

1995/98(26)(29)

Σ PAHs**B[a]p Fl.thene

UK 27.8

2.32.62.5

6.00.11.1

83.87.54

1994 (27)1989 (33)1991 (29)

Σ PAHs*** Switzerland(municipal)

0.50 0.35 0.04 1.83 (29)

B[a]pI[1,2,3-cd]p

Fl.thene

Italy <0.05<0.05<0.05

2000 (34)

Σ PAHs Poland 72.3 74.4 32.7 114.3 1999 (2)

EUB[a]p

Fl.theneUSA 13.8

9.954.7 154

1541988

(30/31)

Limits Agricultural Soils Sewage SludgeΣ PAHs**** EU 6 (proposed) (9)

Pyrene WHO 480 (5)B[a]p USEPA 21.4 (25)

B[a]p: Benzo[a]pyrene; Fl.thene: Fluoranthene I[1,2,3-cd]p: indeno(1,2,3-c,d)pyrene * sum of 16 USEPA priority list PAHs (see Appendix B) ** sum of naphthalene, acenaphthylene, acenaphthene, fluorene, phenanthrene, anthracene, fluoranthracene,pyrene, chrysene, benzo(a)anthracene *** sum of benzo(b+k)fluoranthene, benzo(ghi)perylene, benzo(a)pyrene, fluoranthene, indeno(1,2,3-c,d)pyrene

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**** sum of acenapthene, phenapthene, fluorine, fluoranthene, pyrene, benzo(b+j+k)fluoranthene,benzo(a)pyrene, benzo(ghi)perylene, indeno(1,2,3-c,d)pyrene

Table 3.14b) Survey of organic pollutants in sewage sludge: mg kg-1 DS (b)PCBs

PCBs Country Mean Median Min. Max. Year/s ofSurvey

PCB (28,52,101,138, 153,180)

Austria 0.07 0.05 0.02 0.27 1994/95(24)

Germany 0.01-0.040.5 0.05 15

1991-96(13)

1985-87 (7)Denmark 0.05 0.03 <0.03 0.2 (28)

Spain 0.05 0.93 (20)PCB

(101,118,138)France 0.03 0.4 1994 (6)

Sweden 0.1 0.1 1995/98(26)

UK 0.34 0.01 21.5 (18)

EUUSA 1.46 1.48 14.8 1988

(30/31)

Limits Agricultural Soils Sewage SludgeEU 0.8 (proposed) (9)

WHO 30 (5)USEPA 6.6 (25)

Table 3.14c) Survey of organic pollutants in sewage sludge: mg kg-1 DS (c)DEHPDEHP Country Mean Median Min. Max. Year/s of

Survey

Austria 23.4 34.4 (11)Bis-(2ethylhexyl)-

phthalateGermany 20-60

<2.4 3201991-96

(13)(7)

Denmark 38 25 3.9 170 (28)Bis-(2ethylhexyl)-

phthalateSweden 6.7 28 (29)

EUBis-(2ethylhexyl)-

phthalateCanada 68.0 11 959 (3)

USA 110 17 891 1988(30/31)

Limits Agricultural Soils Sewage SludgeEU 100 (proposed) (9)

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Table 3.14e) Survey of organic pollutants in sewage sludge: mg kg-1 DS (e)LAS

LAS Country Mean Median Min. Max. Year/s ofSurvey

Austria 8107 7579 2199 17955 1994/95(24)Germany 5000 50 16000 1985-87(7)Denmark 2700 530 11 16100 (28)

AerobicAnaerobic

Spain 10012100

50017800

(1)(21)

Finland 9700 (17)Italy 11500 14000 (4)UK 8700 10400 60 18800 (12)

EUAerobic Anaerobic USA 152

4680 1680 7000(16)(22)

Limits AgriculturalSoils

Sewage Sludge

EU 2600 (proposed) (9)

Table 3.14f) Survey of organic pollutants in sewage sludge: mg kg-1 DS (f)NPE

NPE Country Mean Median Min. Max. Year/s ofSurvey

Austria 24 12 69 1994/95(24)

NP1EONP2EO

Germany 60-120512010

3.85<3

96.38080

1988/89(13)1996 (10)(7)(7)

Denmark 15 8 0.3 67 (28)Sweden

13-2740010-26

26 1100 1990 (32)1995/98(26)

UK 326-638 256 824 (27)EUUSA

Limits Agricultural Soils Sewage SludgeEU 50 (proposed) (9)

Table 3.14g) Survey of organic pollutants in sewage sludge: mg kg-1 DS (g)PCDD/FDIOXINS &

FURANS (NG

TEQ/KG DS)

Country Mean Median Min. Max. Year/s ofSurvey

Germany 15-45 1991-96(13)

Spain 55620

42 729

1608300

1994-98 (8)1979-87 (8)

Sweden 24 23 25 (23)UK 40.2 7.6 192 (29)

EU

DioxinsUSA 82.7

90.437.4 0.49 2321

18201988

(30/31)

Limits Agricultural Soils Sewage Sludge

EU 100 (proposed) (9)

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References1. Berna JL et al, 19892. Bodzek, B. et al,

1999.3. Bridle, T.R. et al 19834. Cavelli L, et al 19935. Chang, A.G. et al

1995.6. Conseil supérieur

d'hygiène publique deFrance, 1998

7. Drescher-Kaden et al1992,

8. Eljarrat. E, et al 1999.9. European Union, 200010. Hessische

Landesanstalt furUmwelt (1991-96).

11. Hohenblum, P, et al2000.

12. Holt MS et al 1992.13. Leschber, R. 199714. Litz. N, et al, 1998.15. Manoli, E. et al 1999. 16. McAvoy DC, et al

1994.17. McEvoy & Giger 198618. McIntyre. A,E, et al

198419. Moreda, JM, et al

(1998a)20. Moreda, JM, et al

(1998b)21. Prats D, et al 1993.

22. Rapaport RA, et al1990.

23. Rappe et al 198924. Scharf, S, et al. 199725. Smith, S.R. 200026. Statistika

meddelanden 199827. Sweetman 199428. Tørsløv J, et al 1997.29. UKWIR 199530. USEPA 199231. USEPA 199932. Wahlberg,.C, et al

199033. Wild. S,R, et al 1989.34. Braguglia et al 2000

Table 3.15 shows the occurrence of certain organic pollutants in sewage sludge in Germany.

Table 3.15 Occurrence of certain organic substances in sewage sludge, Germany[Priority list USEPA and 6/464/EEC of EG].

Compound Occurrencein sludge

Benzo(a)anthracene +++Benzo(a)pyrene +++

Benzo(k)fluoranthene +++Dibenzo(a,h)anthracene +++Indeno(1,2,3-cd)pyrene +++

PCB-1242 ++PCB-1254 +++PCB-1221 ++PCB-1232 ++PCB-1248 ++PCB-1260 +++PCB-1016 +

2,3,7,8-Tetrachlordibenzo-p-dioxin ++Frequency of occurrence: +++ frequent (90-100%), ++ less frequent, + low frequency

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4. HEALTH AND ENVIRONMENTAL EFFECTS OF POLLUTANTS IN UWW AND SS

4.1 Potentially toxic elements

Copper, Cr, Zn, Se, are essential trace elements, however, they are potentially toxicelements, and above certain concentrations, may interfere with or inhibit the actions ofcellular enzymes. Effects are summarised in Table 4.1, together with the MaximumConcentration Limits (MCL) in drinking water (Source: EPA and WHO).

Potentially toxic elements at high concentrations, are acutely toxic to humans. Highconcentrations are rare in urban waste water, but could possibly result from accidental spills,although there is limited exposure from this route. The major concern is exposure to lowconcentrations over longer time periods. This is chronic exposure and may have more subtleeffects.

The chemical form and corresponding bioavailability of potentially toxic elements to plants,fungi, micro-organisms and animals are also important, affecting regulation regarding thecompound. Complex dietary interactions are inherently protective limiting the accumulationof metals in animal body tissues (see Case Study a). The phytotoxic concentrations of Znand Cu in plant leaves less than the zootoxic concentration of these elements in plantproduce consumed by animals and humans. Technically based, precautionary soil limitvalues for these metals protect crop yields and the human diet from the potentially toxiceffects of metals in sewage sludge, see Case Study (a).

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Health effects of potentially toxic elements

Table 4.1: Summary of the acute and chronic effects of the pollutantsDrinking WaterStandards

PTE

EU WHO USEPA

Acute Health Effects Chronic Health Effects Carcinogenicity Notes

Cu 0.1-3mg.l-1

2mg.l-1

1mg.l-1

Irritation of mouth and throat,headaches, dizziness, nausea,diarrhoea, gastric ulcers,jaundice, renal damage, death

Liver and kidney damage, "pink disease",cirrhosis

No evidence

Zn 0.1-5mg.l-1

5mg.l-1

Stomach and digestionproblems, dehydration, impairedmuscular coordination

Immune system damage, and interfereswith the body's ability to take in and useother essential elements such as copperand iron

No evidence Taken up by plants,phytotoxic

Cd 0.005mg.l-1

0.003mg.l-1

0.005mg.l-1

Digestive tract irritation, colitis,vomiting, diarrhoea, death.

Half-life = 10-40 years. Lung, kidney, andhematopoietic system damage due to buildup, fragile bones, anaemia, nerve or braindamage in animals

Strong evidence inanimals, weakevidence inhumans

Taken up by plants,phytotoxic

CrVI

0.05mg.l-1

0.05mg.l-1

0.1mg.l-1

Allergic responses in skin,chromium VI irritates nose,lungs, stomach, and intestines,convulsion, death

Damage nose and lungs, increases risks ofnon-cancer lung diseases, ulcers, kidney,and liver damage. Birth defects andreproductive problems in mice

Evidence inhumans andanimals

High potential forbioaccumulation inaquatic organisms

Hg 0.001mg.l-1

0.001mg.l-1

0.002mg.l-1

Nausea, vomiting, diarrhoea,increase in blood pressure, skinrashes, eye irritation, renalfailure

Brain, lung, kidney, and damage todeveloping foetus, neurological disorders,depression, vertigo, and tremors

Evidence in mice Binds to dissolvedmatter notphytotoxic, Highpotential forbioaccumulation inaquatic organisms

Ni 0.05mg.l-1

0.02mg.l-1

0.04mg.l-1

Allergic reactions, lung damage Chronic bronchitis and reduced lungfunction, lung disease. Affects blood, liver,kidney, immune system, reproduction anddevelopment in mice and rats

Evidence of lungand nasal sinuscancers inhumans

Phytotoxic, verymobile in water

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Pb 0.05mg.l-1

0.01mg.l-1

0.015mg.l-1

Anaemia, constipation, colic,wrist and foot drop, renaldamage. In children, symptomsare irritability, loss of appetite,vomiting, and constipation.

Non-specific: damage to nervous system,kidneys, and immune system. In children,can cause decrease mental ability andreduced growth. In adults, can cause adecrease in reaction time and affectmemory, miscarriage, premature births, anddamage to male reproductive system

Evidence inanimals

Not usuallyphytotoxic,bioconcentration inshellfish

As 0.05mg.l-1

0.01mg.l-1

0.05mg.l-1

Nausea, vomiting, diarrhoea,damage to tissues includingnerves, stomach, intestines andskin,

Skin keratosis, decrease in the productionof blood cells, bone marrow suppression,abnormal heart function, liver/kidneydamage, impaired nerve function, damageto foetus in animals

Evidence inhumans, ofincreased risk ofliver, bladder,kidney, and lungcancer, may inhibitsome DNA repairmechanisms

Taken up by plants,which are sensitiveto lowerconcentrations thananimals,bioaccumulation infish and shellfish,persistent in theenvironment

Ag 0.01mg.l-1

- 0.05mg.l-1

Breathing problems, lung andthroat irritations, and stomachpains, allergic reactions,necrosis, haemorrhage, andpulmonary oedema

Argyria and may affect brain and kidneys No evidence

Se 0.01mg.l-1

0.01mg.l-1

- Dizziness, fatigue, irritation,collection of fluid in lungs,bronchitis, rashes, swelling,

Brittle hair, deformed nails, and loss offeeling and control in arms and legs,reproductive effects in rats and monkeys,and birth defects in birds

Suspected humancarcinogen: liverand lung tumoursobserved in ratsand mice

EU: EC drinking water directive (1998)WHO: WHO (2000) and Guidelines for drinking water quality Vol.2 (1996).USEPA: ATSDR (2000) and Standards for maximum permissible values in sewage sludge/soils. Estimating concern levels for concentration ofchemical substances in the environment. Washington DC (1984).

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4.2 Organic pollutants

Over 6,000 organic compounds have been detected in raw water sources many of which aredue to commercial activities. These compounds are present in many industries, processesand by-products, and are wide spread in a number of domestic products. They may alsoenter wastewater from surface run off. LAS and NPE are widely used in detergents and areubiquitous in the environment.

Examinations of the surfactant toxicity revealed that anionic and non-ionic surfactants arenon-toxic to man upon oral intake, but can be very toxic for marine life. Table 4.2 shows thetoxicity of anionic and non-ionic surfactants [Berth and Jeschke, 1987]

Table 4.2 Toxicity of anionic and non-ionic surfactants [Berth and Jeschke, 1987]

Surfactant ToxicityLD50 fish (mg/l) LD50 Daphnia (mg/l)

LAS 3 - 10 8.9 - 14NPE 1.5 - 11 4 - 50

Effect of Sludge Treatment Processes on Organic Contaminants

Sewage sludge is treated to reduce its fermentability, nuisance (particularly odour) andvector attraction potential as well as to reduce bulk and improve its physical characteristicsto aid handling, dewatering and acceptability for use on agricultural land. These techniquesinclude physical, chemical and microbiological manipulation of the sludge and may involvemechanical dewatering, heating, chemical reactions and microbiological transformationprocesses that may influence the loss, or potential formation, of organic contaminants. Lossmechanisms include (Wild and Jones, 1989; Rogers, 1996):

• Volatilisation;• Biological degradation;• Abiotic/chemical degradation eg hydrolysis;• Extraction with excess liquors;• Sorption onto solid surfaces and association with fats and oils.

Some chemicals are more susceptible to these transformation/loss mechanisms than othersand certain treatment processes can increase the concentration of conservative compoundsthat are retained as volatile solids are removed during stabilisation action are due to theformation of stable reaction intermediate compounds.

Sewage sludge potentially contains thousands of organic compounds derived from industrial,domestic, atmospheric and natural dietary sources. The total organic matter content ofsludge is typically in the range 60-80%, depending on the extent of stabilisation treatmentand volatile solids destruction. The vast majority of the organic content is environmentallybenign material that confers part of the agronomic benefit associated with sewage sludgeuse on farmland by improving the physical and structural characteristics of treated soil.However, certain anthropogenic compounds are potentially toxic and would be detrimental tohuman health if they were to transfer from sludge-treated soil to the foodchain intoxicologically significant amounts. The range of organic compounds known to exist insludge is extensive and diverse. For example, Drescher-Kaden et al. (1992) reported that332 organic substances, with the potential to exert a health or environmental hazard, hadbeen identified in German sludges and 42 of these were regularly detected in sludge.

The main route of entry of environmental contaminants into the human foodchain is byuptake into the edible parts of crop plants. However, despite the increasing intensity andextent of scientific investigation into the potential environmental consequences of organic

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contaminants applied to farmland in sewage sludge, there is no evidence for soil-croptransfer. This is because those compounds exhibiting some solubility and potential for plantuptake are also susceptible to rapid degradation processes in soil or are lost throughvolatilisation, whereas other more persistent compounds usually have very low solubilitiesand are strongly adsorbed by the soil matrix in non-bioavailable forms. The principal concernand theoretical mechanism of entry to the human foodchain of organic contaminants insewage sludge applied to agricultural land is from the surface application of liquid sludgeand the intake of organic contaminants by livestock grazing treated pasture and theaccumulation of lipophilic compounds in meat fat and milk.

However, in practice, there has been no demonstrable relationship between sludgeapplication and transfer of organics to animal tissues or milk. Furthermore, this theoreticalexposure route will diminish with the additional precautions being introduced to minimise therisk of animal infections with enteric pathogens in sludge that prevent the application ofconventionally treated sludge products, such as liquid digested sludge, to grazed pasture.Improved pasture hygiene obtained by sub-surface injection of liquid sludge also eliminatesthe theoretical exposure of grazing livestock to organic contaminants in sludge. Currenttrends suggest that, in future, surface application to grassland is likely to be in the form ofenhanced treated biosolids, such as thermally dried sludge, which do not adhere to grassthus mitigating the intake of contaminants in sludge by ruminants.

New and recently identified organic compoundsUptake of organic contaminants by crop plants from sludge-treated soil has beendemonstrated experimentally to be negligible and the human dietary intake of organics fromplant foods is not considered to be an important route of exposure. The theoretical potential risk to humans from these compounds in sewage sludge arises from the direct ingestion ofsurface applied sludge adhering to grazed pasture and the potential for transfer to milk andmeat fat consumed as food. However, it is emphasised that surface application ofconventionally treated sludges to grazing land has been phased out in the UK, it is notpractised in a number of European countries and may not be permitted in the revised theSludge Directive. The 3rd Draft of the Working Document on Sludge (CEC 2000) states thatdeep injection followed by a no grazing interval of six weeks after application is acceptablefor sludge treated by conventional processes (eg mesophilic anaerobic digestion). Deepinjection burial of sludge below the soil surface improves pasture hygiene and avoids thepotential for sludge ingestion by grazing livestock. Advanced treated sludges may be spreadon grazing land, but current trends suggest thermal drying of sludge may expand andpasture contamination and animal ingestion is significantly reduced by the application ofdried, or finely divided solid products, which do not adhere to grass. These developments significantly reduce the potential risk of human dietary exposure to organic contaminants insewage sludge applied to farmland. A comprehensive list of compounds that could bepresent in sludge and maintain their integrity during treatment and following application toagricultural land has been developed by Alcock et al. (1999). The compounds selectedincluded:

Chlorinated paraffins – There are >200 commercial chlorinated paraffin formulations in useas plasticisers in PVC and other plastics, extreme pressure additives, flame retardants,sealants and paints. Total production of short chain paraffins in the EU is approximately15000 t y-1. Illicit disposal of oils containing paraffins into the wastewater system may be asignificant route of entry into the environment. Approximately 50 % of the lubricating oilsused by industry in Sweden may be released into the air or WWTS.

Brominated diphenyl ethers (PBDEs) - This group of compounds is used, in descendingorder of importance, for flame retardation in high impact polystyrene, and in flexiblepolyurethane foam, textile coatings, wire and cable insulation and electrical connectors.Their use has expanded in the last decade in response to more stringent fire regulations andthe increased use of plastic material and synthetic fibres. Total annual global consumption iscurrently 40,000 t. A survey in Sweden (Sellstrom, 1993) indicated relatively high

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concentrations of the TeBDE and PeBDE congeners were present in sludge (15 and 19 µgkg-1, respectively).

Polychlorinated naphthalenes (PCNs) – The production of PCNs ceased in the US in 1977and by the mid-1980s in western Europe. This group of 75 compounds was used for cableinsulation, wood preservatives, engine oil additives, capacitors and as feedstock for dyeproduction, similar to that for PCBs. The principal releases of PCNs into the environment areprobably waste incineration and landfill disposal of items containing PCNs. PCNconcentrations in sewage sludge may be in a similar range to individual PCB congeners. InSwedish sludge, for example, PCN concentrations were in the range 0.15 – 0.16 µg kg-1 ds(Nylund, 1992). Some PCN congeners have dioxin like activity and have been assignedTCDD toxic equivalent values similar to those for coplanar PCBs (Hanberg, 1990) and sohave toxicological interest.

Quintozene (pentachloronitrobenzene) – A small amount of this compound is produced inthe EU (21.5 t y-1) and it is registered for use in the UK, Spain, Greece and Cyprus. It haslow water solubility and half-lives in soil and are reported to be in the range 5 – 10 months.

Polydimethylsiloxanes (PDMS) – These are nonvolatile silicone polymers used in industrialand consumer products including lubricants, electrical insulators and antifoams. They arehydrophobic and partition onto the sludge solids and concentrations in sludge were reportedin the range 290 – 5155 mg kg-1 in a survey of N American WWTS (Fendinger, 1997).However, PDMS do not bioconcentrate or exhibit significant environmental toxicity, but theyare relatively persistent in soil and can take months to years to fully degrade.

Nitro musks (chloronitrobenzenes) – This is a group of synthetic dinitro- trinitro-substitutedbenzene derivatives used as substitutes for natural musk in perfume and body careproducts, washing agents, fabric softeners, air freshners etc. Musk xylene and musk ketoneand the most extensively used forms and current world production of musk xylene is morethan 1000 t y-1. In Europe, current consumption is estimated to be 124 t y-1 for musk ketoneand 75 t y-1 for musk xylene and the release of these compounds to the environment isdominated by domestic discharges to sewer. Some biodegradation is likely duringwastewater and sludge treatment and the compounds will also biodegrade in soil with areported half-life of 12 days. There is only limited knowledge of toxicity to humans, but thepotential risks from sewage sludge use on farmland are probably minor relative to thegeneral exposure that is received from the use of these compounds in body care andwashing products.

Oestrogenic compounds – The endogenous oestrogens (17β-oestradiol and oestrone) andsynthetic steroids such as ethinyloestradiol, which is the active oestrogenic component inoral contraceptives, are considered to be far more significant in relation to impacts onaquatic ecosystems in discharged effluents from WWTS rather than from partitioning ontoparticulates and associating with sewage sludge (Alcock et al., 1999).

Pharmaceuticals – Compounds designed for specific biological effects in medical practicecan enter the wastewater system by excretion or residues and metabolites in urine and fromintentional disposal. This is regarded as acceptable practice because of the dilution receivedwithin the sewer system. Interestingly, 30 – 90 % of antibiotics administered to humans andanimals are excreted in active forms in urine. The potential toxicological and ecotoxicologicalactivities of these substances in the wider environment are generally unknown, but, becauseof their biological function, they are generally designed to be rapidly metabolised anddegraded. Many commonly used analgesic drugs are rapidly biodegraded during sewagetreatment including aspirin, ibuprofen and paracetamol (Richardson and Bowron, 1985).Most of these are soluble and exist primarily in the aqueous phase and transfer to sewagesludge is probably of only minor concern, although it is not possible to predict partitioningand fate during sewage treatment due to the absence of physico-chemical data for many ofthe compounds.

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Synthetic fats – In addition to Alcock’s list there have been recent concerns in the US aboutsynthetic fat substitutes that are used in food products. These compounds are not toxic tohumans, but are persistent and are not biodegraded in sewage and sludge treatment or insoil (pers. com. T.J. Logan, N-Viro, USA).Pharmaceuticals are covered in more detail in Case Study (d). Some of the bodycareproducts are covered in more detail in Case Study (e).

PolyelectrolytesPolyelectrolytes based on polyacrylamide and cationic copolymers are used extensively insludge treatment to aid dewatering. The polyelectrolyte concentration in mechanicallydewatered cakes is relatively high and typically in the range 2500 – 5000 mg kg-1 ds.Furthermore, they only degrade relatively slowly by abiotic processes in cultivated soil at arate of 10 % per year (Azzam et al., 1983). Acrylamide is a common monomer associatedwith polyelectrolytes and is potentially toxic to humans and is a reported carcinogen (IARC,1994). Concern about the implications for human health from the residual monomers inpolyelectrolytes used in drinking water treatment has resulted in them being withdrawn fromthis application in Japan and Sweden, and stringent controls on their use are in place inGermany and France (Letterman and Pero, 1990). These factors give rise to potentialenvironmental concerns about the long-term accumulation of polyelectrolytes in sewagesludge and sludge-treated soil.

The limited available evidence suggests that no environmental problems have beenidentified from the application of polyelectrolytes to soil. Acrylamide monomer residuals arerapidly biodegraded in the environment and are unlikely to represent a risk to human healthin sludge (Gustavsen, 1998). Furthermore, polyelectrolytes are frequently applied directly tosoil as soil conditioning agents to maintain soil structural conditions and as a carrier gel forfluid drilling pre-germinated seeds in agricultural and horticultural practice (Fordham andBiggs, 1985). In the fluid drilling technique seedlings of sensitive pregerminated cropsemerge directly from the injected polyacrylamide gel. This practical example illustrates thatpolyelectrolytes do not have a phytotoxic action in soil. An investigation of the potentialimpacts of polyelectrolytes on nitrification processes in soil (Cartmell et al., 1998) showed noeffect on nitrate accumulation at ten times the normal rate of addition of polyelectrolyte tosoil in sludge, but nitrification was partially inhibited when the dose was increased to 100xthe normal input. Polyelectrolyte was added directly to soil in this study to assess the worse-cased effects on nitrification processes in a controlled laboratory incubation. In practice,however, polyelectrolyte compounds are strongly sorbed to the sludge solids and this islikely to significantly reduce their availability and potential toxicity to soil microorganisms.

Compared to metals, organic pollutants have only recently been identified as havingpotential adverse human health effects. Most organic pollutants are present in theenvironment at very low concnetrations. Exposure to these compounds, for example throughdrinking water, is very low. However, as some of these compounds may bioaccumulate orhave effects at low concentrations chronic health effects are starting to be investigated forsome of these compounds. Table 4.3 presents acute and chronic health effects of the majororganic pollutants covered in this report, and the limits set for these compounds in drinkingwater.

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Table 4.3 Summary of health effects for organic pollutantsMCL in Drinking waterCompound

EU WHO USEPA

Acute Health effects Chronic Health effects Carcinogenicity Notes

DEHP 10 µg.l-1

EmulsifiedHydro-carbons

8 µg.l-1 6 µg.l-1 Usually low acutetoxicity: mildgastrointestinaldisturbances,nausea, vertigo

Damage to liver andtestes, reproductiveeffects, birth defects

Some evidence inrats and mice,mutagenic at highdoses

Synthetic substance,adsorbs to soil andsediments,bioconcentrates due tolipophilicity

PCBs 1.0 µg.l-1

Organo-chlorine

0.5 µg.l-1 Acne-like eruptionsand pigmentation ofthe skin, hearing andvision problems,spasms

Similar to acutepoisoning: irritation ofnose, throat, and GItract, changes in liverfunction andreproductive systems

Some evidence inrats

Adsorbs to soil andsediments, persistent,bioconcentrate, andaccumulate in fat

PAHs 0.2 µg.l-1 0.7µg.l-1

Benzo(a)-pyrene

0.2 µg.l-1

Benzo(a)-Pyrene

Red blood celldamage, anaemia,suppressed immunesystem

Developmental andreproductive effects,birth defects

Some evidence fromhumans and animals

Adsorbs to soil andsediments, weakdegradation, lipophilic,bioaccumulate

LAS 200 µg.l-1

SurfactantsGastrointestinalproblems

Kidney and liver toxicant Synthetic compound, toxicto marine life: LC50= 3-10mg/l, biodegrades easily inaerobic conditions, noaccumulation in sedimentsand biological tissues

NPE 200 µg.l-1

Surfactants1-5 µg.l-1 Allergies, endocrine

disrupting chemicalHas an oestrogeniceffect on fish

Lipophilic, mainly affectsaquatic organisms,partially biodegraded inthe environmentmetabolites are very toxicand persistent

PCDD/PCDF 0.00003µg.l-1

USEPAwebsite

Chloracne, rashes,discoloration,excessive body hair,large speciesdifferences in toxicity,very specific

Thymic atrophy, damageto the hormonal andimmune systems,spontaneous abortions,alteration in glucosemetabolism

Lymphoma,melanoma inhumans, canincrease risk ofseveral types ofcancer in man.

Highly volatile, binds tosoil and particles,accumulates in fat,persistent, birth defects inanimals

EU: Water treatment directiveWHO:Guidelines for drinking water quality Vol.2 (1996).USEPA:Standards for maximum permissible values in sewage sludge/soils. Estimating concern levels for concentration of chemical substances in the enviroment. WashingtonDC (1984). USEPA website: http://www.epa.gov/safewater/mcl.html

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5. REVIEW OF EU AND NATIONAL MEASURES TO REDUCE THE POTENTIALLYTOXIC ELEMENT AND ORGANIC POLLUTANT CONTENT OF WASTEWATER ANDSEWAGE SLUDGE.

5.1 Approaches to reducing pollutant content of wastewater and sludgeAs has been shown so far in this report UWW treatment is not sufficient to deal with allpollution. Most of the pollutants are concentrated in sewage sludge although differentcontaminants end up to varying degrees in effluent from WWTPs. This effluent ends up inaquatic environments. Concentrating pollutants in sewage sludge, while it may protectreceiving waters, presents a number of problems for its disposal or use on agricultural land.

Because of this, a variety of instruments have been attempted at national level, with varyingsuccess, on different levels to reduce UWW pollution closer to at source. These instrumentsare consistent with an overall strategy of waste minimisation, polluter pays, and reduction atsource and include individual regulatory, economic and voluntary and educationalinstruments.Table 5.1 Summary of Available Instruments

Instruments Regulation Economic Voluntary Educational

DirectMeasures

(Applying todirect inputs

to UWW)

Licensing ofdisposal to

UWWLicensing ofproduct use

Integration ofWater

Managementand UWW intoLand Use and

Spatial Planning

“Downstream”Charges and

taxes ondisposal to

UWWSubsidies for

alternativetechnologies,substances or

collectionservices

Trading ofpollution

reduction credits

Agreements andManagement

Systemsdirected at

wasteminimization or

alternativecollection

Research intoalternative

products andprocesses,

Managementand Waste

MinimisationTraining

IndirectMeasures

(applying toindirect or

diffusesources)

Product andSubstance Bans

Product andSubstance Use

LicensingIntegration of

WaterManagement

and UWW intoLand Use and

Spatial Planning

“Upstream”Taxes andCharges on

Particular InputsSubsidies for

alternativemanagementand products

Trading ofpollution

reduction credits

Agreements,ManagementSystems and

Codes ofPractice aimedat alternativemanagement

strategies

Research intoalternative

products andprocesses,

Managementand Waste

MinimisationTraining

Labelling andSpecific and

GeneralPublicity

Measures

The following selected examples demonstrate strategies for the minimisation of waste atsource through these instruments:

• Regulatory Instruments:• The Prohibition of Mercury Clinical Thermometers in France• Economic Instruments:• Wastewater Tax in Germany• Charges on Cadmium in Fertilisers in Sweden

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• Subsidies to Waste Collection in the Dry Cleaning Industry in France• Voluntary Initiatives:• Targeted Waste Collection in France• Waste Minimisation Strategy in Leather Industry in Italy• The Anjou Recherche Programme and Special Conventions in France• Local Initiatives, Promotion of Environmental Management and Cleaner Production in

Denmark.• Educational Initiatives• Provision of Consumer Information in France• Eco-labeling and LAS in Scandinavia• Awards Company Innovation in Waste Management and Minimisation in France

These measures may be differently targeted, to general and particular discharges or toproduct use and substitution, to specific processes and industries and households, or tospecific localities and catchment areas.

5.2 Regulatory Instruments

The use of traditional command and control instruments involves licensing of the use anddischarge of material to UWW. This approach is notoriously difficult to monitor and controleffectively. In the case of particularly difficult pollutants, targeted product bans can be a morecost effective solution where clear alternatives are available and the production of thematerial in question is relatively easy to regulate.

The Prohibition of Mercury Clinical Thermometers in France

Mercury is used in a broad area of applications. Industrial uses of mercury may be found inthe electro-technical industry to produce batteries or fluorescent tubes or in dental practicesfor fillings. Other industries require mercury for the production of instruments such asmanometers, barometers and thermometers. It is estimated that the total production ofmercury in France for industrial productions amount to an average of 60 tonnes per year.

Impact: Mercury is of particular health concern for not only are humans exposed to mercurydirectly when the product breaks, but they also may be exposed after it is disposed andaccumulates in UWW and sludge. It can also leach from landfills through the water wastepathway or be directly released in the wastewater system after disposal in sinks and toilets –raising concern for the risk of human exposure and environmental contamination. It is alsoestimated that before the introduction of the ban on the marketing of clinical mercurythermometers, an average of 15 – 20 million mercury-thermometers were in circulation inFrance. These are equivalent to 12 tonnes of mercury – about 20% of the yearly amountproduced. It is also estimated that an average 1.5 – 5 million of these thermometers werebroken every year, releasing between 3 and 10 tonnes of mercury in the environment yearlyand causing significant health and environmental concern. (Romp, 1993 in Öko Institut,Berlin).

Measure: Under the new French regulation [Arreter of 24 December 1999, Journal Officiel,31 Decembre], the marketing of clinical mercury thermometers is prohibited both nationallyand at EU level, as from March 1, 1999. However, the enactment does not prohibit export tonon-EU nations. The main objectives of the ban being that of minimising mercury waste andprevent its release into the environment, hence reducing risk to humans.

Costs: Although difficult to evaluate precisely, the implementation of the regulation resultedin two types of costs: 1) Reprocessing Costs associated to the reprocessing mercurythermometers and 2) Replacement Costs for the purchase of new (electronic or eardrum)devices. The average cost of an electronic thermometer ranges between 11 and 30 EURO

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per unit, whilst a professional device costs around 305 EURO. The cost of an eardrumthermometer (an infra-red device) ranges around 69 EURO. The professional timpanicdevice costs between 305 and 610 EURO. Assuming that at least 15% of mercurythermometers are collected in the first year, the total cost of the scheme amounts to about3.81 million EURO.Effectiveness: Although still at its early stage, the ban has been very successful and othernations such as Germany and Sweden are already following similar policy measures. It isestimated that the ban will reduce mercury waste by 12 tonnes a year – a reduction of 20%of the total annual mercury production in France. However, such action should also be takenfor other industrial sectors using mercury - and French policy is seemingly moving towardsthis direction – so as to further minimise the quantity of mercury present in the environmentand reduce the risk to human health and the environment.

Economic Instruments

Economic instruments may be used to provide a continuing incentive for reduced use of amaterial and even product substitution. Nevertheless; evidence suggests that taxes andcharges need to be carefully designed and targeted so as to promote continuing wastereduction. The social and economic implications of taxes and charges may make theirintroduction controversial, and the costs of their administration can make them as complex toadminister as command and control regulation.

Wastewater Tax in Germany

This measure is a “downstream” charge on discharges, and demonstrates how the creationof economic incentives alone may not be more effective in promoting reductions thancommand and control measures.

Measure: The Federal Wastewater Tax was adopted in September 1976, comprising a taxon sewage discharges; the amount payable was established on the basis of the amount andharmfulness of the waste. The objective of the tax was better regulation of discharges and topromote a reduction in wastewater pollution over all

Effectiveness: In terms of offering an overall incentive to reduce discharges, the impact ofthe measure was found to be limited. A study carried out by Jass (1990) conclude that whilethe charge had lead to considerable improvements of the purification of sewage and theenvironmental technology, it had not constituted any incentives for further improvement.Sewage taxes only created an incentive to comply with the command-and-control regulation,but nothing more (Lubbe-Wolf, 1996 in Cremer and Fisahan, 1998)

Charges on Cadmium in Fertilisers in Sweden

This is an example of a specifically targeted measure. Sweden has had environmentalcharges on nitrogen and phosphorus in commercial fertilizers since 1984. The “downstream”charge on phosphorus was abolished in January 1994 and replaced by an “upstream”charge on cadmium.

Measure: An important reason for the introduction of the cadmium charge was that it createsan ongoing incentive to reduce the concentrations (Swedish EPA, 1991) in fertilizers. SinceNovember 1994, the charge rate is SEK 30 (EURO 3.3) per gramme of cadmium if thecadmium content exceeds 5 mg per kg phosphorus (about 2.2 mg Cd per kg P2O5) (SwedishEPA, 1997).

Impact: According to the Swedish Board of Agriculture, the content of cadmium in fertiliserhas gradually fallen from 35-40 mg Cd per kg phosphorus (before the introduction of the

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charge) to about 23 mg in 1994/95 and 16 mg in 1995/96. The Board conclude that the Cdcharge in combination with the demand (for a low content of Cd in fertiliser) by theagricultural sector has kept Cd levels on a low level (Jörnstedt, 1998). Drake and Hell-Strand(1998) concluded that the combination of governmental policy (including the charge and astandard) and voluntary efforts has been successful in reducing the content of cadmium inphosphorus fertilisers. However, it was not possible to estimate the relative importance ofthe different measures.

Costs: State tax gross revenues in 1996 were around 10 million SEK (1 million EURO), andthe administrative costs are estimated to be around 1% of the gross revenues (Drake andHellstrand, 1998). The tax is administrated by the National Tax Board, together with the taxon nitrogen. Importers and producers report quantities and contents every month. The onlycontrol made seems to be ‘tax audits’ (concerning the accounting of the firms involved). In1999, 25 such audits were performed (corresponding to about one ‘man-year’). Theauthorities do not measure the actual cadmium content of fertilisers. Some problems withillegal imports by small firms of fertilisers from Poland and the Baltic states (probably ofRussian origin) are reported (Jörnstedt, 2000).

Subsidies to Waste Collection in the Dry Cleaning Industry in France

A very good example of an economic instrument is where competent authorities find itcheaper or more effective to subsidise alternative production processesses or wastecollection strategies than to treat a particular pollutant.

For the businesses using chlorinated solvents, currently only two main options are availableto deal with the waste: regeneration of the solvent (10%) and incineration (90%). Theproblem with regeneration is that the process may affect the added stabilisers, which arerequired for the stability of the solvent at high temperature and humidity. Hence, it isnecessary to treat the regenerated solvent with new stabilisers. Nevertheless, it is usuallycheaper than destruction of the waste, which in France costs around 4100 FF per tonne.However, generally only larger firms employ this solution.

Measure: Associations have been created in the dry cleaning sector in order to organisewidespread regeneration of solvents, and these are starting to cover other solvent usingsectors. These organisations co-ordinate the collection and treatment of wastes. TheAgences de l'Eau subsidise dry cleaners who adhere to these associations in order to helpfinance the cost of the collection of their wastes.

Costs: A typical dry cleaning facility produces 400 kg of chlorinated waste per year and thecost of collection and treatment is around 3500 FF per year, which represents about 1% oftheir revenue [ADEME, 1995].

5.4 Voluntary Instruments and Government Industry Cooperation:

“Voluntary” instruments encompass a range of policy initiatives, from industry initiativesbased on EMAS and other management and auditing systems, to agreements betweenorgans of the state and industry. Voluntary Instruments however are very rarely just that, andoften operate as part of a broader scheme of regulation, economic incentives (includingsubsidies) and public education exercises. The division of responsibility and costs betweengovernment and industry varies from case to case.

Some instruments amount to an indirect subsidy to industry, others enforce the polluter paysprinciple by creating an incentive to manage waste through the adoption of producerresponsibility, as in the packaging waste directive. The following examples include a rangeof instruments adopted by government and industry in cooperation, including targeted

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collection of problematic wastes, voluntary and regulatory or economically inspired wasteminimisation strategies. All imply some proactive governmental intervention.

Targeted Waste Collection in France

This measure constitutes a specific drive and effort by authorities to collect dangerous andharmful waste from homes. While effective in its own terms it is not a long-term solution tothe problem of discharges to UWW. It may be effective to deal with continuing risks ofcontamination from smaller and diffuse sources, and be used in connection with the adoptionof a longer-term waste minimisation and collection strategy and public education campaign.

Measure: One of the first targeted waste collection initiatives carried out in France was in1989, where 11,500 kg of waste products were collected over 16 days (see Table 5.2).

Table 5.2 Quantities of products containing pollutants collected at Savie in 1989[ADEME, 1995]

Product Amount (kg] Percentage of totalWaste Collected

Paint 5590 48.6%Solvents 2360 20.5%

Medicines 1010 8.8%Pesticides 1005 8.7%

Chemical laboratory products 70 0.6%Thermometers 65 0.5%

Costs: A more recent example is given by the 2 day collection in 1994 in the area ofBoisset-Gaujac (Gard) where 114 kg of chemical products, 169 kg of paint, 12 kg ofaerosols, 27 kg of solvents, and 29 kg of pesticide products were collected and dispatchedseparately for treatment at a cost of about 12,000 French francs [ADEME, 1997]1.

Effectiveness: Experience has shown that is it more efficient to appropriately equip wastereception centres for handling special wastes rather than organising random collection days.In contrast to targeted waste collection, the adoption of a comprehensive waste minimisationstrategy directed at particular sectors has demonstrated longer term and verifiable results.

Waste Minimisation Strategy in Leather Industry in Italy

This is another example of targeted intervention by public authorities, this time engaging aparticular sector in the long-term management of industry specific problems.

Italy is the principal European location for the leather and tannery sector in terms ofestablishments, employment and production. Tanneries are mostly small and medium sizedenterprises, with only 10% of them employing more than 20 people. The industrial areacovered by the plan is located in the Valle del Chiampo, in the province of Vicenza. The areaembraces 10 municipalities, of which Arzignano is the most important with more than half ofthe tanneries. The tannery area is the largest in Europe and supplies 50% of the Italianproduction. Manufacturing includes shoes, furniture and other leather goods (Vicenza Dept.of Environment, 1997)

Impacts: The tannery industry is considered to have important environmental effects onboth water and air, primarily because of toxic wastes generated by the large amounts ofchemicals employed during the various phases of the tanning process. The main pollutants

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being chromium III, sulphur, chlorides, solvents and organic wastes. (Vicenza Dept. ofEnvironment, 1997). The tanning process comprises two phases - covering and fixing, wherethe pelts are sprayed with paints and chemicals for treatment. It is estimated that around80% of the solvents used become hazardous waste and 1997data from the Dept. ofEnvironment of Vicenza indicate that the solvents used for the finishing process amount to20.000 tonnes per year, 85% for the covering phase, 15% for the finishing phase. Metoxypropanol is the solvent most used in tanneries and is classified as Class III hazardsubstance under Italian law, which has 5 classes of hazardous substances for whichmaximum allowed emissions are set.

Measure: In 1997, the Dept. of Environment of Vicenza introduced new measures to tacklethe environmental problems of the spray finishing process. The intervention focused on twomain aspects: 1) the introduction of innovative machinery, such as the more efficient lowpressure airbrushes or high pressure airless brushes instead of high pressure devices; plus2) the use of chemicals dissolved in solution. ‘Gruppo Conciario Veneto’ a tanneryassociation comprising 4 companies (La Veneta, Conciaria Adriatica, Sacpa and VenetaConciaria Valle Agno), with a total of 340 employees, 26.500 m2 of indoor plant and a 1997turnover of 103 million. EURO, adopted a scheme reducing use of, and ultimately replacing,the solvent metoxy propanol with alternative substances. Other substitutes are currentlybeing tested.

Effectiveness: As a result of the waste minimisation scheme, the use of Metoxy wassignificantly reduced from 7,300 kg in August 1998 to 1,300 kg in December 1998 (- 82%),but has been replaced by the less hazardous isopropyl alcohol (Class IV), which has in turnincreased in consumption. However, the total quantity of solvent used decreased from10,650 kg to 5,007 kg per month (- 53%) and therefore less toxic waste has been generatedand released by the tannery (no quantitative data available).

Costs: The reduction in utilised solvents also brought significant economic benefits, with areduction in monthly consumption equivalent to 64%, from 10,590 EURO in August 1998 to3,794 EURO in December 1998. These economic savings appear to be the main incentiveand persuasive tool to company innovation.

The Anjou Recherche and ‘Eaux Industrielles Initiatives’: Special Conventions inFrance

These cases are further examples of targeted intervention at the local level, representing aprogramme of sources and discharge identification, and the first example of a truly“voluntary” initiative through the adoption of negotiated approaches to the reduction ofproblematic discharges.

Measure: Anjou Recherché has studied five regions of France which have full statistical dataand 34 UWW collecting systems. This research programme allows many SME polluters tobe identified. Following the identification of industries that are potential polluters, a contractis drawn up between local industry and the mayor of the area. The contracts are calledspecial conventions for discharge, and about 2000 have now been set up, mainly in theSouth West region of France around Toulouse. The contract determines pollution reductionsand modifies discharge conditions into the UWW collecting system.

In Tours the "eaux industrielles" initiative was established in 1988. This initiative aimed tocontrol the quantity of discharges into the UWW collecting system based on monthlyanalyses on the sludge quality are carried out to control for pollution. In 1996 it was foundthat copper and mercury concentrations were very close to the recommended limit and muchhigher than the averages for France (Table 5.3).

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Table 5.3 Average concentrations (g t-1) in potentially toxic elements in the Toursludges [from ADEME, 1997] *French Standard NF U 44 041

Averageconcentration

(Tours)

Averageconcentration

(France)

French Standards

References LimitsCd 5.3 5.3 20 40Cr 96 80 1000 2000Cu 805 334 1000 2000Hg 7.3 2.7 10 20Ni 106 39 200 400Pb 153 133 800 1600Se 3.5 7.4 100 200Zn 977 921 3000 6000

Visits to the businesses potentially capable of releasing copper were undertaken and it wasfound that the problem initiated from a mirror manufacturing company. Effluents werecomposed of degreasing waters, discharges from the silver and copper adding steps.Certain observations were made: the neutralising pH in the treatment of the degreasingwaters was not alkaline enough to precipitate the metals and the pH was not checkedregularly enough. Thirty days after this intervention, the copper concentration showedsignificant reduction. For mercury, medical activities were investigated, as seen with theAnjou Recherché Case Study and it was discovered that the pollution was mainly due todischarges from two hospitals and one analytical laboratory.

Effectiveness: As the WWTP receives less contaminated wastewater, the manufacturer paysless for the treatment of wastewater. Furthermore, it avoids the cost of landfilling the sludge,as it can then be used in agriculture, decreasing overall costs.

Local Initiatives, Promotion of Environmental Management and Cleaner Production inDenmark.

In Denmark, about 45 industries perform wet treatment of textiles, i.e. pre-treatment, dying,printing and/or after treatment. The majority of these companies are located in the County ofRingkjøbing. This type of business consumes large amounts of water, energy andchemicals.

Impacts: Total consumption of chemicals is approximately 22,000 tonnes per annum(1998), of which approximately 18,000 tonnes per annum consists of basic chemicals(especially salts, acids and bases). The dye stuffs make up approximately 900 tonnes perannum and the residue consists of excipients such as detergents, phthalates etc. Thewastewater from wet treatment is typically heavily dyed and has large contents of salt,detergents, post-treatment agents and other chemicals. The total amount of wastewaterwithin this business makes up 6.6 billion cubic metres per annum. Most of the dye worksdischarge the wastewater at the municipal WWTPs but there are 4 dye works which havetheir own water treatment plants with a subsequent discharge to recipient waters. (Genbrugaf procesvand fra reaktivfarvning af bomuld, Miljøprojekt nr. 374 1998, p. 15. Miljø- ogEnergiministeriet)

Measure: The County of Ringkjøbing, in close collaboration local authorities and theFederation of Danish Textile and Clothing Industries set up a working party to launch and co-ordinate evaluation of chemicals used in the textile dye process. This working partyestablished a score system to govern chemicals. The score system is based on four

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parameters. Parameter A represents the estimated amount of the chemicals discharged intothe environment as wastewater. Parameter B is a score on biodegradability, C is a score onbioaccumulation, and D is a score on toxicity. Importantly, a lack of information about acertain product will automatically result in a high score. The score system was implementedin 1992-1993 and the dye works in the County of Ringkjøbing undertook to inform thesupervisory authorities about their consumption of chemicals according to the score system.

The reporting system allows both administrators and company managers to select prioritychemicals and subject them to close examination, on the basis of information filled in thechemical supplier’s specification sheets. For example, in "Egetæpper a/s" a company whichdyes and manufactures tufted carpets produces an annual score report. This report is usedby the company and the local environmental authorities scope environmental problems,analyze the consumption patterns, and to find opportunities for product substitution andcleaner production technologies. In particular, the company's purchasing policy requires anenvironmental impact assessment , including the score allotted, before products are bought..

Impacts: The number of products with a high score used in the production at Egetæppera/s has been reduced considerably between 1992 and 1997. This is partly because some ofthe products are no longer used by the company and partly because huge efforts have beenmade in order to procure missing data on the effects these products have on theenvironment. Egetæpper a/s' environmental report from 1995, the consumption of dye stuffsand excipients has been listed and in this report it is stated that:

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5.5 Education and Information

Education and Information Campaigns can be general or specific, comprising publiceducation campaigns and specific labelling initiatives. The first example is a generalinformation campaign addressing discharges to water the impact of which is difficult toassess. The second is a consumer information campaign that seeks to promote productsubstitution. The third is both an incentive and an information measure in the form of a prizefor effective environmental performance.

Provision of Consumer Information in France

Measure: Centre d’Information sur l’Eau based in Paris launched a campaign to alert peopleto the problems caused by disposing of compounds down the sink or the toilet. In order to doso, they distribute a leaflet where they note that the following substances should not bedisposed of into UWW which include;

• leftover weed killers or fertilisers• out of date or opened medicines• car oils• hydrocarbons• leftover paint or varnish• pesticides and other protecting products

The leaflet also stresses that collection systems are in place in many areas and thatinformation about these is available, and that using a waste disposal unit when connected tothe UWW collecting system is prohibited, as is discharging rainwater from gutters into theUWW collecting systems when a separate system is in place in the area. The centre hadover 7000 enquiries in 1998.

Eco-labeling and LAS in Scandinavia

Measure: Eco-labeling. Initiatives such the ‘Nordic Swan’ and ‘Good Environmental Choice’-were developed by the Swedish Society for Nature Conservation (SSNC) and the DanishSociety for the Conservation of Nature (DSCN) respectively. These were aimed at phasingout undesirable chemicals in washing powders and have significantly reduced the use of pLAS (Linear Alkyl Benzene Sulphonate), and Nonylphenol. Eco-labelling campaigns werelaunched by both the SSNC and DSCN, against washing powders containing LAS, whilst atthe same time promoting the use of ‘Swan’ or ‘Good Environmental Choice’ labelled washingpowders. In the past decade, the eco-label ‘Good Environmental Choice’ has become one ofSSNC most important tools for dealing with environmental problems. The initial campaignwas launched in 1988 with the publication of one of many green consumer guides, whichincluded information about household chemicals - some of which were found to bereplaceable with more suitable alternatives.

Use: LAS is the most important surfactant used in Europe and is found in considerablequantities in washing powders (up to 50%). Some 300000 tonnes of LAS were sold inEurope in 1999, but concerns for its environmental threat have risen since LAS is toxic toaquatic animals and plants. LAS is only degradable in the presence of oxygen, which is notalways present in wastewater treatment plants. In addition, it is estimated that about 15-20%of the amount of LAS entering a water treatment plant ends up in sludge with no furtherdegradation (S. Hagenfors, 2000).

Effectiveness: Products which do not contain LAS or Nonylphenol and have gainedconsiderable market share since the launch of the campaign. “In Sweden, products with thislabel accounted for 95% of sales by 1997 and consumer choice lead to a decline in use of

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LAS from 6,300 tonnes per year to 260 tonnes per year” (Danish Environmental Agency2000). The published guide was a success and prompted consumers to demand for moredetailed information about domestic products and their potential environmental impact.Coincidentally one the main environmental issues in the region at the time was that ofwidespread eutrophication of Scandinavian water bodies which subsequently induced ademand by consumers for environmentally friendly products. In brief, public awarenessmade the market for eco-labelled products grow in Sweden, so that after the firstmanufacturer (Unilever) launched its eco-labelled detergent, other multi-nationals had tofollow.

In Sweden, eco-labelling has however been limited in terms of product range and multi-national manufacturers have not introduced other eco-labelled products, with the exceptionof a few minor producers that specialise in ‘green’ products. In other Scandinavian countries,the market share of ‘Swan’ products is significantly smaller, with 15% in Finland and less inNorway and Denmark; the difference in share thought to be due to less public awarenesscampaigning. Yet, in the last ten years, the Swedish market for eco-labelled products hasundergone important modifications. It is estimated that presently 60% of the chemicalingredients in soap, shampoo, detergents have been substituted or removed, the remaining40% of chemicals falling within the list of substances approved by the ‘Swan’ and ‘GoodEnvironmental Choice’ guidelines. The 38000 tons of chemicals subject to substitution havebeen replaced by 29000 tons of approved substances, the difference (9000 tons) havingbeen removed. The substitute chemicals are all approved by eco-labels and are morebiodegradable in sewage treatment plants and sludge, meaning less environmental risk andless treatment costs. As a result, it is estimated that the use of LAS in Sweden has beenreduced by 95% (S. Hagenfors, 2000).

In addition, the development of the eco-label ‘Swan’ has prompted Stockholm Water Co. toinitiate a plan for the identification of measures to reduce hazardous household wastes beingflushed down the drain and into WWTP. This lead to the establishment of ‘environmentalstations’ or collection points and to the launch of a thorough public awareness andinformation campaign about the impacts of household products on the aquatic environment

Awards for Company Innovation in Waste Management and Minimisation

In 1996, the trophy ADEME "Economic and clean technologies" went to the STEN society,which is a metal finishing company which managed zero cadmium discharges byconcentrating the cadmium-containing effluents through evaporation and recovered themetal through electrolysis [ADEME, 1997].

5.6 Conclusion

The review of National Practices with respect to UWW demonstrated a range of problems,and identified some strategies for dealing with diffuse and small-scale sources of waterpollution.

Without upstream control, UWW treatment cannot treat all wastewater pollutants, resulting ineffluent and sludge quality problems. A range of measures are likely to be most effective inmitigating problems, and the appropriate policy mix will vary according to the pollutant,industry structure, local conditions and social factors, and there should be room for localdiscretion.

Product Bans in particular may be appropriate where product substitution is a possibility.Economic Instruments such as taxes and charges can, if carefully designed and targeted,provide appropriate incentives for alternative approaches by industry and producers ofwaste. Where the costs of treatment are transparent, it is more likely that least cost

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alternatives will be pursued through the use of subsidies to waste reduction or collection.Targeted Voluntary and Educational Initiatives may assist but the full costs and benefits ofspecific instruments are difficult to assess.

Full Cost Recovery for Waste and Water Services (under the Water Framework Directive)will assist in making the costs of dealing with particular pollutants more transparent.Nevertheless, the use of economic and alternative instruments and strategies should berecognized as the legally appropriate implementation measures in both EU and NationalLegislation, and the competent authorities should be legally empowered to establish andparticipate in alternative measures, including substitution and recovery schemes.

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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

Removal efficiency 64%, max. effluent 1.2µgl-1, insurface waters 0.28µgl-1.

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

Removal efficiency 90%, max. effluent 3.4µgl-1, insurface waters 0.53µgl-1.

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.

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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).

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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).

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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

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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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7. REPORT SYNOPSIS, DISCUSSION AND CONCLUSIONS

7.1 Report Synopsis

7.1.1 Scope of the study

The European Union is introducing an integrated policy to promote sustainable water usethrough the Water Framework Directive-2000/60/EC (WFD) based on a long-termprotection of water resources. A strategic aspect of this is the progressive decrease incontaminant discharges to the aquatic environment. Allied to this water policy is theamended proposal by the European Commission (COM/2001/17) listing 32 prioritysubstances to be phased out and 11 hazardous substances that will be subject toemission controls and quality standards in accordance with Article 16 of the WFD.Furthermore, Directive 86/278EEC concerning the use of sewage sludge in agriculture(USSA) is undergoing major revision of the quality standards for potentially toxicelements and organic contaminants in sludge and metal concentrations in sludge-treatedsoil. The urban wastewater (UWW) collection system is one of the main facilities used forthe disposal of many types of commercial and domestic wastes that contain potentiallytoxic elements and organic pollutants. It also receives unintentional diffuse inputs ofcontaminants in storm run-off from paved surfaces. However, wastewater treatmentprocesses are effective in mitigating the discharge of many substances to surface waterswith the treated effluent. This is achieved principally through aerobic biologicaltransformations or because a significant proportion of the contaminant load is transferredto the sewage sludge. Centralised collection and treatment of urban wastewater andsewage sludge is therefore a pivotal link in defining the pathways and fate ofcontaminant releases to the environment.

A major review of scientific, operational and regulatory information and literature frompublished sources, research institutions, industrial operators and environmental agencieshas pooled current knowledge and developed a data-base of the sources and pathwaysof pollutants in the wastewater treatment system (WWTS) within the European context.

The main objectives of this study were to determine the multiple sources of potentiallytoxic elements and organic pollutants entering UWW, to determine quantitatively theamounts of pollutants passing into the influent of wastewater treatment plant (WWTP),and to assess the effects of treatment processes on the fate of both inorganic andorganic pollutants in sewage sludge and treated effluents. This information was to beused to identify problems, to make recommendations and enable optimum reduction ofpollutant inputs, and to identify areas where data are lacking and where further researchis required. The significance of the inputs and types of contaminants from differentsectors connected to the wastewater collection system and potential opportunities andpractical measures to reduce the extent of contaminant discharges to sewer arediscussed. Potentially toxic elements are routinely measured in wastewater and sludgeand an extensive data-base of information has accumulated on the concentrations andfate of these elements in environmental media. Several important groups of organiccontaminant have also received considerable attention (eg PAHs, PCBs, PCDD/Fs), but,in general, the amount of published information on organics is limited compared withpotentially toxic elements due to the high cost of analysis, and need for specialistanalytical facilities and the absence of standardised methodologies.

More information has been identified for potentially toxic elements, which have beensubject to more detailed monitoring, as they are relatively easy to analyse using routineand inexpensive tests and environmental or food chain impacts are much easier toquantify where high soil loadings have occurred. Much less information is available on

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organic pollutants where some six thousand or more compounds have been identified.Many of these cannot be routinely determined in the majority of laboratories due to lackof appropriate instrumentation, the absence of a unified methodology and expense.

The type of pollutants and the magnitude of loadings to the WWTS system are acomplex function of:

• Size and type of conurbation (commercial, residential, mixed);• Plumbing and heating systems;• Domestic and commercial product formulation and use patterns;• Dietary sources and faeces;• Atmospheric quality, deposition and run-off;• Presence and type of industrial activities;• Use of metals, and other materials in construction;• Urban land use;• Traffic type and density;• Urban street cleaning;• Maintenance practices, for collecting systems and stormwater controls;• Accidental releases.

Quantitative information on both metal and organic pollutants in urban runoff arising fromanthropogenic activities is difficult to evaluate, due to the lack of information onbackground levels of these substances in the environment. Background concentrationsrelate to natural geochemical sources and biological sources, and include amounts insoils, dusts and waters derived from historical pollution.

The specific potentially toxic elements (PTEs) and organic contaminants considered inthis report are listed in Table 7.1.

Table 7.1 Potentially toxic elements and organic contaminants examined in thereview of pollutants in urban waste water and sewage sludge

Potentially toxic elements Organic contaminants

Zinc (Zn) Linear alkylbenzene sulphonates (LAS)

Copper (Cu) Nonylphenolethoxylates (NPE)

Nickel (Ni) Di-(2-ethylhexyl)phthalate (DEHP)

Cadmium (Cd) Polycyclic aromatic hydrocarbons (PAHs)

Lead (Pb) Polychlorinated biphenyls (PCBs)

Chromium (Cr) III and VI Polychlorinated dibenzo-p-dioxins (PCDDs)

Mercury (Hg) Polychlorinated dibenzo-p-furans (PCDFs)

Platinum group metals (PGMs) Nitro musks (chloronitrobenzenes)

Other PTEs* Pharmaceuticals

Oestrogenic compounds:

Endogenous forms: 17β-oestradiol, oestrone

synthetic steroids: ethinyloestradiol

Polyelectrolytes (polyacrylamide)

Other organics***- such as Arsenic, Silver, Molybdenum and Selenium **- such as Adsorbable organo halogens (AOX) and chlorinated paraffins

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Specific case studies on the following key issues are included in greater detail in Section6

• Platinum group metals (PGMs),• Sustainable urban drainage,• Artisanal activities,• Pharmaceuticals,• Body care products and fragrances,• Surfactants,• Polyelectrolytes in sludge treatment and dewatering,• Influence of chemical phosphorus removal on potentially toxic element content in

sewage sludge.

The review has emphasised that an assessment of the contaminants entering UWW andsludge cannot be divorced from an understanding of their potential impacts on theenvironment, and particularly in relation to the use of sewage sludge in agriculture, asthis provides the context for defining the significance of UWW and sludge contamination.

Finally, areas where there is a deficit of information concerning particular sources orcontaminant groups, or their behaviour in the WWTS are identified andrecommendations for further investigations to inform decisions about the direction offuture research requirements have been formulated.

7.1.2 Potentially toxic elements

Sources of potentially toxic elements entering the WWTS

The sources of potentially toxic element contamination in the wastewater system havebeen classified into three main categories:

• Domestic;• Commercial;• Urban runoff.

The information that was collected identified domestic inputs as the largest overallsources of Cu, Zn and Pb entering the UWW system, whereas commercial sourcesrepresent the major inputs of Hg and Cr (Table 7.2). Commercial discharges contributemoderately larger inputs of Cd (30 – 60 %) compared with domestic sources (20 – 40%). Estimates of Ni loadings from domestic sources are highly variable and thiscontribution may be similar to, or greater than, the commercial input. The size of the run-off contribution depends on climate and traffic intensity, and can be a significantproportion (>20 %) of the total metal load for Cd and Pb, but it is a relatively minorsource of Cu and Hg. However, storm discharge was identified as an important source ofHg in a survey of metal inputs to the river Rhine (French region), contributing 15 % of thetotal load of this element to the river. Chromium, Ni and Zn represent an intermediategroup and run-off typically contributes <20 % of the total input of these elements toUWW.

There is a degree of uncertainty in the mass balance, however, and a significantproportion of the load is from unidentified sources, which represent ≥50 % of the totalinput of Cr, Ni and Zn. Unclassified inputs of Cu, Hg and Pb represent 20-40% of thetotal loading of these elements, whereas >80 % of the Cd is from identified sources.Nevertheless, the relative magnitude of the domestic and commercial inputs of metals toUWW is highly consistent with the estimated total loadings from these sources to majorriver systems in Europe (Table 7.3). Strategies to reduce discharges of metals to sewer

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can only focus on those sources that can be identified and quantified. Therefore, giventhe extent of the uncertainty in establishing the sources and inputs of certain elements, apriority area for research would be to determine these contributions and to develop acomplete mass balance for potentially toxic elements entering to the UWW system indifferent European countries.

Cadmium and Hg have been proposed as priority hazardous substances by theEuropean Commission and will be subject to the controls required by the WFD to end allreleases of these elements to the water environment within a 20-year period. The statusof Pb as a priority substance is currently under review. Nickel is not designatedhazardous, but is classified as a dangerous substance and emission controls and qualitystandards will also apply to this element as required by the WFD.

Table 7.2 Provisional potentially toxic element load from different sources enteringUWW in EU countries (% of total input)

PTE DomesticWastewater

CommercialWastewater

Urban Runoff

Zn 30-50 5-35 10-20Cu 30-75 3-20 4-6Ni 10-50 30 10-20Cd 20-40 30-60 3-40Pb 30-80 2-20 30Cr 2-20 35-60 2-20Hg 4-5 50-60 1-5

The ranges in Table 7.2 are estimates from various sources in this study and may notadd up to 100%. Quite a high proportion of the Zn, Ni and Cr are not identified andattributed to certain sources, and more research is needed in calculating the massbalances of these PTEs in the urban environment.

Table 7.3 Diffuse and point sources of potentially toxic elements entering the riverRhine reported in 1999 (% of total input)

Point sources(1) Total diffuse sources(2)PTEDomestic Industry

Zn 30 10 60Cu 20 15 65Ni 20 20 60Cd 10 10 80Pb 15 10 75Cr 15 25 60Hg 20 20 60

(1)Points sources include discharges through the UWW system(2)Diffuse inputs enter surface water directly

Despite uncertainties in identifying all the metal inputs there is ample evidencedemonstrating the significant reduction in metal discharges to the WWTS, principally dueto implementation of effective trade effluent controls, more efficient processes andchange in industrial base. This is reflected in the declining concentrations of metalspresent in sewage sludge and in influent wastewaters (Section 2.1) and was reported inall countries where data could be abstracted on temporal trends (The Netherlands,France, UK and Sweden). Examples of the general declining trends in metalconcentrations in sludge over the past 20 years in European countries are shown inFigure 7.1 for Zn and Cd for a major WWTP in the UK. In Sweden, Cd inputs declined by60 % during the period 1992 – 1998 and Cr, Hg and Pb were reduced by 40 – 50 %. Zn

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and Ni declined by 10 % during the same period, and there was no change in Cugenerally reflecting the importance of the domestic contribution of these elements inUWW. Metal concentrations in rivers receiving treated UWW have also shown a markeddecline. For example, concentrations in the river Thames have declined on average by50 % in the period 1986 – 1995.

Figure 7.1 Reduction in (a) zinc and (b) cadmium concentrations (untransformedand log10 transformed data, respectively) in sewage sludge from Nottingham STW,UK during the period 1978 – 1999

Measures to reduce metals discharges from domestic sources

Faeces comprise a significant proportion (>20 %) of the load of Cd, Cu, Zn and Ni(equivalent to 60-70 % of the total amount of these elements in domestic contribution towastewater). Faeces typically contain 250 mg Zn kg-1, 70 mg Cu kg-1, 5 mg Ni kg-1, 2 mgCd kg-1 and 11 mg Pb kg -1 dry solids. The residual sewage sludge from WWTS generallyhas significantly larger concentrations of metals than those present in human faecesbecause of the contributions from other sources.

For Cd, however, which is the most labile zootoxic element in UWW and sludge and itsrelease is of particular concern in the environment, the weighted average content insludge is comparable to the concentration in human faeces. Whilst other sources of thisimportant element can be identified, solids inputs and secondary sludge productionduring WWT may have a significant dilution effect on the mass input and transfer to thesewage sludge. Consequently, a critical balance between input concentration inwastewater and dilution processes may control the final Cd content in sludge to withinthe normal range for faeces, which represents the minimum potential backgroundconcentration value. The precise nature of these mechanisms could be identified tobetter understand the behaviour of Cd in WWTS and in relation to effects of measures toreduce Cd inputs on sludge quality. Whilst the opportunities for further reducing Cdinputs would generally appear to be limited, removing phosphate from detergentformulations has successfully reduced Cd discharges from domestic sources in TheNetherlands and Sweden (Section 2.1.2).

Tot

al Z

n (m

g kg

-1 d

s)

400

600

800

1000

1200

1400

1600

1800

2000

1978 1983 1988 1993 1998

(a) Zinc (b) Cadmium

Tot

al C

d (L

og m

g kg

-1 d

s)0

0.2

0.4

0.6

0.8

1

1.2

1.4

1.6

1.8

2

1978 1983 1988 1993 1998

Year

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However, the principal sources of metals in domestic wastewater are body careproducts, pharmaceuticals, cleaning products and liquid wastes (eg paint). Plumbing isthe main source of Cu in hard water areas, contributing >50 % of the Cu load. It mayalso be a principal source of Pb where this was used historically as a means of drinkingwater conveyance, accounting for 25 % of the input to WWTS. Lead solder in copperpipes may also be an important source.

One approach to reducing the domestic input of contaminants is for homeowners to beadvised on the proper disposal of household wastes, for example the ‘bag it and bin it’campaign run by some UK water companies . This will require co-operation between thewastewater collector and regional authority responsible for municipal waste disposal toestablish and publicise schemes and accessible facilities for the disposal of liquid wastesby homeowners. Metals (eg Zn) are also important active ingredients in other commonlyused household products, but it may be impracticable to eliminate these althoughmanufacturers minimise their excessive use, for example in packaging. An assessmentand survey of the practical implementation of voluntary collection schemes for liquidwaste and a programme of pilot studies is recommended to assess attitudes and theextent of homeowner participation.

Commercial sources and inputs of potentially toxic elements to UWWRegional studies of metal emissions from commercial premises indicate that furtherreductions in discharges of most elements of concern could be potentially achieved fromthis sector (Section 2.1.2). The main sources that may be the focus of furtherinvestigation include health establishments, manufacturing industry and hotel/catering.Approximately 30 % of medical establishments and 20 % of the other activities may bedischarging significant amounts of potentially toxic elements in wastewater. In particular,Hg in UWW and sludge is largely attributable to dental and medical practices. Dentalamalgam separators are effective at reducing Hg emissions where their use iscompulsory and there is evidence, for example following the Danish ban, thatalternatives to Hg used in thermometers and other products can reduce discharges tothe WWTS.

Large industrial installations are required to treat wastewater to an appropriate standardbefore discharging to the WWTS and are subject to rigorous trade effluent controlstandards. However, small commercial, artisanal enterprises that are connected toWWTS may also be a potentially significant source of contaminants. One of the largestagglomerations of artisanal activities in Italy was investigated in Case Study (c) . ThisCase Study identified vehicle repair shops, metal processing and jewellerymanufacturers as principal artisanal activities potentially responsible for dischargingsignificant amounts of PTEs to the UWW system (table 7.4). However, many smallenterprises pretreat wastewater before discharge and this was found to markedly reducethe extent of actual metal discharges to sewer. In the Vicenza Case Study, for example,pretreatment reduced the input of Zn, Cu and Pb from artisanal activities to <10 % of thetotal load entering the WWTS and Cr and Ni inputs were reduced to <2 and 0.5 % of thetotal input. Cadmium achieved the smallest overall reduction and approximately 40 % ofthe total load of this element to the WWTS, was identified as coming from artisanal andcommercial activities, principally goldsmiths and car repairers. Further detailedassessment of wastewater management by these types of small enterprises isnecessary in other European countries to determine the extent and effectiveness ofwastewater treatment prior to discharge and whether specific measures andrequirements are necessary and warranted to reduce potentially toxic elementcontamination from these sources.

Incomplete information regarding these sources could account for a major proportion ofthe unidentified inputs of metals to UWW (Table 7.2). The identification and control of

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these inputs is practiced in a number of regions in France by operating registrationschemes and inventories of all discharging business premises/activities connected to theWWTS within specific catchment areas. Premises are inspected and the necessarylevels of remedial action are agreed with the owner. Small business enterprises may bean important source of potentially toxic elements and the extent of these inputs to UWW,and the level of process wastewater treatment or segregation, should be examined inother European countries to assess the effectiveness and practicability of targetingartisanal activities for specific remedial action.

Other investigations in Sweden and Norway confirm the importance of motor industries,vehicle workshops and washing facilities (particularly heavy goods vehicles) as majorsources of potentially toxic elements in UWW. Oil separating devices may be fitted as ameasure to reduce wastewater contamination from these activities. However, theformation of stable emulsions in the wastewater by detergent microemulsion formulationsused for vehicle washing tend to limit the effectiveness of these systems at reducingpollutant emissions.

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Table 7.4 Total metal loads (mean values) from artisanal activities (in discharged and segregated wastewater) in Northern Italy

Activity(number ofenterprises(1))

Zn% total load toWWTS (mean)

Cu% total load toWWTS (mean)

Ni% total load toWWTS (mean)

Cd% total load toWWTS (mean)

Pb% total load toWWTS (mean)

Cr(III)(4)

% total load toWWTS (mean)

Food (52) 0.28 0.09 <LOD <LOD <LOD <LODCar repair (174) 6.4 8.5 0.3 6.3 34.0 1.0Ceramic (24) 0.01 0.001 <LOD 0.14 3.1 1.1Galvanic (17) 0.07 0.84 0.67 0.39 1.5 0.15Printing (141) 0.24 2.0 < LOD 2.8 0.59 0.6Wood (92) 0.29 0.46 <LOD <LOD 5.5 0.76Metallurgist (155) 8.7 5.3 15.6 2.3 25.7 4.7Dental (88) 0.16 0.02 <LOD 0.25 <LOD 0.05Gold (253) 2.1 7.6 <LOD 32.2 <LOD 0.14Hairdressing (316) 0.99 0.69 <LOD <LOD <LOD <LODLaundry (88) 0.49 0.09 0.04 2.3 0.45 <LODTotal(2) 19.9 25.5 16.6 46.6 71.1 8.9Total(3) 8.8 9.5 0.46 36.9 7.9 1.8

(1)Total number = 1579 enterprises (2)Calculated on basis of no pre-treatment and discharge directly to UWW system.(3)Based on discharge of pre-treated wastewater to the UWW system (car repair, ceramic, galvanic, printing, wood processor and metallurgy enterprisesusually segregate and pre-treat process water to comply with the standards for discharge of industrial wastewater in Italy). (4) Cr VI was discharged in the effluent from galvinising shops, but was not detected any other sample or in the influent to the WWTS. Cr VI is rapidlytransformed to Cr III on contact with organic matter in sewage and sludge. Therefore the Cr VI input from galvanising shops was calculated as a proportion ofthe total Cr load to the UWWS (as Cr III). Loadings >5 % of the total input to the UWWS system are highlighted in bold and <LOD=below limit of detection

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Urban run-offMetal inputs to WWTS in urban run-off are highly variable and depend on traffic,surface and roofing materials and age, and environmental factors (Section 2.1.3).The principal sources of PTEs in urban run-off include:

• Wet and dry atmospheric deposition;• Degradation of roofing materials;• Construction;• Road and vehicle related pollution;• Litter, vegetation and associated human activities;• Soil erosion.

Urban run-off is potentially a significant source of Cd and Pb, contributing up to 40 %of the total load of these elements entering the WWTS (Table 7.2). The removal ofPb from fuel additives has significantly reduced runoff inputs of this element.Precipitation represents a comparatively small proportion of the total bulk input ofmetals in urban run-off. Rainfall contributes approximately 10 % of the Cd, Pb and Cuand only about 1 % of the Zn load in run-off. The majority of elemental inputs arederived from dry deposition onto paved surfaces, surface degradation processes aswell as from vehicle emissions.

Roads and vehicles have been investigated as potential sources of PTEs enteringWWTS. Lead, Cu, Cd and Zn are present in high concentrations in brake linings andamounts of Cd and Zn are also significant in tyre rubber. Tyre abrasion and brake-lining wear appear to be more important as sources of Cd and Zn transfer to runoffwater. Whilst these metals may be released during brake-lining wear, the amounts ofPb and Cu entering the drainage system in runoff is relatively small. Run-off is aminor source of Cu entering the UWW compared with the domestic load fromplumbing. Nickel is the main element released from the abrasion of road surfaces.Roofing materials influence concentrations in runoff and Zn is significantly increasedby galvanised surfaces. Lead in window frames and roof sheeting and also Pbpainted surfaces can contribute significant amounts of this element to runoff water.Interestingly, the concentrations of Cd, Cu, Pb and Zn in roof runoff may becomparable to, or greater than, those in road run-off suggesting that atmosphericdeposition onto paved surfaces may be more important as a source of metals in run-off than vehicle or road related inputs. Concrete structural materials transfernegligible amounts of potentially toxic elements to run-off. The relative size of theinputs of different potentially toxic elements to UWW in run-off can be ranked inincreasing order as: Cd<<Ni<Cu<Pb<<Zn. Although there is more Zn in urban run-offthan any other metal, this represents a relatively minor input to the WWTS (<20 %)compared with domestic (50 %) and commercial (<35 %) discharges of Zn to sewer(Table 7.2).

The available mass balance information suggests that the overall contribution ofurban run-off to the metal flux in UWW is generally comparatively small, andintermittent depending on rainfall, relative to other sources (Table 7.2). However, thevariable nature and uncertainty about the extent of these inputs could partly explainthe apparent discrepancy between the identified sources of PTEs and the actual loadentering the WWTS. The significance of potentially toxic elements in urban run-offcould be quantitatively examined by relating rainfall frequency, duration and intensitywith temporal trends in the metal content of sewage sludge sampled from WWTSserving different catchments types (industrial, domestic, mixed). There are a numberof possible measures that can potentially reduce contaminant entry into UWW fromrun-off. For industrial and other paved areas, for example, where it is practicable,interception systems to remove contaminants bound to suspended solids by

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sedimentation or filtration can reduce inputs to the WWTS. However, these are onlymoderately effective because the smallest particles containing the largestconcentrations of metals do not settle efficiently. The design and management ofthese systems are described in Case Study (b). In general, however, it is difficult topractically control or minimise run-off contamination due to the diffuse nature of thesources of metals, although targeting Pb painted surfaces for source preventiveaction would further curtail inputs of this element into the environment.

Metal transfers to sewage sludge during WWTThe majority of the polluting load in UWW, including potentially toxic elements, istransferred to the sewage sludge during WWT and a relatively small soluble orsuspended fraction may be discharged to the environment in the treated effluent.

Empirical and mechanistic models of metal transfer (TOXCHEM) have beendeveloped. Both modelling approaches can be used to predict the metalconcentration in wastewater on the basis of the content in sludge. Metal removalduring primary sedimentation is a physico-chemical process, dependent on thesettlement of precipitated, insoluble metal or the association of metals with settleableparticulate matter. The removal of metals during secondary wastewater treatment isdependent upon the uptake of metals by the microbial biomass and the separation ofthe biomass during secondary sedimentation. Approximately 60 – 80 % of mostelements in raw sewage are removed and transferred to the sludge. Nickel is themost soluble element and removals of 40 % are more typical for this element. Thearithmetic mean metal concentrations in sewage sludge in European countries arelisted in Table 7.5.

Mass balance of metals in the WWTS must take into account metal contents andtheir fates in sewage sludge. Detailed environmental, toxicological andecotoxicological assessments of metals provide assurance that the concentrations ofCd, Cu, Ni, Cr, Hg and Pb in sewage sludge are well below values that constitute apotential risk to soils and the environment when sludge is applied to agricultural land.Zinc, on the other hand, presents the principal potential risk to crop yields fromphytotoxicity to affecting soil microbial processes. However, it is also an importanttrace element and the environment can be protected from the potentially toxic effectsof Zn in sludge by technically based, but precautionary soil limits.

Table 7.5 Arithmetic mean metal content in sludge and maximum permissiblelimits in soil and sludge (mg kg-1ds) in the EU(1)

Element Mean 86/278/EECmaximum in

sludge(range)

Proposed EUmaximum in

sludge

86/278/EECmaximum insoil (range)

ProposedEU

maximum insoil (pH 6-7)

Zn 863(2) 2500-4000 2500 150-300 150Cu 337 1000-1750 1000 50-140 50Ni 37 300-400 300 30-75 50Cd 2.2(3) 20-40 10 1-3 1Pb 124 750-1200 750 50-300 70Cr 79(4) 1000 60Hg 2.2 16-25 10 1-1.5 0.5

(1)Data are reported for 13 countries: Austria, Denmark, Finland, France, Germany,Greece (Athens), Ireland, Luxembourg, Norway, Poland, Sweden, The Netherlands and UK.(2)Excludes Poland and Greece (represented by Athens WWTS); the mean Zn content inPolish sludge and sludge from Athens WWTP is 3641 and 2752 mg kg-1 ds, respectively. Themean European value including Poland and Greece is 1222 mg Zn kg-1.

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(3)Excludes Poland; the mean Cd content in Polish sludge is 9.9 mg kg-1 value. The meanEuropean value including Poland is 2.8 mg Cd kg-1 ds.(4)Excludes Greece; the mean Cr content in sludge from Athens WWTP is 886 mg kg-1. Themean European value including Greece is 141 mg Cr kg-1.

Variations in metal concentrations in sewage sludges observed between differentcountries are difficult to reconcile because all European governments, localenvironmental agencies, water utilities and municipal authorities actively enforcetrade effluent control. Metal concentration data are most widely reported for sludgethat is used in agriculture and lower values may be apparent where nationalregulations enforce stringent metal limits for sludge, although this approach tends todiminish the opportunities for utilising sludge on agricultural land. These data aretherefore specific to particular sludge sources and do not provide an overallassessment of sludge contamination with potentially toxic elements disposed of byalternative routes such as incineration. However, certain countries report lower metalcontents in sludge than in other states, even when sludge limit values do not appearto be generally restrictive of sludge recycling to land.

A possible explanation may be related to the statistical characteristics of metalconcentrations and whether data are reported as arithmetic or weighted averages.The data are usually skewed and the median values are usually lower than themeans. This would give rise to large discrepancies in amalgamated data fromdifferent member states. Variations could also be related to the efficiency of differentanalytical methods at extracting metals from the sludge matrix. The size of WWTPmay also be important since sludges from large works contain more metals than fromsmaller works. This could be interpreted as the result of higher industrial inputs ofmetals to large works, although it may also be explained by the greater interceptionof atmospheric deposition of metals by paved areas in urban centres served by thelargest sewage treatment works. A more detailed assessment and examination of thesludge qualities and quantities from different types of treatment centre in relation todisposal outlet in European Member States is necessary, and a consistent statisticalformat should be developed, for the comparative analysis of concentration data fromdifferent countries.

Enhancing metal partitioning between treated wastewater and sewage sludgeThe chemical treatment of wastewater to remove phosphorus is increasinglypracticed to control P discharges in the treated effluent as a measure to reduceeutrophication of sensitive water courses. Chemical precipitation with Al or Fe saltscan also enhance the efficiency of metal removal and transfer to the sewage sludge,thus potentially increasing the metal content of the sludge (Case Study (i)). Forexample, Cu and Zn removal from UWW can be raised by 50 % compared toconventional sedimentation without chemical enhancement; Pb removal mayincrease by 80 % and a 3-fold increase in Cr (III) removal by chemical precipitationhas been reported. However, potentially the most important effect of chemicalprecipitants on metal concentrations in sewage sludge is associated with the actualquality and metal content of commercially available precipitant formulationsthemselves. For example, Fe-based precipitants marketed for use in wastewatertreatment can be industrial by-products from titanium oxide production and maycontain significant amounts of potentially toxic elements.

Theoretical calculations (Case Study (i)) suggest that typical dosing rates of FeSO4

salt may increase the Cd content of sludge by 4.5 mg kg-1. The Ni concentration insludge may increase by 130 % compared with national weighted average values andZn and Pb contents may be typically raised by 10 %. Indeed, the use of low-gradeprecipitants could erode the significant progress that has been achieved in reducing

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metal emissions to sewer through controlling trade effluent discharge. In response tothese concerns, certain European countries (e.g. Germany) have introduced controlson the composition of Fe and Al-based coagulants used for UWW treatment andpotable water grade Fe salts are recommended to avoid potential problemsassociated with potentially toxic elements. However, relatively pure sources of Feand Al salts can be recovered from other types of waste, such as acid mine drainage,and are also effective P precipitants. The use of secondary resources for Pprecipitation during UWW treatment is intuitively attractive and also alleviates otherpotential environmental problems associated with the disposal of those wastes.However, the metal content of waste-derived products should be established, and thepotential consequences of their use on sludge quality determined, before a particularproduct is accepted for use as a chemical precipitant in UWW treatment.

Significance of Platinum Group Metals in UWW

The platinum group metals (PGMs) are a group of rare elements including platinum(Pt), palladium (Pd), rhodium (Rh), ruthenium (Ru), iridium (Ir) and osmium (Os) andthe current status of the sources and knowledge of the fate and environmentalconsequences of PGMs are described in Case Study (a). The commercial use of Ptand Pd in particular has expanded significantly in recent years in the manufacture ofcatalytic converters to reduce atmospheric emissions of carbon monoxide,hydrocarbons and nitrous oxide from internal combustion engines. The discharge ofPt from excreted anti-neoplastic drugs used in the treatment of cancers is anotheridentified source entering the WWTS. However, reliable quantitative estimates ofmajor Pt and Pd sources are available and show that hospitals contribute a relativelyminor input, equivalent to 6 – 12 % of Pt discharged to sewer, compared with vehicleexhaust catalysts. Glass, electronics and jewelery manufacturing are other potentialsources of PGMs and can represent important local inputs of these elementsentering particular WWTS. The rate of removal of PGMs from UWW by sewagetreatment processes is generally within the range of most other potentially toxicelement species and approximately 70 % of the Pt in wastewater is transferred to thesewage sludge. Reported concentrations of Pt in sludge from 2 WWTP in Munichwere in the range 86 – 266 µg kg-1 ds.

Platinum group metals emitted as autocatalyst particles behave inertly and havelimited mobility in soil so there would appear to be negligible risk to health,groundwater and the environment. However, it is possible for transformations tosoluble, bioactive forms to occur and as the commercial use of PGMs continues torise, there is a case for a limited investigation of their environmental significance.These studies should focus on the factors controlling the solubility and bioavailabilityof PGMs and on the behaviour of PGMs in surface waters receiving treated sewageeffluents. Information is also needed on the uptake of PGMs by crops from theagricultural use of sludge and direct deposition onto urban garden soils to quantifythe potential transfer by multiple exposure routes to the human foodchain.

7.1.3 Organic compounds

GeneralCompared with the small number of PTEs of concern in wastewater and sludgewhich are routinely monitored and controlled, the range of organic contaminantspresent in these media, with the potential to exert a health or environmental hazard,are significantly more diverse. For example, approximately 140 organic compoundshave been identified in UWW in Sweden and more than 330 organic substanceshave been determined in German sewage sludge. Forty two organic compounds areregularly detected in sludge.

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Persistent organic contaminantsSource and emission controls on persistent organic contaminants were introducedbetween 1980-90 to curb the extent of environmental releases as concern increasedabout the extent of their occurrence and potential toxicity. This legislation has beenrelatively effective and there are several reliable examples in the literature illustratingsignificant reductions in the primary sources of PAHs, PCBs and PCDD/Fs and thishas lowered inputs to the UWW system and reduced concentrations in sewagesludge (Table 7.6). For example, discharges of PCBs to UWW declined by >99 % inthe Rhine region of France between 1985 to 1996 due to stricter industrial sourcecontrol and PAHs were also reduced by >90 % over the same time period. Theprincipal inputs of these contaminants to UWW are from atmospheric deposition ontopaved surfaces and run-off.

Controls on combustion and incineration emissions and the production of certainpotentially contaminated chlorinated pesticides have also markedly reduced therelease of PCDD/Fs to the environment. A review of PCDD emissions in Austria in1998 show small consumers including household, trade and administration activitiesare the principal sources of release contributing almost 60 % of the totalcontemporary load to the environment (Figure 7.2). In the UK, incineration ofmunicipal waste is the largest emitter of PCDD/Fs representing about 40 % of totalemissions to the environment. The emission controls have had a positive effect onthe quality of sewage and PCDD/F concentrations have declined significantly. Forexample, the average PCDD/F concentration in sludge sampled from SpanishWWTS declined from an average value of 620 ng kg-1 TEQ reported for the period1979 – 1987 to 55 ng kg-1 TEQ in 1999. In the case of sludge from a major LondonWWTS (Figure 7.3), PCDD/F concentrations decreased by >97 % in the past 40years from 166 ng kg-1 TEQ in 1960 to 4.2 ng kg-1 TEQ in 1998. Typical PCDD/Fconcentrations reported in sewage sludge are significantly below the highlyprecautionary standard for agricultural utilisation (100 ng kg-1 TEQ) proposed by theEU (Table 7.6). Indeed the numerical limit is unlikely to restrict land application yetthe cost of PCDD/F analysis is high and routine monitoring of sludge for its PCDD/Fcontent is impractical. The implementation of quality standards for PCDD/Fs insewage sludge should be reviewed in terms of environmental effects, analysisfrequency and cost.

Historic pollution levels are important for organic pollutants as well as potentially toxicelements. For example PCBs, which are no longer used, largely result fromvolatilisation from soil and are found in similar concentrations in all the WWTS. Thissuggests either that sources are diffuse and spread evenly or that this is due togeneral background concentrations. Soil is an effective scavenger and sorptivemedium for organic pollutants and acts as a long-term and major repository, althoughbiodegradation also takes place. Contemporary remobilisation by volatilisation fromsoil and redeposition onto surfaces and consequential collection by UWW systems isa major source of these compounds entering sewage sludge.

Controls on the use and emissions of persistent organic contaminants havesignificantly reduced industrial inputs of these substances to sewer. Therefore, theconcentrations present in sludge principally reflect:

• Background inputs to the sewer from normal dietary sources ;• Background inputs by atmospheric deposition due to contemporary

remobilisation/volatilisation from soil and cycling in the environment (egPCBs, PCDD/Fs and PAHs);

• Atmospheric deposition from waste incineration (eg PCDD/Fs);

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• Atmospheric deposition from domestic combustion of coal;• The extent of biodegradation of organic contaminants during sludge

treatment, which is limited for most of these compound types;• The extent of volatile solids destruction during sludge treatment, which

increases the concentration of conservative persistent organic compounds insludge.

Polycyclic aromatic hydrocarbons are proposed as priority hazardous substances(COM(2001) 17 final) within the European Water Framework Directive and the aimwill be to achieve cessation of emissions, discharges and losses of these compoundsby 2020. However, curbing the emissions of PAHs and PCDD/Fs from domestic coalburning would be technically difficult and incinerators are already subject to stringentair quality emission standards. Consequently, there is probably little opportunity tofurther reduce the inputs and concentrations of PAHs, PCDD/Fs as well as PCBs inUWW and sewage sludge. Furthermore, the compounds are effectively removed bywastewater treatment processes because they strongly bind to the sludge solidsminimising the discharge in treated effluent. The increasing amount of scientificinvestigation also shows there are no significant environmental consequences fromPAHs, PCBs or PCDD/Fs when sludge is used on farmland as a fertiliser. In the lightof such developments and their physico-chemical behaviour, it may be argued thatthe importance of these substances as major pollutants of UWW and sludge hasbeen significantly diminished. Soil is a major repository for persistent organiccontaminants and further investigations are necessary to improve understanding ofthe remobilisation and cycling processes in the environment that control diffuseinputs of these organic compounds to UWW.

Table 7.6 Concentrations and proposed limit values of selected organiccontaminants in sewage sludge (mg kg-1 ds)

Organic compounds Mean contentmg kg-1 ds

Proposed EUmaximum in

sludgemg kg-1 ds

Halogenated organics (AOX) 200(1) 500Linear alkylbenzene sulphonates (LAS) 6500 2600

Di(2-ethylhexyl)phthalate (DEHP) 20 – 60 100Nonylphenol and ethoxylates (NPE) 26 (UK: 330 – 640) 50

Total polycyclic aromatic hydrocarbons(PAH)

0.5 – 27.8 6

Total polychlorinated biphenyls (PCB) 0.09 0.8Polychlorinated dibenzo-dioxins

and –furans (PCDD/Fs)36(2) 100(2)

(1)German sludge only (2)Units ng kg-1 TEQ

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Figure 7.2 Sources of dioxin emissions (% of total emitted) in (a) Austria in1998 and (b) UK in 1991

Figure 7.3 Dioxin content of archived samples of sewage sludge from a WWTSin West London, UK

Other organic contaminants

Emissions of other organic contaminants and entry to the WWTS are associated withdirect or indirect discharges resulting from their use in commercial and domesticactivities. For example, the principal emissions of DEHP occur from the use offinished products and major domestic inputs to UWW are floor and wall coveringsand textiles with PVC prints. In Sweden, for example the domestic contribution wasequivalent to approximately 70 % of the total load of phthalates to GothenburgWWTS. Numerically, detergent surfactants and residues (LAS and NPE) are themost significant contaminants in sewage sludge and the LAS content is typicallymore than twice the proposed European limit for this pollutant in sludge (Table 7.6).Unless action is taken to reduce the use of the surfactant LAS in detergentformulations, the proposed standard will prevent use of sludge on agricultural land.The UK also appears to have a particular problem with NP and NPEs asconcentrations of these compounds in UK sludge are an order of magnitude largerthan those reported in average sludges in other European countries Non-

Year

0

50

100

150

200

250

300

350

400

450

1944 1949 1953 1956 1958 1960 1998Dio

xin

co

nce

ntr

atio

n (

ng

TE

Q k

g d

s)

US EPA proposedlimit

Germanlimit

Small consumer

Industry - combustion

Industry - processes

Waste and landfills

Municipal wasteincinerationIndustrial

Domestic

Clinical wasteincinerationVolatilisation fromchlorophenolVehicle

(a) Austria (b) United Kingdom

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Governmental organisations in Sweden and Denmark have been successful inindependently introducing eco-labelling schemes persuading consumers againstdetergents containing LAS or NPE.

Linear alkylbenzene sulphonate and NPE can be substituted in detergentformulations and alternatives may also need to be sought for DEHP in plasticsmanufacture if emissions of this substance to water are to be controlled. Nonylphenolis a proposed priority hazardous substance within the Water Framework Directiveand the use of NPEs in detergents is therefore likely to end with the implementationof European legislation to phase out discharges, losses and emissions ofalkylphenolic compounds to the aquatic environment. A comprehensive independentreview, also involving the detergent and plastics manufacturers, regarding the fate,behaviour, degradability, toxicity and environmental consequences of these and thealternatives compounds would provide information on the advantages anddisadvantages of product substitution in detergent formulations and plasticsmanufacture.

A group of emerging compounds of potential significance in UWW were identified bythe review due to their persistence during sewage or sludge treatment, persistence insoil or toxicity in the environment:

1. Little is known about the fate and behaviour in UWW of the large number (>200) ofcommercial chlorinated paraffin formulations in use as plasticisers in PVC andother plastics, extreme pressure additives, flame retardants, sealants and paints.

2. The brominated diphenyl ethers (PBDEs) are a group of compounds used forflame retardation in furnishings, textiles and electrical insulation and their use hasexpanded due to fire regulation requirements and the increased use of plasticmaterial and synthetic fibres. A survey in Sweden indicated concentrations of PBDEcongeners in sludge were between 15 and 19 µg kg-1, respectively. PBDEs are onthe list of proposed priority hazardous substances.

3. Polychlorinated naphthalenes (PCNs) are released into the environment bywaste incineration and landfill disposal of items containing PCNs. Concentrations insewage sludge may be in a similar range to individual PCB congeners. A survey ofSwedish sludges indicated that PCN concentrations were in the range 1.6 mg kg-1

ds. Some PCN congeners have dioxin like activity and have been assigned TCDDtoxic equivalent values similar to those for coplanar PCBs and so have toxicologicalinterest.

4. A small amount of Quintozene (pentachloronitrobenzene) is produced in the EU(21.5 t y-1) and it is registered for use in the UK, Spain, Greece and Cyprus. It haslow water solubility and half-lives in soil are reported to be in the range 5 – 10months.

5. Polydimethylsiloxanes (PDMS) are nonvolatile silicone polymers used inindustrial and consumer products including lubricants, electrical insulators andantifoams. They are hydrophobic and partition onto the sludge solids andconcentrations in N American sludge reported to be in the range 290 – 5155 mg kg-1.However, PDMS do not bioconcentrate or exhibit significant environmental toxicity,but they are relatively persistent in soil and can take months to years to degrade.

6. Nitro musks (chloronitrobenzenes) are a group of synthetic dinitro- trinitro-substituted benzene derivatives used as substitutes for natural musk in perfumedproducts (Case Study (e)). In Europe, current consumption is estimated to be 124 t y-

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1 for musk ketone and 75 t y-1 for musk xylene and the release of these compoundsto the environment is dominated by domestic discharges to WWTS.

Further studies on the biodegradation and transformations of these compoundsduring secondary wastewater and sludge treatment are required as a preliminarystage in understanding whether the release of musk compounds to the environmentis significant from WWT.

7. Endogenous oestrogens (17b-oestradiol and oestrone) and synthetic steroidssuch as ethinyloestradiol, which is the active oestrogenic component in oralcontraceptives, are discharged in trace amounts in the effluent from WWTS and areprimarily a concern due to the possible impact on the aquatic ecosystem. Thesecompounds also partition onto particulates and may be associated with sewagesludge, but this is unlikely to have significant environmental implications for use ofsludge in agriculture. Recent investigations show that approximately 90 % ofpotential oestrogenic activity (based on 17b-oestradiol equivalency) in UWW isreduced by sewage treatment and that <3 % may be transferred to the sewagesludge.

8. Pharmaceutical compounds, designed for specific biological effects in medicaland veterinary practice, can enter the wastewater system by excretion or residuesand metabolites in urine and from intentional disposal (Case Study (d)). Disposal intothe UWW system is common practice, although the rationale for justifying this on thegrounds of the dilution received within the sewer system should be reviewed incontext of subsequent environmental fate and behaviour of individual compounds.Significant amounts of administered drugs are excreted from the body and 30 – 90 %of antibiotic doses to humans and animals enter UWW in active forms in the urine.The potential toxicological and ecotoxicological activities of these substances in thewider environment are generally unknown, but, because of their biological function,they are generally designed to be rapidly metabolised and degraded, although theyare often lipophilic and potentially bioaccumulate. Removal efficiencies during WWTof >60 % are typical for most types of drug, although a wide range of removal rates isreported (7 – 96 %) and drug compounds are routinely found in treated effluents andsurface waters. Many commonly used analgesic drugs are rapidly biodegradedduring sewage treatment including aspirin, ibuprofen and paracetamol. Most of theseare soluble and exist primarily in the aqueous phase and transfer to sewage sludge isprobably of only minor concern, although it is not possible to predict partitioning andfate during sewage treatment due to the absence of physico-chemical data for manyof the compounds. The occurrence of drugs in soil is widespread from veterinaryadministration to livestock. A risk assessment on the fate and biodegradation of drugcompounds could form an integrated part of the approval procedure forpharmaceutical products. This would improve understanding of their behaviour andfate to balance the benefits to health with the potential consequences of their releaseinto the environment. Collection systems for unwanted drugs should be encouragedto reduce disposal into the wastewater system.

Polyelectrolytes and sludge treatmentPolyelectrolytes based on polyacrylamide and cationic copolymers are usedextensively in sludge treatment to aid dewatering, see Case Study (g). Thepolyelectrolyte concentration in mechanically dewatered cakes is relatively largetypically in the range 2500 – 5000 mg kg-1 ds and they only degrade relatively slowlyby abiotic processes in cultivated soil at a rate of 10 % per year. Acrylamide is acommon monomer associated with polyelectrolytes and is potentially toxic to humansand is a reported carcinogen. Concern that residual monomers in polyelectrolytesused in drinking water treatment may have implications for human health has

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resulted in their withdrawal from this application in Japan and Sweden, and stringentcontrols on their use are also in place in Germany and France. These factors havedrawn attention to the possible environmental implications of sludge dewateringpractices with polyelectrolytes and long-term accumulation in sludge-treated soil.

Attenuation and transformations during wastewater and sludge treatmentSewage sludge is treated to reduce its fermentability, nuisance and vector attractionand to aid its management and acceptability for use on agricultural land. Theseprocesses include physical, chemical and microbiological treatment of the sludge thatmay influence the loss, or potential formation, of organic contaminants (Section 3.2).Loss mechanisms include:

• Volatilisation;• Biological degradation;• Abiotic/chemical degradation (e.g. hydrolysis);• Extraction with excess liquors;• Sorption onto solid surfaces and association with fats and oils.

The sorption of organic contaminants onto the sludge solids is determined byphysico-chemical processes and can be predicted for individual compounds by theoctanol-water partition coefficient (Kow). During primary sedimentation, hydrophobiccontaminants may partition onto settled primary sludge solids and compounds can begrouped according to their sorption behaviour based on the Kow value as follows:

Log Kow < 2.5 low sorption potentialLog Kow > 2.5 and < 4.0 medium sorption potentialLog Kow > 4.0 high sorption potential

Many sludge organics are lipophilic compounds that adsorb to the sludge matrix andthis mechanism limits the potential losses in the aqueous phase in the final effluent.A proportion of the volatile organics in raw sludge, including benzene, toluene andthe dichlorobenzenes, may be lost by volatilisation during wastewater and sludgetreatment at thickening, particularly if the sludge is aerated or agitated, and bydewatering. As a general guide, compounds with a Henry’s Law constant >10-3 atm(mol-1 m -9) can be removed by volatilisation. The significance of volatilisation lossesof specific organic compounds during sewage treatment can be predicted based onHenry’s constant (Hc) and Kow:

Hc > 1 x 10-4 and Hc/Kow > 1 x 10-9 high volatilisation potentialHc < 1 x 10-4 and Hc/Kow < 1 x 10-9 low volatilisation potential

Biodegradation may be more important than air-stripping in removing volatile organiccontaminants during secondary biological WWT.

Mesophilic anaerobic digestion is the principal sludge stabilisation process adoptedin most European countries and many organic contaminants are biodegraded underanaerobic conditions and this is enhanced by increasing retention time and digestiontemperature. Biodegradation during anaerobic digestion may eliminate certainorganic contaminants from sewage sludge, but in general the destruction achieved istypically in the range of 15 – 35 %. Aromatic surfactants including LAS and 4-nonylphenol polyethoxylate (NPnEO) are not fully degraded during sewage treatmentand there is significant accumulation in digested sludge. For example, mass balancecalculations suggest that approximately 80 % of LAS is biodegraded during theactivated sludge process and 15-20 % is transferred to the raw sludge, although nearcomplete biodegradation (97-99 %) is also reported in some cases. Approximately 20

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% of LAS in raw sludge may be destroyed by mesophilic anaerobic digestion. Thecompounds, nonylphenol monoethoxylate (NP1EO) and nonylphenol diethoxylate(NP2EO) are formed during sewage treatment from the microbial degradation ofNPnEO. These metabolites are relatively lipophilic and accumulate in sludge and arealso discharged in the treated effluent from WWT. Approximately 50 % of the NPnEOin raw sewage is transformed to NP by sludge digestion.

Lower molecular weight phthalate esters and butyl benzyl phthalate are completelydegraded in 7 days by anaerobic digestion at 35°C and should be removed by mostmunicipal anaerobic digesters. The extent and rate of biodegradation of organiccompounds during anaerobic digestion is apparently related to the size of alkyl sidechains and compounds with larger C-8 groups are much more resistant to microbialattack. Therefore, di-n-octyl and DEHP are more persistent and are not removed byconventional anaerobic treatment of sludge. Phthalate esters are rapidly destroyedunder aerobic conditions, however, and biological WWT (eg activated sludgeprocess) can usually achieve >90% removal in 24 h. In soil, the reported half-life ofDEHP is <50 d.

Composting is a thermophilic aerobic stabilisation process and has the potential tobiodegrade relatively persistent organic compounds in sludge. Thermophilic aerobicdigestion processes and sludge storage for three months can achieve similar overallremoval rates for organic contaminants as those obtained with mesophilic anaerobicdigestion. Thermal hydrolysis conditioning of sludge prior to conventional anaerobicstabilisation may have a significant influence on the removal of organic contaminantsfrom sludge, but this is a comparatively new enhanced treatment process and effectson the destruction of organic contaminants have yet to be investigated.

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Table 7.7 Assessment of the significance of PTEs entering UWW and SS

(3)Significance of input sources (L, Low; Moderate, M; High, H)(4)Commercial Run-off Domestic

PTE Tradeeffluentcontrol(Yes, Y;No, N)

(1)Priorityhazardous

substance (Yes,Y;

No, N)

(2)Environmentalsignificance insludge-treated

soil (short–term)

(Yes, Y; No, N)

Relativeimportance

Opportunityto reduce

Relativeimportance

(5)Opportunity to reduce

Relativeimportance

(6)Opportunity to reduce

Zn Y N N* M M L - M L H MCu Y N N L - M M L L H LNi Y N N M M L – M L M LCd Y (7)Y N M - H M L – M L (8)M LPb Y (Y) N L – M M M L H L - MCr Y N N M - H M L – M L L LHg Y Y N H H L L L LPt N N N M L H L L LPd N N N M L H L L L

(1)Proposed priority hazardous substance in COM(2001) 17 final: Amended proposal for a Decision of the European Parliament and of the Councilestablishing the list of priority substances in the field of water policy. Lead is in parentheses as a proposed priority substance under review.(2)Reported environmental effects at current maximum permissible limit concentrations in sludge-treated agricultural soil (86/278/EEC) for Zn, Cu, Ni, Cd, Pb,Hg; There are no reported environmental effects of Cr, Pt or Pd in sludge-treated soil. *Long-term potential for Zn, to soil microbial community at maximumlimit.(3)Relative to total identified inputs according to the following approximate ranges: L = <10 %, M = 10 – 50 %; H > 50 %.(4)Medical activities, manufacturing industries and small artisanal enterprises including: car repair, galvanising, metal work and goldsmiths are identified assources of metals discharging to sewer, where further reductions may be achieved.(5)Run-off is difficult to mitigate in practice; sedimentation ponds capturing run-off from paved areas of known deposition risk (eg industrial sites) can reducepollutant inputs bound to suspended solids.(6)Mainly voluntarily through product ecolabelling and education concerning appropriate use and disposal of liquid wastes, body care products, cleaningagents and detergents by homeowners. A reduction in Pb may be possible in the long-term by replacing historical leaded pipework used for waterconveyancing.(7)Discharges may be indirect as Cd is an impurity associated with P (eg used in detergent formulations) and Zn used in industrial processes and roofingmaterials.(8)Human faeces are the main domestic source of Cd representing the background concentration in UWW and sewage sludge.

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Table 7.8 Assessment of the significance of organic contaminants entering urban wastewater and sewage sludge

Contaminant (1)Content inWW/sludge

(2)Priorityhazardoussubstance

(3)Destruction in treatment Accumulation(Yes, Y; No, N)

Backgroundinputs

(4)Overallsignificance

(Yes, Y; No, N)

Wastewater Sludge Biological Soil (Yes, Y; No, N) Wastewater (5)Sludge

LAS H N H L AnaerobicH Aerobic

N N N H L

NPE M - H Y M L AnaerobicH Aerobic

N N N H L

DEHP M (Y) M L AnaerobicM Aerobic

N N N H L

PAHs L - M Y L L Y Y Y L LPCBs L N L L Y Y Y L LPCDD/Fs L N L L Y Y Y L LPharmaceuticals L N M M N N N M LOestrogenic:Endogenous L N M - H M - H Y N Y H LSynthetic L N L - M L - M Y N N H L

(1)Concentration ranges for sludge: L< 1 mg kg-1 ds; M<100; H >100 mg kg-1 ds. Concentrations in wastewater are small (mg l-1) and highlyvariable, but will follow a similar general pattern to sewage sludge; published values are listed in Table 3.14 in the Main Report.(2)European Commission Amended Proposal for a European Parliament and Council Decision establishing the list of priority substances in the field ofwater policy (COM(2001)17 final of 16 January 2001).(3)Approximate indicative ranges: L < 20 %; M = 20 – 60 %; H > 60 %(4)Significance rating: Low, L; Moderate, M; High, HLAS Linear alkylbenzene sulphonatesNPE NonylphenolethoxylatesDEHP Di-(2-ethylhexyl)phthalatePAHs Polycyclic aromatic hydrocarbonsPCBs Polychlorinated biphenylsPCDDs Polychlorinated dibenzo-p-dioxinsPCDFs Polychlorinated dibenzo-p-furans

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Table 7.9 Sources and control of organic contaminants entering urban waste water and sewage sludge

Contaminant Commercial Run-off DomesticRelative

importanceOpportunity

to reduceRelative

importanceOpportunity

to reduceRelative

importanceOpportunity

to reduceLAS H M L L H MNPE H M L L H L

DEHP H M L L M MPAHs L L H L L LPCBs L L H L L L

PCDD/Fs L L H L L LPharmaceutical H M L L H M

LAS Linear alkylbenzene sulphonatesNPE NonylphenolethoxylatesDEHP Di-(2-ethylhexyl)phthalatePAHs Polycyclic aromatic hydrocarbonsPCBs Polychlorinated biphenylsPCDDs Polychlorinated dibenzo-p-dioxinsPCDFs Polychlorinated dibenzo-p-furans

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7.2 COMMENTS, CHALLENGES AND STRATEGIES FOR THE NEXT 5 TO 10 YEARS

The significance of the identified sources of metals entering the UWW system has beenreviewed and opportunities for further reducing metal inputs are considered. This may bepossible through, for example, controlling discharges from a number of specified commercialsectors and the introduction of remedial measures to reduce inputs from certain domesticand diffuse sources. Therefore, a feasibility study of the most effective approaches to furtherreducing metal inputs is recommended so that appropriate remedial strategies can beimplemented.

The contributions by the three major sources of pollutants to UWW is predicted to change.Commercial sources are expected to become less important as the regulatory control ismore widely implemented and tightened, while the percentage contribution from domesticsources is expected to increase. This is important as it is less easy to regulate domesticdischarges to UWW. However this can be achieved to some extent by controlling theproducts used in homes; for example cadmium emissions from households could bereduced by a ban on the use of phosphates in washing powders. Increasing consumerawareness of pollution in UWW, for example through eco labelling schemes, can lead toreductions in load to UWW from some sources such as the domestic use of LAS, howeverthis requires much wider adoption by manufacturers of eco labelling schemes. Industry andstakeholder involvement need to be encouraged. Mechanisms to provide an incentive forreductions in pollution and alternative product development should be developed.

There is probably little scope for further reductions in the inputs and concentrations ofpersistent organic contaminant types in sewage sludge. For example, it is unlikely that muchcan be done to curb emissions of PAHs and PCDD/Fs from domestic coal burning, wherethis is widely practiced, and incinerators are already subjected to stringent air qualityemission standards. Losses of PCBs from electrical transformers is rapidly declining asthese are phased out of use. However, the importance of soil as a repository for persistentorganic contaminants is emphasised and the focus of further investigations to minimiseinputs to sludge could be directed towards better understanding of the remobilisation andcycling of these substances in the environment.

Due to the variability of existing data on potentially toxic elements and organic pollutants,common methodologies and standards of chemical and statistical analysis are necessaryand should be implemented at a European level as soon as feasible.

Feasibility of reuse of treated wastewater for non-household purposes, such as for irrigationof agricultural crops, parkland, and golf courses needs to be considered but must involveassessment of risks to human , environmental and economic strategic issues. Reuse ofdomestic grey waters in the EU needs to be re-evaluated. Development of commonstandards are necessary.

Return of unused pharmaceuticals to dispensaries/hospitals needs to be encouraged andmade more widespread throughout Europe. Hospitals and medical centres effluents shouldbe monitored closely for discharge to the urban wastewater network. Possible segregationand pre-treatment of hospital, medical centre and laboratory effluents, for the reduction ofPGMs, pharmaceuticals and potentially toxic elements is recommended.

Sustainable urban drainage, discussed in detail, Case Study (b), may have great potential interms of reduction of urban runoff pollution. However it is recommended that each caseshould be assessed individually, and an incremental approach containing both high tech andlow-tech solutions is the most likely development scenario (Butler and Parkinson, 1997).

Financial incentives/grants to encourage and facilitate removal of lead piping fromhouseholds would reduce lead emissions to UWW and should initially target areas with softwater.

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There are some areas where restrictions need to be tailored to specific uses: for example,the urban use of pesticides may need separate guidelines, than those applicable toagriculture. Urban use of pesticides and herbicides in gardens and allotments is oftenindiscriminate and can result in major uncontrolled inputs to the UWWS.

For mercury, dentists are important sources and improved dental practices may reduce thissource, but other sources, such as chloralkali processes, can in some instances emit more.The use of thermo-reactive liquids in thermometers has been introduced in some regions asa replacement for mercury. Mercury recycling schemes have also been a success in manyregions and could be extended to other potential pollutants used in domestic andcommercial settings.

Adjusting the pH of tap water may be useful in reducing corrosion, for example, fromdomestic and commercial heating systems, but may be limited by practical and economicfactors and it can still be a problem in drinking water in some regions. For example, leadpassing into drinking water supplies from old lead piping and then into the sewage systemcan be controlled by adjusting the pH and hardness of the water supply.

This report benefits from the presentation of several specific case histories, which providedetailed information in support of the above conclusions. Some specific potential problemareas have been reviewed. For example, the differences in cadmium concentrationsreported in sewage sludge in the United Kingdom and Germany have been shown to be dueto different approaches to data treatment and presentation rather than real variations incomposition. (see Section 2.3)

In the WWTS the fate of compounds often differs, for example, Ni is removed less well fromtreated wastewater than other metals. This means that more Ni will pass out in the effluentfrom the WWTS and less will be concentrated in the SS. The behaviour of compoundsshould be borne into consideration when considering acceptable levels in the influent.

It is also important to note that some of the metals in this study, for example zinc andcopper, are essential trace elements to plants and animals in low concentrations. It isimportant that levels in treated UWW and SS are not set so low by regulation that mineraldeficiency could arise due to removal of these sources where UWW and SS might havebeen used, for example, on nutrient deficient land.

Domestic and industrial product formulation can have an important role in determining thenature and amount of pollutants that are likely to enter the WWTS. Risk assessment ofdomestic products should be based on their input of pollutants to the urban wastewater andthe biodegradability of the pollutants. Health and environmental effects should be consideredin the risk assessment process. Efforts to remove phosphate in detergents have had a knockon effect in lowering the cadmium input to the WWTS. NPE bans and regional LASreduction through eco-labelling schemes have also been proven effective. It would also bepossible to pre-treat high LAS concentration in UWW from commercial sources, which wouldthen reduce levels reaching the WWTP.

Advice on household refurbishment (e.g. for old lead paint, piping) and advice on disposal ofpotential pollutants (e.g. down sinks) and eco-labels and public education should beprovided and extended where possible to raise awareness of ecological impacts of variousprocesses and products in urban wastewater.

The inclusion of Cu, Zn and LAS on the proposed list of priority substances is recommendedbecause they are the most limiting substances to utilising sludge in agriculture within thelimits proposed in the revision of Directive 86/278/EEC and they are ubiquitous inwastewater. Therefore, these substances should also be the focus of measures to reduce

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discharges and emissions to further improve sludge quality and support the recycling routefor sewage sludge.

In order to reduce inputs to UWW, it is important to target the sources contributing most tothe system, especially the diffuse sources. Trade effluent discharge controls have beensuccessful in lowering emissions but are still necessary for certain industries. Reductions ofmany pollutants including lead and dioxins have been seen in sewage sludge since the1960s. Increasing control on emissions, not only to water but also to air and land, are likelyto continue this reduction. This study concludes that small users, hospitals, garages and carwashes, dental and medical practices need a close monitoring and control for connection tothe urban wastewater systems.

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7.3 INFORMATION GAPS AND RESEARCH RECOMMENDATIONS

Relatively few regional surveys of the sources and significance of potentially toxic element(PTE) inputs to UWW (Section 2.1) have been reported and recently published work, insome cases, quote data from considerably earlier studies. There is a general lack of recent,quantitative information regarding the sources and inputs of PTEs to UWW for the EU andthis is identified as a priority area for further data gathering.

There are very few studies available on the mass balance of potentially toxic elements andorganic pollutants through WWTS. More work is needed to understand the partitioning andspeciation of pollutants, particularly for compounds, such as NPE, which may become moretoxic through wastewater treatment. Some pollutant sources still need further identification.In fact there are also many instances where the majority of pollutants sources remainunidentified, for example the ADEME study in France estimated that over 50% of some ofthe metals came from unidentified sources.

Identification and monitoring of trace organics is still sparse and more detailed work, thoughcostly, is urgently required. Better source inventory data is essential. In some casesspecific industrial sources are difficult to locate. For example, platinum uses and losses fromindustry into the wastewater system are subject to the protection of commercial interests.

Information on organic pollutants comes mainly from France, the United Kingdom, Benelux,Scandinavia and Germany but is very limited from the other EU countries. More research isneeded for estimates and quantification of diffuse sources of organic pollutants especially inIberia, Italy and Greece. Some data especially from Iberia is not readily available and anintegrated data collection system is needed across the EU 15 for PTEs and organicpollutants.

Information is still needed in order to asses health and environmental effects. UWW effluentand sewage sludge use on land usually involves exposure to very small quantities of mixedpollutants over a prolonged period of time, and there is little health data about chronicexposure, particularly to organic pollutants. A particular area of current concern is thepossible impact of cocktail effects where several contaminants are present at the same time.Interactions between metals, organics, and metals and organics may be synergistic orantagonistic, are complex and far from clearly understood.

Another area where knowledge is lacking is that concerning the effects of both potentiallytoxic elements and trace organics on components of the ecosystem, both in soils and insurface and subsurface waters. The prime effects to be considered are those on sensitivereceptors, including micro-organisms, invertebrates and plants. At present there is a lack ofbase line information in this area. Ecotoxocity tests that are applied are undertaken inlaboratory conditions assuming 100% bioavailability of the pollutant, and are not appropriatefor field conditions. Research into the transfer of organic pollutants through uptake intopasture plants and into crops and thus into the food chain is limited and should beconsidered.

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APPENDIX AURBAN WASTEWATER TREATMENT SYSTEMS (WWTS) AND SEWAGE SLUDGE TREATMENT (SST)-REGIONAL ASPECTS

Figure A.1 Water and pollutant sources and pathways in urban catchments[after Ellis, 1986, note that this figure could be extended to include sludge output from the WWTS andalso other potential inputs such as from urban use of pesticides* a gully pot (also known as catchbasin) is a chamber or well, usually built at the kerb side, for theadmission of surface water to a sewer or sub-drain. It has a sediment sump at its base to trap grit anddetritus below the point of overflow.]

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TREATMENT PROCESS OUTLINE:

Urban wastewater and sewage treatment is comprised of unit operations to separate,modify, remove and destroy objectionable, hazardous and pathogenic substances carried bywastewater in solution or suspension in order to render water fit and safe for discharge andintended uses. Stringent water quality and effluent standards have been developed thatrequire reduction in suspended solids, biochemical oxygen demand (BOD, related tobiodegradable organic compounds), COD (chemical oxygen demand) and to some extentcoliform organisms (indicators of faecal pollution), control of pH as well as theconcentrations of certain organic compounds, together with some potentially toxic elementsand non-metals.

Sewage sludge consists of residues originating from mechanical, biological, chemical andphysical treatment of wastewater in sewage plants. The quantity and nature of the arisingsewage sludge are subjected to strong fluctuations depending on the wastewatercomposition, the kind of wastewater purification process and the purification degree. Twodifferent sewage sludge types can mainly be distinguished:

• Primary sludge: becomes physically or chemically separated from wastewater inprimary treatment

• Secondary sludge: arises from the biological step (surplus activated sludge, sewagesludge from trickling filters) and tertiary treatment (often nutrient removal).

Primary and secondary sludges are usually combined to create a composite sludge, whichoften goes for further treatment in sludge digestion and dewatering. Residues resulting fromscreening in preliminary treatment are not considered as sludge, consisting mainly of coarsesolid particles, grits, sands and grease [Magoarou, 1998].

Although the ‘nitrate’ and ‘phosphate’ ions in sewage are beyond the realm of this project, ithas to be noted that pollutants such as potentially toxic elements, organics, sulphides andresiduals, can form insoluble phosphates in the course of treatment processes. These havegreat propensity for sedimentation in all stages of the sewage treatment process, thusbecoming part of the sewage sludge that has to be managed.Figure A.2 shows the schematics of urban wastewater and sewage sludge treatment.

Figure A.2 Schematic Urban Wastewater and Sewage Sludge Treatment

Prelim

inaryT

reatment

Prim

aryT

reatmen

t

Secon

dary T

reatmen

t

Ad

vanced

Tream

ent

Sceening,

Girt R

emoval

& G

rease Trap

Physical/C

hemical

Liquid and S

olidS

eparation

Activated S

ludgeB

asin

Secondary

Settlem

ent

Nutrient R

emoval,

Colour R

emoval

PR

IMA

RY

SL

UD

GE

SE

CO

ND

AR

YS

LU

DG

E

RE

SID

UE

S

TE

RT

IAR

YS

LU

DG

E

Anaerobic/A

erobicM

esophilic/Therm

ophilicC

omposting

Stabilisation ponds

Advanced T

reatments

Sludge

Conditioning

Slu

dge D

igestion

Dew

ateringL

and Application

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REGIONAL ASPECTS:

Protection of the receiving waters from pollution by harmful effluent is the primary goal forthe treatment of urban wastewaters at WWTS. Urban wastewater is defined by the CouncilDirective 91/271/EEC of 21 May 1991 concerning urban wastewater treatment (and asamended by Commission Directive 98/15/EC of 27 February 1998) as, domestic wastewateror the mixture of domestic wastewater with industrial wastewater and/or run-off rainwater.The figures for water supply, consumption and treated wastewater are very variable acrossthe European Union. The water supply (litres/inhabitant/day) fluctuates greatly accross theEuropean urban areas (EUROSTAT data,1998-2000, Ginés, 1997). Not all collectedwastewater is treated; the percentage of urban wastewater not receiving treatment rangesfrom 3% in Germany, up to 77% in Greece. This may include unplanned wastewatercollection, such as septic systems or leakage of UWW collection systems. Cities such asMilan and Brussels do not yet have a centralised WWTS.

Table A.1 Population and household access to sewerage and wastewater treatmentfacilities

Percentage of population with access to sewerage and public WWTfacilities

Access tosewerage

No treatment P P + S P + S +T

Country Population(1998)

thousands

(%) (%) (%) (%) (%)

Year

Austria 8,075 76 1 1 39 35 1995Belgium 10,192 78 37 1997Denmark 5,295 87 18 1996Finland 5,147 78 0 7 71 1997France 58,727 81 4 0 0 77 1994

Germany 82,057 92 >31 >47 1995Greece 10,511 70 16 0 0 54 1996Ireland 3,694 68 32 23 12 1 1995

Italy 57,563 84 16 3 38 26 1996Luxembourg 424 88 0 19 57 11 1995Netherlands 15,654 98 2 0 68 28 1994

Portugal 9,957 55 34 9 11 0 1990Spain 39,348 62 13 11 34 3 1995

Sweden 8,848 86 0 0 5 81 1995UK 59,090 96 9 9 64 14 1996

- England &Wales

- 96 10 0 0 86 1995

-N. Ireland - 83 1996-Scotland - 94 1996

TOTAL EU 374,582 84 - - - 48 -

[after OECD Report, 1999, EUROSTAT, 1998-2000] P=primary, S= secondary, T=tertiarytreatment

Table A.1 shows the data in EU15 for the total population, population access to UWWcollecting systems and to urban wastewater treatment facilities. The EU15 average showsthat 84 % of the households have access to UWW collecting systems, whereas only 48 % ofthe urban wastewater is treated in primary + secondary + tertiary treatment facilities.

According to the European Waste Water Group (1997) report on urban wastewatertreatment in the EU and accession countries, there are marked differences between variousEuropean regions in terms of primary, secondary and tertiary treatment as shown in FigureA.3. The four represented groups of countries are: EU10 (all EU countries minus Austria,Belgium, Denmark, Ireland and Sweden), EU south (France, Greece, Italy, Portugal, Spain),EU north (Germany, Finland, Netherlands, Luxembourg and United Kingdom) and AC10(accession countries, Bulgaria, Czech Republic, Estonia, Hungary, Lithuania, Latvia, Poland,Romania, Slovenia, Slovakia).

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The percentage of population not connected to the UWW collecting systems (considered"rural") increases in the order:

EU north (4%) < EU south (18%) < AC10 (40%)

The percentage of untreated urban wastewater increases in the order:

EU north (7%) < AC10 (18%) <EU south (18%)

In terms of primary, secondary and nutrient removal treatment the order is:

EU north (57%) > EU south (3%) > AC10 (2%)

These differences show a north-south and an west-east divide and point towards the needfor greater investment in the urban wastewater treatment in the southern region and theaccession countries.

Figure A.3 Urban wastewater treatment in EU and accession countries (minus Cyprus)[after EEA, 1999, chapter 3.5 Water Stress]

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Northern region (Sweden, Denmark, Finland and Norway):Environmental awareness in this region is high compared to many other countries in Europe(Germany and the Netherlands being the major exceptions to this). This is in part due to theenvironmental damage experienced in the past decades in this region from trans-boundaryacid deposition due to the burning of high sulphur fossil fuels. A number of national pollutionevents, such as mercury poisoning of fish due to discharges from chloralkali plants alsoincreased awareness of the damage emissions can cause.

This led to increased recognition of the need for national and international agreements tolimit local and global environmental perturbations and has led Sweden and Norway inparticular to be among the most pro-active nations in terms of setting environmentalstandards and education amongst the general population.

Finland, Norway and Sweden have low population densities and there are similarities in thecommercial activities in these regions, for example the oil production, metals andengineering and paper manufacturing industries. Denmark has a higher population densitythan other countries in the Northern region and therefore more of an urban environment butalso has a high level of environmental concern about water pollution, although all thedrinking water in this region is from groundwater sources. This compares with the UK, forexample, where approximately one third of potable drinking water is abstracted from surfacewater sources that receive effluent from wastewater treatment.

The Central Region (Germany, Austria and CEE countries):

In addition to Germany and Austria, data are gathered where available for Switzerland andfor Central and Eastern European countries (Bulgaria, Czech Republic, Estonia, Hungary,Latvia, Lithuania, Poland, Slovakia, Slovenia), all associated EU countries. Environmentalawareness in the Central Region is high in Germany, Austria and Switzerland whereas theCEE countries experience a high level of environmental damage due to the long years ofneglect while part of the Soviet bloc. Transboundary pollution issues are very important inthe accession CEE countries and their neighbours. Very little data is available for the CEEcountries regarding the pollutants in UWW and SS. An inventory of organic pollutants in theenvironment in the CEE countries was done by a team at Brno University, the CzechRepublic. (Persistent, Bioaccumulative and Toxic Chemicals in Central and EasternEuropean Countries - State-of-the-art Report-TOCOEN REPORT No. 150, 1999). The mainorganic pollutants investigated are the PAHs, PCBs, PCDD/PCDFs and their ocurrence inthe urban environments in the CEE countries. Most data available is from the CzechRepublic, Poland, Slovakia and Hungary.

The Southern Region (Greece, Italy, Portugal and Spain):

The environmental awareness in the region is on the increase in the recent years. Waterresources are limited in the Southern Region and therefore recycling of treated waste waterand pollution reduction at source are important issues. Data regarding sources of pollutantsis less abundant in the 'Southern Region' especially regarding specific organic pollutants.The region has among the lowest percentage of population access to UWW collectingsystems in the EU (70% for Greece, 84% for Italy, 55% for Portugal and 62% for Spain). Thepercentage of non-treated wastewater is among the highest in the EU (16% for Greece andItaly, 34% for Portugal and 13% for Spain) [after OECD Report, 1999, EUROSTAT, 1998-2000].

Typically, the metal content of sewage sludge in Italy is low, according to the literature[Garcia-Delgado et al., 1994; Lang et al., 1988] suggesting that the sludge is mainly ofdomestic origin, with negligible contribution from urban and industrial wastewater. Data forSouthern Italy were gathered in a two years pilot campaign [Braguglia et al., 2000; Marani etal., 1998; Mininni et al., 1999].

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Metal Concentration rangeSouthern Italy Literature

As 1.1-1.8 0.3-20Cd 3-23 1-50Co 1-4 5-30Cr 227-535 40-1500Cu 258-373 160-1600Hg 1.1-3.1 1-12Mn 80-109 240-600Ni 34-57 20-240Pb 95-137 80-850Zn 1650-4213 900-4200

Table A.2. Metal concentrations in sewage sludge (mg/kg of DS)

Wastewater treatment systems in Italy may treat 'non domestic' wastewater if the followingrequirements are fulfilled:

- the plant has a residual capacity of treatment;- wastewater meets the limits for discharge in UWW collecting systems;- wastewater derives from the same territorial area;- the same treatment tariff valid for the UWW discharge is applied.

In Spain, which is organised in 17 comunidades autonomas (autonomous communities)there are more than 300,000 point sources of water discharges (both to superficial and togroundwater) out of which 240,000 are to UWW collecting systems [Ministerio de MedioAmbiente, 1997]. However, the quantification of the pollutants discharged is very limited.Routine controls are generally limited to those established in Royal Decree 509/1996 andwhich must be published every two years: BOD5, COD, suspended solids, total phosphorusand total nitrogen in the case of treatment plants located in sensitive areas. However, thesecontrols are usually employed only for the discharges (not for the wastewater coming intothe treatment plant). As the discharges into the UWW collecting systems are under thecontrol of the local governments (municipalities), the limited information on levels ofpollutants in wastewater that exists comes from the enforcement inspection controls and isnot publicly available nor published in any form [Palerm and Singer, 2000].

There is almost no information on sources of pollutants and levels of pollutants in urbanwastewater and sewage sludge in Iberia [Palerm and Singer, 2000]. Water management inSpain is completely de-centralised and the discharges into the UWW collecting system areunder the competence of the municipal authorities, which must meet the establishedparameters for discharge into the public waterways.

There is very little information sources of pollutants and levels of pollutants in urbanwastewater and sewage sludge in Portugal [Palerm and Singer, 2000]. Most of theinformation that exists was obtained prior to designing the wastewater treatment plants, butthis information was gathered at a local level and is not publicly available.

Data are available in Greece for sewage sludge content of potentially toxic elements, andmore limited for pollutants in urban wastewater. The Hellenic Ministry for Environment,Physical Planning and Public works co-ordinates the data management in terms ofpollutants load in urban areas. The research studies tend to centre on the cities ofThessaloniki or Athens.

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The Western Region (UK, Ireland, France and the Benelux countries, Belgium Netherlandsand Luxembourg):

In this region, wastewater treatment is among the most advanced in the EU. France hasover 160,000 km of sewers and 11,300 treatment plants. The WWTS are operated by publicor private bodies. The ownership of WWTS and sewers are always public, owned bymunicipalities or associations of municipalities. France is divided into six Water Agencies,mainly around hydrographic areas, Adour-Garonne, Artois-Picardie, Loire-Bretagne, Rhin-Meuse, Rhone-Méditéranée-Corse and Seine-Normandie. The municipalities operate thesewer systems and often, private companies run the plants [EWWG, 1997].

In the Netherlands, the municipalities are responsible for collection of sewage and disposalof sludge from sewers. The treatment of wastewater and disposal of effluent and sludge isthe responsibility of 27 waterboards. Sludge disposal is partially privatised. In Belgium, themunicipalities are responsible for the sewerage systems in both Vlanderen and Wallonie.Similarly, in Luxembourg, the 118 municipalities are responsible for the collection andtreatment of urban wastewater and sludge. The management of sludge is partially sharedwith operators of solid waste [EWWG, 1997].

In UK there are over 300,000 km sewers and 7,600 WWTS. The collection of sewage, itstreatment and disposal of effluent and sludge are the responsibility of privately-financedwater service companies in England and Wales (10), public water and sewerage authoritiesin Scotland (3), and the Water Service of the Department of Environment in Northern Ireland[EWWG, 1997]. In the Republic of Ireland, the urban wastewater collection and treatment isthe responsibility of local councils. Private-Public Partnerships (PPP) schemes are in placefor modernising and upgrading the WWTS [Dept. of Environment and Local Govt, Ireland,1999]

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239

APPENDIX B Physical and chemical properties of selected pollutants

In order to understand the behaviour of organic pollutants in the urban environment theknowledge of their physical and chemical properties is very important. Vapour pressure,boiling-point, water solubility and distribution coefficients describe the distribution betweensolid, liquid and gaseous phase. The adsorption coefficient KOC is important for transitionsbetween soil or sewage sludge particles and water. The transition water/air is governed bythe Henry coefficient KAW. The distribution coefficient octane/water KOW is a measure for thelipophilic or lipophobic qualities of a chemical compound. For substances with similarphysical properties, their water solubility is a crucial factor for the liquid phase transport.

Anionic and Non-ionic Surfactants

A1. Linear Alkylbenzenesulphonate

Linear alkylbenzenesulphonate (LAS) is a synthetic compound utilized as surfactant indetergents, washing-up liquids and cleaning agents. The most important LAS qualities arerepresented in Table II.22 Figure B.1 shows the LAS structure formula.

Table B.1 LAS properties [Bürgermann 1988].Abbreviation Molecular

formulaMolar mass Colour Solubility Vapor

pressureDensityat 20°C

[g/mol] [g/l] [g/cm3]LAS C18H30O3S ca. 326 colourles

s1.1 extremely low 1

SO3-Na+

CH3-CH2-CH-(CH2)8-CH3

Figure B.1: LAS structure formula.

A2. NPEs or nonylphenol polyethoxylates and APEs alkylphenol ethoxylates

Polyethoxylated nonylphenols are important surfactants used comercially and in somehousehold products for many years, also as emulsifiers and solubilisers in industrialprocessing, as well as household cleaning products. They have the general formula:

R-C6H4-(OCH2-CH2)nOH, where R=C9H19 and n=6-18.

B. Polychlorinated Dibenzo-p-dioxins and Dibenzofurans (PCDD/PCDF)

PCCDs and PCDFs are tricyclic, aromatic, almost planar built ethers with comparablephysical, chemical and biological qualities. They differ from each other in the position andnumber of chlorine atoms and the symmetry of the basic structure. Figure B.2 shows thestructure formulas of the polychlorinated dibenzo-p-dioxins and polychlorinated dibenzofurans (PCDD/F).

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240

1

4

2

3

8

7

9

6Clx Cly

PCDD1

4

2

3

8

7

9

6Clx Cly

PCDFFigure B.2 PCDD/F structure formulas.

PCDD/F are extremely heavy-volatile components (high melting and boiling-points). Thevapour pressure drops with chlorine substitution, so that low-chlorinated dioxins and furansare quite volatile. Highly-chlorinated compounds are found in solid state adsorbed onparticles. The greatest mobility is in the air. In water, PCDD/F are almost completely boundto particles (high octane/water distribution coefficient). The affinity to organic carboncompounds is strongly pronounced (high adsorption coefficients). Due to their physico-chemical qualities PCDD/F are very firmly bound to soil and sediments.

PCDD/F do not react with acids and bases and are quite chemically inert. They are thermallystable at temperatures up to 600-800 °C. The extraordinary stability and the low photolyticreduction render them to be very persistent in the environment. Because of the long life timeand the lipophilic qualities PCDD/F can accumulate in organisms: they accumulatepredominantly in animal fatty substances [Mahnke, 1997].

C. Polychlorinated Biphenyls (PCBs)

The polychlorinated biphenyls (PCBs) form a group of over 209 chlorinated, aromaticcompounds with the same structural features. They differ with the degree of chlorinesubstitution and with their structure. PCBs are substances with low electrical conductibility,high thermal and chemical stability and low water solubility. They are characterised by a highlipophilic character, so they accumulate in the food-chain. Their biological degradability inthe environment depends on the complexity and chlorination of each particular compound.Figure B.3 shows the PCBs structure formula. The physical-chemical qualities of selectedPCBs compounds are specified in Table B.2.

3

4

2

6Clx Cly5

2'

6'

3'

5'

4'

Figure B.3 PCBs structure formula.

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Table B.2 Physical-chemical data of selected PCBs.Substance Henry

constantSolubility[g/m3]

Log KOW

2,4,4´-Trichlorbiphenyl n.a. 8.5E-02 5.74/5.692,5,2´,5´-Tetrachlorbiphenyl 4.9E-03 4.6E-02 6.26/6.093,4,3´,4´-Tetrachlorbiphenyl 3.1E-04 1.8E-01 6.52/5.622,4,5,2´,5´-Pentachlorbiphenyl 1.2E-0.3 3.1E-0.2 6.85/7.072,3,4,2´,4´,5´-Hexachlorbiphenyl 5.3E-04 n.a. n.a./7.442,4,5,2´,4´,5´-Hexachlorbiphenyl 6.9E-04 8.8E-03 7.44/7.752,3,4,5,2´,4´,5´-Heptachlorbiphenyl 1.3E-04 n.a. n.a./n.a.

(Kow distribution coefficient octane-water) [Bürgermann 1988].

D. Polycyclic Aromatic Hydrocarbons

The polycyclic aromatic hydrocarbons (PAHs) are of major public concern because of theirubiquitous occurrence and high carcinogenic potential. PAHs are multi-core aromatic ringsystems with 5 and 6 rings. They represent benzene condensation products. PAHs are solid,mostly colourless compounds. They have a strong lipophilic character and their watersolubility decreases with the increase of ring numbers. Low-molecular PAHs are relativelyvolatile. PAHs with a boiling-point below 400°C exist in the air in gaseous state. The higherboiling compounds are adsorbed to particles. The physico-chemical features of some PAHsare indicated in Table B.3 and Figure B.4 shows the structural formula of selected PAHs.The list of 16 USEPA PAHs is shown in Table B.5.

Table B.3 Physico-chemical data of selected PAHs [Bürgermann 1988].Name Chemical

formulaVapour

pressureSolubility KOW KOC Henry constant

[Pa] [mg/l] [cm3/g]

Benzo(k)fluoranthene C10H12 0.1E-07 0.68E-06 6.84 2843420 0.2E-05Benzo(a)anthracene C18H12 0.67E-06 0.12E-04 5.61 167433 0.54E-05

Benzo(a)pyrene C20H12 0.0 0.38E-05 6.04 450651 0.18E-07

(Kow distribution coefficient octane-water, Koc adsorption coefficient)

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Indeno(1,2,3-cd)pyren Dibenzo(a,h)anthracen

Benzo(b)fluoranthen Benzo(g,h,i)perylen

Figure B.4: Structure formula of selected PAHs. [Bürgermann 1988]

Table B.4 List of 16 PAH group [USEPA, IARC]PAHs Vapour

pressure(Torr at 20°C)

Solubility inwater (mg.l-1)

Kow Carcinogenic potency

IARC/USEPA*

classificationAcenaphthene, Ace 10-3-10-2 3.4 at 25°C 21000Acenaphthylene, Acy 10-3-10-2 3.93 12000Fluorene, Flu 10-3-10-2 1.9 15000Naphtalene, Np 0.0492 32 2300Anthracene, An 2.10-4 0.05 – 0.07 at

25°C28000 3

Fluoranthene, Fl 10-6 to 10-4 0.26 at 25°C 340000 3Phenanthrene 6.8.10-4 1.0 to 1.3 at

25°C29000 3

Benzo[α] anthracene, B[α]An 5.10-9 0.01 at 25°C 4.105 2A/B2Benzo[ß]fluoranthene, B[ß]Fl 10-11 to 10-6 - 4.106 2B/B2Benzo[k]fluoranthene, B[k]Fl 9.6.10-7 - 7.106 2BChrysene, Chry 10-11 to 10-6 0.002 at 25°C 4.105 3/B2Pyrene 6.9.10-9 0.14 at 25°C 2.105 3Benzo[ghi]perylene, B[ghi]Pe ~10-10 0.00026 at

25°C107 3

Benzo[α]pyrene, B[α]Py 5.10-9 0.0038 at 25°C 106 2A/B2Dibenzo[α,h]anthracene, dB[α,h]An ~10-10 0.0005 at 25°C 106 2A/B2Indeno[1,2,3-cd]pyrene,I[1,2,3-cd]Py

~10-10 - 5.107 2B/B2

2A/B2:Probably carcinogenic to humans/Probable human carcinogen; 2B:Possibly carcinogenic tohumans; 3: Not classifiable as to human carcinogenicity; Blank:Not tested for human carcinogenicity.*IARC: International Agency for Research on Cancer; USEPA: US Environmental Protection Agency.

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Di-(2-ethyhexyl)phthalate (DEHP)

Di-(2-ethyhexyl)phthalate (DEHP) appears at 25°C as a colourless, almost odourless, oilyliquid and is fat-soluble (lipophilic). It is transported almost exclusively with fatty substancesand accumulates in sediments. DEHP forms water-soluble complexes with humic and fulvicacids. Figure II.10 shows the DEHP structure formula. The more important physico-chemicalqualities of DEHP are represented in Table II.30.

C=O

O

CH2

HC-C2H5

C4H9

C-O-CH2-CH-C4H9

O C2H5

Figure B.5 DEHP structure formula.

Table B.5 Physical-chemical data of DEHP [Bürgermann 1988].Abbreviation Sum

formulaVaporpressure

Solubility KOW KOC Henryconstant

[Pa] [mg/l] [cm3/g]

DEHP C24H38O4 0.60E-05 0.23E-04 4.88 35,567 0.53E-05

(Kow distribution coefficient octane-water, Koc adsorption coefficient)

Polycyclic musk compoundsThree representatives of the polycyclic musk compounds with abbreviation name, tradename, chemical formula and molecular weight are shown in Table B.6.

Table B.6 Polyclyclic musk compounds.Abbreviation Trade name Chemical

formulaMolecularweight

HHCB Galaxolide C18H26O 258.40AHTN Tonalide C18H26O 258.40ADBI Celestolide C17H24O 244.38

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Nitro-musk Compounds

Properties. Nitro-musk compounds are nitro aromatic bonds with a high stability to chemicaland biochemical reduction, high persistancy and lipophillic behaviour.Musk ambrette, musk xylene, musk ketone, musk tibetene and musk moskene belong to thenitro-musk compounds. Trade names and formulas of the nitro-musk compounds are listedin Table B7.

Table B7 Trade name and formula of the nitro-musk compounds.Trade name FormulaMusk ambrette 1-tert.-butyl-2-methoxy-4-methyl-3,5-

dinitrobenzeneMusk xylene 1-tert.-butyl-3,5-dimethyl-2,4,6-trinitrobenzeneMusk ketone 1-tert.-butyl-3,5-dimethyl-2,6-dinitro-4-

acetylbenzeneMusk tibetene 1-tert.-butyl-3,4,5-trimethyl-2,6-dinitrobenzeneMusk moskene 1,1,3,3,5-pentamethyl-4,6-dinitroindan

Musk Xylene and Musk Ketone

Properties. Musk xylene and musk ketone are nitrobenzene compounds. They arepersistent, lipophile and accumulate in the food chain. Musk xylene has a biologicalaccumulation factor (concentration in fatty tissues / concentration in the environment) of 4.1,Musk ketone of 1.1.

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8 REFERENCES

8.1 Electronic databases searched (www and academically-networked)

American Chemical Society Pubs http://pubs.acs.org/ANTEnet Abstracts in New Technologies and Engineering http://www.antenet.co.ukAqualine academically-networkedASFA-Aquatic Sciences and Fisheries Abstracts http://www.csa1.co.uk/ (CSA)Biotechnology and Bioengineering Abstracts http://www.csa1.co.uk/ (CSA)CELEX (EU Legal database) http://europa.eu.int/celex/Chemical Engineering and Biotechnology Abstracts http://www.rsc.bids.ac.uk/CORDIS (EU database) http://www.cordis.lu/Current Contents http://wos.mimas.ac.uk/ccclogin.htmlEcology Abstracts http://www.csa1.co.uk/ (CSA)EIS-Environmental Impacts Statements http://www.csa1.co.uk/ (CSA)EMBASE (Medicine and Pharmacology) http://www.bids.ac.uk/Environment Abstracts academically-networkedEnvironmental Engineering Abstracts http://www.csa1.co.uk/ (CSA)Environmental Fate Database http://esc.syrres.com/efdb.htmESPM-Environmental Sciences and Pollution Mgmt. http://www.csa1.co.uk/ (CSA)Health and Safety Science Abstracts http://www.csa1.co.uk/ (CSA)International Civil Engineering Abstracts http://www.anbar.com/cgi-bin/ce/CEdbIndex to Scientific and Technical Proceedings http://wos.mimas.ac.uk/istpcgi/login.cgiIngenta Journals (bids) http://www.ingentajournals.bids.ac.uk/Medical Pharmaceutical Biotechnology Abstracts http://www.csa1.co.uk/ (CSA)MEDLINE http://www.csa1.co.uk/ (CSA)Ovid Biomedical Service http://biomed.niss.ac.uk/ovidweb/ovidweb.cgiPollution Abstracts http://www.csa1.co.uk/ (CSA)Risk Abstracts http://www.csa1.co.uk/ (CSA)Science Citation Index http://wos.mimas.ac.uk/Science Direct (integral articles database) http://www.sciencedirect.com/Toxicology Abstracts http://www.csa1.co.uk/ (CSA)Toxline http://www.csa1.co.uk/ (CSA)Wasteinfo academically-networkedWater Resources Abstracts http://www.csa1.co.uk/ (CSA)

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an agricultural soil with a known history of sewage sludge amendments: Polynuclear aromatichydrocarbons. Environmental Science and Technology 24, 1706-1711.

• Wilderer, P., A., Kolb, F., R.: (1997): "Abwasserexfiltration undNiederschlagswasserversickerung“, Studie im Auftrag der Landeshauptstadt München, Juli 1997.

• Wilmanski, K. and van Breeman, A.N. (1990) “ Competitive adsorption of trichloroethylene andhumic substances from groundwater on activated carbon” Water Research, 24(6), 773-779.

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• Witte, H. (1988): 'Belastung des Klärschlamms durch organische Umweltkemikalien“,Umweltkemikalien, Korr. Abwasser

• WMS (1994). World Mineral Statistics, 1988-92. British Geological Survey.• Xanthopoulos, C., Hahn, H. H., (1993): “Anthropogene Schadstoffe auf Straßenoberflächen und

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• Yamada, M., Dazai, M., and Tonomura, K. (1959) Change of mercurial compounds in activatedsludge. Journal of Fermentation Technology 47, 155.

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• Yamagishi T, Miyazaki T, Horii S, Kaneko S (1981). Identification of musk xylene and muskketone in freshwater fish collected from the Tama River, Tokyo. Bull Environ Contam Toxicol26:656-662.

• Yeoman, S., Lester, J.N. and Perry, R. (1993) Phosphorus removal and its influence on metalspeciation during wastewater treatment. Water Research 27, 389-395.

• Yeoman, S., Stephenson, T., Lester, J.N. and Perry, R. (1988) The removal of phosphorus duringwastewater treatment. A review. Environmental Pollution 49, 183-233.

• Ying, W., Duffy, J.J, Tucker, M.E., (1988) “Removal of humic acid and toxic organic compoundsby iron precipitation” Environmental Progress, 7(4), 262-269.

• Zeisel SH. (1999). Regulation of "nutraceuticals." Science 285: 1853-1855• Zereini, F., Dirksen, F., Skerstupp, B., Urban, H. (1998). Sources of anthropogenic PGE:

automotive catalysts versus PGE-processing industries. ESPR – Environ. Sci. Poll. Res. 5 (4)223-230.

• Zereini, F., Skerstupp, B., Alt, F., Helmers, E. and Urban, H. (1997). Geochemical behaviour ofplatinum-group elements (PGE) in particulate emissions by automobile exhaust catalysts:experimental results and environmental investigations. Science of the Total Environment, 206,137-146.

• Zereini, F., Zientek, C., Urban, H. (1993). Concentration and distribution of platinum groupelements (PGE) in soil – platinum emission by abrasion of catalytic converter material. UWSF-ZUmweltchem. Ökotox. 5 (3) 130-134.

• Zhao, H. and Vance,G.F. (1998) “Sorption of trichloroethylene by organoclays in the presence ofhumic substances” Water Research 32(12), 3710-3716.

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8.4 REFERENCES (REPORTS)• Abwasserverordnung vom 9. Februar (1999), Bundesrepublik Deutschland: 'Verordnung über

Anforderungen an das Einleiten von Abwasser in Gewässer“, (AbwV).• Abwasserverordnung vom Februar (1999): "Verordnung über Anforderungen an das Einleiten von

Abwasser in Gewässer“ (Abwasserverordnung - AbwV).• Académie de l'Eau: (2000) Paris, la situation actuelle et l'avenir, Website, Pages 1-15, May.• ADEME (1995a) Les micropolluants métalliques dans les boues résiduaires des stations

d'épuration urbaines, Pages 17-79,. (2)• ADEME: (1995b) Les micropolluants organiques dans les boues résiduaires des stations

d'épuration urbaines, Pages 26-32. (3)• ADEME (1997a) Actes des journées techniques des 5 et 6 Juin 1997: aspects sanitaires et

environnementaux de l'épandage des boues résiduaires, Pages 1-319, June. (1)• ADEME (1997b)Convention de recherche: Cartographie de la pollution de l'air par certains

métaux lourds sur le littoral Calais-Dunkerque au travers de l'analyse des lichens et des boues detoiture, Numéro 9693028,. (4)

• Agence de l'eau Rhin-Meuse (1985) Essai d'évaluation de l'apport global de métaux lourds parles eaux usées domestiques, Pages 1-10, April 1985.

• Agences de l'Eau (1997a) La mesure des micropolluants dans le cadre du réseau national debassins, Issue 54, 1997

• Agences de l'Eau (1997b) Seuils de qualité pour les micropolluants organiques et minéraux dansles eaux superficielles, Synthese 53, 1997

• Agences de l'Eau, (1992) Etude qualitative et quantitative des sources diffuse de mercure, Pages1-47, November 1992

• Agences de l'Eau: (1993) Etude qualitative et quantitative des sources diffuses de solvantschlorés, Pages 1-59, April 1993

• AGHTM, Groupe de travail (2000): TSM dossier: Les déchets mercuriels en France, Volume 7-8,Pages 24-48, July-August 1999, and Volume 3, Pages 17-53.

• ATV (1986): "Lehr- und Handbuch der Abwassertechnik. Band VI: Organisch verschmutzteAbwässer sonstiger Industriegruppen“, Dritte, überarbeitete Auflage.

• ATV (1999): "ATV-Handbuch. Industrieabwasser. Grundlagen“, 4. Auflage.

• ATV-A 115 (1994): 'Einleiten von nicht häuslichem Abwasser in eine öffentlicheAbwasseranlage“, ATV-Regelwerk Abwasser/Abfall.

• Berücksichtigung von Schadstoffen im Klärschlamm, Berichte aus Wassergütewirtschaft undGesundheitsingenieurwesen, Technische Universität München, Heft 38, Eigenverlag München

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Report- Effect on municipal treatment plants of heavy metals : some aspects,• Chemikalien Verbotsverordnung vom 19. Juli (1996) : "Verordnung über Verbote und

Beschränkungen des Inverkehrbringens gefährlicher Stoffe, Zubereitung und Erzeugnisse nachdem Chemikaliengesetz“, (ChemVerbotsV).

• Chemikalien Verbotsverordnung vom 19. Juli (1996): 'Verordnung über Verbote undBeschränkungen des Inverkehrbringens gefährlicher Stoffe, Zubereitung und Erzeugnisse nachdem Chemikaliengesetz“, (ChemVerbotsV).

• Chemikaliengesetzt vom 25. Juni (1994), Bundesrepublik Deutschland: “Gesetz zum Schutz vorgefährlichen Stoffen”, (ChemG).

• Chemikaliengesetzt vom 25. Juni (1994), Bundesrepublik Deutschland: “Gesetz zum Schutz vorgefährlichen Stoffen”, (ChemG).

• Commission Internationale pour la Protection du Rhin (1999): Inventaire des apports desubstances prioritaires dans le Rhin 1996, Pages 1-87, December 1999. (10)

• Commission Internationale pour la Protection du Rhin contre la Pollution: (1987). Programmed'action du Rhin, Pages 1-26, September (9)

• Conseil de l'Union Européenne: (1999) Dossier interinstitutionnel 97/0067 (COD), 22nd October.• Conseil supérieur d'hygiène publique de France, section des eaux: (1998) Risques sanitaires liés

aux boues d'éuration des eaux usées urbaines, Pages 21-107,.• Consejería de Medio Ambiente (1998) Informe 1998 Medio Ambiente en Andalucía (State of the

Environment in Andalucia .• Dachverband Agrarforschung (DAF) (1995): "Stichstoff und Phosphateintrag in Fließgewässer

Deutschlands unter besonderer Berücksichtigung des Eintragsgeschehens imLockergesteinsbereich der ehemaligen DDR“, Band 22, Verlagsunion Agrar Frankfurt.

• Department of the Environment, UK. (1991) The use of herbicides in non-agricultural situations inEngland and Wales, HMSO, London.

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• Department of Environment and Local Government, Ireland (1999), water and SewerageSystems Annual Report.

• EBAV (1996) “Scarichi da attività artigianali-impatto sugli impianti municipali di trattamento deiliquami”

• EEA (1999), State of the Environment, chapter 3.5 Water Stress.• EHPA (1992). An environmental standard for car washing detergents. Environmental and Health

Protection Agency, City of Göteborg, Report 1992:15• ENDS report 256 - Feburary (1997), ENDS environment daily [accessed 25/05/00]

http://www.ends.co.uk/envdaily - Permanent EU Phthalates ban moves closer and Danish clashover ecolabelled washing powders,

• Environment Agency (1999). Tyres in the environment. Report, http://www.environment-agency.gov.uk//envinfo/tyres/index.htm. [8th June, 2000]

• Environment Agency-UK (1999)- Endocrine - disrupting substances in the environment: theenvironment agency's strategy

• Environment Agency-UK, (1998), Environmental Issues Series, Endocrine-disrupting substancesin the environment: what should be done/ Consultative report 1998

• ETC/IW (1998), Technical Report in the EEA from the European Topic Centre on Inland Waters,December 1998.

• EWWG-European Waste Water Group (1997), European waste water catalogue.• German Federal Statistical Agency (1994). Öffentliche Wasserversorgung und

Abwasserbeseitigung 1991, Metzler-Poeschel, Weisbaden.• Hessische Landesanstalt für Umwelt, (1997), Heft 233, Landesweuite untersuchungen auf

organische Spurenverunreiningungen in hessischen Fliessgewässern, Abwässern undKlärschlämmen, 1991-1996.

• Indirekteinleiterverordnung vom April (1999), Baden-Wüttemberg: " Verordnung des Ministeriumsfür Umwelt und Verkehr über das Einleiten von Abwasser in öffentliche Abwasseranlagen“,(Indirekteinleiterverordnung-IndVO).

• Junta de Sanejament (1996) Plan de Saneamiento de Cataluña (Sewerage Plan of Catalonia).• Junta de Sanejament (1996) Programa de Saneamiento de Aguas Residuales Urbanas (Urban

Wastewater Treatment Programme).• Junta de Sanejament (1996) Programa de Tratamiento de los Fangos de las Depuradoras de

Aguas Residuales Urbanas (Sewerage Sludge Treatment Programme).• Junta de Sanejament (1997) Informe de la Gestió dels Biosòlids de Depuració – Any 1997

(Sewerage Sludge Management Report – 1997).• Junta de Sanejament (1998) Memòria d’Activitats – 1998, Departament de Medi Ambient,

Generalitat de Catalunya.• Junta de Sanejamente (1997) Memòria d’Activitats – 1997, Departament de Medi Ambient,

Generalitat de Catalunya.• Klärschlammverordnung vom 15. April (1992): Bundesrepublik Deutschland: "Klärschlamm-

verordnung“, (AbfKlärV).• Kreislaufwirtschafts- und Abfallgesetz vom 27. September (1994), Bundesrepublik Deutschland:

"Gesetz zur Förderung der Kreislaufwirtschaft und Sicherung der umweltverträglichenBeseitigung von Abfällen“, (Kreislaufwirtschafts- und Abfallgesetz-KrW-/AbfG).

• Kreislaufwirtschafts- und Abfallgesetz vom 27. September (1994), Bundesrepublik Deutschland:'Gesetz zur Förderung der Kreislaufwirtschaft und Sicherung der umweltverträglichen Beseitigungvon Abfällen“, (Kreislaufwirtschafts- und Abfallgesetz-KrW-/AbfG).

• Law Lecture Notes (1999) Waste Management Law and Policy: Key Regulation Issues, ImperialCollege, UK.

• LIFE Report 96ENV/F/410, (1999), Garantir la qualite des boues par la maitrise globale dusysteme d'assainissement, Anjou Recherche, Rapport Final, Juin 1999.

• LUFA Hameln - Landwirtschaftliche Untersuchungs- und Forschungsanstalt, (1996),Klärschlammstatistik 1991-1995.

• Ministerio de Medio Ambiente (1997) Estado del Medio Ambiente en España 1997 (State of theEnvironment in Spain 1997), Madrid.

• Ministerio de Medio Ambiente (1998) Libro Blanco del Agua en España (White Paper for Water inSpain), Madrid.

• Ministerio de Medio Ambiente (2000) Actuaciones Públicas en Materia de Medio Ambiente (21 deFebrero de 2000) (Public Activities regarding the Environment).

• Ministry of the Environment and Energy (1999)- Denmark, Action Plan for reducing and phasingout phthalates in soft plastics June 1999

• Niedersächsisches Landesamt für Ökologie, (2000), Landwirtschaftliche Klärschlammverwertungin Niedersachsen

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• OECD (1994) Applying Economic Instruments to Environmental Policy in OECD and DynamicNon-member Economies, OECD, Paris.

• OECD (1999a) Economic Instruments for Pollution Control and Natural Resource Management inOECD Countries: a survey, Working Party on Economic and environmental Policy Integration.

• OECD (1999b) Household water pricing in OECD countries, ReportENV/EPOC/GEEI(98)12/FINAL.

• Office fédéral de l'environnement des forêts et du paysage: Informations concernant la protectiondes eaux: les prescriptions de la Confédération en matière de boues d'épuration, Issue 14, Pages47, 1994.

• Öko-Institute. V, (1999) Waste Prevention and Minimisation – Final Report, Institute for AppliedEcology, Berlin, Germany

• Rapport / Naturvårdsverket (1999), Lead party report on combined municipal and industrialdischarges (Swedish), Baltic Sea Joint Comprehensive Environmental Action Programme ,Stockholm : Swedish Environmental Protection Agency.

• RNDE (1999): Les micropolluants dans les cours d'eau: 3 années d'observations 1995-97,.• RNDE (1999): Les principaux rejets d'eaux résiduaires industrielles: données 1997, Pages 1-36,

1999.• SFT-Norwegian Pollution Control Authority (1997a), 97:27, Sources of heavy metals in municipal

wastewater-smaller industry.• SFT-Norwegian Pollution Control Authority (1997b), 97:28, Sources of heavy metals in municipal

wastewater-households.• SFT-Norwegian Pollution Control Authority (1998a), 98:22, Sources of organic environmentally

hazardous substances in municipal wastewater- smaller industry.• SFT-Norwegian Pollution Control Authority (1998b), 98:23, Sources of organic environmentally

hazardous substances in municipal wastewater- households.• SFT-Norwegian Pollution Control Authority (1999), 99:11, Sources of heavy metals in municipal

wastewater- intensive survey of chosen smaller industry.• SPEED (1993) (superadministerial project effective emissions reduction diffuse sources)

Document - Heavy metals in surface waters and abatement, RIZA report no 93012, RVIM reportnumber 773003001

• Statistika centralbyrån Sweden (1998), Discharges to water and sludge production in 1998. Municipal waste water treatment plants and some coastal industry.

• UAB (1997): "Umweltverträglichkeit von Chemikalien zur Abwasserbehandlung“,Umweltbundesamt, Band Nr. 39/97.

• UAB (1998): "Effects of Endocrine Disruptors in the Environment on Neuronal Development andBehavior“, UAB-TEXTE-Band 50/98.

• UBA (2000) Umweltbundesamt; http://www.umweltbundesamt.de/uba-info-daten/index.htm.• UBA TEXTE (1995), 3/96, ISSN 0722-186X, Umweltbundesamt, Expert Round - Endocrinically

Active Chemicals in the Environment, Berlin, 9. and 10. March 1995• UBA, (1998), Texte 35/98, Technische, analytische, organisatorische und rechtliche Massnamen

zur Verminderung der Klärschlammbelastung mit relevanten organischen Schadstoffen, Band 1und 2.

• USEPA (2000), REPORT OF THE MEETING TO PEER REVIEW “THE INVENTORY OFSOURCES OF DIOXIN IN THE UNITED STATES”—Final Report—Prepared for the NationalCenter for Environmental Assessment Office of Research and Development

• USEPA-R2-72-081(1972). Sartor J.D., and G.B. Boyd Water pollution aspects of street surfacecontaminants.

• USEPA-Report 440/1 (1982): Burns & Roe Industrial Services Corp., Paramus New Jersey USA'Fate of Priority Toxic Pollutants in Publicly Owned Treatment Works“,-82-303-Vol-1 and Vol-2(Final report, PB 83-122788/PB 83-1222796).

• Wasserhausgesetzt vom 18. November (1996), Bundesrepublik Deutschland: 'Gesetz zurOrdnung des Wasserhaushalts“, (Wasserhaushaltsgesetz-WHG).

• Wasserhausgesetzt vom 18. November (1996), Bundesrepublik Deutschland: "Gesetz zurOrdnung des Wasserhaushalts“, (Wasserhaushaltsgesetz-WHG).

• World Wildlife Federation (WWF) Canada Report (1997), Nonylphenol Ethoxylates-Reduction andPhase-Out Initiatives, Ottawa 1997.

• WRc (1993). Surface water outfalls, quality and environmental impact management. Report no.VM 1400. Swindon

• WRc (1994), Diffuse sources of heavy metals to sewers, final report to the Department of theEnvironment DoE 3624, June 1994

• WRc; Water Research Centre (1994) The Destruction of Organic Micropollutants in SludgeDigestion Processes. WRc Report No. UM 1441. WRc, Swindon.

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8.5 Abbreviations

ADBI celestolide

ADP Antecedent Dry Period (dry/wet deposition)

AHTN tonalide

AOX Adsorbable Organohalogen (Organochlorine) Compounds

APEO Alkylphenolethoxylates

AT Austria

BE Belgium

BOD Biochemical Oxygen Demand

CDO Chemically Dissolved Oxygen

CEE Central and Eastern European countries

CEP 2-Chloroethanol phosphate

CH Switzerland

DBT Dibutyltin

DE Germany

DEHP Di (2-ethylhexyl) phthalate

DEP diethyl phthalate

DK Denmark

DMDTAR Dimethyl di-tallowammonium chloride

DO Dissolved Oxygen

DSDMAC Distearyl, dimethylammonium chloride

EI Ireland

EPA or USEPA, United States Environment Protection Agency

ES Spain

EU European Union

FI Finland

FR France

g/t grams per tonne

GR Greece

HGV Heavy Goods Vehicle

HHCB galaxolide

HM Heavy Metals

IKW Industry association personal hygiene and detergents (Frankfurt)

IT Italy

kg/t kilogram per ton

LAN Long-chain alkylnitriles

LAS Linear alkyl-(dodecyl-)benzenesulphonate

LU Luxembourg

LV Light Vehicle

MBAS Methylene Blue Adsorbable Substances

MBT Monobutyltin

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MCL Maximum Contaminant (Contamination) Level

NL Netherlands

NO Norway

NP Nonylphenol

NP1EO Nonylphenol monodiethoxylate

NP2EO Nonylphenol diethoxylate

NPE Nonylphenol ethoxylates

NPEC Nonylphenol carboxylic acids

OP Organic Pollutants

p.e. population equivalent

PAHs Poly aromatic hydrocarbons

PCBs Poly chlorinated biphenyls

PCDD Polychlorinated dibenzodioxines

PCDF Polychlorinated dibenzofurans

PEG Polyethylene glycol

PGM Platinum Group Metals

POPs Persistent Organic Pollutants

PPG Polypropylene glycol

PT Portugal

SE Sweden

SS Sewage Sludge

SSM Suspended Solid Matter

SST Sewage Sludge Treatment

t/a tonnes per year

TAMs Trialkylamines

TBP Tri-n-butyl phosphate

TBT Tributyltin

TCDD Tetrachlorine dibenzo-p-dioxin

TE Toxicity equivalent

UWW Urban Waste Water

WWTS Urban Waste Water Treatment Systems

WWTP Urban Waste Water Treatment Plant

VEC Vehicle Exhaust Catalysts

WHO World Health Organisation

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